The Marine Life Information Network

Global warming (extreme)

Extreme emission scenario (by the end of this century 2081-2100) benchmark of:

• A 5°C rise in SST and NBT (coastal to the shelf seas),

• A 6°C rise in surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 5°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

The distribution of seaweeds is climatically defined (Breeman, 1990). Northern boundaries are set by lethal winter temperatures or summer temperatures too low for growth and/or reproduction, whilst southern limits are set by high lethal summer temperatures or winter temperatures too high for induction of a crucial step in the life cycle (Breeman, 1990). 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 that is thought to be due to their temperature tolerances.  Lithothamnion glaciale, a key ecosystem engineer for these habitats, is found in mid- to high-latitude fjordic shallow-water environments (primarily from 55 to 80°N) (Burdett et al., 2015). In the NE Atlantic, Lithothamnion glaciale is distributed from the UK to the Russian Arctic, including Iceland and the Western Baltic. In the UK it is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man.

Temperature appears to primarily confine Lithothamnion glaciale to northern parts of the UK, possibly because reproductive conceptacles are only produced in winter when water temperatures are below 9°C (Hall-Spencer, 1994), although maerl species are known to propagate mainly by fragmentation (Wilson et al., 2004). 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).

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). 

Sensitivity assessment.  Lithothamnion glaciale is primarily found in the sea lochs of Scotland, with occasional records in England. In Loch Sween, in Scotland, temperatures range from 5 - 16°C (Kamenos et al., 2013), and lochs further north may experience even colder temperatures than this. Under the middle and high emission and extreme scenarios seawater temperatures are expected to temperatures rise by 3-5°C to potential summer sea loch temperatures of 19-21°C and winter temperatures of 8-10°C. Lithothamnion glaciale is generally at its southernmost rage in Scotland, although a few, small isolated patches occur around the rest of the UK, suggesting that an increase in temperature at levels expected for the middle emission scenario (3°C) will lead to significant negative impacts on Lithothamnion glaciale and it may be almost completely lost from Scotland as its range moves northwards. Therefore, under the middle and high emission and extreme scenarios, a resistance of ‘None’ is suggested at the benchmark level. Resilience is assessed as ‘Very low’, due to both the slow-growing nature of this species and the long-term nature of ocean warming, giving a sensitivity assessment of 'High’ under all three scenarios.

Global warming (high)

High emission scenario (by the end of this century 2081-2100) benchmark of:

• A 4°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 3°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

The distribution of seaweeds is climatically defined (Breeman, 1990). Northern boundaries are set by lethal winter temperatures or summer temperatures too low for growth and/or reproduction, whilst southern limits are set by high lethal summer temperatures or winter temperatures too high for induction of a crucial step in the life cycle (Breeman, 1990). 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 that is thought to be due to their temperature tolerances.  Lithothamnion glaciale, a key ecosystem engineer for these habitats, is found in mid- to high-latitude fjordic shallow-water environments (primarily from 55 to 80°N) (Burdett et al., 2015). In the NE Atlantic, Lithothamnion glaciale is distributed from the UK to the Russian Arctic, including Iceland and the Western Baltic. In the UK it is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man.

Temperature appears to primarily confine Lithothamnion glaciale to northern parts of the UK, possibly because reproductive conceptacles are only produced in winter when water temperatures are below 9°C (Hall-Spencer, 1994), although maerl species are known to propagate mainly by fragmentation (Wilson et al., 2004). 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).

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). 

Sensitivity assessment.  Lithothamnion glaciale is primarily found in the sea lochs of Scotland, with occasional records in England. In Loch Sween, in Scotland, temperatures range from 5 - 16°C (Kamenos et al., 2013), and lochs further north may experience even colder temperatures than this. Under the middle and high emission and extreme scenarios seawater temperatures are expected to temperatures rise by 3-5°C to potential summer sea loch temperatures of 19-21°C and winter temperatures of 8-10°C. Lithothamnion glaciale is generally at its southernmost rage in Scotland, although a few, small isolated patches occur around the rest of the UK, suggesting that an increase in temperature at levels expected for the middle emission scenario (3°C) will lead to significant negative impacts on Lithothamnion glaciale and it may be almost completely lost from Scotland as its range moves northwards. Therefore, under the middle and high emission and extreme scenarios, a resistance of ‘None’ is suggested at the benchmark level. Resilience is assessed as ‘Very low’, due to both the slow-growing nature of this species and the long-term nature of ocean warming, giving a sensitivity assessment of 'High’ under all three scenarios.

