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

Little direct evidence was found to assess the effects of increased temperature on piddocks, and the assessment is based on distribution records and evidence for spawning in response to temperature changes. Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007). Temperature influences the timing of reproduction in Pholas dactylus, which usually spawns between July and August. Increased summer temperatures in 1982 induced spawning in July on the south coast of England (Knight, 1984). Species distribution models suggested that the distribution of Pholas dactylus could expand northward in the next century due to ocean warming (Schultz et al., 2024).

Similar observations have been made for other piddock species. Spawning of the piddock Petricolaria pholadiformis was initiated by increasing water temperature (>18 °C) (Duval, 1963a), so elevated temperatures outside of usual seasons may disrupt normal spawning periods. The spawning of Barnea candida was also reported to be disrupted by changes in temperature. Barnea candida normally spawns in September when temperatures are dropping (El-Maghraby, 1955). However, a rise in temperature in late June of 1956, induced spawning in some specimens of Barnea candida (Duval, 1963b).  Disruption from established spawning periods, caused by temperature changes, may be detrimental to the survival of recruits as other factors influencing their survival may not be optimal, and some mortality may result. Established populations may otherwise remain unaffected by elevated temperatures.

Sensitivity assessment. The global distribution of Pholas dactylus suggests that this species can tolerate warmer waters than currently experienced in the UK and may therefore be tolerant of a chronic increase in temperature. Short-term acute increases may, (depending on timing) interfere with spawning cues which appear to be temperature driven. The effects will depend on seasonality of occurrence and the species affected. Adult populations may be unaffected, and, in such long-lived species, an unfavourable recruitment may be compensated for in a following year. Therefore, resistance to an acute change in temperature is therefore assessed as ‘High’ and resilience as ‘High’ (no impact to recover from) and the biotope is considered ‘Not Sensitive’. It should be noted that the timing of acute changes may lead to greater impacts, temperature increases in the warmest months may exceed thermal tolerances whilst changes in colder periods may stress individuals acclimated to the lower temperatures.

Temperature decrease (local)

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

Evidence

Little direct evidence was found to assess the effects of increased temperature on piddocks and the assessment is based on distribution records and evidence for spawning in response to temperature changes. Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007).    Pholas dactylus spawning appears to be temperature dependent and so a long-term drop in temperature may cause Pholas dactylus to be replaced by piddocks tolerant of cooler water such as Barnea candida and Zirfaea crispata so the overall nature of the biotope is unlikely to change significantly.

Sensitivity assessment. The global distribution of Pholas dactylus suggests that this species can tolerate cooler waters than currently experienced in the UK and may therefore be tolerant of a chronic decrease in temperature. Short-term chronic increases may, (depending on timing) interfere with spawning cues which appear to be temperature driven.  The effects will depend on seasonality of occurrence and the species affected. Adult populations may be unaffected and, in such long-lived species, an unfavourable recruitment may be compensated for in a following year. Resistance to an acute change in temperature is therefore assessed as ‘High’ and recovery as ‘High’ (no impact to recover from) and the biotope is considered ‘Not Sensitive’.  It should be noted that the timing of acute changes may lead to greater impacts, temperature increases in the warmest months may exceed thermal tolerances whilst changes in colder periods may stress individuals acclimated to the lower temperatures.

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 has only been recorded from conditions of full salinity (Connor et al., 2004; JNCC, 2015, 2022). No direct empirical evidence was found to assess this pressure for Pholas dactylus, so the assessment is based on the reported distribution of the biotope and other piddocks.

Pholas dactylus has been recorded in salinities of 30 to 35 PSU, with a small number of records from 35 to 40 PSU (OBIS, 2025). Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985) suggesting tolerance of brackish waters. Zenetos et al. (2009) suggest that at all sites where Petricolaria pholadiformis has been found has some freshwater inflow into the sea. According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009) the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It thus appears that reduced salinity facilitates its establishment (Zenetos et al., 2009).

Sensitivity assessment. This biotope occurs in waters with full salinity (30 to 40 PSU). However, there was no evidence of resistance to the pressure at the benchmark level for the characteristic piddock species. Therefore, there is Insufficient Evidence to form the basis of an assessment of the sensitivity of this biotope to this pressure at the benchmark level.

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

No direct empirical evidence was found to assess this pressure for Pholas dactylus, and the assessment is largely based on the reported distribution of the biotope and other piddocks which are recorded in full salinity habitats (Connor et al., 2004).

Pholas dactylus has been recorded in salinities ranging from 30 to 35 PSU, with a small number of records from 35 to 40 PSU (OBIS, 2025). No information was found for the salinity tolerance of Pholas dactylus. Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985) suggesting tolerance of brackish waters. Zenetos et al. (2009) suggest that at all sites where Petricolaria pholadiformis has been found has some freshwater inflow into the sea. According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009) the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It thus appears that reduced salinity facilitates its establishment (Zenetos et al., 2009).

Sensitivity assessment. Based on reported distributions of the biotope and piddocks it is considered that the benchmark decrease in salinity from full to reduced would result in decreased abundance of Pholas dactylus in biotopes that were previously fully marine. Resistance is therefore assessed as 'Low' and resilience (following return to full salinity) as 'Medium', the biotope is therefore considered to have ‘Medium’ sensitivity to this pressure.

