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

Phyllophora crispa is widely distributed on the coasts of the British Isles, except in the east of England. It is widely distributed in the east Atlantic with a northern limit in Iceland and a southern limit in North Africa but is also present in the Mediterranean and Black Sea (Newroth, 1971; Dixon & Irvine, 1977; Guiry & Guiry, 2015; Bunker et al., 2012). Phyllophora crispa has been recorded in temperatures ranging from 5 to 20°C, with the vast majority of observations recorded in the 10 to 15°C range (OBIS, 2025). Kooistra et al. (1989) noted that it was limited to lower shore tide pools, and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity. However, Gallon et al. (2014) reported that changes in red seaweed assemblages across Brittany were correlated with a 0.7°C increase in coastal water temperature over the prior twenty years. Species varied in their response but the occurrence of several species of red algae, including Phyllophora crispa, increased.

Phyllophora crispa showed resistance to elevated temperatures combined with elevated light levels, in an experiment on Mediterranean macroalgae (Hesse et al., 2025). Temperature (21, 26, and 30°C) was fully-crossed with light levels (180, 320, and 760 µmol photons/m²/s), and photosynthesis and respiration were measured via oxygen flux. Net photosynthesis (µmol O₂ /m²/s) was reduced by roughly half in the medium light/medium temperature treatment but remained similar to controls in the high light/high temperature treatment. Photosynthesis to respiration ratio (P:R) ranged from 1.9 in the low light/medium heat treatment to 3.69 in the high light/low heat treatment. P:R values above 1 indicate good health, as the alga is producing more energy through photosynthesis than it is consuming. Hesse et al. (2025) concluded that Phyllophora crispa in the Mediterranean would be relatively resistant to increasing temperature and light regimes, especially compared to Cystoceira spp. which is also a dominant species in Mediterranean macroalgal assemblages.

Sensitivity assessment. Phyllophora crispa is distributed to the north and south of the British Isles and, therefore, is probably tolerant of a long-term 2°C change in temperature for a year. It is also likely to tolerate a 5°C change in the short-term. Therefore, a resistance of ‘High’ is suggested so that resilience is ‘High’ (by default) and the biotope is assessed as ‘Not sensitive’ at the benchmark level.

Temperature decrease (local)

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

Evidence

Phyllophora crispa is widely distributed on the coasts of the British Isles, except in the east of England. It is widely distributed in the east Atlantic with a northern limit in Iceland and a southern limit in North Africa but is also present in the Mediterranean and Black Sea (Newroth, 1971; Dixon & Irvine, 1977; Guiry & Guiry, 2015; Bunker et al., 2012). Phyllophora crispa has been recorded in temperatures ranging from 5 to 20°C, with the vast majority of observations recorded in the 10 to 15°C range (OBIS, 2025) . Kooistra et al. (1989) noted that it was limited to lower shore tide pools and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity.

Sensitivity assessment. Phyllophora crispa is distributed to the north and south of the British Isles and, therefore, is probably tolerant of a long-term 2°C change in temperature for a year. It is also likely to tolerate a 5°C change in the short-term. Therefore, a resistance of ‘High’ is suggested so that resilience is ‘High’ (by default) and the biotope is assessed as ‘Not sensitive’ at the benchmark level. However, confidence in the assessment is ‘Low’ as it is based on expert judgment and proxies for evidence.

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

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

Phyllophora crispa is recorded from shady places in the lower littoral, lower littoral pools and subtidally to approx. 30m (Dixon & Irvine, 1977; Bunker et al., 20102). Kooistra et al. (1989) noted that Phyllophora crispa was limited to lower shore tide pools and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity. Maximova (2013; summary only) reported that ‘morphological and biological changes’ in Phyllophora crispa from the Black Sea changed in experiments where the ‘normal’ salinity was raised from 18 PSU to 25, 32 and 39, but no further details were available. Phyllophora crispa is found in seawater with salinities ranging from 15 to 40 PSU, with most records in the 30 to 35 PSU range (OBIS, 2025) .

A comparative study of salinity tolerances of macroalgae collected from North Zealand and the South Kattegat (Denmark) where salinity is 16 PSU showed that species generally had a high tolerance (maintained more than half of photosynthetic capacity in short-term exposures of four days) to salinities lower than 3.7 PSU. However, tolerances varied between species with Phyllophora pseudoceranoides exhibiting greater tolerance than Phycodrys rubens, which was the least resistant species tested (Larsen & Sand-Jensen, 2006).

