The Marine Life Information Network

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year. Further detail

Evidence

Maerl beds occur from the tropics to polar waters (Foster, 2001; Hinojosa-Arango & Riosmena-Rodriquez, 2004).  Maerl beds in the North East Atlantic range from Norway to the African coast, although the component maerl species vary in temperature tolerance (Birkett et al., 1998a; Wilson et al., 2004).  There are four species of biotope creating maerl beds in the UK.  These species vary in their distribution within the UK, a phenomenon which is thought to be due to their temperature tolerances.  Similarly, the associated communities within the maerl habitat differ and represent a diverse sample of species within the local area.  Lithothamnion coralloides is absent from Scottish waters.  This is due, either to winter temperatures dropping below the minimum survival temperature (between 2 and 5°C) or because the temperatures don’t allow a suitable growing season (Adey & McKibbin, 1970; cited in Wilson et al., 2004).  Lithothamnion corallioides had a higher minimum survival temperature; dying at 2°C and surviving without growth at 5°C (Adey & McKibbin, 1970).  In laboratory conditions, Phymatolithon calcareum survived down to 2°C, died at 0.4°C, and had a recorded optimum temperature for growth of 12-13°C (Adey & McKibbin 1970 cited in Wilson et al., 2004).  Phymatolithon calcareum showed no significant difference on photosynthetic activity at 9°C (the control), 17°C or 25°C for 4-5 weeks but were judged to be dead after 90 minutes at 40°C (Wilson et al., 2004). Temperature appears to confine Lithothamnion glaciale to northern parts of the British Isles, possibly because reproductive conceptacles are only produced in winter when temperature fall below 9°C (Hall-Spencer, 1994 cited in Wilson et al., 2004).  In addition, Adey (1970) found optimal growth rates of Lithothamnion glaciale between 10-12°C and that development of reproductive conceptacles in Lithothamnion glaciale requires winter temperatures of between 1-5°C (Adey, 1970). Blake & Maggs (2003) reported that the growth rate of Phymatolithon calcareum in the laboratory was only slightly affected by temperature treatments (10, 14 and 18°C), with an optimum at 10°C while the growth rate of Lithothamnion corallioides was significantly affected by temperature with an optimum at 14°C, at which temperature it grew faster than Phymatolithon calcareum. Martin et al. (2006) reported that primary productivity in Lithothamnion corallioides was twice as high in August as in January to February in the Bay of Brest. They found that primary productivity, calcification and respiration rates of Lithothamnion corallioides increased as temperature rose from 10 to 16°C (Martin et al., 2006).

The main maerl forming species have wide geographic ranges and their range indicates the limits of their temperature tolerance. Phymatolithon calcareum is a cold temperate species that ranges from Norway to the Mediterranean (Wilson et al., 2004; Martin et al., 2006) and tolerates high temperatures better than many subtidal temperate red algae (Wilson et al., 2004). Lithothamnion corallioides is a warm temperate species ranging from Ireland and the south of Britain to the Mediterranean, while Lithothamnion glaciale and Lithothamnion erinaceum are cold temperate species that replace Lithothamnion corallioides in northern waters of the UK and the North East Atlantic (Melbourne et al., 2017). Martin & Hall-Spencer (2017) noted that a 3°C increase in temperature above that normally experienced by tropical or warm-temperate coralline algae caused bleaching and adversely affected heath, rates of calcification and photosynthesis and survival. Current trends in climate change driven temperature increases have already caused shifts in seaweed biogeography, as the tropical regions widen polewards, to the detriment of the warm-temperate region, and the cold-temperate region shrinks (Martin & Hall-spencer, 2017). 

Neopentadactyla mixta is not reported from shallow water, and it is only likely to be exposed to acute temperature changes due to thermal effluents. It is likely to withdraw into the sediment, away from the thermal plume, and be protected by the temperature of the interstitial waters.  Only long-term acute change (greater than the benchmark) is likely to adversely affect the population. In winter months, it is probably too deep to be affected by significant decreases in temperature as it burrows to a depth of 30-60 cm into the substratum (Smith & Keegan, 1985). Smith & Keegan (1985) suggested that light or winter temperature might be one cue for seasonal torpor but noted that winter turbulence and increased turbidity, due to water movement, may also induce Neopentadactyla mixta to overwinter at depth.

Sensitivity assessment.  An increase in temperature at the benchmark level is unlikely to affect Phymatolithon calcareum (Wilson et al., 2004) and given the slow growth rates exhibited by maerls, no effect is likely to be perceived within the duration of the benchmark.  Therefore, the SS.SMp.Mrl.Pcal biotope and its sub-biotopes probably have a ‘High’ resistance to an increase in temperature at the benchmark level.  Resilience is, therefore, ‘High’, and sensitivity 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. Further detail

Evidence

Maerl beds occur from the tropics to polar waters (Foster, 2001; Hinojosa-Arango & Riosmena-Rodriquez, 2004).  Maerl beds in the North East Atlantic range from Norway to the African coast, although the component maerl species vary in temperature tolerance (Birkett et al., 1998a; Wilson et al., 2004).  There are four species of biotope creating maerl beds in the UK.  These species vary in their distribution within the UK, a phenomenon which is thought to be due to their temperature tolerances.  Similarly, the associated communities within the maerl habitat differ and represent a diverse sample of species within the local area.  Lithothamnion coralloides is absent from Scottish waters.  This is due, either to winter temperatures dropping below the minimum survival temperature (between 2 and 5°C) or because the temperatures don’t allow a suitable growing season (Adey & McKibbin, 1970; cited in Wilson et al., 2004).  Lithothamnion corallioides had a higher minimum survival temperature; dying at 2°C and surviving without growth at 5°C (Adey & McKibbin, 1970).  In laboratory conditions, Phymatolithon calcareum survived down to 2°C, died at 0.4°C, and had a recorded optimum temperature for growth of 12-13°C (Adey & McKibbin 1970 cited in Wilson et al., 2004).  Phymatolithon calcareum showed no significant difference on photosynthetic activity at 9°C (the control), 17°C or 25°C for 4-5 weeks but were judged to be dead after 90 minutes at 40°C (Wilson et al., 2004). Temperature appears to confine Lithothamnion glaciale to northern parts of the British Isles, possibly because reproductive conceptacles are only produced in winter when temperature fall below 9°C (Hall-Spencer, 1994 cited in Wilson et al., 2004).  In addition, Adey (1970) found optimal growth rates of Lithothamnion glaciale between 10-12°C and that development of reproductive conceptacles in Lithothamnion glaciale requires winter temperatures of between 1-5°C (Adey, 1970). Blake & Maggs (2003) reported that the growth rate of Phymatolithon calcareum in the laboratory was only slightly affected by temperature treatments (10, 14 and 18°C), with an optimum at 10°C while the growth rate of Lithothamnion corallioides was significantly affected by temperature with an optimum at 14°C, at which temperature it grew faster than Phymatolithon calcareum. Martin et al. (2006) reported that primary productivity in Lithothamnion corallioides was twice as high in August as in January to February in the Bay of Brest. They found that primary productivity, calcification and respiration rates of Lithothamnion corallioides increased as temperature rose from 10 to 16°C (Martin et al., 2006).

The main maerl forming species have wide geographic ranges and their range indicates the limits of their temperature tolerance. Phymatolithon calcareum is a cold temperate species that ranges from Norway to the Mediterranean (Wilson et al., 2004; Martin et al., 2006) and tolerates high temperatures better than many subtidal temperate red algae (Wilson et al., 2004). Lithothamnion corallioides is a warm temperate species ranging from Ireland and the south of Britain to the Mediterranean, while Lithothamnion glaciale and Lithothamnion erinaceum are cold temperate species that replace Lithothamnion corallioides in northern waters of the UK and the North East Atlantic (Melbourne et al., 2017).