Global warming (middle)

Middle emission scenario (by the end of this century 2081-2100) benchmark of:

• A 3°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf.

• A 2°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

The distribution of seaweeds is climatically defined (Breeman, 1990). Northern boundaries are set by lethal winter temperatures or summer temperatures too low for growth and/or reproduction, whilst southern limits are set by high lethal summer temperatures or winter temperatures too high for induction of a crucial step in the life cycle (Breeman, 1990). 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 that is thought to be due to their temperature tolerances.  Lithothamnion glaciale, a key ecosystem engineer for these habitats, is found in mid- to high-latitude fjordic shallow-water environments (primarily from 55 to 80°N) (Burdett et al., 2015). In the NE Atlantic, Lithothamnion glaciale is distributed from the UK to the Russian Arctic, including Iceland and the Western Baltic. In the UK it is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man.

Temperature appears to primarily confine Lithothamnion glaciale to northern parts of the UK, possibly because reproductive conceptacles are only produced in winter when water temperatures are below 9°C (Hall-Spencer, 1994), although maerl species are known to propagate mainly by fragmentation (Wilson et al., 2004). 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).

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). 

Sensitivity assessment.  Lithothamnion glaciale is primarily found in the sea lochs of Scotland, with occasional records in England. In Loch Sween, in Scotland, temperatures range from 5 - 16°C (Kamenos et al., 2013), and lochs further north may experience even colder temperatures than this. Under the middle and high emission and extreme scenarios seawater temperatures are expected to temperatures rise by 3-5°C to potential summer sea loch temperatures of 19-21°C and winter temperatures of 8-10°C. Lithothamnion glaciale is generally at its southernmost rage in Scotland, although a few, small isolated patches occur around the rest of the UK, suggesting that an increase in temperature at levels expected for the middle emission scenario (3°C) will lead to significant negative impacts on Lithothamnion glaciale and it may be almost completely lost from Scotland as its range moves northwards. Therefore, under the middle and high emission and extreme scenarios, a resistance of ‘None’ is suggested at the benchmark level. Resilience is assessed as ‘Very low’, due to both the slow-growing nature of this species and the long-term nature of ocean warming, giving a sensitivity assessment of 'High’ under all three scenarios.

Marine heatwaves (high)

High emission scenario benchmark: A marine heatwave occurring every two years, with a mean duration of 120 days, and a maximum intensity of 3.5°C. Further detail.

Evidence

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Coralline algae have been shown to be sensitive to marine heatwaves. 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 health, rates of calcification and photosynthesis and survival. In Western Australia, anomalously high seawater temperatures in 2012-2013, which were up to 2°C above the long term average, led to significant mortality of coralline crustose algae (Short et al., 2015).

When the temperature of the co-occurring maerl species Phymatolithon calcareum was experimentally exposed to in the laboratory was increased from 9°C to 25°C for five weeks, photosynthesis was maintained (Wilson et al., 2004), suggesting this species can cope with these temperatures in the short term. In contrast, there was a 30% decrease in growth when  Lithothamnion coralloides was exposed to temperatures of 18°C compared to those exposed to 14°C (Blake & Maggs, 2003), which suggested that this species may be less tolerant to high temperatures. Shubert et al. (2019) investigated the impact of an experimental heatwave (increase of 5°C from 23 to28°C) on the photosynthesis and calcification of two species of maerl, Lithothamnion crispatum and Melyvonnea erubescens, and found that whilst photosynthesis decreased only in Melyvonnea erubescens, calcification was negatively affected in both species.

