|Researched by||Emilia d'Avack, Dr Harvey Tyler-Walters & Catherine Wilding||Refereed by||Dr Leigh Jones|
Mid and upper shore wave-sheltered muddy fine sand or sandy mud with narrow-leafed eelgrass Zostera noltei at an abundance of frequent or above. This is similar to polychaetes and Cerastoderma edule (LMS.Pcer) since it is most frequently found on lower estuary and sheltered coastal muddy sands with a similar infauna. Exactly what determines the distribution of the Zostera noltei is, however, not entirely clear. Zostera noltei is often found in small lagoons and pools, remaining permanently submerged, and on sediment shores where the muddiness of the sediment retains water and stops the roots from drying out. A black layer is usually present below 5 cm sediment depth. The infaunal community is characterized by polychaetes Pygospio elegans and Arenicola marina, mud amphipods Corophium volutator and bivalves Cerastoderma edule, Limecola balthica and Scrobicularia plana. Typically an epifaunal community is found that includes the mud snail Hydrobia ulvae, shore crabs Carcinus maenas and the green alga Ulva sp. This biotope should not be confused with IMS.Zmar which is a Zostera marina bed on the lower shore or shallow sublittoral clean or muddy sand. (Information taken from the Marine Biotope Classification for Britain and Ireland, Version 97.06: Connor et al., 1997a, b).
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Zostera beds are naturally dynamic and may show marked seasonal changes. Leaves are shed in winter, although Zostera noltei retains its leaves longer than Zostera marina. Leaf growth stops in September/October (Brown, 1990). Leaves are lost, or removed by grazing or wave action over-winter. For example, in the Wadden Sea, Nacken & Reise (2000) noted that 50% of leaves fell off, while Brent geese removed 63% of the plant biomass.
Zostera noltei over-winters as rhizome and shoot fragments, resulting in 'recruitment' of several cohorts in the following spring (Marta et al., 1996). However, Nacken & Reise (2000) noted that the Zostera noltei beds recovered normal shoot density and that grazing wildfowl helped to maintain a balance between accretion and erosion within the bed, without which recovery was inhibited. The rhizome of Zostera noltei has limited carbohydrate storage capability, Marta et al. (1996) and Dawes & Guiry (1992) regarded this species as ephemeral, taking advantage of seasonal increases in nutrients and light especially to grow rapidly in spring and early summer.
Where present, Arenicola marina spawns synchronously either once or twice a year; the precise timing depending on location (Howie, 1959; Clay, 1967; Bentley & Pacey, 1992). Cerastoderma edule spawns between March - August with a peak in summer, Limecola balthica spawns in February - March with another peak in autumn, whilst Scrobicularia plana spawns in summer (Fish & Fish, 1996).
Settlement of spat in intertidal bivalves is generally sporadic (see Cerastoderma edule for review). While Limecola balthica may be protected from low winter temperatures by its depth in the sediment, Cerastoderma edule is vulnerable to low temperatures in winter, especially in severe winters. Therefore, cockle mortality is likely over winter due to low temperatures, lack of food and predation, especially from wildfowl such as the oystercatcher (Haematopus ostralegus). Further mortality is likely in year one cockles due to exhausted energy reserves and predation by the shore crab Carcinus maenas. Epifaunal species, such as Littorina littorea and Hydrobia ulvae may suffer additional wildfowl predation over winter without the refuge provided by Zostera noltei leaves, however, being mobile they are able to seek alternative food sources.
Leaves slow current and water flow rates under the canopy, which encourages settlement of fine sediments, detritus and larvae (Orth, 1992). Seagrass rhizomes stabilize sediment and protect against wave disturbance. Presence of seagrass increases species diversity by favouring sedentary species that require stable substrata (Orth, 1992; Davison & Hughes, 1998).
Zostera noltei provides shelter or substratum for a wide range of species, especially epiphytes and periphyton. Epiphytic species may be grazed by other species (Davison & Hughes, 1998) such as the mobile epifauna, Hydrobia ulvae and Littorina littoreapresent in seagrass beds. The sediment supports a rich infauna of polychaetes, bivalve molluscs and the mud amphipod Corophium volutator. Cockle beds (Cerastoderma edule) are often associated with intertidal seagrass beds. The sediment also includes a diverse meiofauna, for example many species of free-living turbellarians, ostracods and copepods (Asmus & Asmus, 2000b). In addition, intertidal seagrass beds are visited by several fish species when immersed.
Seagrass beds are characterized by high productivity and biodiversity and are considered to be of great ecological and economic importance (Davison & Hughes, 1998; Asmus & Asmus, 2000b). Primary production is derived from phytoplankton, microphytobenthos and Zostera sp. In addition, organic carbon is derived from the input of detritus into the system (for estimates of g C/m²/year see Asmus & Asmus, 2000b). Asmus & Asmus (2000b) reported that seagrass beds are sediment traps and nutrient sinks, which under storm conditions may become nutrient sources for the surrounding ecosystems, and are, therefore, important for the material flux in the ecosystem. For example, in the Sylt-Rømø Bight, Asmus & Asmus (2000b) estimated that the seagrass beds contributed significantly to material flux within the total intertidal system even though the seagrass beds only covered 12% of the intertidal area.
In periods when Zostera noltei dies off (winter), epiphytic algae and periphyton contribute significantly to the overall community productivity and above ground biomass (Welsh et al., 2000; Philippart, 1995b). Philippart (1995b) estimated that by May on an intertidal mudflat off Terschelling, the Netherlands, periphyton biomass equalled Zostera noltei biomass, declining to 20% of the total above ground biomass by the end of September.
Detritus food chains within the seagrass beds are driven by bacterial decomposition of dead seagrass tissue and other detritus. Dissolved organic matter (DOM) leaching from seagrass and bacterial decomposition supports high numbers of heterotrophic protists. Seagrass detritus is rich in micro-organisms, e.g. 1 g (dry weight) may support on average 9 mg of bacteria and protists, including heterotrophic flagellates and ciliates (Davison & Hughes, 1998). Dead seagrass leaves can be transported by currents to great depths or washed up on the shore; hence supporting detritus based food chains and communities in distant areas of the coast (Davison & Hughes, 1998).
Although primary production is high, secondary production is similar in un-vegetated areas and seagrass beds (Asmus & Asmus, 2000b). Asmus & Asmus (2000b) presented a general food web for intertidal Zostera spp. beds, noting that loss of intertidal seagrass beds resulted in profound changes in the food web of the total ecosystem.
Zostera sp. are perennials but may be annuals under stressful conditions. Seedlings rarely occur in seagrass beds except in areas cleared by storms, blow-out or excessive herbivory (Phillips & Menez, 1988). Seed mortality is very high (Phillips & Menez, 1988; Fishman & Orth, 1996). Seed may be dispersed through the gut of wildfowl (Fishman & Orth, 1996), or float long distances (up to 200 m in some cases) on attached gas bubbles. The generative stalk may be released and can also float long distances. Displaced pieces of shoot or rhizome float and may root if they settle on suitable substrata. Vegetative reproduction probably exceeds seedling recruitment except in areas of sediment disturbance (Phillips & Menez, 1988; Reusch et al., 1998).
Seagrasses are capable of both sexual and asexual reproduction, and therefore beds can be established by either clonal growth or flowering events and seeds (Alexandre et al., 2006). Hootsmans et al. (1987) noted that potential recruitment was maximal (32% of seeds) at 30 °C and 10 psu, no recruitment occurred at 30 psu and they estimated that, in 1983 <5% of Zostera noltei plants in their study area originated from seed. Recruitment of Zostera noltei is, therefore, most likely to occur by dispersal of shoots and rhizomes rather than seed.
Potential recruitment may be hampered by competition with infauna such as the ragworm Hediste diversicolor or blow lug Arenicola marina (Philippart, 1994a; Hughes et al., 2000). Hughes et al. (2000) noted that Hediste diversicolor consumed leaves and seeds of Zostera noltei by pulling them into their burrow, therefore reducing the survival of seedlings.
The distribution of Zostera noltei can be restricted by burrowing and bioturbation of infauna such as Hediste diversicolor and Arenicola marina. Philippart (1994a) concluded that the blow lug populations in the Wadden Sea may have contributed to the decline in the Zostera noltei beds over the previous 25 years. The rhizome mat of the seagrass can inhibit burrowing and colonization of the seagrass bed by burrowing infauna (Hughes et al., 2000; Philippart, 1994a). At low densities, blow lug may be beneficial as they increase nutrient flux and oxygenation in the sediment. Corophium volutator has been reported to inhibit colonization of mud by Salicornia sp. (Hughes et al., 2000) and where present may also inhibit Zostera noltei recruitment.
While Zostera noltei has been reported to recover from seasonal grazing and loss of leaf biomass over-winter (Nacken & Reise, 2000), beds in the Wadden Sea have continued to decline (Philippart, 1994b). The scarce distribution of Zostera noltei in the UK suggests that recovery or recruitment is slow and in the region of 5-10 years (Holt et al., 1997; Davison & Hughes, 1998).
Epifaunal species such as Hydrobia ulvae are widely distributed, mobile, occur at high densities, and have a planktonic life cycle suggesting that they would recruit rapidly. Similarly Littorina littorea is likely to recruit rapidly.
Development of both Arenicola marina and Pygospio elegans starts in the female's tube. Larvae of Pygospio elegans are pelagic, while Arenicola marina larvae migrate up the shore. Recruitment in Arenicola marina is rapid, especially where there are adjacent populations present.
Recruitment in infaunal bivalve populations is sporadic due to variation in larval supply and post-settlement mortality. For instance, although recruitment in Cerastoderma edule is likely to occur annually, significant recruitment to the population may take up to five years.
Zostera noltei is able to recover relatively quickly compared to other seagrass species (Marbà et al., 2004). Nacken & Reise (2000) noted that Zostera noltei beds had returned to the previous abundance within a year following leaf loss and grazing by wildfowl. The majority of species associated with intertidal seagrass beds are not restricted to the biotope (Asmus & Asmus, 2000b), with the exception of Zostera sp. Specific epiphytes, and are likely to be present in the sediment or migrate into the developing bed. Zostera noltei is regarded as a relatively ephemeral species (Dawes & Guiry, 1992).
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|Depth Range||Upper shore, Mid shore|
|Water clarity preferences|
|Limiting Nutrients||Nitrogen (nitrates), Phosphorus (phosphates)|
|Salinity preferences||Full (30-40 psu), Variable (18-40 psu)|
|Physiographic preferences||Enclosed coast / Embayment|
|Biological zone preferences||Eulittoral|
|Substratum/habitat preferences||Muddy sand, Sandy mud|
|Tidal strength preferences|
|Wave exposure preferences||Extremely sheltered, Sheltered, Very sheltered|
|Other preferences||No text entered|
Seagrass requires a particular light regime to net photosynthesize and grow. The intertidal is likely to be more turbid than the shallow subtidal occupied by Zostera marina due to runoff and re-suspension of sediment by wave and tidal action. Turbidity decreases light penetration and reduces the time available for net photosynthesis. However, intertidal Zostera noltii 'escapes' this turbidity since it is able to take advantage of the high light intensities at low tide (Vermaat et al., 1996).Seagrass beds act as sinks for nutrients (Asmus & Asmus, 2000b) and as such, nitrogen may not be limiting in sparse intertidal seagrass beds. In sandy sediments phosphate may be limiting where it is adsorbed onto particles (Short, 1987; Jones et al., 2000).
Zostera noltei is the main species creating this habitat as removing Zostera marina plants would result in the disappearance of this biotope. Although a wide range of species are associated with seagrass beds which provide habitat and food resources, these species occur in a range of other biotopes and were therefore not considered to characterize the sensitivity of this biotope (d'Avack et al., 2014). Zostera noltei is not dependent on associated species to create or modify habitat, provide food or other resources. The sensitivity assessments are thus based on Zostera noltei alone and do not consider the sensitivity of associated species that may live in or around seagrass beds. Effects on other components of the community will however be reported where relevant.
Zostera noltei is the smallest of British seagrasses. The species occurs on sedimentary substrata, in sheltered or extremely sheltered locations with low current velocity. It is predominantly found in the intertidal region but can also be found subtidally. However, where water cover is permanent, Zostera noltei is often out competed by Zostera marina (Borum et al., 2004).
d’Avack et al. (2014) reviewed the recovery and resilience rates of seagrass habitats. The report determined that although seagrass species are fast-growing and relatively short-lived, they can take a considerable time to recover from damaging events if recovery does occur at all. Every seagrass population will have a different response to pressures depending on the magnitude or duration of exposure pressure as well as the nature of the receiving environment. In general terms, the resilience of seagrass biotopes to external pressures is low, as shown by the very slow or lack of recovery after the epidemic of the wasting disease in the 1930s. Seagrass recolonization of a disturbed area can occur through sexual (seed supply) and asexual (vegetative growth from adjacent rhizomes. However, Boese et al. (2009) found that natural seedling production was not of significance in the recovery of seagrass beds but that recovery was due exclusively to rhizome growth from adjacent perennial beds. Zostra plants are monomorphic, restricted to the horizontal growth of roots and, hence, unable to grow rhizomes vertically. This restriction to horizontal elongation of the roots makes the recolonisation of adjacent bare patches difficult and explains why large beds are only found in gently sloping locations. A depression of the seabed caused by disturbance of the sediment can thus restrict the expansion of the bed. The size and shape of impacted areas will also have a considerable effect on resilience rates (Creed et al., 1999). Larger denuded areas are likely to take longer to recover than smaller scars, for example, seagrass beds are likely to be more resilient to physical damage resulting from narrow furrows left after anchoring because of large edge to area ratio and related availability of plants for recolonization.