Sensitivity assessment. The presence of Phyllophora crispa in the lower intertidal suggests that it might be exposed to changes in salinity due to evaporation or rainfall but only for very short periods. This biotope (KSwSS.Pcri) is only recorded from full salinity so that a decrease in salinity at the benchmark level would expose the biotope to reduced salinity conditions (18 to 30 PSU). The observations from the Black Sea, the South Kattegat and OBIS (2025) suggest that Phyllophora crispa could survive reduced salinity conditions but the biotope would probably experience a reduction in species richness and less resistant species left or were lost from the biotope. Therefore, a resistance of ‘High’ is suggested, resilience is ‘High’ (no impact to recover from) so that sensitivity is assessed as ‘Not sensitive’ at the benchmark level. However, confidence in the assessment is ‘Low’ as it is based on expert judgment and proxies for evidence.

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

Phyllophora crispa was recorded from moderately strong to very weak tidal flow (Connor et al., 2004). It has been recorded to regenerate after erosion (Dixon & Irvine, 1977) while Molenaar & Breeman (1994) noted that Phyllophora pseudoceranoides exhibited annual growth and die back patterns where growth was removed annually by abrasion or water action.  However, this biotope is unusual because the very weak to weak tidal streams and very wave sheltered conditions allow Phyllophora crispa to grow abundantly on fine sediments (muddy sands and sandy muds).  It is presumably attached to small stones  within the sediment.  A significant increase in water flow may winnow away the mud surface or even remove the mud habitat and hence the biotope if prolonged. An increase of 0.2 m/s may begin to erode the mud surface where the site is already subject to flow (e.g. weak flow at the seabed), based on sediment erosion deposition curves (Wright, 2001). Therefore, an increase in water flow could result in the loss of the ‘loose-lying’ mat of Phyllophora crispa.  However, an increase of only 0.1-0.2 m/s may only affect example of the biotope already in weak flow, rather than very weak flow and a resistance of Low is suggested with Low confidence. Resilience is probably Medium so that sensitivity is assessed as Medium.

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

Changes in emergence are ‘Not relevant’ to this biotope, which is restricted to fully subtidal habitats. The pressure benchmark is relevant only to littoral and shallow sublittoral fringe biotopes.

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

Phyllophora crispa was recorded from wave exposed to very wave sheltered sites (Connor et al., 2004). It has been recorded to regenerate after erosion (Dixon & Irvine, 1977) while Molenaar & Breeman (1994) noted that Phyllophora pseudoceranoides exhibited annual growth and die back patterns where growth was removed annually by abrasion or water action.  However, this biotope is unusual because the very weak to weak tidal streams and very wave sheltered conditions allow Phyllophora crispa to grow abundantly on fine sediments (muddy sands and sandy muds).  It is presumably attached to small stones within the sediment.  A further decrease in wave exposure is unlikely to affect the biotope. However, an increase in wave exposure is likely to remove the loose-lying mat of Phyllophora crispa but a 3-5% change in significant wave height (the benchmark) is unlikely to have a significant effect. Therefore, the biotope is probably Not sensitive (resistance and resilience are High) at the benchmark level. 

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

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 was found

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

Kooistra et al. (1989) noted that Phyllophora crispa was limited to lower shore tide pools but concluded that temperature and salinity were not the limiting factors but that oxygenation and competition were possible limiting factors. However, no direct evidence was found.

Hypoxia caused by a severe eutrophication event in the north-western Black Sea was reported as a constraint on the recovery of ‘Zernov’s Phyllophora field’ (Stevens et al., 2019). This area had undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The field was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the field, and the ecosystem of fish and other algae declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (Black Sea Commission, 2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that deoxygenation could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

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

‘Zernov’s Phyllophora field’ in the north-western Black Sea has undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The ‘field’ was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the ‘field’, and the ecosystem of fish and other algae decreased or declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (Black Sea Commission, 2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that nutrient enrichment and resultant algal blooms could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

Organic enrichment

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

Evidence

‘Zernov’s Phyllophora field’ in the north-western Black Sea has undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The ‘field’ was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the ‘field’, and the ecosystem of fish and other algae decreased or declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that organic enrichment due to algal blooms or organic wastes, together with turbidity could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

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

If sedimentary substrata were replaced with rock substrata the biotope would be lost, as it would not longer be a sedimentary habitat.