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

Established adult piddocks are, to a large extent, protected from direct effects of increased water flow, owing to their environmental position within the substratum. Increases or decreases in flow rates may affect suspension feeding by altering the delivery of suspended particles or the efficiency of filter feeding.  This biotope has been recorded from areas where tidal flows vary between 0.5 -1.5 m/s (Connor et al., 2004), suggesting that changes in  flow rates (increase or decrease) within this range will not negatively impact the biotope.  Adult piddocks may become exposed should physical erosion of the clay and chalk substratum occur at a greater rate than burrowing, and lost from the substratum. At higher densities bioerosion by piddocks may destabilise the substratum increasing vulnerability to erosion.  

 The most damaging effect of increased flow rate would be the erosion of the clay or soft chalk substratum as this could eventually lead to loss of the habitat. No evidence was found to assess the water velocities at which erosion of clay or chalk occurs. Some erosion will occur naturally and storm events and wave action may be more significant in loss and damage of substratum than surface water flow.

Sensitivity assessment. No direct evidence was found to assess this pressure at the benchmark. Based on the exposure of piddocks in this biotopes to water flows between 0.5 and 1.5 m/s (Connor et al., 2004), the biotope is considered to be unimpacted by changes within this range as long as these do not lead to increased erosion of the substratum. Resistance is therefore assessed as 'High' and resilience as 'High' (based on no impact to recover from), so that the biotope is considered to be ‘Not sensitive’.

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

Not relevant to sublittoral habitats.

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

No direct evidence was found to assess sensitivity to this pressure. The biotope typically occurs in moderately wave exposed locations (Connor et al., 2004). The piddocks are unlikely to be directly affected by changes in wave exposure, owing to their environmental position within the substratum, which protects them. On chalk and clay substrata, it is possible however, that wave action actively erodes the substratum at a faster rate than the piddocks can burrow leaving them exposed to predators or displaced. At higher densities bioerosion, by piddocks may destabilise the substratum increasing vulnerability to erosion.

Potentially the most damaging effect of increased wave heights would be the erosion of the substratum as this could eventually lead to loss of the habitat. Increased erosion would lead to the loss of habitat and removal of piddocks. No evidence was found to link significant wave height to erosion. Some erosion will occur naturally and storm events may be more significant in loss and damage of clays than changes in wave height at the pressure benchmark. For example, Micu (2007) observed numerous Pholas dactylus that had been washed out of the clay substratum or exposed due to storm damage to the clay in the Romanian Black Sea. Erosion rates at the Cretaceous chalk cliffs in East Sussex on the south coast of the UK has accelerated by 22 to 32 cm/year due to natural and anthropogenic modification of the coast (Hurst et al., 2016).

Sensitivity assessment.  Wave action and especially storm action has the potential to cause erosion of the substratum that characterizes this biotope, potentially exposing piddocks or removing part of the substratum. However, a change in significant wave height by 3 to 5% is probably not biologically significant, in this moderately wave exposed biotope. Therefore, resistance is assessed as 'High', resilience as 'High' and the biotope is considered to be 'Not sensitive' at the pressure benchmark.

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

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

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

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

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

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

Radionuclide contamination

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

Evidence

No evidence.

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

This pressure is Not assessed.

De-oxygenation

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

Evidence

Specific information concerning oxygen consumption and reduced oxygen tolerances were not found for important characterizing species within the biotope. Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg O₂/l and probable adverse effects below 2mg O₂/l. Duval (1963a) observed that conditions within the borings of Petricolaria pholadiformis were anaerobic and lined with a loose blue/black sludge. Similarly  Knight (1984) observed Pholas dactylus exposed to low oxygen concentrations  in burrows in areas of chalk bedrock, that were overlain with silt and were anoxic below the surface (Knight, 1984).  Piddocks may therefore have some tolerance of this pressure at the benchmark, however, insufficient information has been recorded to develop an assessment.

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.  The benchmark is set at compliance with WFD criteria for good status, based on nitrogen concentration (UKTAG, 2014).  No evidence was found to assess the sensitivity of piddocks to this pressure.  Nutrient enrichment that enhances productivity of phytoplankton may indirectly benefit the suspension feeding piddocks by increasing food supply.

Sensitivity assessment. The pressure benchmark is set at a level that is relatively protective and the biological assemblage is considered to be 'Not sensitive' at the pressure benchmark. Resistance and resilience are therefore assessed as 'High'.

Organic enrichment

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

Evidence

No evidence was found to assess this pressure. 

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 is characterized by the clay or soft chalk substratum which supports populations of burrowing piddocks. A change to a sedimentary, rock or artificial substratum would result in the loss of piddocks significantly altering the character of the biotope. The biotope is therefore considered to have 'No' resistance to this pressure, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years). The biotope is dependent on the presence of clay or soft chalk, when lost restoration would not be feasible and recovery is therefore categorised as 'Very low'. Sensitivity is therefore assessed as 'High', based on the lack of recovery of the substratum. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

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

This biotope is characterized by the clay or soft chalk substratum which supports populations of burrowing piddocks. A change to a sedimentary substratum would result in the loss of piddocks significantly altering the character of the biotope. The biotope is therefore considered to have 'No' resistance to this pressure, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years) but see caveats in the recovery notes. The biotope is dependent on the presence of soft chalk or clayt, when lost restoration would not be feasible and recovery is therefore categorised as 'Very low'. Sensitivity is therefore assessed as 'High', based on the lack of recovery on chalk or clay substratum. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