Neopentadactyla mixta is not reported from shallow water, and it is only likely to be exposed to acute temperature changes due to thermal effluents. It is likely to withdraw into the sediment, away from the thermal plume, and be protected by the temperature of the interstitial waters.  Only long-term acute change (greater than the benchmark) is likely to adversely affect the population. In winter months, it is probably too deep to be affected by significant decreases in temperature as it burrows to a depth of 30-60 cm into the substratum (Smith & Keegan, 1985). Smith & Keegan (1985) suggested that light or winter temperature might be one cue for seasonal torpor but noted that winter turbulence and increased turbidity, due to water movement, may also induce Neopentadactyla mixta to overwinter at depth.

Sensitivity assessment.  A decrease in temperature at the benchmark level is unlikely to affect Phymatolithon calcareum (Wilson et al., 2004).  Lithothamnion glaciale has a more northern distribution in the UK and may benefit from a localised temperature decrease in the long-term, so that the relative abundance of Lithothamnion glaciale and Phymatolithon calcareum may change in the long-term.  However, given the slow growth rates exhibited by maerls, no effect is likely to be perceived within the duration of the benchmark, but long-term climate change effects may be noticed in future.  Therefore, the biotope SS.SMp.Mrl.Pcal and its sub-biotopes probably have a ‘High’ resistance to a decrease in temperature at the benchmark level.  Resilience is, therefore ‘High’, and sensitivity is assessed as ‘Not sensitive’ at the benchmark level.

Salinity increase (local)

Benchmark. A increase in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

The majority of maerl beds occur in full salinity. Joubin, (1910 cited in Wilson et al., 2004) thought that maerl beds were only present in areas with lowered salinity.  Bosence (1976) found that, although surface salinities could be low, the benthic water was mostly fully saline.  The only maerl species currently thought to create beds in biotopes with salinities below fully marine is Lithothamnion glaciale (Connor et al., 2004), although the recently described Lithothamnion erinaceum might also.  Wilson et al. (2004) noted that Phymatolithon calcareum and Lithothamnion coralloides were tolerant up to 40 psu while most subtidal seaweeds can survive up to 50 psu.  The growth of Phymatolithon calcareum is impaired at salinities <24% (Adey & McKibbin, 1970; King & Schramm, 1982).

Echinoderms are restricted to the marine environment and one of the only stenohaline phyla in the animal kingdom (Russell, 2013).  Although some species can acclimatise to hypo/hypersaline conditions, Russell (2013) did not mention Neopentadactyla mixta amongst them. Smith (1983) noted that hypo or hypersaline water caused the animal to withdraw its tentacles. Neopentadactyla mixta is not reported from shallow water, and it is only likely to be exposed due to hypo/hypersaline effluents. Roberts et al. (2010b) reported that hypersaline effluents from desalination plants disperse with tens of metres of the discharge point but reported widespread alteration in seagrass and soft sediment communities in poorly flushed environment. Echinoderms and ascidians were amongst the most sensitive to hypersaline brine in the studies examined (Roberts et al., 2010b).  While hypersaline effluents are likely to sink to the seabed, and potentially penetrate into the sediment, the water movement characteristic of this biotope is likely to disperse the effluent and limit the effect to the immediate vicinity of any discharge point.

Sensitivity assessment. An increase in salinity above full is unlikely, except via the discharge of hypersaline effluents from desalination plants, none of which occur in the UK at present. An increase in salinity above 40 psu is likely to be detrimental to Neopentadactyla mixta and interrupt feeding but if prolonged for a year (see benchmark) may result in the death of individuals in the vicinity of the discharge. Others echinoderms are mobile are may avoid the impact. Maerl does not naturally occur within hypersaline areas, and although it may be able to tolerate a short-term increase in salinity, an increase to hypersaline conditions for a year would probably cause significant negative impacts. However, no evidence was found on which to base an assessment of maerl forming species. Therefore, a precautionary resistance assessment of 'Medium' for SMp.Mrl.Pcal.Nmix, and therefore SMp.Mrl.Pcal is suggested to represent the potential loss in abundance of the characterizing Neopentadactyla mixta population, but with Low confidence. Resilience is probably Medium so that sensitivity is assessed as Medium.

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

The majority of maerl beds occur in full salinity.  Although Joubin, (1910, cited in Wilson et al., 2004) thought that maerl beds were only present in areas with lowered salinity, Bosence (1976) found that although surface salinities could be low, the benthic water was mostly fully saline.  The only maerl species currently thought to survive in biotopes with salinities below fully marine is Lithothamnion glaciale (Connor et al., 2004).  The growth of Phymatolithon calcareum is impaired at salinities <24 ppt (Adey & McKibbin, 1970; King & Schramm, 1982). However, Wilson et al. (2004) noted that Phymatolithon calcareum was more tolerant of low salinity than Lithothamnion glaciale in their experiments.  Both species survived at 3 psu for five weeks but showed significantly reduced photosynthetic activity.  However, at 15 psu, Phymatolithon calcareum recovered from the initial drop in photosynthetic activity, while Lithothamnion glaciale did not.  

Echinoderms are restricted to the marine environment and one of the only stenohaline phyla in the animal kingdom (Russell, 2013).  Although some species can acclimatise to hypo/hypersaline conditions, Russell (2013) did not mention Neopentadactyla mixta amongst them. Smith (1983) noted that hypo or hypersaline water caused the animal to withdraw its tentacles. A reduction is salinity to reduced (18 to <30 psu) for a year is likely to reduce feeding or drive Neopentadactyla mixta into the sediment where it cannot feed. Its seasonal torpor lasts from September to March each year, during which it loses condition significantly so it is unlikely to survive for a year without feeding. 

Sensitivity assessment.  SS.SMp.Mrl.Pcal, SS.SMp.Mrl.Pcal.R, SS.SMp.Mrl.Pcal.Nmix are characterized by Phymatolithon calcareum and are all only found in fully saline conditions.  However, a reduced salinity for an extended period of time would stress Phymatolithon calcareum and could lead to mortality.  No long-term salinity experiments have been carried out on Phymatolithon calcareum. However, reduced salinity may also result in a reduction in the abundance of echinoderms within the biotope, in particular, Neopentadactyla mixta. Therefore, precautionary resistance assessment of ‘Medium’ has been given, the resilience is, therefore, ‘Very low’ and sensitivity is assessed as ‘Medium’.

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s to 0.2 m/s for more than one year. Further detail

Evidence

Maerl requires enough water movement to prevent smothering with silt (Hall-Spencer, 1998).  Therefore, maerl beds are restricted to areas of strong tidal currents or wave oscillation (Birkett et al. 1998a).  For example, Birkett et al. (1998a) quote a flow rate of 0.1 m/s across the maerl bed at spring tides in Greatman’s Bay, Galway, while the UK biotope classification (Connor et al., 2004) reports maerl beds occurring at sites with between moderately strong to very weak tidal streams.  As Birkett et al. (1998a) note, local topography and wave generated oscillation probably result in stronger local currents at the position of the bed.  Hall-Spencer et al. (2006) reported that maerls beds in the vicinity of fish farms became silted with particulates from fish farms even in areas of strong flow. Hall-Spencer et al. (2006) reported peak flow rates of 0.5 to 0.7 m/s at the sites studied, and one site experienced mean flows of 0.11 to 0.12 m/s and maxima of 0.21 to 0.47 m/s depending on depth above the seabed. 

Neopentadactyla mixta occurs in maerl beds and coarse gravel sediments, both of which are associated with water flow either due to tidal streams (moderately strong to weak) or wave mediated water movement (exposed to moderately wave exposed) (Connor et al. 2004).  For example, the beds of Neopentadactyla mixta in coarse sediments examined by Konnecker & Keegan (1973) were found in tidal currents of up to 2.5 knots (ca 1.28 m/s).  Nevertheless, artificially increased current beyond the calm weather, spring tide, maximum of ca 1.5 m/s caused Neopentadactyla mixta to stop feeding a withdraw into its burrow, as did bombardment with dislodged sediment (Smith & Keegan 1985). Similarly, a heavy gale in August caused Neopentadactyla mixta to withdraw deep into the sediment for six to ten days (Smith & Keegan, 1985).  The species regularly undertakes a ca six month long torpor period, during which it loses condition and lipid energy stores.  Smith & Keegan (1985) suggested that the overwinter torpor may be a response to poor food availability coupled with increased turbulence experienced in winter at their study site. 