Genetic diversity also influences the resilience of seagrasses in particular when pressure persists over a long period of time. The genetic diversity of Zostera population is very high, particularly in the NE Atlantic (Olsen et al., 2004). Rice & Emery (2003) showed that evolutionary change in seagrasses can occur within a few generations, suggesting that genetically diverse population would be more resilient to changes in environmental conditions compared to genetically conserved populations. Pressures causing a rapid change in seagrass environments will have a greater impact as the natural ability of the plants to adapt is compromised. Plasticity is a further a key element in determining the resilience of seagrass biotopes. Maxwell et al. (2014) investigated the response of seagrass ecosystems to severe weather events (i.e. flooding) in order to understand the process that promotes acclimation. The study found that phenotypic plasticity (changes in physiological and morphological characteristics) enabled the species to cope with varying degree of stress to avoid mortality. Phenotypic plasticity can thus increase the length of time seagrass can persist in unfavourable environments such as reduced light availability. Different populations will thus have different resilience to external pressures.
Resilience assessment. The resilience of seagrass beds and the ability to recover from human induced pressures is a combination of the environmental conditions of the site, growth rates of the seagrass, the frequency (repeated disturbances versus a one-off event) and the intensity of the disturbance. This highlights the importance of considering the species affected as well as the ecology of the seagrass bed, the environmental conditions and the types and nature of activities giving rise to the pressure. Changes in biological communities after seagrass disappear might impact seagrass resilience. A rise in the abundance of sea urchin, for instance, could prevent the recovery of seagrass beds due to increased herbivory (Valentine & Heck Jr, 1991). The removal of seagrass plants can induce a negative feedback loop inhibiting recovery. Indeed the removal of plants can cause chronic turbidity due to continual resuspension of unconsolidated sediments. When water quality conditions do not return to their original state, recovery of seagrass beds may not occur at all (Giesen et al., 1990). Seagrass species comprise an important winter food for wildfowl. Tubbs &Tubbs (1983) reported that wildfowl were responsible for a reduction of 60 to 100% in Zostera noltei biomass from mid-October to mid-January. The removal of plants by wildfowl is part of the natural seasonal fluctuation in seagrass cover. Similarly, Nacken & Reise (2000) found that in intertidal Zostera noltei beds biomass was reduced by 63% due to wildfowl feeding. Beds, however, recovered by the following year and the authors suggested that this disturbance was necessary for the persistence of intertidal populations. Added anthropogenic disturbance may, however, be detrimental to seagrass beds as soon as the ‘normal’ level caused by grazing birds is exceeded by human activities.
It should be noted that the recovery rates are only indicative of the recovery potential. Recovery of impacted populations will always be mediated by stochastic events and processes acting over different scales including, but not limited to, local habitat conditions, further impacts and processes such as larval-supply and recruitment between populations.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
Temperature is considered the overall parameter controlling the geographical distribution of seagrasses. All enzymatic processes, related to plant metabolism are temperature dependent and specific life cycle events, such as flowering and germination, are also often related to temperature (Phillips et al., 1983). For seagrasses, temperature affects biological processes by increasing reaction rates of biological pathways. Photosynthesis and respiration increase with higher temperature until a point where enzymes associated with these processes are inhibited. Beyond a certain threshold, high temperatures will result in respiration being greater than photosynthesis resulting in a negative energy balance. Increased temperatures do also encourage the growth of epiphytes increasing the burden upon seagrass beds and making them more susceptible to disease (Rasmussen, 1977). Massa et al. (2009) investigated the thermal tolerance limits of Zostera noltei in a coastal lagoon system in Portugal. The study recorded that plant survival at 35°C and 37°C was 95% and 90% respectively. However, at 39°C and above the rate of shoot mortality was close to 100% (Massa et al., 2009).
Other species associated with seagrass habitats are also affected by changes in temperature. For instance, the gastropod Lacuna vincta, an important grazer found in seagrass beds, is near its southern range limit in the British Isles. Long-term increases in temperature due to human activity may limit the survival of the snail and restrict subsequent distribution whilst a short-term acute temperature increase may cause death. The loss of grazers could have detrimental effects on seagrass beds as the leaves provide a substratum for the growth of many species of epiphytic algae. These epiphytes may smother the Zostera plants unless kept in check by the grazing activities of gastropods and other invertebrates. Healthy populations of epiphyte grazers are therefore essential to the maintenance of seagrass beds.
Sensitivity assessment. A 5°C change in temperature over one month or a 2°C change in temperature over the period of a year is unlikely to cause direct mortalities as Zostera noltei is well within its thermal tolerance limits in the British Isles. Resistance is, therefore, considered ‘High’. Recovery will be rapid once conditions return to normal resulting in a ‘High’ resilience score. The biotope is, therefore, considered to be 'Not sensitive' to a change in temperature at the pressure benchmark. However, in areas where seagrasses are already exposed to high temperatures, a change at the level of the benchmark may result in considerable losses.
Populations of Zostera noltei occur from the Mediterranean to southern Norway and Zostera sp. are regarded as tolerant of temperatures between about 5 -30°C. Therefore, they may tolerate the range of temperatures likely in the British Isles (Davison & Hughes 1998). Intertidal populations may be damaged by frost (Den Hartog, 1987) and Covey & Hocking (1987) reported defoliation of Zostera noltei in the upper reaches of mudflats in Helford River due to ice formation in the exceptionally cold winter of 1987. However, the rhizomes survived and leaves are usually lost at this time of year due to shedding, storms or grazing with little apparent effect (Nacken & Reise, 2000). Populations at the edge of the range are likely to be more intolerant of temperature change. Therefore, the biotope probably has a 'High' resistance and 'High' resilience to this pressure and is 'Not sensitive' at the benchmark level.
In general, seagrass species have a wide salinity tolerance. Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt. Zostera noltei is a euryhaline species found in the intertidal and more tolerant to extreme salinities than Zostera marina. Hypersaline conditions can affect the performance of angiosperms as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007). For instance, a study by Vermaat et al. (2000) observed considerable mortalities of Zostera noltei plants at 35 ppt (25% survival for one test population and 60% for a second test population). Similarly, Fernández-Torquemada et al. (2006) found that both growth and survival of Zostera noltei were significantly affected by high levels of salinity (>41 psu). These findings suggest that Zostera noltei is ill-equipped to withstand extreme saline conditions. d’Avack et al. (2014) reported that phenotypic plasticity can play an important role in the ability of seagrasses to withstand external pressures such as changes in salinity. Changes in physiological and morphological characteristics of seagrass plants will enable species to cope with varying degrees of stress for an extended period of time (Maxwell et al., 2014).
Sensitivity assessment. Even though Zostera noltei displays a wide tolerance to a range of salinities, an increase from 35 to above 40 units for the period of one year will cause some mortality of plants. This suggests that Zostera noltei will be adversely affected by activities such as brine discharges from seawater desalination plant. Resistance is thus assessed as ‘Medium’. Recovery by recolonization from surrounding communities will be fairly rapid once conditions return to normal scoring ‘Medium’ resilience. The biotope is therefore considered to have a ‘Medium’ sensitivity to an increase in salinity at the pressure benchmark.
In general, seagrass species have a wide salinity tolerance. Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt, while Den Hartog (1970) reported tolerance to salinity as low as 5 ppt. Zostera noltei is a euryhaline species found in the intertidal and more tolerant to extremes salinities than Zostera marina. Hyposaline conditions can, however, affect the performance of angiosperms as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007). A study by Charpentier et al. (2005) investigated the consequences of a sudden decrease (from 15 to < 5) in water salinity on Zostera noltei over an extended period. The study found that Zostera noltei plants remained dominant for a period of 3 years after the initial drop in water salinity. The subsequent decline of seagrass beds in the area was not directly associated with low salinity but may have been the result of synergetic effects of sediment trapping and suspended particles brought along with decreased saline conditions. Once salinity levels returned to normal, Zostera noltei was able to rapidly recolonize from the shallow borders. Full recovery of Zostera noltei occurred within 10 years of the initial drop in salinity. d’Avack et al. (2014) reported that phenotypic plasticity can play an important role in the ability of seagrasses to withstand external pressures such as changes in salinity. Changes in physiological and morphological characteristics of seagrass plants will enable species to cope with varying degrees of stress for an extended period of time (Maxwell et al., 2014).
Most of the other intertidal species (e.g. Hydrobia ulvae and Littorina littorea) present in this biotope can also tolerate a wide range of salinities. Cardoso (2008) however found that Hydrobia ulvae populations can be negatively impacted by changes in salinity observed during severe flooding. Similarly, both Cerastoderma edule and Arenicola marina have also been reported to be susceptible to a drop in salinities after heavy rains, especially at low tide.
Sensitivity assessment. Zostera noltei is more tolerant to changes in salinity than Zostera marina and a drop in salinity at the level of the benchmark is unlikely to result in mortality. Resistance is thus assessed as ‘High’. Recovery will be fairly rapid once conditions return to normal resulting in a ‘High’ resilience. The biotope is therefore considered to be 'Not sensitive' to a decrease in salinity at the pressure benchmark.
Human activities in coastal waters which alter hydrology have been implicated in the disappearance of seagrass beds. For instance, Van der Heide et al. (2007) noted that the construction of a dam in the Wadden Sea influencing the hydrological regime inhibited the recovery of Zostera plants after their initial decline following the wasting disease in the 1930s. The complex interactions existing between seagrass beds and water flow have been reviewed by d’Avack et al. (2014). Water flow determines the upper distribution of plants on the shore whilst plants mitigate the velocity of the flow by extracting momentum from the moving water. Reducing the flow increases water transparency and causes the deposition and retention of fine sediments. Increased flow rate, on the other hand, is likely to erode sediments, expose rhizomes and lead to loss of plants.
The highest current velocity a seagrass can withstand is determined by a threshold beyond which sediment resuspension and erosion rates are greater than the seagrasses ability to bind sediment and attenuate currents. In very strong currents, leaves might lie flat on the seabed reducing erosion under the leaves but not on the unvegetated edges which begin to erode. High velocity currents can thus change the configuration of patches within a meadow, creating striations and mounding in the seagrass beds. Such turreted profiles destabilise the bed and increase the risk of 'blow outs' (Jackson et al., 2013). Populations found in stronger currents are usually smaller, patchy and more vulnerable to storm damage.
A review by Koch (2001) determined that the range of current velocities tolerated by seagrass lies approximately between a minimum of 5 cm/s and a maximum of 180 cm/s. No exact numerical estimates were found for Zostera noltei. Recovery will depend on the species capacities to adapt to changes in water flow regime. A laboratory study by Peralta et al. (2006) on Zostera noltei demonstrated that plants are able to acclimate to hydrodynamic stresses by changing their architecture. When exposed to a water flow of 35 cm/s for four weeks, Zostera noltei plants had an improved anchoring system and changed leave morphology. The above ground/below ground biomass ratio was thereby reduced and the cross sections of leaves and rhizomes increased leading to a reduced risk of shoot breakage.
Sensitivity assessment: Any changes in hydrology will have a considerable impact on the integrity of seagrass habitat. A change in water flow at the level of the benchmark of 10 to 20 cm/s for more than 1 year would cause some mortality in seagrasses resulting in a ‘Medium’ resistance score. Recovery will depend on the species capacities to adapt to changes in water flow regime but is considered to be fairly rapid. Resilience is thus assessed as ‘Medium’. The biotope scores a ‘Medium’ sensitivity to changes in water flow at the pressure benchmark.