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 removal of substratum to 30cm depth will remove the clay or chalk substratum, piddocks and the associated biological assemblage, in the impact footprint.  Resistance is therefore assessed as ‘None’, recovery of the biological assemblage (fwhere suitable substratum remains) is considered to be 'Medium' (2-10 years). The biotope is dependent on the presence of clay or chalk substratum, when lost restoration would not be feasible and recovery is therefore categorised as 'Very low'. Sensitivity is therefore assessed as 'High', based on the lack of recovery of clay or chalk habitats. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

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

The substratum may be impacted by activities that damaged the surface layers, resulting in removal or increased erosion. Natural erosion processes are likely to be ongoing within this habitat type. Where abundant, the boring activities of piddocks also contribute significantly to bioerosion, which can make the substratum habitat more unstable and can result in increased rates of coastal erosion (Evans 1968a, Trudgill 1983, Trudgill & Crabtree, 1987). Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. Surface erosion is therefore a natural part of the environmental processes the biotope experiences although rates could be enhanced by surface abrasion and disturbance. Erosion rates at the Cretaceous chalk cliffs in East Sussex on the south coast of the UK has accelerated by 22 to 32 cm y^-1 due to natural and anthropogenic modification of the coast (Hurst et al., 2016).

Within this biotope the sparse epifauna could be damaged and removed by surface abrasion. Some species protruding from the surface may also be removed. Although the piddocks are afforded some protection from surface abrasion by living in their burrows, the clay or chalk is soft which leaves many individuals, especially those near the surface of the clay, vulnerable to damage and death through exposure of burrows and habitat damage. Micu (2007), for example, observed that after storms in the Romanian Black Sea, the round goby, Neogobius melanostomus, removed clay from damaged or exposed Pholas dactylus burrows to be able to remove and eat the piddocks.

Sensitivity assessment. Surface abrasion may remove epifauna and result in the loss of some piddocks and damage to habitat, especially clays. Resistance is assessed as ‘Medium’ for piddocks and the substratum. As the substratum cannot recover, resilience is assessed as ‘Very Low’, and sensitivity of the overall biotope is considered to be ‘Medium’.

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

Penetration and disturbance below the surface of the substratum will damage surface fauna and could damage or expose piddocks. Individuals in damaged burrows, or those that are removed from the substratum, are unlikely to be able to rebury and will be predated by fish and other mobile species (Micu, 2007). 

The most significant impact arising from this pressure may be the damage and removal of the chalk and clay substratum. Where abundant the boring activities of piddocks can make the substratum habitat more unstable and can exacerbate erosion (Evans 1968a, Trudgill 1983, Brookes & Stevens, 1985, Trudgill & Crabtree, 1987).  Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. Piddock burrowing can therefore  make the substratum more vulnerable to damage and removal.

Sensitivity assessment. Sub-surface penetration and disturbance will remove and damage the sparse epifauna and result in the loss of piddocks and damage to the habitat. Resistance is therefore assessed as ‘Low’ for the piddocks and substratum. Resilience for the piddocks is assessed as ‘Medium’ where suitable substratum remains, so the sensitivity of the piddocks is ‘Medium’. Clay and chalk habitats are restricted in distribution and have been identified as irreplaceable habitats (Tillin et al., 2022). When removed, there is no mechanism by which the substratum can be replaced. Therefore, when removed in part or entirely, no recovery of habitat is possible, and resilience is assessed as 'Very low' (>25 years).

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

No direct evidence was found to assess this pressure. Increased suspended particles will decrease light penetration, may enhance food supply (where these are organic in origin), or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts). Very high levels of silt may clog respiratory and feeding organs of some suspension feeders.  Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, however, the piddocks are protected from scour within burrows and increased organic particles will provide a food subsidy.

Pholas dactylus occurs in habitats such as soft chalks where turbidity may be high (UK BAP, 2008) and is therefore unlikely to be affected by an increase in suspended sediments at the pressure benchmark. Piddocks, in common with other suspension feeding bivalves, have efficient mechanisms to remove inorganic particles via pseudofaeces. Experimental work on Pholas dactylus showed that large particles can either be rejected immediately in the pseudofaeces or passed very quickly through the gut (Knight, 1984).  Increased suspended sediments may impose sub-lethal energetic costs on piddocks by reducing feeding efficiency and requiring the production of pseudofaeces with impacts on growth and reproduction.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of Pholas dactylus. However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply.  

Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure however, based on the occurrence of Pholas dactylus in turbid areas and evidence for the production of pseudofaeces by piddocks, resistance is assessed as ‘High’ and resilience as High (no impact to recover from). The biotope is therefore considered to be ‘Not sensitive’.  

Smothering and siltation rate changes (light)

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

Evidence

No direct evidence was found to assess this pressure. Increased suspended particles will decrease light penetration, may enhance food supply (where these are organic in origin), or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts). Very high levels of silt may clog respiratory and feeding organs of some suspension feeders.  Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, however, the piddocks are protected from scour within burrows and increased organic particles will provide a food subsidy.

Pholas dactylus occurs in habitats such as soft chalks where turbidity may be high (UK BAP, 2008) and is therefore unlikely to be affected by an increase in suspended sediments at the pressure benchmark. Piddocks, in common with other suspension feeding bivalves, have efficient mechanisms to remove inorganic particles via pseudofaeces. Experimental work on Pholas dactylus showed that large particles can either be rejected immediately in the pseudofaeces or passed very quickly through the gut (Knight, 1984).  Increased suspended sediments may impose sub-lethal energetic costs on piddocks by reducing feeding efficiency and requiring the production of pseudofaeces with impacts on growth and reproduction.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of Pholas dactylus. However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply.  

Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure however, based on the occurrence of Pholas dactylus in turbid areas and evidence for the production of pseudofaeces by piddocks, resistance is assessed as ‘High’ and resilience as High (no impact to recover from). The biotope is therefore considered to be ‘Not sensitive’.  

Smothering and siltation rate changes (heavy)

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

Evidence

The burrowing mechanisms of the piddocks Pholas dactylus and Barnea candida and other Pholads, mean that the burrows have a narrow entrance excavated by the juvenile. As the individual grows and excavates deeper the burrow widens resulting in a conical burrow from which the adult cannot emerge.  Piddocks cannot therefore emerge from layers of deposited silt as other more mobile bivalves can.

No examples of direct empirical evidence or experiments on mortality rates in response to siltation have been found for piddocks. Indirect indications for the impacts of siltation are provided by studies of Witt (2004) on the impacts of harbour dredge disposal. Petricolaria pholadiformis (as Petricola pholadiformis was absent from the disposal area, and Witt (2004) cites reports by Essink (1996, not seen) that smothering of Petricola pholadiformis from siltation could lead to mortality within a few hours.  Hebda (2011) also identified that sedimentation may be one of the key threats to Barnea truncata populations.  At Agigea (Micu, 2007) reported that smothering of clay beds by sand and finer sediments had removed populations of Pholas dactylus. In this area sand banks up to 1m thick frequently shift position driven by storm events and currents (Micu, 2007). Similar smothering was described in the case of Barnea candida populations boring into clay beds (Gomoiu & Muller 1962, cited from Micu, 2007).

Sensitivity assessment. As piddocks are essentially sedentary and as siphons are relatively short, siltation from fine could be lethal.  Siltation at the pressure benchmark is considered to smother most or all of the piddocks and the surface fauna.  Resistance to siltation is therefore assessed as ‘None' although effects could be mitigated where water currents and wave exposure rapidly removed the overburden and this will depend on shore height and local hydrodynamic conditions. Resilience is assessed as ‘Medium’ (2-10 years) for piddocks and sensitivity is therefore assessed as ‘Medium’. 

Litter

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

Evidence

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

No evidence.

Underwater noise changes

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

Evidence

Not relevant.

Introduction of light or shading

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

Evidence

Pholas dactylus can perceive and react to light (Hecht, 1928) however there is no evidence that this pressure would impact the biotope. 

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

Not relevant.

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 grounding vessels is addressed under ‘surface abrasion’. 

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

Pholas dactylus reacts quickly to changes in light intensity, after a couple of seconds, by withdrawing its siphon (Knight, 1984). This reaction is ultimately an adaptation to reduce the risk of predation by, for example, approaching birds (Knight, 1984). However, its visual acuity is probably very limited and it is unlikely to be sensitive to visual disturbance. Resistance and resilience are therefore assessed as ‘High’ by default.

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

The species characterizing this biotope, Pholas dactylus and other piddocks, are not farmed or translocated and therefore this pressure is 'Not relevant' to this biotope.

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

No evidence.

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

Piddocks may be removed as bait and across Europe they have traditionally been harvested for food, however high levels of habitat damage are associated with the removal of boring molluscs (Fanelli et al., 1994) and this practice has largely been banned. The most sensitive component of this biotope to targeted harvesting is the peat substratum which may be damaged and removed if piddocks were excavated from their burrows, this effect is considered through the physical damage pressures, abrasion and penetration and sub-surface damage.

Sensitivity assessment. Removal of piddocks will result in loss of targeted individuals and damage to the habitat. Resistance is assessed as ‘Low’ as piddocks are sedentary and burrow openings are readily detected.  Piddocks are predicted to recover within 2-10 years so that resilience is considered to ‘Medium’ and sensitivity is ‘Medium’. 

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

This biotope is characterized by burrowing piddocks. Their removal would reduce their abundance, or result in their loss from the biotope, resulting in the loss of the biotope as defined in the classification. The removal of epifaunal species due to surface abrasion would reduce the biodiversity of the biotope, although the biodiversity of this biotope is typically low (JNCC, 2015, 2022). Hence, resistance is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’, albeit with ‘Low’ confidence due to the lack of direct evidence. This assessment does not address the potentially significant impact of physical disturbance of the substratum during removal, which is address under ‘abrasion’ above.

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2024; OBIS, 2025).

Abundances at its northern and southern extremes may be low but densities in UK and France are often over 1000/m² and it may carpet the seafloor in the Solent and Essex. In the UK, it was reported to reach abundances of >1000/m² (max. 2,748/m²) in the Milford Harbour Waterway (Bohn et al., 2012), 84 /m² in Portsmouth, 174/m2 in Langstone and 306/m² in Chichester harbours in 2017 (Helmer et al., 2019). In France, it has been reported to reach >4,700/m² in the Bay of Marennes-Oleron, France, 11.6 tonnes/ha in Bay of Mont-Saint-Michel, 8.2 tonnes/ha in the Bay of Brest and 2.8 tonnes/ha in the Bay of Saint-Brieuc (Blanchard, 2009; Bohn et al., 2012, 2015; Powell-Jennings & Calloway, 2018).