Sensitivity assessment.  An increase in water flow to strong or very strong may winnow away the surface of the bed and result in loss of the biotope. A decrease in water flow may result in increased siltation, smothering maerl, and causing the death of maerl and significant change in the associated community (see smothering/siltation below).  The effect will depend on local hydrography and the wave climate.  A change of 0.1-0.2 m/s may have a limited effect in areas of moderately strong flow but may be significant in areas of weak or negligible flow.  Hall-spencer (pers. comm.) noted that any change in water flow is likely to affect maerl beds. Therefore, a resistance of 'Low' is suggested but with 'Low' confidence. Hence, as resilience is likely to be 'Very low', sensitivity is assessed as 'High'.

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. Further detail

Evidence

In the UK, maerl beds do not occur in the intertidal, as maerl is highly sensitive to desiccation (Wilson et al., 2004).  Also, it is very unlikely that a maerl bed would be exposed at low water as a result of human activities or natural events (Tyler-Walters, 2013).  Therefore, this pressure is probably ‘Not relevant’.

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year. Further detail

Evidence

Maerl requires enough water movement to prevent smothering with silt (Hall-Spencer, 1998).  Therefore, maerl beds develop in strong currents but are restricted to areas of low wave action.  For example, in Mannin Bay dense maerl beds were restricted to less wave exposed parts of the bay (Birkett et al., 1998a). In Galway Bay, Keegan (1974) noted the formation of ripples due to wave action and storms, where the ripples were flattened over time by tidal currents.  However, he reported that the rippled area (average crest height 20 cm) had a poor faunal diversity with heavy macroalgal settlement on any firm substratum, including the tubes of Chaetopterus sp.  However, the infauna was a typical ‘Venus’ community, the majority of which was found at depths of more than 20 cm.  Hall-Spencer & Atkinson (1999) noted that mega-ripples at their wave exposed site were relatively stable but underwent large shifts due to storms.  However, the mixed sediments of the subsurface of the bed (>12 cm) were unaffected so that the burrows of the mud shrimp remained in place.  Similarly, Birkett et al. (1998a) noted that in areas where storms affected the maerl at a depth of 10 m, only the coarse upper layer of maerl was moved while the underlying layers were stable.  Following storms, infaunal species renewed burrow linings within a week.  However, the epiflora of maerl beds was severely disturbed by storms in Galway Bay with a marked drop in abundance in winter months.  Deep beds are less likely to be affected by an increase in wave exposure.

Neopentadactyla mixta occurs in maerl beds and coarse gravel sediments, both of which are associated with water flow either due to tidal streams (moderately strong to weak) or wave mediated water movement (exposed to moderately exposed) (Connor et al., 2004). Smith & Keegan (1985) noted that a  heavy gale at their study site in August caused Neopentadactyla mixta to withdraw deep into the sediment for six to ten days.  The species regularly undertakes a ca six month long torpor period, during which it loses condition and lipid energy stores.  Smith & Keegan (1985) suggested that the overwinter torpor may be a response to poor food availability coupled with increased turbulence experienced in winter at their study site. 

Sensitivity assessment.  Maerl beds occur in a range of wave exposures and can survive in areas subject to wave action and storms.  Therefore, an increase in wave exposure is probably detrimental to shallow maerl beds. Similarly, a decrease in wave action may be detrimental where wave action is the main contribution to water movement through the bed, due to the potential increase of siltation and reduction in infaunal diversity.  However, a 3-5% change in significant wave height is unlikely to damage the maerl bed. Both resistance and resilience are assessed as ‘High’, and the biotope is assessed as ‘Not sensitive’ to this pressure at the benchmark.

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of contaminants on maerl-forming species are summarized below.  The full 'Maerl evidence review' should be consulted for details of the studies examined. 

Only sublethal effects were reported after exposure of calcareous red algae to heavy metals (Amiard, 1973 (data only); Wilson et al., 2004). In particular, Wilson et al. (2004) exposed the maerl-forming species Phymatolithon calcareum to a single dose of a mixture of heavy metals in the ratio 37:16:14:11:1, Zn: Pb: Ni: Cu: Cd, where the Cd concentration of ranges from 0.174 to 174 ppb and 1.74 ppb represented standard industrial effluent. Phymatolithon calcareum experienced a significant reduction in photosynthetic capacity depending on concentration but recovered quickly. However, the authors noted that longer-term exposure to heavy metals may have chronic effects. Nevertheless, only sublethal effects were reported. ‘Insufficient evidence’ is recorded as it is imprudent to suggest that all maerl-forming species are ‘Not’ sensitive’ without further evidence.

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of contaminants on maerl-forming species are summarized below.  The full Maerl evidence review should be consulted for details of the studies examined. 

The effect of petrochemical hydrocarbons on encrusting coralline algae was only reported by a single paper using intertidal field mesocosms (Bokn et al., 1993). They reported that the encrusting coralline Phymatolithon lenormandii exhibited a significant decrease in cover in the upper shore mesocosm after exposure to 129.4 µg/l Diesel WAF for two years but an increase in cover in the mid and lower shore mesocosms. This suggests that diesel WAF only affect the species in the more physiologically demanding upper shore. However, the resultant worst-case resistance of Phymatolithon lenormandii could be assessed as ‘Medium’. No information on its recovery was available. If it is similar to other encrusting corallines then its resilience is probably ‘High’ and its sensitivity ‘Low’. However, it is difficult the extrapolate to maerl-forming species with any confidence, especially based on a single study.

The effect of oil spills on coralline species was reported in several papers but no evidence of the effects on maerl-forming species was found. Diaz et al. (2009a) examined changes in macroalgal abundance (inc. Corallina elongata and Lithophyllum incrustans) along the Basque coast after the Prestige oil spill but did not find any significant differences between oiled and non-oiled sites. However, Bowman et al. (1978) reported that 100% of the cover of Lithothamnia was bleached and dead rims of Lithothamnia in lower shore rock pools after the Dounreay oil spill and treatment with BP100X. Similarly, Newey & Seed (1995) reported bleached and dead coralline algae (no species were specified) in mid-shore rockpools close to the wreck of the Braer oil tanker. Jackson et al. (1989) also reported that crustose corallines (no species were specified) and other fleshy algae decreased in cover after the Panamanian oil spill, to levels below those observed before the spill. Crump et al., (1999) reported that encrusting coralline algae, Lithothamnion incrustans, Phymatolithon purpureum, and Corallina officinalis were bleached in West Angle Bay immediately after the Sea Empress oil spill but recovered quickly, which suggested only the surface layers were affected rather than individuals were killed. Crump et al. (1999) also stated that previous literature has shown oil and dispersants to have harmful effects on the pigmentation of red algae in experimental conditions.

Smith et al. (1968) reported that the dispersants used to treat the Torrey Canyon oil spill killed Corallina and Lithophyllum in shallow rock pools, while those in deep pools survived, and that Corallina was killed in high shore pools but appeared healthy in mid-shore pools. Overall, the effects of detergents depended on duration and concentration of exposure, assuming lower shore populations were exposed for shorter periods than higher-shore populations as the tide returned.

Felder et al. (2014) and Fredericq et al. (2014) both examined the effects of the DWH oil spill on deep (50-77 m) ‘rhodolith’ beds in the Gulf of Mexico. They both reported that the ‘rubble’ and ‘rhodoliths’ were visibly bleached and where the encrusting calcareous algae were alive, the diverse seaweed community was lost. They suggested that the rhodolith and encrusting species included Lithophyllum sp., Lithothamnion sp., Mesophyllum sp., and Porolithon sp.

Sensitivity assessment. The above evidence suggests that exposure to oil spills and/or their dispersants can result in bleaching or death of calcareous coralline algae, especially encrusting corallines, depending on the length of exposure, shore height, and type of oil. Therefore, the resistance of encrusting corallines or Corallina sp. to exposure to oil spills and dispersants is assessed as ‘Low’ based on the worst-case scenario reported by Bowman et al. (1978). Hence, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’. Confidence in the assessment is ‘Medium’ due to the variation in the effect between studies.