Seagrasses are generally not tolerant to exposure to aerial conditions, suggesting that the shallowest distribution should be at a depth below mean low water (MLW) (Koch, 2001). Zostera noltei grows predominantly in the intertidal zone and demonstrates a higher resistance to desiccation than Zostera marina which occurs more frequently in the subtidal. To understand the differences in desiccation tolerance between Zostera species, Leuschner et al. (1998) investigated the photosynthetic activity of emerged plants. The study found that after 5 hours of exposure to air during low tide, leaves of Zostera noltei had lost up to 50% of their water content. Decreasing leaf water content resulted in a reversible reduction in light-saturated net photosynthesis rate of the plant. The experiment further showed that photosynthesis was more sensitive to desiccation in Z. marina plants than in Zostera noltei under a given leaf water content. The experiment confirmed that Zostera marina is most susceptible to local changes in emergence regimes by being less tolerant to desiccation pressure. A review by d’Avack et al. (2014) reported that the limited tolerance of seagrass species to aerial exposure means that a decrease in relative sea level could force seagrass to grow deeper to reduce exposure to air. As the depth limit of seagrasses is set by light penetration, this change is likely to reduce the extent of suitable habitat. However Cabaco et al. (2009) found that Zostera noltei displayed considerable plasticity at a physiological-, plant- and population-level along the intertidal zone, indicating the ability of the species to acclimate to the steep environmental gradient of this particular ecosystem. This plasticity will allow plants to cope with changes in emergence regime.
A decrease in emergence, on the other hand, could enable the biotope to expand further up the shore. A potential expansion is, however, dependent on available habitat and will be impossible where barriers such as dams and seawalls are present resulting in the net loss of plants.
Sensitivity assessment. Zostera noltei has a limited tolerance towards aerial exposure; resistance is thus assessed as ‘Medium’. Recovery will be enabled by recolonization from surrounding communities located further down the shore and via the remaining seed bank. Recovery is therefore considered to be fairly rapid resulting in a ‘Medium’ resilience. The biotope is considered to have a ‘Medium’ sensitivity to changes in emergence regime at the pressure benchmark.
An absolute wave exposure limit and maximum wave height for Zostera has not been established (Short et al., 2002) but an increase in wave action can harm plants in several ways. Seagrasses are not robust. Strong waves can cause mechanical damage to leaves and to the rest of plant. By losing above ground biomass due to increased wave action, the productivity of seagrass plants is limited. Small and patchy populations, as well as seedlings, will be particularly vulnerable to wave exposure as they lack extensive rhizome systems to effectively anchor the plant to the seabed. Wave action also continuously mobilises sediments in coastal areas causing sediment resuspension which in turn leads to a reduction in water transparency (Koch, 2001) (see 'changes in suspended sediments' pressure). Photosynthesis can be further limited by breaking waves inhibiting light penetration to the seafloor. Wave exposure can also influence the sediment grain size, with areas of high wave exposure having coarser sediments with lower nutrient concentrations. Coarser sediments reduce the vegetative spreading of seagrasses and inhibit seedling colonisation (Gray & Elliott, 2009). Changes in sediment type can, therefore, have wider implications on the sensitivity of seagrasses on a long-term scale.
Sensitivity assessment. No evidence was available to determine the impact of this pressure at the benchmark level. However, exposure models from Studland Bay and Salcombe, where seagrass beds are limited to low wave exposure, show that even a change of 3% is likely to influence the upper shore limits as well as beds living at the limits of their wave exposure tolerance (Rhodes et al., 2006; Jackson et al., 2013). Change in wave exposure will impact the upper limit of seagrass and thus influence its wider distribution. At the benchmark level, an increase in wave exposure is likely to remove surface vegetation and the majority of the root system causing some mortality. Resistance is thus assessed as ‘Medium’. Recovery will depend on the presence of adjacent seagrass beds but is considered to be fairly rapid scoring a ‘Medium’ resilience. The biotope, therefore, scores a ‘Medium’ sensitivity changes in wave exposure at the pressure benchmark.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
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This pressure is Not assessed.
The effects of oxygen concentration on the growth and survivability of Zostera noltei are not reported in the literature. Zostera sp. leaves contain air spaces (lacunae) and oxygen is transported to the roots where it permeates into the sediment, resulting in an oxygenated microzone. This enhances the uptake of nitrogen. The presence of air spaces suggests that seagrass may be tolerant of low oxygen levels in the short-term, however, prolonged de-oxygenation, especially if combined with low light penetration and hence reduced photosynthesis may have a negative effect. Epifaunal gastropods may be tolerant of hypoxic conditions, especially Littorina littorea and Hydrobia ulvae. Infaunal species are likely to be exposed to hypoxic conditions, especially at low tide when they can no longer irrigate their burrows e.g. Arenicola marina can survive for 9 days without oxygen (Hayward, 1994). Conversely, possibly since it occupies the top few centimetres of sediment, Cerastoderma edule may be adversely affected by anoxia and would probably be killed by exposure to 2 mg/l oxygen for a week.
Sensitivity assessment. Overall de-oxygenation is not likely to adversely affect seagrass beds, especially in the lower intertidal where the biotope would experience periodic exposure to the air. Therefore, resilience is probably High, albeit with Low confidence, so that resistance is also High and the biotope is assessed as Not sensitive at the benchmark level.
During the past several decades, important losses in seagrass meadows have been documented worldwide related to an increase in nutrient load. Seagrasses are typically found in low energy habitats such as estuaries, coastal embayments and lagoons with reduced tidal flushing where nutrient loads are both concentrated and frequent. A typical response to nutrient enrichment is a decline in seagrass populations in favour of macroalgae or phytoplankton (Baden et al., 2003). Nutrient enrichment, especially of nitrogen and phosphorus, can lead to eutrophication. The mechanisms responsible for seagrass decline under eutrophication are complex and involve direct and indirect effects relating to changes in water quality, smothering by macroalgal blooms (Den Hartog & Phillips, 2000), and competition for light and nutrients with epiphytic microalgae and with phytoplankton (Nienhuis, 1996). In the Mondego estuary (Portugal), eutrophication triggered serious biological changes, which led to an overall increase in primary production and to a progressive replacement of seagrass Zostera noltei beds by coarser sediments and opportunistic macroalgae (Cardoso et al., 2004). Nutrients stimulate phytoplankton blooms that compete for nutrients but more importantly increase the turbidity and absorb light, reducing seagrass productivity (discussed in ‘changes in suspended solids’). In general terms, algae are able to out-compete seagrasses for water column nutrients since they have a higher affinity for nitrogen (Touchette & Burkholder, 2000). Short and Burdick (1996) found that excessive nitrogen loading stimulated the proliferation of algal competitors that caused shading and thereby stressing Zostera plants. Many seagrasses have a positive response to nitrogen and/or phosphorous enrichment (Peralta et al., 2003), but excessive loads can inhibit seagrass growth and survival, not only indirectly through light reduction resulting from increased algal growth but also directly in terms of the physiology of the seagrass. Direct physiological responses include ammonium toxicity and water column nitrate inhibition through internal carbon limitation (Touchette & Burkholder, 2000). Indirect effects of nutrient enrichment can accelerate decreases in seagrass beds such as sediment re-suspension from seagrass loss (see ‘changes in suspended solids’ pressure).
Sensitivity assessment. The loss of seagrass beds worldwide has been attributed to nutrient enrichment, due in part to the likeliness of smothering by epiphytes, and the effects of reduced light penetration caused by eutrophication. For instance, a study by Greening & Janicki (2006) found that in Florida, the USA, recovery of seagrass beds was incomplete 20 years after nutrient enrichment causing an eutrophication event. Seagrass beds are regarded as highly intolerant (or of low resistance) to this pressure. However, the benchmark of this pressure (compliance with WFD ‘good’ status) allows for a 30% loss of intertidal seagrass beds under the WFD criteria for good status. Therefore, at the level of the benchmark resistance of seagrass beds to this pressure is assessed ‘Medium’. The resilience of seagrass beds this degree of impact is assessed as ‘Medium’. The sensitivity score is therefore assessed as ‘Medium’.
Organic enrichment may lead to eutrophication with adverse environmental effects including deoxygenation, algal blooms and changes in community structure (see ‘nutrient enrichment’ pressure). Evidence on the effects of organic enrichment on Zostera species is limited but abundant for other seagrass species. Neverauskas (1987) investigated the effects of discharged digested sludge from a sewage treatment on Posidonia spp. and Amphibolis spp. in South Australia. Within 5 years the outfall had affected an area of approximately 1900 ha, 365 ha of which were completely denuded of seagrasses. The author suggests that the excessive growth of epiphytes on the leaves of seagrasses was a likely cause for reduced abundance. A subsequent study by Bryars & Neverauskas (2004) determined that eight years after the cessation of sewage output, total seagrass cover was approximately 28% of its former extent. While these results suggest that seagrasses can return to a severely polluted site if the pollution source is removed, they also suggest that it will take many decades for the seagrass community to recover to its former state. The effects of organic enrichment from fish farms were investigated on Posidonia oceanica seagrass beds in the Balearic Islands (Delgado et al., 1999). The fish culture had ceased in 1991; however, seagrass populations were still in decline at the time of sampling. The site closest to the former fish cages showed a marked reduction in shoot density, shoot size, underground biomass, sucrose concentration and photosynthetic capacities. The shoot also had high P-concentration in tissues and higher epiphyte biomass compared to the other sites. Since water conditions had recovered completely by the time of sampling, the authors suggest that the continuous seagrass decline was due to the excess organic matter remaining in the sediment (Delgado et al., 1999).
It should be noted that coastal marine sediments where seagrasses grow are often anoxic and highly reduced due to the high levels of organic matter and slow diffusion of oxygen from the water column to the sediment. Seagrasses are adapted to these conditions but if the water column is organically enriched, plants are unable to maintain oxygen supply to the meristem and die fairly quickly. The enrichment of the water column could therefore significantly increase the sensitivity of seagrasses to this pressure.
Sensitivity assessment. The organic enrichment of the marine environment increases turbidity and causes the enrichment of the sediment in organic matter and nutrients (Pergent et al., 1999). Evidence shows that seagrass beds found in proximity to a source of organic discharge were severely impacted with important losses of biomass. Although no study was found on the British species, the evidence suggests that Zostera noltei will be negatively affected by organic enrichment. No evidence was found addressing the benchmark of this study. A deposition of 100 gC/m2/year is considerably lower than the amount of organic matter discharged by sewage outlets and fish farms. Resistance to this pressure is thus deemed ‘Medium’. Recovery is assessed as ‘Medium’. The overall sensitivity of seagrasses to this pressure is ‘Medium’.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
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 of '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. Adjacent habitats and species populations may be indirectly affected where meta-population dynamics and trophic networks are disrupted and where the flow of resources e.g. sediments, prey items, loss of nursery habitat etc. is altered.
A change to another seabed type (from sediment to hard rock) will result in a permanent loss of suitable habitat for seagrass species. Resistance is thus assessed as ‘None’. As this pressure represents a permanent change, recovery is impossible as a suitable substratum for seagrasses is lacking. Consequently, resilience is assessed as ‘Very low’. The habitat, therefore, scores a ‘High’ sensitivity. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.
Seagrass beds occur almost exclusively in shallow and sheltered coastal waters anchored in sandy and muddy bottoms. A physical change to another seabed type such as a change in Folk class will, therefore, have a detrimental effect on seagrass beds as they will be excluded from the newly created habitat. A change towards a coarser sediment type would inhibit seagrasses from becoming established due to a lack of adequate anchoring substratum. A more mud dominated habitat, on the other hand, could increase sediment re-suspension and exclude seagrasses due to unfavourable light conditions. The resistance was assessed as ‘Low’. As this pressure represents a permanent change, recovery is impossible without intervention as a suitable substratum for seagrasses is lacking. Consequently, resilience is assessed as ‘Very low’. The habitat, therefore, scores a ‘High’ sensitivity. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.
The extraction of sediments to 30 cm (the benchmark) will result in the removal of every component of seagrass beds. Roots and rhizomes are buried no deeper than 20 cm below the surface (see ‘ ‘penetration and/or disturbance of the substratum below the surface of the seabed’). Resistance is therefore assessed as ‘None’ for and resilience is considered ‘Very low’ resulting in a ‘High’ sensitivity score.
Seagrasses are not physically robust. The leaves and stems of seagrass plants rise above the surface and the roots are shallowly buried so that they are vulnerable to surface abrasion. Activities such as trampling, anchoring, power boating and potting are likely to remove leaves and damage rhizomes. The removal of above-ground biomass would result in a loss of productivity whilst the removal of roots would cause the death of plants. Seagrasses are limited to shallow, protected waters and soft sediments. These areas are often open to public access and are widely used in commercial and recreational activities.
Trampling and vehicles: human wading in shallow coastal waters is a common activity that inherently involves trampling of the substratum. Trampling may be caused by recreational activities such as walking, horse-riding and off-road driving. These activities are likely to damage rhizomes and cause seeds to be buried too deeply to germinate (Fonseca, 1992). Negative effects of human trampling on seagrass cover, shoot density, and rhizome biomass, have been reported by Eckrich & Holmquist (2000) for the seagrass Thalassia testudinum. The study found that recovery occurred within a period of seven months after trampling ceased but the reduced cover was still visually distinguishable 14 months after the experiment. A study by Major et al. (2004) found that trampling impact varied depending on substratum type. A significant decrease in shoot density as a result from trampling was only observed at a site with soft muddy substratum with no impact detected on hard packed sand substratum. Damage from trampling is thus dependent on the substratum type with seagrass beds growing on soft substrata being most vulnerable to this pressure.