Its density and ability to spread within and between sites (e.g., Bays) depends on the availability of suitable habitat, completion with other species, larval retention with the site, human activity (e.g., dredging) and summer and winter temperatures (especially in the intertidal). For example, the Crepidula fornicata population in the Bay of Mont-Saint-Michel grew by 50% between 1996 and 2004 and covered 25% at a high density (51 to 100% cover) aided by local oyster farming and shellfish dredging (Blanchard, 2009). However, in Arcachon Bay, France, Crepidula fornicata was limited to only 155 tonnes in 1999 and 312 tonnes in 2011 (De Montaudouin et al., 2001, 2018). Crepidula was limited to muddy sediments that were only ~8% of the bay and were colonized by Zostera beds and represented only 0.4% of suspension feeder biomass of the oysters Magallana gigas in the bay and did not show signs of increasing biomass at a 12-year scale. In addition, benthic trawling was prohibited in the bay (De Montaudouin et al., 2001, 2018). As a result, De Montaudouin et al. (2018) concluded that Crepidula was not invasive in the Bay of Arcachon.

Crepidula fornicata is recorded from shallow, sheltered bays, lagoons and estuaries or the sheltered sides of islands, in variable salinity (from 18 to 40) although it prefers ~30 (Tillin et al., 2020). It is recorded from the lower intertidal to ~160 m in depth but it 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, 2025; 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 is also recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020).

For example, 75% to 98% of Crepidula larvae settled on dead Crepidula shells, in the eastern Solent harbours of Portsmouth, Langstone, and Chichester, while ~4% settled on stone, 2.5% on live Crepidula, 0.3% oyster shell, 0.6% cockle shell, 0.3% winkle shell and 0.1% perwinkle shell (Preston et al., 2020). However, in the Milford Harbour Waterway, the highest densities of Crepidula were found in areas of sediment with hard substrata (e.g., mixed fine sediment with shell, gravel, or both). While Crepidula density increased with increasing gravel cover in the subtidal zone, the opposite pattern was observed in the intertidal zone (Bohn et al., 2015). 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. The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula functions as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). Bohn et al. (2015) also noted that Crepidula density was low in areas of homogenous fine sediment and absent in 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. However, Hinz et al. (2011) recorded Crepidula off the Isle of Wight in the English Channel, at ~60 m on rough ground in areas of high tidal flow. Tillin et al. (2020) suggested that the effect of oscillatory wave meditated flow might have a greater effect on Crepidula than tidal flow, presumably due to mobilization of the substratum. Similarly, Crepidula was absent from sandy substrata in Swansea Bay but was most abundant in the shelter of the breakwater at Swansea east site (Powell-Jennings & Calloway, 2018).

The density of Crepidula populations in the northern Europe (Germany, Denmark, and Norway) are significantly lower (<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 ~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 become more abundant in northern Europe. Valdizan et al. (2011) noted higher water temperatures between 2000 to 2001 and 2006 to 2007 together with elevated chlorophyll-a corresponded to an increase in gametogenesis and the duration of broods in Crepidula population in Bournerf Bay, France. They suggested that rising temperatures in northern Europe could increase its reproductive success due favourable breeding temperatures and increased phytoplankton (Valdizan et al., 2011). Nehls et al. (2006) noted that the decline in mussel (Mytilus edulis) beds in the Wadden Sea was due to mild winters that favoured non-native oysters (Magellana gigas) and slipper limpets, which co-existed with the mussels.

Crepidula fornicata has one or two reproductive periods per year (depending on location), is highly fecund, and has long-lived pelagic larvae. Hence, dispersal is potentially high. However, Bohn et al. (2012, 2013a, 2013b, 2015) suggested that lack of suitable habitat rather than larval supply, together with local hydrography may limit the northward spread of Crepidula from Milford Harbour Waterway, and that post-settlement mortality is particularly important in the intertidal. Dupont et al. (2007) reported genetic isolation with distance along the English Channel but a high degree of genetic connectivity between the bays of northern France, which were consistent with hydrographic models of larval transport. They noted marked genetic isolation of the population in the semi-enclosed Bay of Brest. Dupont et al. (2007) suggested that Crepidula populations were isolated by hydrographic barriers over distances of ~100 km. Riel et al. (2009) noted that larval supply was low in the Bay of Mont Saint-Michel partly due to larval mortality and larval export out of the bay, although recruitment was still adequate to maintain the population. Bohn et al. (2012) suggested that homogenous sediments and boulders at the entrance to the Milford Harbour Waterway formed a barrier to dispersal and, together with high larval export probably explained the slow of northward expansion of Crepidula along the Welsh coast. Nevertheless, the initial spread of Crepidula was facilitated by human activities such as shipping, shellfish culture (e.g. oysters and mussels), ballast water (Blanchard, 1997) and fisheries (e.g., dredging) (Blanchard, 1997, 2009; De Montaudouin et al., 2018; Kostecki et al., 2011; McNeill et al., 2010; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015).

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.

Crepidula fornicata 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 from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Tillin et al., 2020). Close examination of the literature shows that evidence of its colonization and density on bedrock in the infralittoral or circalittoral was lacking. Tillin et al. (2020) suggested that Crepidula could colonize circalittoral rock due to its presence on tide-swept rough grounds in the English Channel (Hinz et al., 2011). However, Hinz et al. (2011) reported that Crepidula fornicata only dominated one assemblage (with an average of 181 individuals per trawl) on gravel substratum with boulders. Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas dominated by boulders, and Bohn et al. (2013a, 2013b, 2015) and Preston et al. (2020) showed that while Crepidula could settle on slate panels or ‘stone’ it preferred shell, especially that of conspecifics.