Sensitivity assessment. Resistance to the pressure is considered ’None‘, and resilience ’Very low‘ or ‘None’ (as the pressure represents a permanent change) and the sensitivity of this biotope is assessed as ’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

This biotope is recorded from sandy gravelly muds (Connor et al., 2004, sediment matrix). Phyllophora crispa is found on other substrata, including rock and other seaweeds. The low energy environment of the biotope, i.e. low water flow and wave sheltered conditions, determines the nature of the sediment. The muddy sediment is probably inhospitable to most other macroalgae so that Phyllophora can become abundant. A change in sediment from sand muddy gravels to gravel or mud may not affect the biotope adversely. Bunker et al. (2012) note that Phyllophora crispa thrives on rock subject to shell gravel. Therefore, in the very sheltered condition so of this biotope, the mat of Phyllophora would probably survive on gravel or on muds where enough stone for attachment remains. Therefore, the biotope is probably Not sensitive (resistance and resilience are High) at the benchmark level. 

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 biotope is an epifloral mat sitting on the surface of the sediment. Extraction of the sediment to any depth would result in removal of the Phyllophora mat from the affected area.  Therefore, a resistance of None is suggested. Resilience is probably Medium and sensitivity is assessed as Medium but with Low confidence due to the lack of any direct evidence.

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

Dixon & Irvine (1977) noted that Phyllophora crispa regenerates after erosion and animal grazing. Bunker et al. (2012) noted that it tolerated sediment cover and thrived in areas subject to shell gravel. Both observations suggest that it can either resist or regrow from damage due to sediment scour or animal grazing. However, this biotope (KSwSS.Pcri) is an epifloral mat sitting on the surface of sediment; probably loosely attached to small stones or shells in the sediment. Therefore, any passing bottom gear is liable to remove the Phyllophora crispa mat.

In 2012, the Costa Concordia cruise ship crashed into a rocky reef near Giglio Island in the Mediterranean (Piazzi et al., 2020). It took until July 2017 to fully remove the wreckage. Surveys that took place immediately after removal showed that the impacted site and two reference sites had macroalgal assemblages that were dominated by Phyllophora crispa. Percentage cover was initially 98.25 and 94.58% at the two reference sites, compared to 70.67% at the impacted site. Four months later, surveys showed that percentage cover only changed marginally at the reference sites but had declined by a further 56.5% at the impacted site. This decline coincided with an increase in native turf-forming macroalgae, which could colonize faster than Phyllophora crispa.

Sensitivity assessment. A resistance of ‘Low’ is suggested. Resilience is probably ‘High’ so that the sensitivity is assessed as ‘Low’ at the benchmark level but with ‘Low’ confidence due to the lack of any direct evidence.

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

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

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

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

Airoldi (2003) identified a number of morphological, physiological and life history traits that conferred high levels of tolerance to sedimentation. For example, regeneration of upright fronds from a perennial basal crust resistant to burial and scour, calcified thalli, apical meristems, large reproductive outputs, lateral vegetative growth and slow growth rates (Airoldi, 2003). Algae with tough thalli are more resistant to sedimentation and scour (Pedersén & Snoeijs, 2001). Phyllophora crispa was reported to regenerate after erosion and animal grazing (Dixon & Irvine, 1977) and Bunker et al. (2012) noted that it tolerated sediment cover and thrived in areas subject to shell gravel.  Both observations suggest that it can either resist or regrow from damage due to sediment scour.  However, in the wave sheltered, low energy conditions that typify this biotope any deposit of sediment is likely to remain and smother the biotope. 

Sensitivity assessment. Phyllophora crispa grows up to 15 cm in length and would be completely smothered by 30 cm of deposited sediment.  The sediment is likely to remain, and the plants will be removed from light and probably succumb to localise anoxia under the sediment. Therefore, a resistance of None is suggested. Resilience is probably Medium and the sensitivity is assessed as Medium benchmark level.

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

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. No studies investigating the effect of EMFs at the population or community level for benthic organisms were found.