Hodges and Howe (1997) documented the impact of vehicular access on Zostera angustifolia beds in Angle Bay, Wales after the Sea Empress oil spill. Vehicle use, required for the initial clean up, resulted in patchy beds, criss-crossed with wheel ruts up to 1 m deep. Unauthorized activities before the spill, including vehicles associated with bait digging and the use of motorbikes, created ruts that were still visible over a year later.
Boating activities: boats passing in close proximity to seagrass beds can create waves. Turbulences from propeller wash and boat wakes can resuspend sediments, break off leaves, dislodge sediments and uproot plants. The re-suspension of sediments is further assessed in ‘changes in suspended sediment’ pressure. Koch (2002) established that physical damage from boat wakes was greatest at low tide but concluded that negative impacts of boat-generated waves were marginal on seagrass habitats. The physical impact of the engine’s propellers, shearing of leaves and cutting into the bottom, can also have damaging effects on seagrass communities. In severe cases, propellers cutting into the bottom may completely denude an area resulting in narrow dredged channels through the vegetation called propeller scars. Scars might expand and merge to form larger denuded areas. A study in Florida looking at the seagrasses Thalassia testudinum, Syringodium filiforme and Halodule wrightei determined that recovery of seagrass to propeller impact depend on species (Kenworthy et al., 2002). For Syringodium filiforme recovery was estimated at 1.4 years and for Halodule wrightei at 1.7 years, whilst recovery for Thalassia testudinum was estimated to require 9.5 years. Variations in recovery time were explained by different growth rates. However, it is not appropriate to assume that recovery rates are similar from one geographical or climatic region to another and more in-depth research is needed for Zostera species around the British Isles.
Potting: static gear is commonly deployed in areas where seagrass beds are found, either in the form of pots or as bottom set gill or trammel nets. Damage could be caused during the setting of pots or nets and their associated ground lines and anchors, by their movement over the bottom during rough weather and during recovery. Whilst the potential for damage is lower per unit deployment compared to towed gear (see 'penetration and/or disturbance of the substratum below the surface of the seabed' pressure), there is a risk of cumulative damage if use is intensive. Wall et al. (2008) categorized seagrass beds as being highly sensitive to high intensities of potting (pots lifted daily, with a density of over 5 pots per ha) and medium sensitive to lower levels (pots lifted daily, less than 4 pots per ha). However, no direct evidence was found to confirm these estimates.
Grazing: Nacken & Reise (2000) investigated physical disturbance caused by Brent geese (Branta b. bernicla) and wigeon (Anas penelope) feeding on Zostera noltei in the northern Wadden Sea. To graze on leaves and shoots above the sediment and on rhizomes and roots below, birds reworked the entire upper 1cm layer of sediment and excavated pits by trampling. As a result, birds pitted 12% of the seagrass bed and removed 63% of plant biomass. Plants recovered by the following year with the authors suggesting that seasonal erosion caused by herbivorous wildfowl was necessary for the persistence of Zostera noltei beds (Nacken & Reise, 2000). Similarly, Davison & Hughes (1998) suggested that Zostera sp. can rapidly recover from 'normal' levels of wildfowl grazing. Physical disturbance may, however, be detrimental to seagrass beds as soon as the ‘normal’ level caused by grazing birds is exceeded by human activities.
Experimental: Boese et al. (2009) examined the recolonization of experimentally created gaps within intertidal perennial and annual Zostera marina beds in the Yaquina River Estuary, USA. The experiment looked at two zones, the lower intertidal almost continuous seagrass and an upper intertidal transition zone where there were patches of perennial and annual Zostera marina. The study found that recovery began within a month after disturbance in the lower intertidal continuous perennial beds and was complete after two years, whereas, plots in the transition zone took almost twice as long to recover.
Sensitivity assessment. In summary, a wide range of activities gives rise to this pressure with intertidal habitat being more exposed as they are more readily accessible than subtidal beds. Seagrass plants are not physically robust and their root system is located in the upper layer of the sediment making them prone to damage by abrasion. The resilience and recovery of seagrass beds to abrasion of the seabed surface depends on the frequency, persistence and extent of the disturbance. Factors such as the size and shape of the impact will also influence the sensitivity of seagrass to this pressure. There is also considerable evidence that the type of substratum plays a role in determining the magnitude of impact. Soft and muddy substratum is thought to be more easily damaged than harder more compact ground. Finally, temporal effects should also be taken into account. The state of the tide will influence the magnitude of damage as will seasonal effects with damage induced in winter being more likely to have a lesser impact than damage occurring during the growing season. Overall, studies suggest little resistance to abrasion resulting in ‘Low’ resistance. Physical disturbance and removal of plants can lead to increased patchiness and destabilisation of the seagrass bed, which in turn can lead to reduced sedimentation within the seagrass bed, increased erosion, and loss of larger areas of plants (Davison & Hughes, 1998). Recovery will, however, be fairly rapid resulting in ‘Medium’ resilience. Overall this biotope, therefore, has a ‘Medium’ sensitivity to this pressure.
Seagrass species are vulnerable to physical damage. Activities such as digging and raking for clams, anchoring and mooring will penetrate the substratum to an average depth of 5 cm removing plant biomass above and below ground. Penetration to the substratum to a depth greater than 5 cm will directly impact seagrass habitats as the plant is confined to the upper layer of the sediment. All biomass (leaves, rhizomes) will be completely removed leading to the death of the plant. Seagrass beds are often associated with commercially important bivalves. Fisheries targeting these species are therefore likely to impact seagrass habitats and are the most widespread (and best studied) activities giving rise to this pressure on this habitat.
Clam digging and clam raking: Racking and digging for shellfish is a common practice in the intertidal zone. In southern Portugal clam harvesters dig up intertidal sediments dominated by the seagrass Zostera noltei, using a hand-blade, which breaks and removes the shoots and rhizomes of plants. Cabaço & Santos (2007) found that clam harvesting activities change the species population structure by significantly reducing shoot density and total biomass, particularly during August, when the harvest effort is highest. Experimental harvest revealed a short-term impact on shoot density, which however rapidly recovered to control levels during the following month. By experimentally manipulating rhizome fragmentation, the authors determined that plant survival was only reduced when fragmented rhizomes were left with less than two intact internodes; fragmented rhizomes having 2 to 5 internodes were not significantly affected, even though growth and production were lower with fewer internodes. The results of this study suggest that Zostera noltei is adversely affected by clam harvesting, however, the species is able to rapidly recover from this physical disturbance. In the same study area, Alexandre et al. (2005) looked into the effects of clam harvesting on sexual reproduction. Disturbed meadows showed significantly lower vegetative shoot density but significantly higher reproductive effort. These results were confirmed by manipulative experiments and suggest that Zostera noltei responded to clam harvesting disturbance by both increasing its reproductive effort and extending its fertile season. Boese (2002) investigated the effects of manual clam harvesting on Zostera marina by raking and digging for clams in experimental plots in Yaquina Bay, USA. After three monthly treatments, measures of biomass, primary production (leaf elongation), and percent cover were compared between disturbed and undisturbed plots. The study found that clam raking treatments visibly removed large numbers of seagrass leaves and some below-ground rhizomes. However, 2 weeks after the end of the experiment, no statistical difference in percentage cover was observed between disturbed and control plots indicating a fast recovery rate. Clam digging, on the other hand, caused visual differences in percentage cover for 10 months after the end of the experiment, although differences were not statistically significant. Boese (2002) concluded that recreational clamming is unlikely to have a major impact on seagrass beds in the Yaquina estuary. The author calls, however, to view the results with caution as multi-year disturbances were not investigated and differences in sediment characteristics are likely to influence the resistance and resilience of seagrasses to this pressure. Similarly, Peterson et al. (1987) found that hand raking and moderate clam-kicking (a commercial harvesting method in which propeller wash is used to dislodge hard clams) resulted in a reduction in Zostera marina biomass by approximately 25%. No differences between control and experimental areas were apparent one year after the experiment. However at higher intensity, clam-kicking reduced seagrass biomass to about half of control levels and recovery remained incomplete 4 years after the end of the experiment (Peterson et al., 1987).
Anchoring and mooring: an anchor landing on a patch of seagrass can bend, damage and break seagrass shoots (Montefalcone et al., 2006) and an anchor being dragged as the boat moves driven by wind or tide causes abrasion of the seabed. Milazzo et al. (2004) found that the extent of damage depended on the type of anchor with the folding grapnel having the greatest impact. The study further determined that heavier anchors (often associated with larger boats) will sink deeper into the substratum and thereby cause greater damage. A technical paper by Collins et al. (2010) using SCUBA divers found bare patches (typically 1–4 m2) were caused by anchoring by leisure boats in Studland Bay, UK. The study further determined that average shear vane stress was significantly higher in intact seagrass beds compared to scars indicating a less cohesive and more mobile substratum caused by anchors. Axelsson et al. (2012) also investigated anchor damage in Studland Bay. The study did not provide consistent evidence of boat anchoring impacting the seagrass habitat in this location. The study did, however, observe higher shoot density and percentage cover of seagrass in a voluntary anchor zone compared to a control area where anchoring occurred. The authors recommended longer monitoring in order to determine whether the trend was caused by natural variations or the effects of anchor exclusion. Traditional mooring further contributes to the degradation of seagrass habitats. A traditional swing mooring is a buoy on a chain attached to a static anchoring block fixed on the seabed, to buffer any direct force on the permanent block, the chain lies on the seabed where it moves around with wind and tides, as the chain pivots on the block it scours the seabed. In proximity to seagrass beds, the chain usually removes not only the seagrass above ground parts such as leaves and shoots but also the roots anchored in the sediment. Further sediment abrasion may occur in the vicinity to the anchoring blocks due to eddying of currents. The blocks themselves may increase the competition of seagrass with other algae as they provide ideal settlement surfaces. Boats might also moor on intertidal sediments. When the tide goes out, the boat sits directly on top of the soft sediment. Walker et al. (1989) found that boat moorings caused circular or semi-circular depressions of bare sand within seagrass beds between 3 to 300 m2 causing important habitat fragmentation. The scours created by moorings in the seagrass canopy interfere with the physical integrity of the meadow. Though relatively small areas of seagrass are damaged by moorings, the effect is much greater than if an equivalent area was lost from the edge of a meadow. Such mooring scars have been observed for Zostera marina around the UK such as in Porth Dinllaen in the Pen Llyna’r Sarnau Special Area of Conservation, Wales (Egerton, 2011) and at Studland Bay (Jackson et al., 2013).
Trawling: Neckles et al. (2005) investigated the effects of trawling for the blue mussels Mytilus edulis on Zostera marina beds in Maquoit Bay, USA. Impacted sites ranged from 3.4 to 31.8 ha in size and were characterized by the removal of above- and belowground plant material from the majority of the bottom. The study found that one year after the last trawl, Zostera marina shoot density, shoot height and total biomass averaged respectively to 2-3%, 46-61% and < 1% that of the reference sites. Substantial differences in Zostera marina biomass persisted between disturbed and reference sites up to 7 years after trawling. Rates of recovery depended on initial fishing intensity but the authors estimated that an average of 10.6 years was required for Zostera shoot density to match pre-trawling standards.
Dredging and suction dredging: the effects of dredging for scallops on Zostera marina beds were investigated by Fonseca et al. (1984) in Nova Scotia, USA. Dredging was carried out when plants were in the vegetative stage on hard sand and on soft mud substrata. Damage was assessed by analysing the effects of scallop harvesting on seagrass foliar dry weight and on the number of shoots. Lower levels of dredging (15 dredges) had a different impact depending on substrata, with the hard bottom retaining a significantly greater overall biomass than a soft bottom. However, an increase in dredging effort (30 dredges) led to a significant reduction in seagrass biomass and shoot number on both hard and soft bottoms. Solway Firth is a British example for the detrimental effects of dredging on seagrass habitats. In the area, where harvesting for cockles by hand is a traditional practice, suction dredging was introduced in the 1980s to increase the yield. A study by Perkins (1988) found that where suction dredging occurred, the sediment was smoothened and characterized by a total absence of Zostera plants. The study concluded that the fishery was causing widespread damage and could even completely eradicate Zostera from affected areas. Due to concerns over the sustainability of this fishing activity, the impacts on cockle and Zostera stocks, and the effects on overwintering wildfowl, the fishery was closed to all forms of mechanical harvesting in 1994.
Sensitivity assessment. The deployment of fishing gears on seagrass beds results in physical damage to the above surface part of the plants as well as to the root systems. Seagrasses do not have an avoidance mechanism; resistance to this pressure is therefore assessed as ‘None’. The recovery of seagrass beds after disturbance to the sub-surface of the sediment will be slow with the speed depending on the extent of removal. Rates may be accelerated where adjacent seed sources and viable seagrass beds are present but can be considerably longer where rhizomes and seed banks were removed. Using a model simulation, it has been suggested that with favourable environmental conditions, seagrass beds might recover from dragging disturbance in 6 years; conversely, recovery under conditions less favourable to seagrass growth could require 20 years or longer (Neckles et al., 2005). Resilience is thus assessed as ‘Low’. The mechanical harvest of shellfish damaging the sub-surface of the sediments poses a very severe threat to seagrass habitats, yielding a ‘High’ sensitivity score.