Sensitivity Assessment. No evidence of Crepidula fornicata presence in clay or chalk habitats was found. According to Tillin et al. (2020), clay exposures are unsuitable for Crepidula fornicata settlement, although this is stated with low confidence. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

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). A lack of published descriptions and an incomplete historical record has led to the widespread misidentification of Didemnum vexillum, and it is often recorded as Didemnum spp. Hence, the native range of the species is not known conclusively (Lambert, 2009; Stefaniak et al., 2012; McKenzie et al., 2017; Holt, 2024). However, molecular data and limited historical evidence have suggested that the species may be native to Japan, with its native range possibly extending into continental Asia and north-western Pacific (Stefaniak et al., 2012; Tillin et al., 2020; Holt, 2024). Previously unrecorded populations of a colonial ascidian have been recently identified as Didemnum vexillum (Tillin et al., 2020).

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). Zhang et al. (2020) suggested that in the current climate conditions (based on depth, current, temperature and salinity) Didemnum vexillum had not yet occupied their predicted suitable habitats, and predicted that the Northern Atlantic coast is susceptible to invasion by Didemnum vexillum and that climate change could cause a poleward expansion of Didemnum vexillum.

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 localised 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 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-mediated transport via aquaculture facilities, boat hulls, commercial fishing vessels, and ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijkstra et al., 2007; 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). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water, and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables and dredging (Holt, 2024). Recent evidence from genetic material suggests 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, 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 the 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 and 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; McKenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where the highest average temperatures were recorded in February (18 to 22°C), and the lowest average temperatures were recorded in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with a higher level of recruitment (Fletcher et al., 2013a).

Didemnum vexillum requires suitable hard substrata for successful settlement and establishment. It can grow quickly and establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). 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). The extensive mats formed by the invasive species over cobble-pebble substrata can bind or ‘glue’ small pebbles and cobbles together by filling spaces between the sediment particles, which alters the habitat complexity of the seafloor turning it into a more homogenous two-dimensional habitat rather than heterogeneous three-dimensional one (Griffith et al., 2009; Mercer et al., 2009; Lengyel et al., 2009).

In addition, Didemnum vexillum is commonly found associated with artificial hard substrata, being mostly found in harbours and marinas where it covers a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ships' hulls and at aquaculture facilities (Valentine et al., 2007 a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012). Didemnum vexillum was abundant in the marinas at Terschelling, Texel, and Vlieland, in the Wadden Sea (Gittenberger et al., 2015). In the UK, Didemnum vexillum was initially recorded in marinas and adjacent shallow man-made structures (Tillin et al., 2020). In Wales, it was first recorded in Holyhead Marina, then subsequently reported in Plymouth Marina and other marinas around the UK (Griffith et al., 2009; Minchin & Nunn, 2013; Bishop et al., 2015).

Didemnum sp. can colonize both horizontal and vertical surfaces of fouling and benthic communities, commonly occurring on upper horizontal surfaces in benthic habitats (Dijkstra et al., 2007; Tillin et al., 2020). It has been recorded on overhangs or the underside of boulders (Hitchin, 2012) or on the underside of docks, boat hulls, and pontoons (Griffiths et al., 2009; Minchin & Nunn, 2013). In sheltered areas, colonies are lobed and beard-like, forming long tendrils that drop down from the underside of docks or other artificial substrata to establish new colonies if there are suitable substrata available (Valentine et al., 2007a). In areas of stronger current, colonies are low undulating mats (Valentine et al., 2007 a&b).

Didemnum vexillum has the ability to rapidly overgrow and displace on other sessile organisms such as other colonial ascidians (Ciona intestinalis, Styela clava, Ascidiella aspera, Botrylloides violaceus, Botryllus schlosseri, Diplosoma listerianium and Aplidium spp.), bryozoan, hydroids, sponges (Clione celata and Halichrondria sp.), anemone (Diadumene cincta), calcareous tube worms, eelgrass (Zostera marina), kelp (Laminaria spp. and Agarum sp.), green algae (Codium fragile subsp. fragile), red algae (Plocamium, Chondrus crispus and bush weed Agardhiella subulata), brown algae (Ascophyllum nodosum, Sargassum, Halidrys, Fucus evanescens and Fucus serratus), calcareous algae (Corallina officinalis), mussels (Mytilus galloprovincialis, Perna canaliculus and Mytilus edulis), barnacles, oysters (Magallana gigas, Ostrea edulis and Crassostrea virginica), sea scallops (Placopecten magellanicus), or dead shells (Dijkstra et al., 2007; Gittenberger, 2007; Valentine et al., 2007a; Valentine et al., 2007b; Griffith et al., 2009; Carman & Grunden, 2010; Dijkstra & Nolan, 2011; Groner et al., 2011; Hitchin, 2012; Tagliapietra et al., 2012; Minchin & Nunn, 2013; Gittenberger et al., 2015; Long & Groholz, 2015; Vercaemer et al., 2015). Some species have been shown to tolerate overgrowth by Didemnum vexillum. Such as anemones (did not specify species name), which were observed in high densities of 10 to 339 individuals in transects with high percentage cover of Didemnum vexillum (Lengyel et al., 2009). In the Netherlands, the sea anemone Sagartia elegans and Sabella pavonia tubes were not overgrown by Didemnum sp. (Gittenberger, 2007). Botrylloides violaceus can overgrow Didemnum sp. (Gittenberger, 2007), although it was noted to be overgrown in other studies (Valentine et al., 2007a). In addition, Styela clava and Ascidiella aspera survived overgrowth by Didemnum vexillum as long as their siphons remained free (Gittenberger, 2007). However, Gittenberger (2007) stated that the boring sponge Clione celata, the sea anemone Diadumene cincta, Mytilus edulis, Magallana (syn. Crassostrea) gigas, Ostrea edulis, a variety of hydroids, the colonial ascidians Aplidium (Fig. 4) and Diplosoma listerianum and the solitary ascidians Ciona intestinalis start to die on contact with Didemnum sp.  Didemnum vexillum can overgrow bivalve species, such as oysters, scallops and mussels, as the hard shells can provide suitable hard substrata for settlement. It has been described as a ‘shellfish pest’ by the aquaculture industry because it is likely to completely encapsulate bivalves and smother them resulting in death or partially encapsulate and partially smother them resulting in reduced bivalve growth (Auker, 2010; Bullard et al., 2007; Coutts & Forrest, 2007, Valentine et al., 2007a; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020). Didemnum vexillum has been recorded overgrowing mussels in Strangford Lough, Northern Ireland (Minchin & Nunn, 2013) and recorded forming large mats over Blue Mussel beds in the Gulf of Maine, completely covering individuals (Auker et al., 2014).