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C. Under both these scenarios live Lithothamnion glaciale is likely to be lost from most of the UK and move its southern biogeographical limit polewards as a result of rising temperatures (see Global warming). A heatwave of either of these magnitudes is likely to cause mass mortality of Lithothamnion glaciale, which is an Arctic- cold temperate species. Therefore, resistance has been assessed as ‘None’. Resilience has been assessed as ‘Very low’ due to the slow-growing nature of this species. This biotope is assessed as having ‘High’ sensitivity to marine heatwaves under both the middle and high emission scenarios.

Marine heatwaves (middle)

Middle emission scenario benchmark:  A marine heatwave occurring every three years, with a mean duration of 80 days, with a maximum intensity of 2°C. Further detail.

Evidence

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Coralline algae have been shown to be sensitive to marine heatwaves. 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 health, rates of calcification and photosynthesis and survival. In Western Australia, anomalously high seawater temperatures in 2012-2013, which were up to 2°C above the long term average, led to significant mortality of coralline crustose algae (Short et al., 2015).

When the temperature of the co-occurring maerl species Phymatolithon calcareum was experimentally exposed to in the laboratory was increased from 9°C to 25°C for five weeks, photosynthesis was maintained (Wilson et al., 2004), suggesting this species can cope with these temperatures in the short term. In contrast, there was a 30% decrease in growth when  Lithothamnion coralloides was exposed to temperatures of 18°C compared to those exposed to 14°C (Blake & Maggs, 2003), which suggested that this species may be less tolerant to high temperatures. Shubert et al. (2019) investigated the impact of an experimental heatwave (increase of 5°C from 23 to28°C) on the photosynthesis and calcification of two species of maerl, Lithothamnion crispatum and Melyvonnea erubescens, and found that whilst photosynthesis decreased only in Melyvonnea erubescens, calcification was negatively affected in both species.

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C. Under both these scenarios live Lithothamnion glaciale is likely to be lost from most of the UK and move its southern biogeographical limit polewards as a result of rising temperatures (see Global warming). A heatwave of either of these magnitudes is likely to cause mass mortality of Lithothamnion glaciale, which is an Arctic- cold temperate species. Therefore, resistance has been assessed as ‘None’. Resilience has been assessed as ‘Very low’ due to the slow-growing nature of this species. This biotope is assessed as having ‘High’ sensitivity to marine heatwaves under both the middle and high emission scenarios.

Ocean acidification (high)

High emission scenario benchmark: a further decrease in pH of 0.35 (annual mean) and corresponding 120% increase in H+ ions , seasonal aragonite saturation of 20% of UK coastal waters and North Sea bottom waters, and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, occurring at a depth of 400 m by the end of this century 2081-2100. Further detail 

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop by a further 0.35 units by the end of this century, dependent on emission scenario. Approximately 20% of coastal areas (particularly around Scotland) are expected to suffer from seasonal aragonite undersaturation by the end of this century (Ostle et al., 2016).  Coralline algae are thought to be one of the groups of species most vulnerable to ocean acidification due to the solubility of their high magnesium-calcite skeletons (Martin & Hall-Spencer, 2017).

In response to CO₂ enrichment leading to a pH decrease of 0.48 units, Lithothamnion glaciale appears to upregulate daytime calcification rates to counteract any dissolution that occurs during the night, to maintain net calcification (Kamenos et al., 2013). Ragazzola et al. (2012) found that, although Lithothamnion glaciale managed to maintain calcification under increasing levels of carbon dioxide, growth rates were significantly lower at pH levels expected for the end of this century under the high emission scenario (pH 7.7). Both of these studies found that the structural integrity of the skeletons weakened as a result of increasing carbon dioxide, with one finding that wall thickness decreased under both the middle and high emission scenario (Ragazzola et al., 2012). The weaker structural integrity of the maerl skeletons may lead to increased fragmentation and a decrease in ecosystem function and the biodiversity of maerl beds (Kamenos et al., 2013). Büdenbender et al. (2011) found that net dissolution occurred at carbon dioxide levels expected for the high emission scenario during winter conditions, but a positive calcification rate was maintained during summer.