Irradiance decreases exponentially with increasing depth, and the suspended sediment concentration has a direct linear effect on light attenuation (Van Duin et al., 2001). Changes in suspended solids will thus reduce light available for seagrass plants necessary for photosynthesis. Impaired productivity due to a decrease in photosynthesis will affect the growth and reproductive abilities of plants. Turbidity also results in a reduction of the amount of oxygen available for respiration by the roots and rhizomes thus lowering nutrient uptake. The resulting hypoxic conditions will lead to a build-up of sulphides and ammonium, which can be toxic to seagrass at high concentrations (Mateo et al., 2006). Davison & Hughes (1998) reported considerable declines in seagrass populations related to increases in turbidity from dredging in the Wadden Sea.
Water clarity is a vital component for seagrass beds as it determines the depth-penetration of photosynthetically active radiation of sunlight. Seagrasses have light requirements an order of magnitude higher than other marine macrophytes making water clarity a primary factor in determining the maximum depth at which seagrasses can occur. The critical threshold of light requirements varies among species ranging from 2% in-water irradiance for Z. noltei, to 11 to 37% for Z. marina (Erftemeijer & Robin, 2006). These differences in the light requirement for Zostera are reflected by the position of species along a depth gradient with Zostera noltei occurring predominantly in the intertidal and Zostera marina found at greater depth in the subtidal. However, differences in light requirements also vary within species. For example, the minimum light requirement for Zostera marina in a Danish embayment was 11% in-water irradiance, whereas the estimated light requirement for the same species in the Netherlands was 29.4% in-water irradiance (Olesen, 1993). This variability within species is likely attributed to photo-acclimation to local light regimes. A study by Peralta et al. (2002) investigated the effects of reduced light availability on Zostera noltei in Spain. The authors determined that plants were able to tolerate acute light reductions for a short period of time (below 2% of surface irradiance for two weeks) by storing and mobilizing carbohydrates at a low level of irradiance. However, Zostera noltei are likely to be less tolerant to chronic, long-term reductions in light availability. In a six month long experiment in the Dutch Wadden Sea, Philippart (1995) found that shading induced a 30% decrease in the leaf growth rate, a 3-fold increase in the leaf loss rate, and a 80% reduction in the total biomass of Zostera noltei. The decreasing growth rate is most probably the result of a reduction of photosynthesis due to shading. The increased leaf loss may have been the result of enhanced deterioration of leaf material under low light conditions. The study also established that during the summer period, the maximum biomass of Zostera noltei under the control light conditions was almost 10 times higher than those under the low light conditions (incident light reduced to 45% of natural light conditions). The summer is a critical period for maintenance and growth of vegetative shoots. The effects of shading may, therefore, be most severe during summer months.
Changes in suspended solids leading to turbid conditions over a prolonged period of time are, therefore, likely to adversely impact seagrass species. The extent of damage will depend on individual seagrass beds. Older, more established perennial meadows have greater carbohydrate reserves and are thus more able to resist to changes in light penetration than annual plants (Alcoverro et al., 2001). Seagrass plants found in clear waters may be able to tolerate sporadic high turbidity (Newell & Koch, 2004). However, where seagrass beds are already exposed to low light conditions, then losses may result from even short-term events (Williams, 1988).
Sensitivity assessment. Turbidity is an important factor controlling production and ultimately survival and recruitment of seagrasses. Seagrass populations are likely to survive short-term increases in turbidity, however, a prolonged increase in light attenuation, especially at the lower depths of its distribution, will probably result in loss or damage of the population. A score of ‘Low’ was therefore recorded for resistance. A loss of seagrass beds will promote the re-suspension of sediments, making recovery unlikely as seagrass beds are required to initially stabilise the sediment and reduce turbidity levels (Van der Heide et al., 2007). A high turbidity state appears to be a highly resilient alternative stable state; hence return to the seagrass biotope is unlikely resulting in ‘Low’ resilience. Zostera noltei should be considered intolerant of any activity that changes the sediment regime where the change is greater than expected due to natural events, yielding a ‘High’ sensitivity score.
Several studies have documented deterioration of seagrass meadows by smothering due to excessive sedimentation. Consequences of enhanced sedimentation for seagrass beds depend on several factors such as the life history stage as well as the depth and timing of burial. Early life stages of seagrass, smaller in size than adult plants, are most vulnerable to this pressure as even a small load of added sediment will lead to the complete burial. Cabaço & Santos (2007) determined that Zostera noltei is highly sensitive to burial and erosion disturbances due to the small size of this species and the lack of vertical rhizomes. Buried plants, however, produced longer rhizome internodes as a response to burial, suggesting an attempt to relocate the leaf-producing meristems closer to the sediment surface. The carbon content of leaves and rhizomes, as well as the non-structural carbohydrates (mainly the starch in the rhizomes), dropped significantly during the experimental period, indicating an internal mobilisation of carbon to meet the plant demands as a consequence of light deprivation. However, shoots did not survive more than 2 weeks under complete burial. Cabaço & Santos (2007) concluded that Zostera noltei was highly sensitive to burial disturbance and determined that the threshold for total shoot loss was between 4 cm and 8 cm of burial. The study did not observe any recovery within 2 months of the experiment. Han et al. (2012) found that mortality of Zostera noltei plants was related to sediment depth with survival rapidly decreasing when rhizomes were buried deeper than 1 cm. Similar to Cabaço & Santos (2007), Han et al. (2012) observed that Zostera noltei was able to relocate rhizomes to the depth at which the rhizomes of undisturbed plants were most frequently found. However, contrary to the previous study Han et al. (2012) concluded that Z. noltei is well adapted to cope with sediment disturbances of limited amplitude (i.e. ± 6 cm) by rapidly relocating their rhizomes to the preferential depth. Tu Do et al. (2012) investigated the recovery of Zostera noltei beds in Arcachon Bay in France after burial resulting from dredging activities (10 cm, mainly discharged in a main single event). The study found that seagrass beds had completely disappeared within 6 months with plants only partly recovering 5 years after the initial disturbance.
Other factors influencing the sensitivity of Zostera noltei to smothering is the frequency and the timing of deposition of material. The timing of the siltation event plays a particularly important role for intertidal beds. At low tide, for instance, the seagrass bed is exposed with plants lying flat on the substratum. The addition of material would immediately smother the entire plant and have a greater impact on leaves and stem than if added on plants standing upright. The resistance of intertidal beds to this pressure may thus vary with time of day. In addition, sudden burial has a more pronounced negative effect on the survival response of Zostera noltei than continuous burial (Han et al., 2012).
Sensitivity assessment. The above studies suggest that Zostera noltei is intolerant of smothering with some discrepancy between the critical threshold depths of burial. All studies, however, indicate that at the level of the benchmark (5 cm of fine material added to the seabed) some mortalities will occur resulting in a 'Medium' resistance score. Some plants will survive by successfully relocating rhizomes closer to the sediment surface. With the benchmark set at ‘material added to the seabed in a single event’, sensitivity will be greater than if burial occurred in a continuous way. In addition, seagrass beds are restricted to low energy environments, suggesting that once the silt is deposited, it will remain in place for a long period of time so habitat conditions will not reduce exposure. Resilience is therefore assessed as 'Medium'. The biotope is considered to have a ‘Medium’ sensitivity to siltation at the pressure benchmark.
Zostera noltei is sensitive of smothering by excessive siltation (see light smothering above). Studies have found that the seagrass is capable of producing longer rhizome internodes as a response to burial in an attempt to relocate leaf-producing meristems closer to the sediment surface (Cabaço & Santos, 2007; Hall et al., 2012; Tu Do et al., 2012). All studies indicate that seagrass species are sensitive to an increase in sedimentation rates at the benchmark level of 30 cm. In addition, seagrass beds are restricted to low energy environments, suggesting that once the silt is deposited, it will remain in place for a long period of time so habitat conditions will not reduce exposure. Resistance is assessed as ‘None’ as all individuals exposed to siltation at the benchmark level are predicted to die and consequent resilience as ‘Low’ to ‘Very Low’. Sensitivity based on combined resistance and resilience is therefore assessed as ‘High’.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Species characterizing this habitat do not have hearing perception but vibrations may cause an impact, however, no studies exist to support an assessment
Seagrasses have light requirements an order of magnitude higher than other marine macrophytes making water clarity a primary factor in determining the maximum depth at which seagrasses can occur. The critical threshold of light requirements varies among species ranging from 2% in-water irradiance for Zostera noltei, to 11 to 37% for Zostera marina (Erftemeijer & Robin, 2006). These differences in the light requirement for Zostera are reflected by the position of species along a depth gradient with Zostera noltei occurring predominantly in the intertidal and Zostera marina found at greater depth in the subtidal. However, differences in light requirements also vary within species. For example, the minimum light requirement for Zostera marina in a Danish embayment was 11% in-water irradiance, whereas the estimated light requirement for the same species in the Netherlands was 29.4% in-water irradiance (Olesen, 1993). This variability within species is likely attributed to photo-acclimation to local light regimes. A study by Peralta et al. (2002) investigated the effects of reduced light availability on Zostera noltei in Spain. The authors determined that plants were able to tolerate acute light reductions for a short period of time (below 2% of surface irradiance for two weeks) by storing and mobilizing carbohydrates at a low level of irradiance. However, Zostera noltei are likely to be less tolerant to chronic, long-term reductions in light availability. In a six month long experiment in the Dutch Wadden Sea, Philippart (1995) found that shading induced a 30% decrease in the leaf growth rate, a 3-fold increase in the leaf loss rate, and a 80% reduction in the total biomass of Zostera noltei. The decreasing growth rate is most probably the result of the reduction of photosynthesis due to shading. The increased leaf loss may have been the result of enhanced deterioration of leaf material under low light conditions. The study also established that during the summer period, the maximum biomass of Zostera noltei under the control light conditions was almost 10 times higher than those under the low light conditions (incident light reduced to 45% of natural light conditions). The summer is a critical period for maintenance and growth of vegetative shoots. The effects of shading may, therefore, be most severe during summer months.
Overall, the effects of shading could mirror those of reduced water clarity (increased turbidity) depending on the scale of the artificial structure. Therefore, a resistance of 'Low', with a resilience of 'Medium' and sensitivity of 'Medium' is suggested, albeit with low confidence.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
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 the dispersal of seed. But seed dispersal is not considered under the pressure definition and benchmark.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Not relevant to seabed habitats. NB. Collision by grounding vessels is addressed under ‘surface abrasion’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
Translocation of seagrass seeds, rhizomes and seedlings is a common practice globally to counter the trend of decline of seagrass beds. Zostera marina is the seagrass species most commonly translocated. Williams & Davis (1996) found that levels of genetic diversity of restored eelgrass Zostera marina beds in Baja California, USA, were significantly lower than in natural populations.The loss of genetic variation can lead to lower rates of seed germination and fewer reproductive shoots, suggesting that there might be long-term detrimental effects for population ﬁtness. Williams (2001) affirmed that genetic variation was essential in determining the potential of seagrass to rapidly adapt to a changing environment. Transplanted populations are, therefore, more sensitive to external stressors such as eutrophication and habitat fragmentation, with reduced community resilience, compared to natural populations (Hughes & Stachowicz, 2004). Even though restoration efforts tend to focus on Zostera marina, transplantations of Zostra noltei (Martins et al., 2005) have also been undertaken. Similar reductions in genetic diversity are expected, making the transplanted populations particularly sensitive to external stressors.
Translocation also has the potential to transport pathogens to uninfected areas (see 'introduction of microbial pathogens' pressure). The sensitivity of the ‘donor’ population to harvesting to supply stock for translocation is assessed for the pressure ‘removal of target species’. No evidence was found for the impacts of translocated beds on adjacent natural seagrass beds. However, it has been suggested that translocation of plants and propagules may lead to hybridization with local wild populations. If this leads to loss of genetic variation there may be long-term effects on the potential to adapt to changing environments and other stressors.
Sensitivity assessment: Presently, there is no evidence of loss of habitat due to genetic modification and translocation of seagrass species. However, if hybridization occurred and genetic diversity was reduced, then the affected populations may become more susceptible to change and, hence, more sensitive.
The effect of non-native species on seagrass species was reviewed by d’Avack et al. (2014). The review reported several non-native invasive plants as well as invertebrate species negatively impacting British seagrass beds. These included Sargassum muticum, Spartina anglica, Codium fragile ssp. tomentosoides, Didemnum vexillum, Urosalpinx cinerea and Magallana gigas.