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 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 surrounding sand ridges, which have a mobile surface that is moved daily by the strong tidal currents (Valentine et al., 2007b). Evidence also indicates that the species cannot 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; they were absent from the natural seabed beneath the pontoon, composed of silty mud, and from deeper sections of mooring chains that became immersed in mud at low spring tides (Griffiths et al., 2009).

In contrast to Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca. 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated gravel, cobbles and pebbles. 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 are 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). An experiment in the Thames River estuary, Connecticut, found a significant difference between Didemnum vexillum growth rates at different depths, with faster growth rates seen in shallow water (1.0 m) compared to deeper depths (4.0 m) (Bullard & Whitlatch, 2009). It was also found that although Didemnum vexillum grew faster in shallow depths during the experiment, it grew well at all depths examined (1.0 m, 2.5 m and 4.0 m) and there was no significant difference in survival between the depths.

Didemnum vexillum tolerates a wide range of environmental conditions, including temperature and salinity (Herborg et al., 2009; Tillin et al., 2020). Didemnum vexillum can withstand a wide range of salinities from 20 to 44 ppt, is commonly found in marine waters around 33 pp,t but is unable to survive in salinities below 20 ppt (Bullard & Whitlatch, 2009; Groner et al., 2011; Tillin et al., 2020). It has been recorded in estuarine conditions and tidal lagoons (Dijkstra et al., 2007; Tillin et al., 2020). In the Lagoon of Venice, Didemnum vexillum is found in waters at 30 PSU. It was absent in low salinity, such as the estuary and around the salt marshes, but well established in the euhaline and tidally well-flushed zones of the Lagoon of Venice (Tagliapietra et al., 2012). Similar results were found in Connecticut and Rhode Islan,d where Didemnum vexillum was not found in environments with salinity less than 20 ppt (Bullard & Whitlatch, 2009). However, in the Wadden Sea, colonies of Didemnum vexillum were abundant in salinities between 17.91 and 25.97 ppt (Gittenberger, 2007; Gittenberger et al., 2015). Salinity can influence the growth rates of Didemnum vexillum. For example, in an experiment in the Thames River estuary, Connecticut, Bullard & Whitlatch (2009) found growth rates were significantly higher in high salinity areas (26 to 30 ppt) and although survival at different salinities was not significantly different, the Didemnum vexillum colonies in low (10 to 26 ppt) and medium (15 to 28 ppt) salinities were bloated, discoloured and appeared to be dying. In unpublished data from Bullard & Whitlatch (2009), similar results were found in the laboratory, as most colonies appeared to be dying after one week in 20 ppt and healthy in 30 ppt.

A study on Didemnum vexillum colonies from Holyhead Marina, Isle of Anglesey, found colony growth within a week was significantly impaired and reduced by two-thirds at lower salinities (27 PSU and 20 PSU), while in ambient Holyhead Marina salinity (34 PSU), the growth increased and surface area doubled (Groner et al., 2011). Mortality was described as negligible in colonies of Didemnum vexillum in ambient salinity (34 PSU) after two weeks. However, mortality increased as salinity decreased. At the end of the two-week experiment, 72% of invasive colonies survived in 27 PSU and 55% of colonies survived in 20 ppt (Groner et al., 2011). When exposed to severe low salinity of 10 PSU for two hours, Didemnum vexillum showed no mortality, which suggested the duration of exposure influences mortality, not the stress intensity (Groner et al., 2011). Colonies of Didemnum vexillum collected from Anglesey, Wales, experienced more mortality under severe hypo-salinity (20 PSU, 38% colonies survived) compared to moderate hypo-salinity (27 PSU, 82% colonies survived) after two weeks, showing severe hypo-salinity creates more stressful conditions for Didemnum vexillum (Lenz et al., 2011). Therefore, Didemnum vexillum can tolerate a short-term severe decline in salinity, but prolonged exposure over two weeks caused chronic stress and increases in mortality.