No direct evidence of the effect of petrochemical hydrocarbons on maerl-forming species or their beds was found. However, the evidence of effects on similar species, e.g. Phymatolithon spp. and Lithothamnion spp. above suggests that maerl-forming species and their beds may experience similar bleaching and possible death depending on exposure to oil spills. As maerl beds are sublittoral they may be protected from direct exposure, although the evidence from the DWH spill at 50 to 77 metres suggests they are not immune. The potential for smothering by oil was not discussed. However, Tuya et al. (2023) suggested oil spills were a potential threat to ‘rhodoliths’ beds worldwide. Therefore, the resistance of maerl-forming species and their beds to exposure to oil spills and/or dispersants is assessed as ‘Low’ as a precaution, albeit with ‘Low’ confidence due to the lack of direct evidence. Hence, resilience is assessed as ‘Very low’ and sensitivity as ‘High’.

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of contaminants on maerl-forming species are summarized below.  The full Maerl evidence review should be consulted for details of the studies examined. 

The herbicides Diuron, Atrazine and Hexazinone were found to inhibit photosynthesis in crustose coralline algae (Harrington et al. 2005; Negri et al., 2011; McCoy & Kamenos, 2015). Negri et al. (2011) reported that diuron was the most toxic after 24 hours, then hexazione, and then atrazine in Neogoniolithon fosliei. Harrington et al. (2005) reported that diuron was also toxic to Porolithon onkodes when applied alone. Visible bleaching of Porolithon onkodes was observed at 29 µg/l of diuron, likely due to the destruction of chloroplasts and carotenoids that causes its colouration to lighten. However, when exposure to diuron was combined with sedimentation, the toxicity of diuron was greater and the time taken to recover in clean water was increased. However, MacVicar et al. (2022) reported no effects of exposure of Lithothamnion spp. to a high concentration of the UV filter oxybenzone (benzophenone-3; BP-3) after 15 days.

A reduction in photosynthesis is likely to reduce growth rates and increase an individual’s susceptibility to other stresses, such as sedimentation. Nevertheless, only sublethal effects were reported based on short-term experiments and no evidence of long-term effects were found. ‘Insufficient evidence’ is recorded as it is imprudent to suggest that all coralline algae and maerl-forming species are ‘Not’ sensitive’ to exposure to herbicides without further evidence.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels. Further detail

Evidence

Hernández et al. (2011) reported that Corallina elongata and Jania rubens accumulated plutonium (Pu) in granules but did not report any adverse effects on the species. Hence, the evidence does not support an assessment of resistance or sensitivity. 

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of contaminants on maerl-forming species are summarized below. The full Maerl evidence review should be consulted for details of the studies examined. 

Legrand et al. (2022) examined the effects of exposure to hydrogen peroxide (H₂0₂), used as an antiparasitic treatment in Norwegian salmon farms, on photosynthesis in the maerl-forming species Lithothamnion soriferum. Photosynthesis was significantly reduced at concentration >=200 mg/l H₂0₂ after 1-hour exposure but recovered after 48 hours or 28 days. However, they also showed significant bleaching (28% at 200 mg/l or 63% at 2,000 mg/l) after 28 days of recovery. McCoy & Kamenos (2015) noted that bleaching could indicate death in coralline red algae and could result in structural damage but was also reversible, and the long-term effects of bleaching required further study. The authors went on to suggest that prolonged or repeated exposure may have a greater impact. The authors also noted that the crustacean fauna of maerl beds may be particularly sensitive to H₂0₂ so the diversity of affected maerl beds may be threatened, although further study was required (Legrand et al., 2022).

Overall, the evidence suggests that the maerl-forming species Lithothamnion soriferum is adversely affected by hydrogen peroxide exposure >=200 mg/l, resulting in reduced photosynthesis, reduced growth, and significant bleaching. Therefore, resistance is assessed as ‘Medium’ as a precaution but with ‘Low’ confidence as mortality was not reported directly. Hence, resilience is assessed as ‘Very low’ and sensitivity as ‘Medium’. 

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). Further detail

Evidence

Deoxygenation at the benchmark level is likely to be detrimental to the maerl beds and their infaunal community but mitigated.  Water flow experienced by these biotopes suggests that deoxygenating conditions may be short-lived.  However, Hall-Spencer et al. (2006) examined maerl beds in the vicinity of fish farms in strongly tidal areas. They noted a build-up of waste organic materials up to 100 m from the farms examined and a 10-100 fold increase in scavenging fauna (e.g. crabs). In the vicinity of the farm cages, the biodiversity was reduced, particularly of small crustaceans, with significant increases in species tolerant of organic enrichment (e.g. Capitella). In addition, they reported less live maerl around all three of the fish farm sites studied than the 50-60% found at reference sites. Most of the maerl around fish farms in Orkney and South Uist was dead and clogged with black sulphurous anoxic silt. The Shetland farm had the most live maerl but this was formed into mega-ripples, indicating that the maerl had been transported to the site by rough weather (Hall-Spencer et al., 2006).   Eutrophication resulting from aquaculture is cited as one reason for the decline of maerl beds in the North East Atlantic (Hall-Spencer et al., 2010).  In the laboratory, Wilson et al. (2004) noted that burial in black muddy sand, smelling of hydrogen sulphide, was fatal to live maerl. Even thalli placed on the surface of the black muddy sand died within two weeks, together with thalli buried by 0.25 cm and 2 cm of the sediment (Wilson et al., 2004). A study of a phytoplankton bloom that killed herring eggs on a maerl bed in the Firth of Clyde found that the resultant anoxia caused mass mortalities of the burrowing infauna (Napier, in press, cited by Hall-Spencer pers comm.).

Neopentadactyla mixta probably needs coarse sediments to survive, as the open matrix provided by coarse sediments or maerls at depth, together with water flow, ensures that the water is oxygenated at depth in the sediment. Neopentadactyla mixta reduces its metabolism and oxygen consumption from 0.11 ml O₂/ gm dry wt. to 0.03 ml O₂/ gm dry wt. during its overwinter torpor (Smith & Keegan, 1985). Therefore, it might be able to survive lower oxygen levels overwinter than in spring, summer and autumn. Vaquer-Sunyer & Duarte (2008) suggested a median sublethal oxygen concentration of 1.22 mg O₂/l (± 0.25) for a number of echinoderms reviewed in their study.  Echinoderms were neither the most or the least sensitive of the taxonomic groups examined. 

Sensitivity assessment. The available evidence suggests that maerl and its associated community is sensitive to the effects of deoxygenation and anoxia, even in areas of strong water movement. Therefore, resistance has been assessed as ‘Low’, resilience as ‘Very low’, and sensitivity is assessed as ‘High’.

Nutrient enrichment

Benchmark. Compliance with WFD criteria for good status. Further detail

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations.  The nutrient enrichment of a marine environment leads to organisms no longer being limited by the availability of certain nutrients.  The consequent changes in ecosystem functions can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) decreases in dissolved oxygen and uncharacteristic microalgal blooms (Bricker et al., 1999, 2008).

Johnston & Roberts (2009) undertook a review and meta-analysis of the effect of contaminants on species richness and evenness in the marine environment.  Of the 47 papers reviewed relating to nutrients as a contaminant, over 75% found that it had a negative impact on species diversity, <5% found increased diversity, and the remaining papers finding no detectable effect.  None of the 47 papers considered the impact of nutrients on this biotope.  Yet this finding is still relevant as the meta-analysis revealed that the effect of marine pollutants on species diversity was ‘remarkably consistent’ between habitats (Johnston & Roberts, 2009).  It was found that any single pollutant reduced species richness by 30-50% within any of the marine habitats considered (Johnston & Roberts, 2009).  Throughout their investigation, there were only a few examples where species richness was increased due to the anthropogenic introduction of a contaminant.  These examples were almost entirely from the introduction of nutrients, either from aquaculture or sewage outfalls.  However research into the impacts of nutrient enrichment from these sources on intertidal rocky shores often lead to shores lacking species diversity and the domination by algae with fast growth rates (Abou-Aisha et al., 1995, Archambault et al., 2001, Arévalo et al., 2007, Diez et al., 2003, Littler & Murray, 1975).