Sensitivity assessment. Most species of rhodoliths/maerl appear to suffer negative consequences of ocean acidification (Martin & Hall-Spencer, 2017), and calcareous red algae are often conspicuously absent from CO₂ vents with extremely low pH (pH <7), and significantly reduced in areas of pH predicted for the end of this century (Hall-Spencer et al., 2008).

Under the middle emission scenario, aragonite undersaturation is not expected to occur around the coast of the UK by the end of this century, and therefore Lithothamnion glaciale is not expected to suffer dissolution. No mortality is expected under this scenario, although at a pCO₂ of 550 ppm, mean relative calcification will remain positive but is likely to decrease (Büdenbender et al., 2011). Therefore resistance to ocean acidification under the middle emission scenario has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, Lithothamnion glaciale is assessed as ‘Not sensitive’ to ocean acidification at this benchmark level.

Under the high emission scenario, 20% of coastal areas, primarily around Scotland, are expected to suffer from seasonal aragonite undersaturation.  Annual mean net dissolution of Lithothamnion glaciale was estimated to start at an aragonite saturation state between 1.1 and 0.9 (Büdenbender et al., 2011), therefore this species is likely to be significantly negatively impacted by ocean acidification at levels expected for this scenario and net dissolution is likely to occur during the winter months. Under this scenario, significant loss of Lithothamnion glaciale is expected, although some individuals may manage to upregulated their calcification rates if conditions allow. Therefore resistance has been assessed as ‘Low’, whilst resilience is assessed as ‘Very Low’ due to the long term nature of ocean acidification. Under the high emission scenario, sensitivity to ocean acidification is assessed as ‘High’.

Ocean acidification (middle)

Middle emission scenario benchmark: a further decrease in pH of 0.15 (annual mean) and corresponding 35% increase in H+ ions with no coastal aragonite undersaturation and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, at a depth of 800 m by the end of this century 2081-2100. Further detail.

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop by a further 0.35 units by the end of this century, dependent on emission scenario. Approximately 20% of coastal areas (particularly around Scotland) are expected to suffer from seasonal aragonite undersaturation by the end of this century (Ostle et al., 2016).  Coralline algae are thought to be one of the groups of species most vulnerable to ocean acidification due to the solubility of their high magnesium-calcite skeletons (Martin & Hall-Spencer, 2017).

In response to CO₂ enrichment leading to a pH decrease of 0.48 units, Lithothamnion glaciale appears to upregulate daytime calcification rates to counteract any dissolution that occurs during the night, to maintain net calcification (Kamenos et al., 2013). Ragazzola et al. (2012) found that, although Lithothamnion glaciale managed to maintain calcification under increasing levels of carbon dioxide, growth rates were significantly lower at pH levels expected for the end of this century under the high emission scenario (pH 7.7). Both of these studies found that the structural integrity of the skeletons weakened as a result of increasing carbon dioxide, with one finding that wall thickness decreased under both the middle and high emission scenario (Ragazzola et al., 2012). The weaker structural integrity of the maerl skeletons may lead to increased fragmentation and a decrease in ecosystem function and the biodiversity of maerl beds (Kamenos et al., 2013). Büdenbender et al. (2011) found that net dissolution occurred at carbon dioxide levels expected for the high emission scenario during winter conditions, but a positive calcification rate was maintained during summer.

Sensitivity assessment. Most species of rhodoliths/maerl appear to suffer negative consequences of ocean acidification (Martin & Hall-Spencer, 2017), and calcareous red algae are often conspicuously absent from CO₂ vents with extremely low pH (pH <7), and significantly reduced in areas of pH predicted for the end of this century (Hall-Spencer et al., 2008).