Sensitivity assessment. Given the lack of data at the level of individual biotopes, resistance and resilience to EMFs cannot be robustly assessed. Sensitivity is therefore recorded as ‘Insufficient 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

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

The pressure definition is not directly applicable to seabed biotopes so Not relevant has been recorded.  Collision via ship groundings or terrestrial vehicles is possible but the effects are probably similar to those of abrasion above.

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

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

No evidence was found of the translocation, breeding or species hybridization of the important characterizing species.

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 was found.

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

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

Accidental physical disturbance due to access (e.g. trampling), grounding, or passing fishing gear is examined under abrasion above. However, the accidental removal of the Phyllophora mat would result in a significant change in the biological character of, and loss of, the biotope. Therefore, a resistance of None is suggested. Resilience is probably Medium so that sensitivity is assessed as Medium but with 'Low' confidence.

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., 2011b; 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). 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., 2011b; 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% periwinkle shell (Preston et al., 2020).

Bohn et al. (2015) noted that Crepidula density was low in areas of homogenous fine sediment and absent in areas dominated by boulders. They also suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. However, Hinz et al. (2011b) 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).

Crepidula fornicata is recorded from the lower intertidal to ~160 m in depth but 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., 2011b; OBIS, 2025; Tillin et al., 2020).

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 (Magallana 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. 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; McNeill et al., 2010; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015).

High densities of Crepidula fornicata cause ecological impacts on sedimentary habitats. The species can smother the seabed in shallow bays, changing and modifying the habitat structure (Blanchard, 1997, 2009; Helmer et al., 2019; Tillin et al., 2020). 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 to 50 Crepidula /m². However, Blanchard (2009) noted that sandy areas in the Bay of Mont Saint-Michel were not colonized by Crepidula due to sediment mobility, although adjacent areas were colonized. Thieltges et al. (2003) noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate.

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 (~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 to 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, 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). Bohn et al. (2015) suggested that high densities of Crepidula in high-energy environments were possible in the subtidal but not the intertidal. Hinz et al. (2011b) reported a substantial increase in the occurrence of Crepidula off the Isle of Wight, between 1958 and 2006, at a depth of ~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 (Tillin et al., 2020).

Crepidula creates more muddy substrata, this impacts the larval settlement and survival of other species such as the King scallop (Pecten maximus) and Queen scallop (Aequipecten opercularis), causing a decrease in stocks (Stiger-Pouvreau & Thouzeau, 2015). This impact is more significant to the environment in more densely colonized areas (Blanchard, 2009).

Crepidula invasion on sediment also affects the hydrodynamics and transport properties of the benthic boundary layer. Results have suggested that seabed erosion and velocity measurements of flows over an artificial Crepidula shell bed decreases as roughness density increased, suggesting a sheltering effect by the shells (Stiger-Pouvreau & Thouzeau, 2015). Higher particle resuspension was observed in the study in muddy sand substrates with few stacks of Crepidula when compared with higher density areas (Stiger-Pouvreau & Thouzeau, 2015).

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., 2011b; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Tillin et al., 2020).

Sensitivity assessment. No evidence was found on the interaction of Crepidula and kelp and seaweed-dominated communities on sublittoral sediments. Tillin et al. (2020) suggested that Crepidula was unlikely to colonize kelp and seaweed-dominated communities on sublittoral sediment. De Montaudouin et al. (2001, 2018) noted that abundance of Crepidula was limited in areas of muddy sediment occupied by Zostera in Arcachon Bay, France, and concluded it was not invasive in the bay. Tillin et al. (2020) suggested that the sediment type (muds) and the sweeping of Zostera leaves probably limited Crepidula abundance and hence suggested that kelp and seaweeds could also limit colonization of sublittoral sediment by Crepidula, albeit with ‘Low’ confidence’.

However, the presence of kelp and other seaweeds in this biotope group (SS.SMp.KSwSS) is dependent on the availability of hard substrata provided by the mixed sediment, or gravels, cobbles, shell, etc., in soft sediments (Connor et al., 2004, JNCC, 2022). Hence, the suitability of the biotope for colonization by Crepidula probably varies depending on the underlying sediment type, and hence wave action, while its subsequent abundance may also be limited by the presence of kelps and other seaweeds.