Grall & Glemarec (1997) noted that increased turbidity and eutrophication due to agricultural runoff in Brittany prevented the establishment of many algal species resulting in domination of ubiquitous species (e.g. Ceramium sp. and Ulva sp.), while localised eutrophication due to fish and mussel farming (aquaculture) in a sheltered area resulted in bacterial mats of Beggiatoa.  Hall-Spencer et al. (2006) examined maerl beds in the vicinity of fish farms in strongly tidal areas.  They noted a build-up of waste organic materials up to 100 m from the farms examined and a 10-100 fold increase in scavenging fauna (e.g. crabs). In the vicinity of the farm cages, the biodiversity was reduced, particularly of small crustaceans, with significant increases in species tolerant of organic enrichment (e.g. Capitella). In addition, they reported less live maerl around all three of the fish farm sites studied than the 50-60% found at reference sites. Most of the maerl around fish farms in Orkney and South Uist was dead and clogged with black sulphurous anoxic silt (Hall-Spencer et al., 2006). Eutrophication resulting from aquaculture is cited as one reason for the decline of some beds in the North East Atlantic (Hall-Spencer et al., 2010). 

In Brittany, numerous maerl beds were affected by sewage outfalls and urban effluents, resulting in increases in contaminants, suspended solids, microbes and organic matter with resultant deoxygenation (Grall & Hall-Spencer, 2003).  This resulted in increased siltation, higher abundance and biomass of opportunistic species, loss of sensitive species and reduction in biodiversity.  Grall & Hall-Spencer (2003) note that two maerl beds directly under sewage outfalls were converted from dense deposits of live maerl in the 1950s to heterogeneous mud with maerl fragments buried under several centimetres of fine sediment with species poor communities.  These maerl beds were effectively lost. 

Sensitivity assessment.  The effect of eutrophication on maerl beds is difficult to disentangle from the effects of organic enrichment, and sedimentation.  It is likely that nutrient enrichment could adversely affect the infauna and epiflora communities but the direct effect on maerl is unclear. Nevertheless the biotope is 'Not sensitive' at the pressure benchmark of compliance with good status as defined by the WFD.

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr. Further detail

Evidence

The organic enrichment of a marine environment at this pressure benchmark leads to organisms no longer being limited by the availability of organic carbon.  The consequent changes in ecosystem functions can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) and decreases in dissolved oxygen and uncharacteristic microalgae blooms (Bricker et al., 1999, 2008).  Grall & Hall-Spencer (2003) considered the impacts of eutrophication as a major threat to maerl beds.

Hall-Spencer et al. (2006) compared maerl beds under salmon farms with reference maerl beds.  It was found that maerl beds underneath salmon farms had visible signs of organic enrichment (feed pellets, fish faeces and/or Beggiatoa mats), and significantly lower biodiversity.  At the sites underneath the salmon nets, there were 10 – 100 times the number of scavenging species present compared to the reference sites.  Grall & Glémarec (1997) noted similar decreases in maerl bed biodiversity due to anthropogenic eutrophication in the Bay of Brest.  In Brittany, numerous maerl beds were affected by sewage outfalls and urban effluents, resulting in increases in contaminants, suspended solids, microbes and organic matter with resultant deoxygenation (Grall & Hall-Spencer, 2003).  This resulted in increased siltation, higher abundance and biomass of opportunistic species, loss of sensitive species and reduction in biodiversity.  Grall & Hall-Spencer (2003) note that two maerl beds directly under sewage outfalls were converted from dense deposits of live maerl in the 1950s to heterogeneous mud with maerl fragments buried under several centimetres of fine sediment with species poor communities.  These maerl beds were effectively lost. 

Grall & Hall-Spencer (2003) note that two maerl beds directly under sewage outfalls were converted from dense deposits of live maerl in the 1950s to heterogeneous mud with maerl fragments buried under several centimetres of fine sediment with species poor communities.  These maerl beds were effectively lost. 

Sensitivity assessment.  Little empirical evidence was found to directly compare the benchmark organic enrichment of maerl biotopes.  However, the evidence strongly suggests that organic enrichment and resultant increased in organic content, hydrogen sulphide levels and sedimentation may result in loss of maerl beds.  Resistance has been assessed as ‘None’ and resilience has been assessed as ‘Very low’.  This gives an overall sensitivity score of ‘High’.

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

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. Further detail

Evidence

Maerl biotopes can contain a variety of sediment types including gravels, sand and mud.  However, maerl biotopes never contain bedrock.  Therefore, if rock or an artificial substrate was to replace the normal substratum within this biotope the physical conditions required for this biotope would be lost along with the biotope itself.  Therefore, resistance is likely to be 'None', resilience 'Very Low' (permanent change) and sensitivity is assessed as 'High'. 

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification). Further detail

Evidence

The sediment associated with maerl biotopes varies from shell and maerl gravel through to sand and mud.  The characterizing maerl species is also not attached to the substratum, and instead, lies over the top of it.  Therefore, if the substratum were to change this wouldn’t have a negative effect on the characterizing species.  The other species within the associated community depend on different aspects of the sediment.  Those species which are found infaunally may be negatively affected.

Sensitivity assessment.  A change in this pressure at the benchmark will not affect the characterizing species yet may affect other species found infaunally within the biotope.  The loss of an infaunal species will create a niche for another species to become established, therefore continuing the biological function and ensuring the character of the biotope remains.  Resistance and resilience are assessed as ‘High’, resulting in a ‘Not Sensitive’ assessment.

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). Further detail

Evidence

As maerl need to photosynthesise only the top layer of the deposit which has access to light will be alive.  Live maerl also requires good water flow around it, a factor which is likely to be limited 30 cm into the bed.  Maerls requirement for both light and water flow suggest that the majority if not all of the live maerl will be found in the top 30 cm of a maerl bed.  This is also where many of the associated species will be found.  Although long-lived elements of maerl bed fauna are known to burrow up to 72 cm into the substrate (Hall-Spencer & Atkinson, 1999).  The extraction of the substratum to 30 cm within this biotope would remove the vast majority of the biological component of the biotope.

Hauton et al. (2003) undertook experimental suction (hydraulic dredging) in Stravanan Bay, Scotland, a site subject to scallop dredging and recorded as impacted dead maerl by Kamenos et al. (2003).  The suction dredge removed epiflora (burrowing algae and macroalgae), maerl, slow moving epifauna (e.g. starfish, gastropods and clingfish) and mainly infauna.  Large or fragile polychaetes (e.g. Chaetopterus) and Cerianthus lloydii were removed and damaged, while polychaetes with tough bodies or strong tubes survived.  Large infaunal bivalves dominated the catch, including Dosinia exoleta, Tapes rhomboides, Abra alba, and Ensis arcuatus but, while Mya truncata and Lutraria angustior were not caught because of their depth, the catch did include torn siphons from these species; an injury they are unlikely to survive.  The dredge resulted in a visible track that left numerous damaged megafauna, which in turn attracted scavengers.  In addition, the dredging fragmented maerl and resulted in a large plume of fine sediment that settled over the surrounding area.  However, recovery was not examined.  Hall-Spencer et al. (2003) drew attention to the dangers of suction dredging for bivalves in maerl beds, especially as many of the larger infaunal bivalves are long-lived (e.g. Dosinia exoleta), suggesting that the population would take a long time to recover. 

Neopentadactyla mixta lives in the sediment is a ‘U-shaped’ posture with its oral tentacles raised above the surface and its anus just below the surface of the sediment (Smith & Keegan, 1985).  It is usually found in this position in its burrow 15-25 cm deep in the sediment (Könnecker & Keegan, 1973).  However, in the winter months (ca September to March) its burrows into the sediment to a depth of 30-60 cm.  It maintains this depth, even if the surface of the sediment is eroded or accreted (Smith & Keegan, 1985). In spring to autumn, extraction of the sediment to 30 cm is likely to remove the majority of the resident population but in winter, the majority of the population would survive as long as suitable substratum remained after extraction.