Didemnum vexillum is a temperate species that can survive a broad temperature range of -2 to 24°C, with an upper survival limit suggested to be 25°C (Bullard et al., 2007; Valentine et al., 2007a; Herborg et al., 2009; Kleeman, 2009; McKenzie et al., 2017; Holt, 2024). It thrives best at 14 to 20 °C, with optimal growth temperature between 14 and 18°C during summer months (May, June, September, October) (Gittenberger, 2007; Kleeman, 2009; McKenzie et al., 2017). Didemnum vexillum has been recorded surviving in 4 to 15°C in Georges Bank and 5 to 22°C in Holyhead (Bullard et al., 2007; Valentine et al., 2007b; Griffith et al., 2009). In New England, colonies tolerate temperatures as low as -2°C (Bullard et al., 2007), but reports from the Netherlands show colonies “die off” when temperatures drop below 5°C during winter months from November to April (Gittenberger, 2007; 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). Temperature changes are an important factor influencing the seasonal growth cycle and reproduction of Didemnum vexillum (Valentine et al., 2007a).

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

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

Sensitivity Assessment. There is no evidence of Didemnum vexillum colonization on chalk. However, it has been recorded on clay boulders (Hitchin, 2012). According to Tillin et al. (2020), clay exposures are potentially suitable substrates for Didemnum vexillum colonization, although this is stated with low confidence. Resistance is therefore ‘Low’, resilience is ‘Very Low’ as Didemnum vexillum would need to be removed to allow recovery, and sensitivity is assessed as ‘High’, albeit with low confidence due to a lack of direct evidence.

The Pacific oyster, Magallana gigas

Evidence

Wireweed, Sargassum muticum

Evidence

Wakame, Undaria pinnatifida

Evidence

Undaria pinnatifida (Wakame or Asian kelp) 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; Heiser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; Epstein & Smale, 2018; Kraan, 2017; Epstein et al., 2019a,b; Tidbury, 2020). Undaria pinnatifida originates from Japan but is established currently on the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, Holland, Argentina, Mexico, and the USA (De Leij et al., 2017). Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016) and has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it became 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). In Plymouth, UK, De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies, such as Laminaria hyperborea, Laminaria ochroleuca, Laminaria digitata and Saccharina latissima had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria pinnatifida recruits. However, the dense canopies did not always prevent the invasion of Undaria pinnatifida as sporophytes were still recorded within dense Laminaria canopies, so canopy disturbance was not always required (De Leij et al., 2017; Epstein & Smale, 2018).

Undaria behaves as a winter annual, and recruitment occurs in winter followed by rapid growth through spring, maturity and then senescence through summer, with only the microscopic life stages persisting through autumn. It exhibits multiple dispersal strategies, such as short-range spore dispersal, and long-range dispersal as whole drift plants or fragments. Undaria pinnatifida has spread rapidly across the UK and Europe, resulting in community-wide responses and impacts (Vaz-Pinto et al., 2014; Epstein & Smale, 2017). Its impacts are complex and context-specific, depending on space, time, and taxa present in the introduced location (Epstein & Smale, 2017; Teagle et al., 2017; Tidbury, 2020). 

Undaria pinnatifida has a wide physiological niche meaning it can occur in both coastal and estuarine environments showing tolerance for varying salinities, turbidity and siltation (Heiser et al., 2014; Epstein & Smale, 2018). Undaria pinnatifida has a greater preference for 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. No evidence was found of Undaria pinnatifida presence in chalk or clay habitats. Undaria pinnatifida prefers low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018; Epstein et al., 2019a), while this biotope occurs in moderately exposed sites. It is therefore unlikely that Undaria pinnatifida poses a threat to this biotope. Resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

Other INIS

Evidence

The friable nature of the substratum which is subject to on-going erosion means this biotope supports only a sparse epifauna and flora. This biotope is therefore unlikely to be invaded by sessile invasive non-indigenous species. As the biotope occurs subtidally and turbidity levels are often high this biotope likely to be unsuitable for invasive non-indigenous algae.  

The American piddock, Petricolaria pholadiformis is a non-native, boring piddock that was unintentionally introduced from America with the American oyster, Crassostrea virginica, not later than 1890 (Naylor, 1957). Rosenthal (1980) suggested that from the British Isles, the species has colonized several northern European countries by means of its pelagic larva and may also spread via driftwood, although it usually bores into clay, peat or soft rock shores. In Belgium and The Netherlands Petricolaria pholadiformis  almost completely displaced the native piddock, Barnea candida (ICES, 1972). However, this has not been observed elsewhere, and later studies have found that Barnea candida is now more common than Petricolaria pholadiformis in Belgium (Wouters, 1993) and there is no documentary evidence to suggest that Barnea candida has been displaced in the British Isles (J. Light & I. Kileen pers. comm. to Eno et al., 1997). This species is also unlikely to displace Pholas dactylus which is more likely to occur subtidally.

Although not currently established in UK waters, the whelk Rapana venosa, may spread to habitats. This species has been observed predating on Pholas dactylus in the Romanian Black Sea by Micu (2007).

Sensitivity assessment. Based on the lack of records of invasive non-indigenous species in this biotope, and the unsuitability of the habitat for algae and other attached epifauna this biotope is considered to have ‘High’ resistance to this pressure and, by default ‘High’ resilience, this biotope is therefore considered to be ‘Not sensitive’. This assessment may need revising in light of future invasions, e.g. the introduction of the whelk Rapana venosa.