SS.SMp.KSwSS.Pcri occurs in stable, low energy habitats and characterized by sandy mud and gravel but includes boulders, pebbles, cobbles, and shell (Connor et al., 2004), which could provide suitable attachment for Crepidula larvae and allow it to get a foothold. Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution, to represent colonization by Crepidula, although its abundance may be mitigated by the presence of seaweeds. Resilience is likely to be 'Very low’ as Crepidula would need to be removed to allow recovery. Hence, sensitivity is assessed as ‘Medium’ but with ‘Low’ confidence due to the lack of direct evidence.

The carpet sea squirt, Didemnum vexillum

Evidence

The Pacific oyster, Magallana gigas

Evidence

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; GBNNSIP, 2011b, 2012a). 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; GBNNSIP, 2011b, 2012a; 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; GBNNSIP, 2011b, 2012a; 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, 2025; 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 (GBNNSIP, 2011b, 2012a; 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). But adults are long-lived so that 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).

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

Dense macroalgal cover is unsuitable for the Magallana gigas (Herbert et al., 2012, 2016; Tillin et al., 2020), being rarely found under macroalgal cover in Northern Ireland, absent from exposed bedrock or large boulders with macroalgae cover in the Solway Firth, Scotland, and absent in Poole Harbour where there was competition with macroalgae (Kochmann, 2012; Kochmann et al., 2013; McKinstry & Jensen, 2013; Cook et al., 2014; Tillin et al., 2020). Fucus cover significantly reduced larval recruitment of the Pacific oyster in the Wadden Sea (Diederich, 2005). Hence, the Pacific oyster is more likely to colonize bare rock, boulders, or mussel beds without macroalgae (Diederich, 2005; Cook et al., 2014). Kochmann et al. (2013) suggested that macrophyte canopies prevent larvae from settling on the rocks underneath and macroalgal fronds inhibit settlement and recruitment by exuding metabolites.

The majority of the evidence indicates that infralittoral rock and other habitats that occur at depths more than 10 m are unlikely to be suitable for Magallana gigas because it is considered an intertidal and shallow subtidal species rarely recorded below extreme low water (Herbert et al., 2012, 2016; Tillin et al., 2020). However, in suitable situations (e.g. Oosterschelde) it may form beds down to 42 m.

The Pacific oyster may be extending its invasive range subtidally (Tillin et al., 2020), with evidence suggesting it can occupy the shallow subtidal on subtidal sediments, as it does in its native habitat (Padilla, 2010, Herbert et al., 2012). For example, it has been reported at 10 m depth in the Oosterschelde, Netherlands and in the Wadden Sea on subtidal sediments, 1 m to 9 m on subtidal sediments Sweden, and at least 2 m to 3 m below Chart Datum in the Thames estuary and Essex and Kent area on subtidal sediments (Herbert et al., 2012, 2016; Tillin et al., 2020). In Lough Swilly, Ireland Kochmann et al. (2013) found Pacific oysters on a shallow subtidal mussel bed during an exceptionally low spring tide. High densities of Magallana gigas were observed in shallow subtidal areas around the Irish coastline, for example, the oyster is found in Lough Foyle at sites predominately composed of mussels, mud, and sand and in Lough Swilly, at sites composed of mussels, boulders or cobbles and mud (Kochmann, 2012).

It has been suggested that recruitment is enhanced, and abundances are higher in wave-sheltered conditions (Robinson et al., 2005; Ruesink, 2007 cited in Teschke et al., 2020; Tillin et al., 2020). Teschke et al. (2020) found the abundance of Magallana gigas was significantly higher at wave-protected sites within the artificial harbours of Helgoland, North Sea, compared to wave exposed sites outside the harbours. The authors suggested that the successful colonization in wave-protected sites could be due to the relative retention of water masses in the harbours that reduces larval drift and whiplash effect on newly settled larvae. In addition, better growth and higher survival rates were observed at wave-protected sites, whereas mortality rates increased at wave exposed sites, due to the wave exposure causing dislodgement or detachment from the settlement substratum (Teschke et al., 2020; Tillin et al., 2020). Similarly, Bergstrom et al. (2021) noted that the occurrence of high densities of both Ostrea edulis and Magallana gigas decreased with increasing wave exposure.

Temperature and salinity affect spawning and recruitment of Magallana gigas populations. While Pacific oyster larvae are vulnerable to environmental change and less adaptable, it has been suggested that Magallana gigas adults and established populations are more resilient (GBNNSIP, 2011b, 2012a; Hansen et al., 2023). The broad geographical spread of Magallana gigas indicates the invasive species has a wide environmental tolerance.