Sensitivity assessment.  The resistance of the biotope to this pressure at the benchmark is ‘None’ and the resilience is assessed as ‘Very low’, giving the biotope a ‘High’ sensitivity to 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). Further detail

Evidence

Physical disturbance can result from; channelization (capital dredging), suction dredging for bivalves, extraction of maerl, scallop dredging or demersal trawling.  The effects of physical disturbance were summarised by Birkett et al. (1998a) and Hall-Spencer et al. (2010), and documented by Hall-Spencer and co-authors (Hall-Spencer, 1998; Hall-Spencer et al.,  2003; Hall-Spencer & Moore, 2000a,b; Hauton et al., 2003; and others).  For example, in experimental studies, Hall-Spencer & Moore (2000a, c) reported that the passage of a single scallop dredge through a maerl bed could bury and kill 70% of living maerl in its path.  The passing dredge also re-suspended sand and silt that settled over a wide area (up to 15 m from the dredged track) and smothered the living maerl.  The dredge left a 2.5 m track and damaged or removed most megafauna within the top 10 cm of maerl (Hall-Spencer & Moore, 2000a).  For example; crabs, Ensis species, the bivalve Laevicardium crassum, and sea urchins.  Deep burrowing species such as the tube anemone Cerianthus lloydii and the crustacean Upogebia deltaura were protected by depth, although torn tubes of Cerianthus lloydii were present in the scallop dredge tracks (Hall-Spencer & Moore, 2000a).  Neopentadactyla mixta may also escape damage due to the depth of its burrow, especially during winter torpor.  Hall-Spencer & Moore (2000a) reported that sessile epifauna or shallow infauna such as Modiolus modiolus or Limaria hians, sponges and the anemone Metridium senile where present, were significantly reduced in abundance in dredged areas for 4 years post-dredging.  Other epifaunal species, such as hydroids (e.g. Nemertesia species) and red seaweeds are likely to be removed by a passing dredge.  The tracks remained visible for up to 2.5 years.  In pristine live beds, experimental scallop dredging reduced the population densities of epibenthic species for over 4 years.  However, in previously dredged maerl beds, the benthic communities recovered in 1-2 years. 

Abrasion may break up maerl thalli into smaller pieces, reduce structural heterogeneity and lower the diversity of species (Kamenos et al., 2003).  Hall-Spencer et al. (2003) noted that certain maerl beds in the Bay of Brest have been dredged for scallops and Venus verrucosa for over 40 years, yet remain productive with high levels of live maerl.  Although they suggest that this is due to local restrictions that limit the activity to one scallop dredge per boat.  Nevertheless, scallop dredging, demersal trawling and extraction have been reported to contribute to declines in the condition of maerl beds in the North East Atlantic and the UK (Barbera et al., 2003, Hall-Spencer et al., 2010, Hall-Spencer et al., 2003).  Irish maerl are considered to be in generally good condition but some are deteriorating due to commercial extraction, mariculture, demersal fishing and the localized effects of boat mooring chains (Vize, 2005).

Sensitivity assessment.  Physical disturbance is likely to result is drastic changes in and loss of components of the community within the maerl bed.  Fragmentation of the maerl will not kill the maerl directly but subsequent death is likely due to a reduction in water flow caused by compaction and sedimentation (Hall-spencer & Moore, 2000a; 2000c).  Dredging can create plumes of sediment that can settle on top of the maerl, and overturn and bury maerl, causing it to be smothered, a pressure to which maerl is highly intolerant (see smothering and siltation (light) pressure).  The evidence from Hall-Spencer & Moore (2000a; 2000c) alone strongly suggests that resistance to physical disturbance and abrasion is ‘None’. Therefore, resilience is probably ‘Very low’, resulting in a sensitivity assessment of ‘High’. 

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

As maerl need to photosynthesise only the top layer of the deposit which has access to light will be alive.  Live maerl also requires water flow around it.  Maerl beds become less structurally complex if they have been affected by dredging (Kamenos et al., 2003).  A lack of structural complexity will restrict the niches for other species, reducing biodiversity and will also restrict water flow through the bed.  Penetration and disturbance both have the capacity to break up maerl into smaller fragments. The evidence provided within the abrasion and disturbance pressure shows that maerl is intolerant of the abrasion of any form. Penetration of the maerl bed will exacerbate the negative effect by damaging more of the underlying maerl.

Sensitivity assessment.  Based on the evidence provided within the 'abrasion and disturbance' assessment the resistance of the biotope to this pressure at the benchmark is considered ‘None’ and the resilience is assessed as ‘Very low’ so that sensitivity is assessed as ‘High’.

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. Further detail

Evidence

Maerl's requirement for light restricts live maerl to depths shallower than 32 m in the relatively turbid waters of northern Europe (Hall-Spencer, 1998).  An increase in suspended sediments in the water column will increase light attenuation and decrease the availability of light to the biotope.  A decrease in light availability will alter the ability of the maerl to photosynthesise. This could be detrimental to maerl beds found towards their bottom depth limit (i.e. 32 m).  An increase in suspended solids within this biotope is likely to also increase scour, as there are characteristically high levels of water movement through maerl beds.  Scour is known to induce high mortality in early post-settlement algal stages and prevents the settlement of propagules owing to the accumulation of silt on the substratum (Vadas et al., 1992).  Increased particulates may provide additional food for filter feeders.  However, an increase in suspended sediment may increase the fines within the bed, decreasing water flow and oxygenation through the bed, and hence the depth of the surface epifauna.  It may result in an increase in burrowing species compared to filter feeding species.  However, De Grave (1999) noted that sedimentary heterogeneity within maerl beds (including maerl debris with mud, sand or gravel) resulted in only minor changes in the community of amphipods and crustaceans present.

A decrease in suspended solids will increase light levels, which could benefit maerl.  However, a decrease in the suspended matter is likely to reduce the quantity of food available for filter feeders.  This could change the species present within the community.

Sensitivity assessment.  Any factor which decreases the ability for the characterizing maerl species to photosynthesise will have a negative impact.  Examples of the biotope found at the very bottom depth limit may experience high levels of mortality of the characterizing species.  The resistance of this biotope is assessed as ‘Medium’ and the resilience is ‘Very low’.  Hence, the sensitivity is assessed as ‘Medium’ to the pressure at the benchmark level.

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. Further detail

Evidence

Smothering results from the rapid deposition of sediment or spoil, which may occur after dredging (suction or scallop), capital dredging (channelization), extreme runoff, spoil dumping etc.  The effects depend on the nature of the smothering sediment.  For example, live maerl was found to survive burial in coarse sediment (Wilson et al., 2004) but to die in fine sediments.  Phymatolithon calcareum survived for 4 weeks under 4 and 8 cm of sand or gravel but died within 2 weeks under 2 cm of muddy sand.  Wilson et al. (2004) suggested that the hydrogen sulphide content of the muddy sand was the most detrimental aspect of burial since even those maerl nodules on the surface of the muddy sand died within two weeks. They also suggested that the high death rate of maerl observed after burial due to scallop dredging (Hall-Spencer & Moore, 2000a,c) was probably due to physical and chemical effects of burial rather than a lack of light (Wilson et al., 2004).

In addition, detrimental effects on Fucus embryos were reported in fine sediments, presumably as fine sediment restricts water flow.  Similarly, fine sediment is likely to prevent settlement of algal propagules, so that the effects are potentially greater during their settlement period.  Kranz (1972; cited in Maurer et al. (1986) reported that shallow burying siphonate suspension feeders are typically able to escape smothering with 10-50 cm of their native sediment and relocate to their preferred depth by burrowing.  Dow & Wallace (1961) noted that large mortalities in clam beds resulted from smothering by blankets of algae (Ulva sp.) or mussels (Mytilus edulis).  In addition, clam beds have been lost due to smothering by 6 cm of sawdust, thin layers of eroded clay material, and shifting sand (moved by water flow or storms) in the intertidal.