Under the middle emission scenario, aragonite undersaturation is not expected to occur around the coast of the UK by the end of this century, and therefore Lithothamnion glaciale is not expected to suffer dissolution. No mortality is expected under this scenario, although at a pCO₂ of 550 ppm, mean relative calcification will remain positive but is likely to decrease (Büdenbender et al., 2011). Therefore resistance to ocean acidification under the middle emission scenario has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, Lithothamnion glaciale is assessed as ‘Not sensitive’ to ocean acidification at this benchmark level.

Under the high emission scenario, 20% of coastal areas, primarily around Scotland, are expected to suffer from seasonal aragonite undersaturation.  Annual mean net dissolution of Lithothamnion glaciale was estimated to start at an aragonite saturation state between 1.1 and 0.9 (Büdenbender et al., 2011), therefore this species is likely to be significantly negatively impacted by ocean acidification at levels expected for this scenario and net dissolution is likely to occur during the winter months. Under this scenario, significant loss of Lithothamnion glaciale is expected, although some individuals may manage to upregulated their calcification rates if conditions allow. Therefore resistance has been assessed as ‘Low’, whilst resilience is assessed as ‘Very Low’ due to the long term nature of ocean acidification. Under the high emission scenario, sensitivity to ocean acidification is assessed as ‘High’.

Sea level rise (extreme)

Extreme scenario benchmark: a 107 cm rise in average UK by the end of this century (2018-2100). Further detail.

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). In the North East Atlantic (NEA), maerl beds occur in ocean-facing coastal waters <20-30 m deep, that are tideswept and have little terrigenous sediment supply (Bosence & Wilson, 2003). This biotope occurs at 0 –10 m depth and is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man (JNCC, 2015).

Understanding how sea-level rise will affect exposure or tidal energy, is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Mossman et al., 2015; Lowe et al., 2018; Palmer et al., 2018).

Sensitivity assessment. It is difficult to assess the effect of sea-level rise scenario on exposure or tidal energy, as evidence predicts that any changes will be site-specific.  Growth rates of maerl species reduce with depth. For example, in Phymatolithon calcareum when depth is increased from 5 m to 10 m, growth decreases by a third (Blake & Maggs, 2003). As this biotope (Mrl.Lgla) can occur from 0-20 m in depth, it is assumed that a sea-level rise of 50 cm or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme scenario) might result in loss of part of the deeper extent of the biotope in some sites, with little likelihood of landward migration. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’.  But resistance is possibly ‘Medium’ under the extreme scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

Sea level rise (high)

High emission scenario benchmark: a 70 cm rise in average UK by the end of this century (2018-2100). Further detail.

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). In the North East Atlantic (NEA), maerl beds occur in ocean-facing coastal waters <20-30 m deep, that are tideswept and have little terrigenous sediment supply (Bosence & Wilson, 2003). This biotope occurs at 0 –10 m depth and is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man (JNCC, 2015).

Understanding how sea-level rise will affect exposure or tidal energy, is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Mossman et al., 2015; Lowe et al., 2018; Palmer et al., 2018).

Sensitivity assessment. It is difficult to assess the effect of sea-level rise scenario on exposure or tidal energy, as evidence predicts that any changes will be site-specific.  Growth rates of maerl species reduce with depth. For example, in Phymatolithon calcareum when depth is increased from 5 m to 10 m, growth decreases by a third (Blake & Maggs, 2003). As this biotope (Mrl.Lgla) can occur from 0-20 m in depth, it is assumed that a sea-level rise of 50 cm or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme scenario) might result in loss of part of the deeper extent of the biotope in some sites, with little likelihood of landward migration. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’.  But resistance is possibly ‘Medium’ under the extreme scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

Sea level rise (middle)

Middle emission scenario benchmark: a 50 cm rise in average UK sea-level rise by the end of this century (2081-2100). Further detail.

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). In the North East Atlantic (NEA), maerl beds occur in ocean-facing coastal waters <20-30 m deep, that are tideswept and have little terrigenous sediment supply (Bosence & Wilson, 2003). This biotope occurs at 0 –10 m depth and is most common in the sea lochs of Scotland, with occasional records along the south coast, Lundy and the Isle of Man (JNCC, 2015).