For example, the cord grass Spartina anglica is non-native grass, which was recorded to have negative effects on seagrass beds. This hybrid species of native (Spartina alterniflora) and an introduced cord grass species (Spartina maritima) colonizes the upper part of mud flats, where due to its extensive root system, it effectively traps and retains sediments. Spartina anglica has rapidly colonised mudflats in England and Wales due to its fast growth rate and high fecundity. Deliberate planting to stabilise sediments accelerated its spread throughout Britain (Hubbard & Stebbings 1967). By consolidating the sediments the plant is responsible for raising mud flats as well as reducing sediment availability elsewhere. Butcher (1934) raised concerns that its pioneering consolidation may result in the removal of sediments from Zostera beds. Declines in Zostera noltei due to the encroachment of Spartina anglica were observed in Lindisfarne National Reserve in north-east England (Percival et al. 1998). The reduction in Zostera noltei beds had a direct impact on wildfowl populations as the food availability for the wildfowl was reduced on the top of the shore. This pressure will affect the upper limits of the intertidal rather than subtidal biotopes.
Sensitivity assessment: d’Avack et al. (2014) found that invasive flora had the greatest impact on seagrass bed but pointed out extensive knowledge gaps on how invasive species influence the health of Zostera beds in UK waters More research is thus needed in order to fully comprehend this pressure. Resistance is assessed as 'Low'. Return to ‘normal’ conditions is highly unlikely if an invasive species would come to dominate the biotope. Indeed recovery would only be possible if the majority of the NIS were removed (through either natural or unnatural process) to allow the re-establishment of other species. Therefore actual resilience is assessed as ‘Low’ resulting in an overall ‘High’ sensitivity score.
Historic records show that seagrass species, in particular, Zostera marina, are highly susceptible to microbial pathogens. During the 1930s, a so-called ‘wasting disease’ decimated Zostera marina populations in Europe and along the Atlantic Coast of North America with over 90% loss (Muehlstein, 1989). Wasting disease resulted in black lesions on the leaf blades which potentially lead to loss of productivity, degradation of shoots and roots, eventually leading to the loss of large areas of seagrass (Den Hartog, 1987). Wasting disease is caused by infection with a marine slime mould-like protist called Labyrinthula zosterae (Short et al., 1987; Muehlstein et al., 1991). Recovery of seagrass beds after the epidemic has been extremely slow or more or less absent in some areas such as the Wadden Sea (Van der Heide et al., 2007).
The disease is less likely at low salinities, however, and Zostera noltei was little affected (Rasmussen, 1977; Davison & Hughes, 1998). Hence, Zostera noltei populations did not suffer to the same extent even though the disease also occurs in this species (Vergeer & Den Hartog, 1991).
Sensitivity assessment. Zostera noltei is susceptible to microbial pathogens but unlikely to suffer the level of mortality experienced by Zostera marina. Therefore, a resistance of 'Medium' is recorded, with a resilience of 'Medium, resulting in a sensitivity of ‘Medium’.
Seagrass is not targeted by a commercial fishery. Seeds and shoots are, however, harvested for extensive transplantation project aimed at promoting seagrass populations in areas denuded by natural or anthropogenic causes. Divers are most commonly employed to remove material from the source population, an activity with a low overall impact on seagrass habitats. In the USA, however, a mechanical seed harvesting technique was invented and put into practice (Orth & Marion, 2007). The mechanised harvester was able to drastically increase the number of Zostera seed collected from a source population (1.68 million seeds in one day compared to 2.5 million seeds collected by divers in one year). However, the large-scale removal of seeds, the productive output of seagrasses, can affect the integrity of the natural seagrass beds. To date, no mechanical harvesting has been employed in the UK. The ecological impact of seed collection by divers is low; the harvesting of Zostera in British waters has, therefore, a minimal effect on natural seagrass habitats. The effect of the translocation of species is covered in the pressure ‘genetic modification and translocation of indigenous species’. The direct physical effects on seabed habitats from activities are described below in ‘abrasion/disturbance’ of the substratum on the surface of the bed’ and ‘penetration and/or disturbance of the substratum below the surface’.
Harvesting of seagrasses as craft material is a small but growing industry. However, the present legislation for the conservation of seagrasses will discourage expansion of this industry (see Jackson et al. (2013) for a full list on the political framework for seagrass protection in the UK).
Seagrass beds are not considered dependent on any of the organisms that may be targeted for direct removal e.g. oysters, clams and mussels. However, an indirect effect of fisheries targeting bivalves is a change in the water clarity, crucial for the growth and development of Zostera species. Indeed bivalves have been shown to significantly contribute to the clearance of the water column which subsequently increases light penetration, facilitating the growth and reproduction of Zostera species (Wall et al., 2008). Newell & Koch (2004) using modelling, predicted that when sediments were resuspended, the presence of even low numbers of oysters (25 g dry tissue weight m2) distributed uniformly throughout the domain, reduced suspended sediment concentrations by nearly an order of magnitude. A healthy population of suspension-feeding bivalves thus improves habitat quality and promotes seagrass productivity by mitigating the effects of increased water turbidity in degraded, light-limited habitats (see 'changes in suspended solids' pressure). Bivalves also contribute pseudofaeces to fertilise seagrass sediments (Bradley & Heck Jr, 1999).
Seagrass plants may be directly removed or damaged by static or mobile gears that are targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope.
Sensitivity assessment. Seagrass beds have no avoidance mechanisms to escape targeted harvesting of leaves, shoots and rhizomes. Resistance to this pressure is, therefore, assessed as ‘Low’. A study by Nacken & Reise (2000) investigated the removal of Zostera noltei plants caused by Brent geese (Branta b. bernicla) and widgeon (Anas penelope) in the northern Wadden Sea. Birds removed 63% of plant biomass. Plants recovered by the following year with the authors suggesting that seasonal erosion caused by herbivorous wildfowl may be necessary for the persistence of Zostera noltei beds (Nacken & Reise, 2000). The study suggests that recovery from the removal of target species will be rapid resulting in 'High' resilience score. Added anthropogenic disturbance may, however, be detrimental to seagrass beds as soon as the ‘normal’ level caused by grazing birds is exceeded by human activities. Overall the sensitivity of this biotope is deemed ‘Low’ to this pressure.
Filter-feeders such as mussels, clams and scallops are often associated with seagrass beds. Fisheries targeting these bivalves employ methods such as trawling, dredging, digging and raking which all result in the non-targeted removal of seagrass species. The direct physical effects of such fishing methods on seagrass are described in detail for the pressures ‘abrasion’ and ‘penetration and/or disturbance of the substratum’.
Seagrasses may be directly removed or damaged by static or mobile gears that are targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species in this biotope.
Sensitivity assessment. Seagrass habitats are not dependent on any other organisms but the incidental removal of seagrass as by-catch could be detrimental and could remove the biotope (see evidence presented under 'penetration and/or disturbance of the substratum' above). Therefore, resistance is considered to be 'None', resilience 'Low' and sensitivity 'High'.
Alcoverro, T., Manzanera, M. & Romero, J., 2001. Annual metabolic carbon balance of the seagrass Posidonia oceanica: the importance of carbohydrate reserves. Marine Ecology Progress Series, 211, 105-116.
Alexandre, A., Santos, R. & Serrão, E., 2005. Effects of clam harvesting on sexual reproduction of the seagrass Zostera noltii. Marine Ecology Progress Series, 298, 115-122
Asmus, H. & Asmus, R., 2000a. ECSA - Workshop on intertidal seagrass beds and algal mats: organisms and fluxes at the ecosystem level. (Editorial). Helgoland Marine Research, 54, 53-54.
Asmus, H. & Asmus, R., 2000b. Material exchange and food web of seagrasses beds in the Sylt-Rømø Bight: how significant are community changes at the ecosystem level? Helgoland Marine Research, 54, 137-150.
Axelsson, M., Allen, C., Dewey, S. , 2012. Survey and monitoring of seagrass beds at Studland Bay, Dorset – second seagrass monitoring report. Report to The Crown Estate and Natural England by Seastar Survey Ltd.
Baden, S., Gullström, M., Lundé n, B., Pihl, L. & Rosenberg, R., 2003. Vanishing Seagrass (Zostera marina, L.) in Swedish Coastal Waters. Ambio, 32(5), 374-377.
Bentley, M.G. & Pacey, A.A., 1989. A scanning electron microscopial study of sperm development and activation in Arenicola marina L. (Annelida: Polychaeta). Invertebrate Reproduction and Development, 15, 211-219.
Bester, K., 2000. The effects of pesticides on seagrass beds. Helgoland Marine Research, 54, 95-98.
Boese, B.L., 2002. Effects of recreational clam harvesting on eelgrass (Zostera marina) and associated infaunal invertebrates: in situ manipulative experiments. Aquatic Botany, 73 (1), 63-74.
Boese, B.L., Kaldy, J.E., Clinton, P.J., Eldridge, P.M. & Folger, C.L., 2009. Recolonization of intertidal Zostera marina L. (eelgrass) following experimental shoot removal. Journal of Experimental Marine Biology and Ecology, 374 (1), 69-77.
Bradley, J. & Heck Jr, K.L., 1999. The potential for suspension feeding bivalves to increase seagrass productivity. Journal of Experimental Marine Biology and Ecology, 240 (1), 37-52.
Brown, R.A., 1990. Strangford Lough. The wildlife of an Irish sea lough. The Institute of Irish Studies, Queens University of Belfast.
Bryan, G.W. & Gibbs, P.E., 1991. Impact of low concentrations of tributyltin (TBT) on marine organisms: a review. In: Metal ecotoxicology: concepts and applications (ed. M.C. Newman & A.W. McIntosh), pp. 323-361. Boston: Lewis Publishers Inc.
Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters, vol. 5. Ocean Management, part 3, (ed. O. Kinne), pp.1289-1431. New York: John Wiley & Sons.
Bryars, S. & Neverauskas, V., 2004. Natural recolonisation of seagrasses at a disused sewage sludge outfall. Aquatic Botany, 80 (4), 283-289.
Buchsbaum, R.N., Short, F.T. & Cheney, D.P., 1990. Phenolic-nitrogen interactions in eelgrass Zostera marina: possible implications for disease resistance. Aquatic Botany, 37, 291-297.
Burkholder, J.M., Mason, K.M. & Glasgow, H.B. Jr., 1992. Water-column nitrate enrichment promotes decline of eelgrass Zostera marina: evidence from seasonal mesocosm experiments. Marine Ecology Progress Series, 81, 163-178.
Burton, P.J.K., 1961. The Brent goose and its winter food supply in Essex. Wildfowl, 12, 104-112.
Butcher, R., 1934. Zostera. Report on the present condition of eel grass on the coasts of England, based on a survey during August to October, 1933. Journal du Conseil, 9 (1), 49-65.
Butcher, R.W., 1941. Zostera: report on the present conditions of eelgrass on the coasts of England, based on a survey during August to October, 1933. The International Wildfowl Inquiry, 1, 49-65.
Cabaço, S., Machás, R. & Santos, R., 2009. Individual and population plasticity of the seagrass Zostera noltii along a vertical intertidal gradient. Estuarine, Coastal and Shelf Science, 82 (2), 301-308.
Cabaço, S. & Santos, R., 2007. Effects of burial and erosion on the seagrass Zostera noltii Journal of Experimental Marine Biology and Ecology, 340, 204-212
Cardoso, P., Pardal, M., Lillebø, A., Ferreira, S., Raffaelli, D. & Marques, J., 2004a. Dynamic changes in seagrass assemblages under eutrophication and implications for recovery. Journal of Experimental Marine Biology and Ecology, 302 (2), 233-248.
Cardoso, P.G., Raffaelli, D. & Pardal, M.A., 2008. The impact of extreme weather events on the seagrass Zostera noltii and related Hydrobia ulvae population Marine Pollution Bulletin, 56, 483-492
Charpentier, A., Grillas, P., Lescuyer, F., Coulet, E. & Auby, I. 2005. Spatio-temporal dynamics of a Zostera noltii dominated community over a period of fluctuating salinity in a shallow lagoon, Southern France Estuarine, Coastal and Shelf Science, 64, 307-315
Clay, E., 1967a. Literature survey of the common fauna of estuaries, 2. Arenicola marina Linnaeus. Imperial Chemical Industries Limited, Brixham Laboratory, PVM45/A/395.
Collins, K., Suonpää, A. & Mallinson, J., 2010. The impacts of anchoring and mooring in seagrass, Studland Bay, Dorset, UK. Underwater Technology, 29 (3), 117-123.
Connor, D.W., Brazier, D.P., Hill, T.O., & Northen, K.O., 1997b. Marine biotope classification for Britain and Ireland. Vol. 1. Littoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 229, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report No. 230, Version 97.06.
Covey, R. & Hocking, S., 1987. Helford River Survey. Report for the Heinz, Guardians of the Countryside and World Wide Fund for Nature, 121 pp.