Smothering by 5 cm of sediment (the benchmark) is likely to clog or reduce water flow through the surface of the bed, and directly smother small non-mobile members of the epifauna and epiflora, while larger species e.g. sea squirts, anemones, some sponges and macroalgae would protrude above the smothering sediment. Mobile small burrowing species (e.g. amphipods and polychaetes) would probably burrow to safety. However, non-motile epifauna (e.g. encrusting bryozoans and small hydroids) and small or prostrate algal will probably be reduced in abundance.  Deep burrowing bivalves may experience some mortality due to loss of water flow through the bed, deoxygenating and lack of food depending on their depth. But large burrowing anemones and mud shrimp would probably just burrow through the smothering material.  In Galicia, France, ongoing deterioration of maerl has been linked to mussel farming which increases sedimentation, reducing habitat complexity, lowering biodiversity, and killing maerl (Pena & Barbara, 2007a, b;  cited in Hall-Spencer et al., 2010). Wilson et al. (2004) also point out that the toxic effect of fine organic sediment and associated hydrogen sulphide explain the detrimental effect on maerl beds of Crepidula fornicata in Brittany, sewage outfalls, and aquaculture (Grall & Hall-spencer, 2003).

Neopentadactyla mixta lives in the sediment is a ‘U-shaped’ posture with its oral tentacles raised above the surface and its anus just below the surface of the sediment (Smith & Keegan, 1985).  It is usually found in this position in its burrow 15-25 cm deep in the sediment (Könnecker & Keegan, 1973).  However, in the winter months (ca September to March) its burrows into the sediment to a depth of 30-60 cm.  It maintains this depth, even if the surface of the sediment is eroded or accreted (Smith & Keegan, 1985).The tentacular crown can expand to 140 cm² (Smith & Keegan, 1985) and probably extends to ca 4-5 cm above the substratum (expert opinion). However, the deposit of 5 cm of fine sediment would probably discourage Neopentadactyla mixta from feeding and it would probably withdraw into its burrow.

Sensitivity assessment.  Even though these biotopes occur in areas of tidal or wave mediated water flow, the fine smothering material would penetrate the open matrix of the maerl bed rather than sit on top of the bed. At the pressure benchmark this biotopes resistance is assessed as ‘None’, and the resilience is ‘Very low’, resulting in an overall sensitivity of ‘High’. 

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. Further detail

Evidence

A deposit at the pressure benchmark would cover all species with a thick layer of fine materials.  The pressure is significantly higher than light smothering discussed above.  Therefore, resistance is assessed as ‘None’, and the resilience is ‘Very low’, resulting in an overall sensitivity of ‘High’.

Litter

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

Evidence

Not assessed.

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT. Further detail

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

Species characterizing this habitat do not have hearing perception but vibrations may cause an effect, however, no studies exist to support an assessment.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means. Further detail

Evidence

Maerl forming species require light, which restricts them to depths shallower than 32 m in the relatively turbid waters of northern Europe (Hall-Spencer, 1998).  This suggests that maerl is intolerant of long-term reductions in light availability.  However, in the short-term maerl exhibits little stress after being kept in the dark for 4 weeks (Wilson et al., 2004). In addition, Fazakerley & Guiry (1998; cited in Blake & Maggs, 2003) noted that the growth rate of Phymatolithon calcareum approximately doubled at very shallow depth due to a 10% reduction in ambient light. This observation indicated that the growth of Phymatolithon calcareum was photo-inhibited at shallow depth, and explained its faster growth at 10 m than 5 m (Blake & Maggs, 2003).

Sensitivity assessment.  Artificial light is unlikely to affect any but the shallowest biotopes. There is a possibility that shading by artificial structures could result in the loss of live maerl in deep examples of the biotope, but only where shading was long-term or permanent. There is insufficient information regarding to assess the effect of this pressure at the benchmark on this biotope.  The sensitivity of this biotope is given as ‘No 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. Further detail

Evidence

Not relevant – this pressure is considered applicable to mobile species, e.g. fish and marine mammals rather than seabed habitats. Physical and hydrographic barriers may limit propagule dispersal.  But propagule dispersal is not considered under the pressure definition and benchmark.

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. Further detail

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. Further detail

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 that may result in changes in the genetic structure of local populations, hybridization, or change in community structure. Further detail

Evidence

This pressure is not relevant to the characterizing species within this biotope. 

Introduction or spread of invasive non-indigenous species

Benchmark. The introduction of one or more invasive non-indigenous species (INIS). Further detail

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

Although a widespread invader, Didemnum vexillum has a limited ability for natural dispersal since the pelagic larvae remain in the water column for a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localized populations (Herborg et al., 2009; Vercaemer et al., 2015; Holt, 2024). However, Bullard et al. (2007) suggested that Didemnum vexillum can form new colonies asexually by fragmentation. Colonies can produce long tendrils from an encrusting colony, which can fragment, disperse and settle, attaching to suitable hard substrata elsewhere (Bullard et al., 2007; Lambert, 2009; Stefaniak & Whitlatch, 2014). A fragmented colony can spread naturally for up to three weeks transported by ocean currents, attached to floating seaweed, seagrass or other floating biota, or as free-floating spherical colonies (Bullard et al., 2007; Lengyel et al., 2009; Stefaniak & Whitlatch, 2014; Holt, 2024). Fragments can reattach to suitable substrata within six hours of contact. Fragments have the potential to disperse around 20 km before reattachment (Lengyel et al., 2009). Valentine et al. (2007a) reported that colonies of Didemnum vexillum enlarged by 6 to 11 times by asexual budding after 15 days and enlarged 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-meditated 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; Griffith 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 that human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics and its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

The seasonal growth cycle of Didemnum vexillum 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 water 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 water (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in Sandwich tide pool (temperature range from -1 °C to 24 °C, with daily fluctuations), probably does not occur in deep offshore subtidal habitats in Georges Bank (annual temperature range from 4 °C to 15°C, and daily fluctuations are minimal) (Valentine et al., 2007a). Larval release and recruitment typically occur between 14 to 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; Mckenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where 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 the establishment of colonies. 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). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems because of its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders or artificial substrata such as a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ship 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; 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). 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).

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).

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

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

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. Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities which is enough to cause dislodgement (Reinhardt et al., 2012). 

Once established, Didemnum vexillum can expand rapidly, taking over most available hard substrata, which studies have hypothesized may alter species diversity and community composition and may decrease species abundance and biodiversity. Gittenberger (2007) stated that at Oosterschelde and the Grevelingen, The Netherlands, Didemnum sp. could cover around 95% of hard substrata locally leaving little space for recruitment and growth of other species. On Georges Bank, USA, Didemnum vexillum altered the benthic community (Lengyel et al., 2009; Tillin et al., 2020). The pebble gravel substrata on Georges Bank are 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. Benthic 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 extenuate, changing the species composition. The increase in abundance of polychaetes Nereis zonata and Harmothoe extenuate 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 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). The predator protection could explain the high abundance of infaunal invertebrates found under the mats as well as the reduced abundance of crabs and demersal fish predators in areas dominated by Didemnum vexillum compared to uncolonized areas (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 comparison 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).

Small colonies of Didemnum sp. have been reported in rhodolith (maerl) banks formed of Mesophyllum erubescens, Lithophyllum stictaeforme, Lithophyllum margaritae, Titanoderma sp. and Neogoniolithon sp. in Santa Catarina, Brazil. However, no evidence was found on the effects of Didemnum sp. on the rhodolith beds (Rocha et al., 2004). Järnegren et al. (2023) suggested that as maerl species are calcareous algae with calcium carbonate in the cells, they are vulnerable to acidification (Cornwall et al., 2022 cited in Järnegren et al., 2023). As Didemnum vexillum has a low pH on its body surface this may conflict with calcification rates and thereby growth in the algae when colonizing the surfaces. 

Grall & Hall-Spencer (2003) reported that beds of the invasive slipper limpet Crepidula fornicata grew across maerl beds in Brittany.  As a result, the maerl thalli were killed, and the bed clogged with silt, faeces and pseudo-faeces, drastically changing the associated community. The high biodiversity of maerl beds is dependent on water flow through the bed (Hall-Spencer et al., 2003) so siltation is likely to drastically reduce biodiversity and well as kill the maerl. Densities of 400 individuals per m² were found on maerl beds in the Bay of Saint-Brieuc by 1989 while a 4 km² area of maerl bed had been smothered by Crepidula by 1994 (Thouzeau, 1989; Hamon & Blanchard, 1994; in French, cited from Grall & Hall-Spencer, 2003). Bivalve fishing was also rendered impossible because the dredges became clogged with Crepidula and scallop densities were lower on the Crepidula beds than on maerl beds (Thouzeau et al., 2000, in French, cited from Grall & Hall-Spencer, 2003). 