Understanding how sea-level rise will affect exposure or tidal energy, is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Mossman et al., 2015; Lowe et al., 2018; Palmer et al., 2018).

Sensitivity assessment. It is difficult to assess the effect of sea-level rise scenario on exposure or tidal energy, as evidence predicts that any changes will be site-specific.  Growth rates of maerl species reduce with depth. For example, in Phymatolithon calcareum when depth is increased from 5 m to 10 m, growth decreases by a third (Blake & Maggs, 2003). As this biotope (Mrl.Lgla) can occur from 0-20 m in depth, it is assumed that a sea-level rise of 50 cm or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme scenario) might result in loss of part of the deeper extent of the biotope in some sites, with little likelihood of landward migration. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’.  But resistance is possibly ‘Medium’ under the extreme scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

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). 

Sensitivity assessment.  An increase in temperature at the benchmark level is unlikely to affect Phymatolithon calcareum present in this biotope (Wilson et al., 2004).  However, an increase in temperature may be detrimental to Lithothamnion glaciale and Lithothamnion erinaceum especially at the southernmost extent of their range.  An increase in temperature of 2°C for a year is likely to reduce growth and reproduction in these species. Long-term climate-induced change in temperatures may result in loss of the extent of SS.SMp.Mrl.Lgla.  Further research is required to determine the effect of a 2-5°C change in temperature on Lithothamnion glaciale.  The northern species of maerl (Lithothamnion glaciale and Lithothamnion erinaceum) are likely to be vulnerable to sea surface warming in Scotland (Hall-Spencer, pers comm.). Therefore, a resistance of ‘Medium’ is suggested at the benchmark level but with 'Low' confidence. Resilience is, therefore, ‘Very low’, and a sensitivity assessment of 'Medium’ is recorded.

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).

Sensitivity assessment.  A decrease in temperature at the benchmark level is unlikely to affect Phymatolithon calcareum (Wilson et al., 2004) when present in this biotope.  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, SS.SMp.Mrl.Lgla probably has 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).

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. Where the biotope was found in areas of reduced or variable salinity, an increase in salinity may result in an increase in biodiversity and a shift in the community to one more representative of full salinity. 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.

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.  

Sensitivity assessment.  Where present, reduced salinity for an extended period of time would stress Phymatolithon calcareum and could lead to mortality.  SS.SMp.Mrl.Lgla is characterized by Lithothamnion glaciale and can be found in both fully marine and variable salinity regimes.  A decrease in salinity could result in a 'reduced' salinity regime.  Experiments show that Lithothamnion glaciale does not recover from a relatively short-term exposure to 15 psu. This suggests that at the level of the benchmark prolonged exposure (for one year) to this pressure could cause significant mortality.  Resistance has been assessed as ‘Low’ and, as resilience is probably ‘Very low’, sensitivity is assessed as ‘High’. There is no information on the responses of other maerl species to a reduction in salinity.

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. 

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.

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.).

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 is assessed as ‘None’ and resilience assessed as ‘Very low’ so that sensitivity is assessed as ‘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’, and sensitivity as ‘Not Sensitive’.

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. 

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

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 resulting in easier displacement by wave action, resulting in a reduced structural heterogeneity and lower 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 in 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; 2000 c) 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 good 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 abrasion. 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’, giving the biotope a ‘High’ sensitivity.

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 them 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 the bottom depth limit in Europe (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 suspended matter is likely to reduce the quantity of food available for filter feeders.  This could lead to a minor change species present within the community but is unlikely to cause a major shift in the biotope.

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 5cm 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).

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 biotope's resistance is assessed as ‘None’, and the resilience as ‘Very low’ so that sensitivity is assessed as ‘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 assessed as ‘None’, the resilience as ‘Very low’, and sensitivity is assessed as ‘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 32m in the relatively turbid waters of northern Europe (Hall-Spencer, 1998).  Suggesting 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). 

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 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

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’.

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.

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’.