Cox, J., 1991. Dredging for the American hard-shell clam - implications for nature conservation. Ecosystems. A Review of Conservation, 12, 50-54.
Creed, J.C., Filho, A. & Gilberto, M., 1999. Disturbance and recovery of the macroflora of a seagrass Halodule wrightii (Ascherson) meadow in the Abrolhos Marine National Park, Brazil: an experimental evaluation of anchor damage. Journal of Experimental Marine Biology and Ecology, 235 (2), 285-306.
d’Avack, E.A.S., Tillin, H., Jackson, E.L. & Tyler-Walters, H. , 2014. Assessing the sensitivity of seagrass bed biotopes to pressures associated with marine activities. JNCC Report No. 505. Joint Nature Conservation Committee, Peterborough. Available from www.marlin.ac.uk/publications.
Dankers, N. & de Vlas, J., 1992. Multifunctioneel beheer in de Waddenzee, integratie van natuurbeheer en schelpdiervisserij. Institute for Forestry and Nature Research, RIN Report, no. 92/15, 18pp.
Dauvin, J.C., Bellan, G., Bellan-Santini, D., Castric, A., Francour, P., Gentil, F., Girard, A., Gofas, S., Mahe, C., Noel, P., & Reviers, B. de., 1994. Typologie des ZNIEFF-Mer. Liste des parametres et des biocoenoses des cotes francaises metropolitaines. 2nd ed. Secretariat Faune-Flore, Museum National d'Histoire Naturelle, Paris (Collection Patrimoines Naturels, Serie Patrimoine Ecologique, No. 12). Coll. Patrimoines Naturels, vol. 12, Secretariat Faune-Flore, Paris.
Davison, D.M. & Hughes, D.J., 1998. Zostera biotopes: An overview of dynamics and sensitivity characteristics for conservation management of marine SACs, Vol. 1. Scottish Association for Marine Science, (UK Marine SACs Project)., Scottish Association for Marine Science, (UK Marine SACs Project),Vol. 1., http://www.english-nature.org.uk/uk-marine
Dawes, C.J. & Guiry, M.D., 1992. Proximate constituents in the seagrasses Zostera marina and Z. noltii in Ireland. Marine Ecology, 13, 307-315.
de Jonge, V.N.& de Jonge, D.J., 1992. Role of tide, light and fisheries in the decline of Zostera marina. Netherlands Institute of Sea Research Publications Series no. 20, pp. 161-176.
Delgado, O., Ruiz, J., Pérez, M., Romero, J. & Ballesteros, E., 1999. Effects of fish farming on seagrass (Posidonia oceanica) in a Mediterranean bay: seagrass decline after organic loading cessation. Oceanologica Acta, 22 (1), 109-117.
Den Hartog, C. & Phillips, R., 2000. Seagrasses and benthic fauna of sediment shores. In Reise, K. (ed.) Ecological Comparisons of Sedimentary Shores. Berlin: Springer, pp. 195-212.
Den Hartog, C., 1970. The sea-grasses of the world. Amsterdam: North Holland Publishing Company.
Den Hartog, C., 1987. "Wasting disease" another dynamic phenomena in Zostera beds. Aquatic Botany, 27, 3 -14.
Den Hartog, C., 1994. Suffocation of a littoral Zostera bed by Enteromorpha radiata. Aquatic Botany, 47, 21-28.
Eckrich, C.E. & Holmquist, J.G., 2000. Trampling in a seagrass assemblage: direct effects, response of associated fauna, and the role of substrate characteristics. Marine Ecology Progress Series, 201, 199-209.
Egerton, J., 2011. Management of the seagrass bed at Porth Dinllaen. Initial investigation into the use of alternative mooring systems. Report for Gwynedd Council, Gwynedd Council, Bangor.
Elliot, M., Nedwell, S., Jones, N.V., Read, S.J., Cutts, N.D. & Hemingway, K.L., 1998. Intertidal sand and mudflats & subtidal mobile sandbanks (Vol. II). An overview of dynamic and sensitivity for conservation management of marine SACs. Prepared by the Scottish Association for Marine Science for the UK Marine SACs Project.
Eno, N.C., Clark, R.A. & Sanderson, W.G. (ed.) 1997. Non-native marine species in British waters: a review and directory. Peterborough: Joint Nature Conservation Committee.
Erftemeijer, P.L. & Robin, L.R.R., 2006. Environmental impacts of dredging on seagrasses: A review. Marine Pollution Bulletin, 52 (12), 1553-1572.
Fernández Torquemada, Y., Lizaso, S. & Luis, J., 2006. Effects of salinity on growth and survival of Cymodocea nodosa (Ucria) Ascherson and Zostera noltii Hornemann Biologia Marina Mediterranea, 13, 46-47
Fish, J.D. & Fish, S., 1996. A student's guide to the seashore. Cambridge: Cambridge University Press.
Fishman, J.R. & Orth, R.J., 1996. Effects of predation on Zostera marina L. seed abundance. Journal of Experimental Marine Biology and Ecology, 198, 11-26.
Fonseca, M.S., Thayer, G.W., Chester, A.J. & Foltz, C., 1984. Impact of scallop harvesting on eelgrass (Zostera marina) meadows. North American Journal of Fisheries Management, 4 (3), 286-293.
Giesen, W.B.J.T., Katwijk van, M.M., Hartog den, C., 1990. Eelgrass condition and turbidity in the Dutch Wadden Sea. Aquatic Botany, 37, 71-95.
Gray, J.S. & Elliott, M., 2009. Ecology of marine sediments: from science to management, Oxford: Oxford University Press.
Greening, H. & Janicki, A., 2006. Toward reversal of eutrophic conditions in a subtropical estuary: Water quality and seagrass response to nitrogen loading reductions in Tampa Bay, Florida, USA. Environmental Management, 38 (2), 163-178.
Han, Q., Bouma, T.J., Brun, F.G., Suykerbuyk, W. & Van Katwijk, M., 2012. Resilience of Zostera noltii to burial or erosion disturbances. Marine Ecology Progress Series, 449.
Hawker, D., 1994. Solway Firth Zostera survey. Monitoring Report. , Scottish Natural Heritage, Aberdeen.
Hayward, P.J. 1994. Animals of sandy shores. Slough, England: The Richmond Publishing Co. Ltd. [Naturalists' Handbook 21.]
Hiscock, S., 1987. A brief account of the algal flora of Zostera marina beds in the Isle of Scilly. In Sublittoral monitoring in the Isles of Scilly 1985 & 1986 (ed. R. Irving). Nature Conservancy Council, Peterborough.
Hodges, J. & Howe, M., 1997. Milford Haven waterway monitoring of eelgrass, Zostera angustifolia, following the Sea Empress oils spill. Report to Shoreline & Terrestrial Task Group. Sea Empress Environmental Evaluation Committee,
Holden, P. & Baker, J.M., 1980. Dispersant-treated compared with untreated crude oil. Experiments with oil and dispersants on the seagrass Zostera noltii. Report to the Advisory Committee on Pollution of the Sea, Field Studies Council.
Holt, T.J., Hartnoll, R.G. & Hawkins, S.J., 1997. The sensitivity and vulnerability to man-induced change of selected communities: intertidal brown algal shrubs, Zostera beds and Sabellaria spinulosa reefs. English Nature, Peterborough, English Nature Research Report No. 234.
Hootsmans, M.J.M., Vermaat, J.E. & Vierssen, van W., 1987. Seed-bank development, germination and early seedling survival of two seagrass species from the Netherlands: Zostera marina L. and Zostera noltii Hornem. Aquatic Botany, 28 (3), 275-285
Howard, S., Baker, J.M. & Hiscock, K., 1989. The effects of oil and dispersants on seagrasses in Milford Haven. In Ecological Impacts of the Oil Industry,(ed. B. Dicks), pp. 61-96. Chichester: John Wiel & Sons Ltd. for the Institute of Petroleum, London.
Howie, D.I.D., 1959. The spawning of Arenicola marina (L.). I. The breeding season. Journal of the Marine Biological Association of the United Kingdom, 38, 395-406.
Hubbard, J. & Stebbings, R., 1967. Distribution, dates of origin, and acreage of Spartina townsendii (sl.) marshes in Great Britain. Proceedings of the Botanical Society of the British Isles, 7 (1), 1-7.
Hughes, A.R. & Stachowicz, J.J., 2004. Genetic diversity enhances the resistance of a seagrass ecosystem to disturbance. Proceedings of the National Academy of Sciences of the United States of America, 101 (24), 8998-9002.
Hughes, R.G., Lloyd, D., Ball, L., Emson, D., 2000. The effects of the polychaete Nereis diversicolor on the distribution and transplantation success of Zostera noltii. Helgoland Marine Research, 54, 129-136.
Huxham, M., Raffaelli, D. & Pike, A.W., 1995. The effect of larval trematodes on the growth and burrowing behaviour of Hydrobia ulvae (Gastropoda: Prosobranchia) in the Ythan estuary, N.E. Scotland. Journal of Experimental Marine Biology and Ecology, 185, 1-17.
Jackson, E.L., Griffiths, C.A., Collins, K. & Durkin , O., 2013. A guide to assessing and managing anthropogenic impact on marine angiosperm habitat - part 1: literature review. Natural England Commissioned Reports NERC111 Part I, Natural England and MMO Peterborough, UK. http://publications.naturalengland.org.uk/publication/3665058
Jackson, M.J. & James, R., 1979. The influence of bait digging on cockle Cerastoderma edule, populations in north Norfolk. Journal of Applied Ecology, 16, 671-679.
Jacobs, R.P.W.M., 1980. Effects of the Amoco Cadiz oil spill on the seagrass community at Roscoff with special reference to the benthic infauna. Marine Ecology Progress Series, 2, 207-212.
Jones, L.A., Hiscock, K. & Connor, D.W., 2000. Marine habitat reviews. A summary of ecological requirements and sensitivity characteristics for the conservation and management of marine SACs. Joint Nature Conservation Committee, Peterborough. (UK Marine SACs Project report.). Available from: http://www.ukmarinesac.org.uk/pdfs/marine-habitats-review.pdf
Katwijk van, M.M., Schmitz, G.H.W., Gasseling, A.P., & Avesaath van, P.H., 1999. Effects of salinity and nutrient load and their interaction on Zostera marina. Marine Ecology Progress Series, 190, 155-165.
Kenworthy, W.J., Fonseca, M.S., Whitfield, P.E. & Hammerstrom, K.K., 2002. Analysis of seagrass recovery in experimental excavations and propeller-scar disturbances in the Florida Keys National Marine Sanctuary. Journal of Coastal Research, 37, 75-85.
Koch, E.W., 2001. Beyond light: physical, geological, and geochemical parameters as possible submersed aquatic vegetation habitat requirements. Estuaries, 24 (1), 1-17.
Koch E.W., 2002. Impact of boat-generated waves on a seagrass habitat. Journal of Coastal Research, 37, 66-74.
Kuelan, van M., 1999. Human uses of seagrass. http://possum.murdoch.edu.au/~seagrass/seagrass_uses.html, 2000-01-01
Leuschner, C., Landwehr, S. & Mehlig, U., 1998. Limitation of carbon assimilation of intertidal Zostera noltii and Zostera marina by desiccation at low tide. Aquatic Botany, 62 (3), 171-176.
Levell, D., 1976. The effect of Kuwait Crude Oil and the Dispersant BP 1100X on the lugworm, Arenicola marina L. In Proceedings of an Institute of Petroleum / Field Studies Council meeting, Aviemore, Scotland, 21-23 April 1975. Marine Ecology and Oil Pollution (ed. J.M. Baker), pp. 131-185. Barking, England: Applied Science Publishers Ltd.
Madsen, J., 1988. Autumn feeding ecology of herbivorous wildfowl in the Danish Wadden Sea and impact of food supplies and shooting on migration. Danish Review of Game Biology, 13, 1-32.
Major, W.W., III, Grue, C.E., Grassley, J.M. & Conquest, L.L., 2004. Non-target impacts to eelgrass from treatments to control Spartina in Willapa Bay, Washington. Journal of Aquatic Plant Management, 42 (1), 11-17.
Marta, N., Cebrian, J., Enriquez, S. & Duarte, C.M., 1996. Growth patterns of western Mediterranean seagrasses: species specific responses to seasonal forcing. Marine Ecology Progress Series, 133, 203-215.
Martin, C.S., Vaz, S., Ernande, B., Ellis, J.R., Eastwood, P., Coppin, F., Harrop, S., Meaden, G. & Carpentier, A., 2005. Spatial distributions (1989-2004) and preferential habitats of thornback ray and lesser-spotted dogfish in the Eastern English Channel.