Crepidula fornicata was reported in maerl beds in Milford Haven and increased by 2005 (Tillin et al., 2020). It was reported to reach abundances of >1000 /m² (max. 2,748 /m²) in the Milford Harbour Waterway (MHW) (Bohn et al., 2012). The increased siltation in the MHW has been attributed, in part, to Crepidula beds, and it is probably a threat to the maerl beds in Milford Haven (Tillin et al., 2020).  

Peña et al. (2014) identified eleven invasive algal species found on maerl beds in the North East Atlantic. The invasive species included Sargassum muticum, which causes habitat shading (Hall-Spencer pers. comm.).

Sensitivity assessment. The evidence above suggests that smothering and siltation by Crepidula is likely to kill the resident maerl at the surface and dramatically alter the habitat resulting in a significant loss of biodiversity. The removal of the surface layer of Crepidula fornicata is possible but only with the removal of the surface layer of maerl itself, which would be highly destructive on live beds. Therefore, resistance is assessed as ‘None’ and resilience as ‘Very low’ because the Crepidula would need to be physically removed without further damaging the maerl bed. Hence, sensitivity is assessed as ‘High’.

Didemnum vexillum requires hard substrata for successful colonization, therefore, it could colonize the shells and gravel typical of this biotope. Also, Didemnum vexillum prefers the wave-sheltered conditions typical of this biotope and can tolerate moderately strong tidal steams (Valentine et al., 2007b; Mercer et al., 2009). There is no evidence (2024) that Didemnum sp. has colonized maerl beds in the UK, and Didemnum vexillum has not yet spread as far as feared (Holt, 2024). However, it has been recorded from maerls (rhodolith beds) in Brazil (Rocha et al., 2004), so the maerl gravel is a potentially suitable habitat should Didemnum vexillum be transported into the habitat. Didemnum vexillum formed extensive mats (ca 230 km²) over consolidated mixed sediment at 40 to 65 m in the Grand Banks (Valentine et al., 2007b). At this depth, the temperature regime meant that Didemnum did not die back seasonally and smothered the sediment. If such a mat formed over a maerl bed it would probably exclude light and oxygen from the maerl, resulting in its eventual death, and hence loss of the living maerl biotope. Furthermore, Didemnum mats have been reported to bind sediment together and cover the surface, which would probably reduce water flow through the maerl bed. Unimpacted maerl grounds are more structurally complex than those affected by dredging (Kamenos et al., 2003). The interstitial space provided by maerl beds allows water to flow through the bed, and oxygenated water to penetrate at depth so that other species can colonize the bed to greater depths than most other sediments. Hence, maerl beds are highly diverse (Birkett et al., 1998a; Hall-Spencer & Moore, 2000a&b). Didemnum mats have been shown to alter the community composition and diversity subtly depending on the site (Valentine et al., 2007b; Mercer et al., 2009; Lengyel et al., 2009). However, the diversity of maerl beds may be more adversely affected if the Didemnum mats clog the maerl gravel or otherwise prevent water flow through the bed. Moderate wave action, experienced by some examples of maerl beds, might prevent mats of Didemnum from smothering the beds completely. However, the evidence suggests that colonization by Didemnum vexillum could smother the beds and kill living maerl. Therefore, resistance is assessed as 'None', albeit with 'Low' confidence due to the lack of direct evidence. Hence, resilience is assessed as 'Very low' because the Didemnum would need to be physically removed to allow recovery, and sensitivity is assessed as 'High'. 

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). Further detail

Evidence

Coralline lethal orange disease found in the Pacific and could have devastating consequences for maerl beds in Europe.  However, this disease was not known to be in Europe (Birkett et al., 1998a).  Many of the species that make up the biological community within this biotope will be susceptible to disease in the form of viruses or parasites.  However, ‘No evidence’ of the effects of diseases and pathogens on maerl beds was found.

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Maerl is mainly sold dried as a soil additive but is also used in animal feed, water filtration systems, pharmaceuticals, cosmetics and bone surgery.  Maerl beds are dredged for scallops (found in high densities compared with other scallop habitats) where extraction efficiency is very high.  This harvesting has serious detrimental effects on the diversity, species richness and abundance of maerl beds (Hall-Spencer & Moore, 2000c). 

Within Europe, there is a history of the commercial collection and sale of maerl.  Two notable sites from western Europe which from which Maerl has been collected are off the coast of Brittany, where 300,000 – 500,000 t / annum are dredged (Blunden, 1991), and off Falmouth Harbour in Cornwall where extraction is around 20,000 t / annum (Martin, 1994; Hall-Spencer, 1998).

Kamenos et al. (2003) reported that maerl grounds impacted by towed demersal fishing gears are structurally less heterogeneous than pristine, un-impacted maerl grounds, diminishing the biodiversity potential of these habitats.  Birkett et al. (1998a) noted that although maerl beds subject to extraction in the Fal estuary exhibit a diverse flora and fauna, they were less species-rich than those in Galway Bay, although direct correlation with dredging was unclear.  Grall & Glemarec (1997; cited in Birkett et al., 1998a) reported few differences in biological composition between exploited and control beds in Brittany.  Dyer & Worsfold (1998) showed differences in the communities present in exploited, previously exploited and unexploited areas of maerl bed in the Fal Estuary but it was unclear if the differences were due to extraction or the hydrography and depth of the maerl beds sampled.  In Brittany, many of the maerl beds are subject to extraction (Grall & Hall-Spencer, 2003).  For example, the clean maerl gravel of the Glenan maerl bank described in 1969, was degraded to muddy sand dominated by deposit feeders and omnivores within 30 years.  Grall & Hall-Spencer (2003) noted that the bed would be completed removed within 50-100 years at the rates reported in their study.  Hall-Spencer et al. (2010) note that maerl extraction was banned in the Fal in 2005.

The other species of commercial interest found within maerl beds are scallops, for which there are two methods of capture for these organisms.  Firstly the use of a scallop dredge the effect of which is assessed under the abrasion and disturbance pressure.  The second method of removal is diver collection.  There is no evidence to suggest that there is a symbiotic relationship between maerl and scallops.  Consequently, the removal of this species is unlikely to have a significant effect on the health of the biotope.

Sensitivity assessment.  Maerl itself has historically been targeted for commercial collection.  The removal of this characterizing species is highly destructive for this biotope and the resistance is assessed as ‘None’, and the resilience is assessed as ‘Very low’, giving a sensitivity assessment of ‘High’.  However, the practice of removing maerl for industry is now banned in places such as the Fal.

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. Further detail

Evidence

Direct, physical impacts from harvesting are assessed through the 'abrasion' and 'penetration of the seabed' pressures above.  The extraction of maerl itself, scallop dredging for scallops and/or suction dredging for other commercially exploited shellfish would also damage other members of the community. For example, the red seaweed community is likely to be damaged or removed at the surface, and interstitial bivalves damaged within or removed from the bed.  The loss of these species and other associated species would decrease species richness and negatively impact on the ecosystem function. In long-term studies of scallop dredging and subsequent recovery (Hall-Spencer & Moore 2000a, 2000b) deep burrowing species including Neopentadactyla mixta were not impacted and their abundance changed little over the four year period.  However, experimental hydraulic blade dredging removed and damaged deep-burrowing species, including small numbers of Neopentadactyla mixta (Hauton et al. 2003) and affected the maerl bed to a depth of 9 cm.  Hydraulic dredging in coarse sand and gravel may have similar effects.

Sensitivity assessment. Removal of a large percentage of the characterizing species would alter the character of the biotope. The resistance to removal is ‘Low’ due to the easy accessibility of the biotopes location and the inability of these species to evade collection. The resilience is ‘Very low’, with recovery only being able to begin when the harvesting pressure is removed altogether. This gives an overall sensitivity score of ‘High’.