The Pacific oyster can withstand a wide range of salinities (from 11 to 34 PSU) but no oysters were observed in areas which had salinities less than 20 PSU and most abundant populations occur in salinities above 20 PSU on the Swedish west coastline (Wrange et al., 2010; Kochmann, 2012; Chu et al., 1996 cited in Tillin et al., 2020). Bergstrom et al. (2021) noted that in the Skagerrak, Sweden native and Pacific oyster densities increased with rising salinity above 15 to 21 PSU up to the full range measured (27 PSU). Larvae can survive salinities between 19 to 35 PSU (Troost, 2010; Tillin et al., 2020). Kochmann (2012) reported 11 to 35 PSU as the optimal salinity range for Magallana gigas (cited in Wood et al. (2021). Growth of Pacific oysters can occur between 10 to 30 PSU (Troost, 2010).

Carrasco & Baron (2010) suggested that Magallana gigas has successfully adapted to colonize a range of thermal niches. Temperature is important for the life cycle of the Pacific oyster and influences the establishment of feral and wild populations (Alves et al., 2021). Within its native range, Magallana gigas occurs in areas where the sea surface temperature ranges from 14.0°C and 28.6°C in the warmest month of the year, and between -1.9°C and 19.8°C in the coldest month (Carrasco & Baron, 2010).

Pacific oysters have been found to reduce to 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 tide and currents seen in the Bay of Brest and effects of organic enrichment at oyster reefs might be minimal due to wave action (Lejart & Hily, 2011).

Sensitivity assessment. Magallana gigas prefers hard substrata in the intertidal zone, but it has been observed in muddy habitats and it can occur in depths down to 42 m. The availability of hard substrata for settlement (gravel, shell, cobbles and pebbles) , together with the wave sheltered conditions in this biotope is also suitable for Magallana gigas. However, the evidence shows that the presence of macroalgae often stops the settlement of Magallana gigas larvae. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ (no impact to recover from), and sensitivity as ‘Not sensitive’. This assessment was made with ‘Low’ confidence due to the lack of direct evidence of Phyllophora crispa resistance and sensitivity to Magallana gigas.

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 has spread to 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).

Undaria pinnatifida behaves as a winter annual. Recruitment occurs in winter, followed by rapid spring growth, maturation in summer, and senescence by late summer, leaving only microscopic stages to persist 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. Undaria pinnatifida requires hard, rocky substrata for settlement., while this biotope occurs on infralittoral muddy sediment. It is also limited to a depth range of -1 to 4 m. It may be able to colonize in shallow (<5 m) examples of the biotope that also include boulders, but it is unclear what effect, if any that may have on the Phyllophora bed, which occurs on the underlying sediment. There is insufficient evidence on the possible effects of Undaria on this biotope. Further evidence is required.

Other INIS

Evidence

Bonnemaisonia hamifera (and the Trailliella-phase) is a non-native red algae introduced to the British Isles from Japan and first recorded in 1890 (Dixon & Irvine, 1977; Maggs & Stegenga, 1998; Gollasch, 2006). It is thought to have been introduced by shipping or with shellfish and to have dispersed by drifting on water currents (Gollasch, 2006). Bonnemaisonia hamifera (and the Trailliella-phase) has spread around the British Isles and Europe, into the Mediterranean and the Canary Isles and north to the Orkneys and the Norwegian coast (Lüning, 1990, Maggs & Stegenga, 1998; Gollasch, 2006). It grows rapidly, reproduces vegetatively, and can spread by fragmentation and drifting (Maggs & Stegenga, 1998).

Bonnemaisonia hamifera (and the Trailliella-phase) occurs in this biotope with Phyllophora crispa but at a lower abundance than its characteristic biotope (SS.SMp.KSwSS.Tra). SS.SMp.KSwSS.Tra occurs in shallower waters but otherwise similar conditions. Therefore, if the abundance of Phyllophora crispa was reduced by an external factor then the Trailliella might be able to take over the available space, especially in the shallow examples of the biotope (KSwSS.Pcri). However, there is No evidence that this has happened to date. Hence, there is insufficient evidence on which to base an assessment.