Massa, S.I., Arnaud-Haond, S., Pearson, G.A., Serrão, E.A. 2009. Temperature tolerance and survival of intertidal populations of the seagrass Zostera noltii (Hornemann) in Southern Europe (Ria Formosa, Portugal) Hydrobiologia 619,195–201
Mateo, M.A., Cebrián, J., Dunton, K. & Mutchler, T., 2006. Carbon flux in seagrass ecosystems. In Larkum, A.W.D., et al. (eds.). Seagrasses: biology, ecology and conservation, Berlin: Springer, pp. 159-192.
Maxwell, P.S., Pitt K.A., Burfeind, D.D., Olds, A.D., Babcock, R.C. & Connolly, R.M., 2014. Phenotypic plasticity promotes persistence following severe events: physiological and morphological responses of seagrass to flooding. Journal of Ecology, 102 (1), 54-64.
Milazzo, M., Badalamenti, F., Ceccherelli, G. & Chemello, R., 2004. Boat anchoring on Posidonia oceanica beds in a marine protected area (Italy, western Mediterranean): effect of anchor types in different anchoring stages. Journal of Experimental Marine Biology and Ecology, 299 (1), 51-62.
Montefalcone, M., Lasagna, R., Bianchi, C., Morri, C. & Albertelli, G., 2006. Anchoring damage on Posidonia oceanica meadow cover: a case study in Prelo Cove (Ligurian Sea, NW Mediterranean). Chemistry and Ecology, 22 (sup1), 207-S217.
Muehlstein, L., 1989. Perspectives on the wasting disease of eelgrass Zostera marina. Diseases of Aquatic Organisms, 7 (3), 211-221.
Nacken, M. & Reise, K., 2000. Effects of herbivorous birds on intertidal seagrass beds in the northern Wadden Sea. Helgoland Marine Research, 54, 87-94.
Neckles, H.A., Short, F.T., Barker, S. & Kopp, B.S., 2005. Disturbance of eelgrass Zostera marina by commercial mussel Mytilus edulis harvesting in Maine: dragging impacts and habitat recovery. Marine Ecology Progress Series, 285, 57-73.
Nejrup, L.B. & Pedersen, M.F., 2008. Effects of salinity and water temperature on the ecological performance of Zostera marina. Aquatic Botany, 88 (3), 239-246.
Neverauskas, V., 1987. Monitoring seagrass beds around a sewage sludge outfall in South Australia. Marine Pollution Bulletin, 18 (4), 158-164.
Newell, R.I. & Koch, E.W., 2004. Modeling seagrass density and distribution in response to changes in turbidity stemming from bivalve filtration and seagrass sediment stabilization. Estuaries, 27 (5), 793-806.
Nienhuis, P., 1996. The North Sea coasts of Denmark, Germany and the Netherlands. Berlin: Springer.
Ogilvie, M.A., & Matthews, G.V.T., 1969. Brent geese, mudflats and man. Wildfowl, 20, 110-125.
Olesen, B. & Sand-Jensen, K., 1993. Seasonal acclimation of eelgrass Zostera marina growth to light. Marine Ecology Progress Series, 94, 91-99.
Olsen, E.M., Heino, M., Lilly, G.R., Morgan, M.J., Brattey, J., Ernande, B. & Dieckmann, U. 2004. Maturation trends indicative of rapid evolution preceded the collapse of northern cod. Nature, 428, 932-935.
Orth, R.J. & Marion, S.R., 2007. Innovative techniques for large-scale collection, processing, and storage of eelgrass (Zostera marina) seeds. Engineer Research and Development Center Vicksburg, USA.
Orth, R.J., 1992. A perspective on plant-animal interactions in seagrasses: physical and biological determinants influencing plant and animal abundance. In Plant-Animal Interactions in the Marine Benthos, Systematics Association Special Volume no. 46, (ed. D.M. John, S.J. Hawkins & J.H. Price), pp. 147-164. Oxford: Clarendon Press.
Peralta, G., Bouma, T.J., van Soelen, J., Pérez-Lloréns, J.L. & Hernández, I., 2003. On the use of sediment fertilization for seagrass restoration: a mesocosm study on Zostera marina L. Aquatic Botany, 75 (2), 95-110.
Peralta, G., Brun, F.G., Perez-Llorens, J. & Bouma, T.J., 2006. Direct effects of current velocity on the growth, morphometry and architecture of seagrasses: a case study on Zostera noltii. Marine Ecology Progress Series, 327, 135.
Peralta, G., Pérez-Lloréns, J.L., Hernández, I. & Vergara, J.J., 2002. Effects of light availability on growth, architecture and nutrient content of the seagrass Zostera noltii Hornem. Journal of Experimental Marine Biology and Ecology, 269, 9-26.
Percival, S., Sutherland, W. & Evans, P., 1998. Intertidal habitat loss and wildfowl numbers: applications of a spatial depletion model. Journal of Applied Ecology, 35 (1), 57-63.
Percival, S.M. & Evans, P.R., 1997. Brent geese (Branta bernicla) and Zostera; factors affecting the exploitation of a seasonally declining food resource. Ibis, 139, 121-128.
Pergent, G., Mendez, S., Pergent-Martini, C. & Pasqualini, V., 1999. Preliminary data on the impact of fish farming facilities on Posidonia oceanica meadows in the Mediterranean. Oceanologica Acta, 22 (1), 95-107.
Perkins, E.J., 1988. The impact of suction dredging upon the population of cockles Cerastoderma edule in Auchencairn Bay. Report to the Nature Conservancy Council, South-west Region, Scotland, no. NC 232 I).
Peterson, C.H., Summerson, H.C. & Fegley, S.R., 1987. Ecological consequences of mechanical harvesting of clams. Fishery Bulletin, 85 (2), 281-298.
Philippart, C.J.M, 1994a. Interactions between Arenicola marina and Zostera noltii on a tidal flat in the Wadden Sea. Marine Ecology Progress Series, 111, 251-257.
Philippart, C.J.M, 1994b. Eutrophication as a possible cause of decline in the seagrass Zostera noltii of the Dutch Wadden Sea. http://www.nioz.nl/en/deps/mee/katja/seagrass.htm, 2000-10-23
Philippart, C.J.M, 1995b. Seasonal variation in growth and biomass of an intertidal Zostera noltii stand in the Dutch Wadden Sea. Netherlands Journal of Sea Research, 33, 205-218.
Phillips, R.C., McMillan, C. & Bridges, K.W., 1983. Phenology of eelgrass, Zostera marina L., along latitudinal gradients in North America. Aquatic Botany, 1 (2), 145-156.
Phillips, R.C., & Menez, E.G., 1988. Seagrasses. Smithsonian Contributions to the Marine Sciences, no. 34.
Prouse, N.J. & Gordon, D.C., 1976. Interactions between the deposit feeding polychaete Arenicola marina and oiled sediment. In Proceedings of a Symposium of the American Institute of Biological Sciences, Arlington, Virginia, 1976. Sources, effects and sinks of hydrocarbons in the aquatic environment, pp. 408-422. USA: American Institute of Biological Sciences.
Rasmussen, E., 1977. The wasting disease of eelgrass (Zostera marina) and its effects on environmental factors and fauna. In Seagrass ecosystems - a scientific perspective, (ed. C.P. McRoy, & C. Helfferich), pp. 1-51.
Reusch, T.B.H., Stam, W.T., & Olsen, J.C. 1998. Size and estimated age of genets in eelgrass, Zostera marina, assessed with microsatellite markers. Marine Biology, 133, 519-525.
Rhodes, B., Jackson, E.L., Moore, R., Foggo, A. & Frost, M., 2006. The impact of swinging boat moorings on Zostera marina beds and associated infaunal macroinvertebrate communities in Salcombe, Devon, UK. Report to Natural England. pp58, Natural England, Peterborough.
Rice, K.J. & Emery, N.C., 2003. Managing microevolution: restoration in the face of global change. Frontiers in Ecology and the Environment, 1 (9), 469-478.
Rodwell, J.S. (ed.), 2000. British plant communities, vol. 5, Maritime communities and vegetation of open habitats. Cambridge: Cambridge University Press.
Short, F., Davis, R., Kopp, B., Short, C. & Burdick, D., 2002. Site-selection model for optimal transplantation of eelgrass Zostera marina in the northeastern US. Marine Ecology Progress Series, 227, 253-267.
Short, F.T. & Burdick, D.M., 1996. Quantifying eelgrass habitat loss in relation to housing development and nitrogen loading in Waquoit Bay, Massachusetts. Estuaries, 19 (3), 730-739.
Short, F.T., 1987. The effects of sediment nutrients on seagrasses; literature review and mesocosm experiment. Aquatic Botany, 27, 41-57.
Short, F.T., Muehlstein, L.K. & Porter, D., 1987. Eelgrass wasting disease: cause and recurrence of a marine epidemic. The Biological Bulletin, 173 (3), 557-562.
Smit, C.J. & Visser, G.J.M., 1993. Effects of disturbance on shorebirds: a summary of existing knowledge from the Dutch Wadden Sea and Delta area. Wader Study Group Bulletin, 68 (special issue).
Smith, J.E. (ed.), 1968. 'Torrey Canyon'. Pollution and marine life. Cambridge: Cambridge University Press.
Stieglitz, W.O., 1966. Utilization of available foods by diving ducks on Apalachee Bay, Florida. Proceedings of the Southeastern Association of Game and Fish Commissioners, 20, 42-50.
Suchanek, T.H., 1993. Oil impacts on marine invertebrate populations and communities. American Zoologist, 33, 510-523.
Sutton, A. & Tompsett, P.E., 2000. Eelgrass (Zostera spp.) Project 1995-1998. A report to the Helford Voluntary Marine Conservation Area Group funded by World Wide Fund for Nature UK and English Nature.
Touchette, B.W., 2007. Seagrass-salinity interactions: physiological mechanisms used by submersed marine angiosperms for a life at sea. Journal of Experimental Marine Biology and Ecology, 350 (1), 194-215.
Touchette, B.W. & Burkholder, J.M., 2000. Review of nitrogen and phosphorus metabolism in seagrasses. Journal of Experimental Marine Biology and Ecology, 250 (1), 133-167.
Tu Do, V., de Montaudouin, X., Blanchet, H. & Lavesque, N., 2012. Seagrass burial by dredged sediments: Benthic community alteration, secondary production loss, biotic index reaction and recovery possibility. Marine Pollution Bulletin, 64 (11), 2340-2350.
Tubbs, C.R. & Tubbs, J.M., 1983. The distribution of Zostera and its exploitation by wildfowl in the Solent, southern England. Aquatic Botany, 15, 223-239.
Valentine, J.F. & Heck Jr, K.L., 1991. The role of sea urchin grazing in regulating subtropical seagrass meadows: evidence from field manipulations in the northern Gulf of Mexico. Journal of Experimental Marine Biology and Ecology, 154 (2), 215-230.
Van der Heide, T., van Nes, E.H., Geerling, G.W., Smolders, A.J., Bouma, T.J. & van Katwijk, M.M., 2007. Positive feedbacks in seagrass ecosystems: implications for success in conservation and restoration. Ecosystems, 10 (8), 1311-1322.
Van Duin, E.H., Blom, G., Los, F.J., Maffione, R., Zimmerman, R., Cerco, C.F., Dortch, M. & Best, E.P., 2001. Modeling underwater light climate in relation to sedimentation, resuspension, water quality and autotrophic growth. Hydrobiologia, 444 (1-3), 25-42.
Vermaat, J.E., Agawin, N.S.R., Fortes, M.D., Uri, J.S., Duarte, C.M., Marbà, N., Enríquez, S. & Vierssen van, W., 1997. The capacity of seagrasses to survive increased turbidity and siltation: the significance of growth form and light use. Ambio, 26 (8), 499-504.
Vermaat, J.E., Verhagen, F.C.A. & Lindenburg, D., 2000. Contrasting responses in two populations of Zostera noltii Hornem. to experimental photoperiod manipulation at two salinities. Aquatic Botany, 67, 179-189.
Walker, D., Lukatelich, R., Bastyan, G. & McComb, A., 1989. Effect of boat moorings on seagrass beds near Perth, Western Australia. Aquatic Botany, 36 (1), 69-77.
Wall, C.C., Peterson, B.J. & Gobler, C.J., 2008. Facilitation of seagrass Zostera marina productivity by suspension-feeding bivalves. Marine Ecology Progress Series, 357, 165-174.
Williams, S.L., 1988. Disturbance and recovery of a deep-water Caribbean seagrass bed. Marine Ecology Progress Series, 42 (1), 63-71.
Williams, S.L., 2001. Reduced genetic diversity in eelgrass transplantations affects both population growth and individual fitness. Ecological Applications, 11 (5), 1472-1488.
Williams, S.L. & Davis, C.A., 1996. Population genetic analyses of transplanted eelgrass (Zostera marina) beds reveal reduced genetic diversity in southern California. Restoration Ecology, 4 (2), 163-180.
Williams, T.P., Bubb, J.M., & Lester, J.N., 1994. Metal accumulation within salt marsh environments: a review. Marine Pollution Bulletin, 28, 277-290.
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Last Updated: 14/08/2015
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