|Researched by||Emilia d'Avack, Dr Harvey Tyler-Walters & Catherine Wilding||Refereed by||Dr Leigh Jones|
Expanses of clean or muddy fine sand in shallow water and on the lower shore (typically to about 5 m depth) can have dense stands of Zostera marina/angustifolia [Note: the taxonomic status of Zostera angustifolia is currently under consideration]. In Zmar the community composition may be dominated by these Zostera species and therefore characterized by the associated biota. Other biota present can be closely related to that of areas of sediment not containing Zostera marina, for example, Saccharina latissima, Chorda filum and infaunal species such as Ensis spp. and Echinocardium cordatum (e.g. Bamber, 1993) and other bivalves listed below. It should be noted that sparse beds of Zostera marina may be more readily characterized by their infaunal community. Beds of this biotope in the south-west of Britain may contain conspicuous and distinctive assemblages of Lusitanian fauna such as Laomedea angulata, Hippocampus spp. and Stauromedusae. Some examples of Zostera marina beds have markedly anoxic sediments associated with them. (Information taken from the Marine Biotope Classification for Britain and Ireland, Version 97.06: Connor et al., 1997a, b).
|Depth Range||Lower shore, 0-5 m, 5-10 m|
|Water clarity preferences||No information found|
|Limiting Nutrients||Nitrogen (nitrates), Phosphorous (phosphates)|
|Salinity preferences||Full (30-40 psu), Variable (18-40 psu)|
|Physiographic preferences||Enclosed coast / Embayment|
|Biological zone preferences||Upper infralittoral|
|Substratum/habitat preferences||Mud, Mud and sandy mud, Muddy sand, Sand, Sand and muddy sand|
|Tidal strength preferences||Moderately Strong 1 to 3 knots (0.5-1.5 m/sec.), Very Weak (negligible), Weak < 1 knot (<0.5 m/sec.)|
|Wave exposure preferences||Extremely sheltered, Moderately exposed, Sheltered, Very sheltered|
Intertidal Zostera marina beds may be damaged by frost, although rhizomes most likely survive. In carbonate based sediments phosphate may be limiting due to adsorption onto sediment particles. Zostera marina is also found in reduced salinities, for example brackish lagoons (Dyrynda, 1997).
Zostera marina 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 by to characterize the sensitivity of this biotope (d'Avack et al., 2014) . Zostera marina is not dependent on associated species to create or modify habitat, provide food or other resources. The sensitivity assessments are thus based on Zostera marina 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.
There is a current debate whether Zostera angustifolia is a distinct species, a variant of Zostera marina or synonym of Zostera marina. Neither Zostera angustifolia nor Zostera marina var angustifolia are accepted taxonomic names (WorMS, 2015) The current consensus is that Zostera angustifolia is a taxonomic synonym of Zostera marina. Van Lent & Verschuure (1994) suggest that there is a continuum of life history strategies exhibited by Zostera marina for survival in a wider range of environments. Any observed differences in terms of morphology and life history are thus likely to be adaptations to different habitats.
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.
Zostera marina 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 recolonization 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 ration 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. Different populations will thus have different resilience to external pressures. Boese et al. (2009) examined the recolonization of gaps created experimentally within Zostera marina beds. The study looked at two zones, the lower intertidal covered with almost continuous seagrass and an upper intertidal transition zone where there were patches of perennial and annual Zostera marina. Recovery started within a month after the disturbance of the lower intertidal continuous perennial beds and was complete after two years. Plots in the transition zone, however, took almost twice as long to recover.
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). Zostera marina can tolerate temperatures between -1 to 25°C with optimum conditions for growth being around 10 to 15°C, and 10°C for seedling development (Hootsmans et al., 1987). Nejrup & Pedersen (2007) found that temperatures between 25 and 30°C lowered photosynthetic rates by 50% as well as growth (production of new leaves by 50% and leaf elongation rate by 75%). High temperatures also resulted in a 12-fold increase in mortality of Zostera marina plants. Moore et al. (2014) found that short-term exposures to a rapid increase of 4–5°C above normal temperature (25°C) during summer months resulted in widespread diebacks of Zostera marina. Recovery was observed to be minimal as the seagrass was replaced by Ruppia maritima. Similarly, Salo & Peterson (2014) found that exposure to high temperature for 5 weeks led to enhanced mortality, reduced formation of new leaves and a lower number of standing leaves per shoot.
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. High temperatures during hot summer months have caused massive die off events among seagrasses worldwide (Moore & Jarvis, 2008; Reusch et al., 2005). A 5°C change in temperature over one month or a 2°C change over the period of a year is thus likely to result in some Zostera marina mortality. Resistance is therefore considered ‘Medium’. Recovery will be fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. If however, temperatures remain elevated for a prolonged period of time, Zostera marina can be outcompeted and subsequently excluded from the habitat by other species such as Ruppia maritima. The biotope is considered to have a ‘Medium’ sensitivity to an increase in temperature at the pressure benchmark.
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). Zostera marina can tolerate temperatures between -1 to 25°C with optimum conditions for growth being around 10 to 15°C, and 10°C for seedling development (Hootsmans et al., 1987). Nejrup & Pedersen (2007) found that low water temperatures (5°C) slowed down the photosynthetic rate by 75%; growth was also affected, with the production of new leaves reduced by 30% and leaf elongation rate reduced by 80% compared to the control, however, mortality was not affected.
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 change 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, although it may be replaced by other grazers. Healthy populations of epiphyte grazers are therefore essential to the maintenance of seagrass beds.
Sensitivity assessment. Overall, a decrease in temperature is likely to reduce growth rates but not to cause mortality directly. Frost damage could occur to plants exposed at extreme low tides in the winter months but as the seagrass dies back in winter this is unlikely to be significant. Therefore, a 5°C decrease in temperature over one month or a 2°C decrease over the period of a year is thus unlikely to result in some Zostera marina mortality. Resistance is therefore considered ‘High’. Recovery will be rapid once conditions return to normal resulting in a ‘High’ resilience score. Hence, the biotope is considered be ‘Not sensitive’ to a decrease in temperature at the pressure benchmark.
In general, seagrass species have a wide salinity tolerance. Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt. Hypersaline conditions can affect seagrass performance as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007). Den Hartog (1997) stated that Zostera noltei has a greater tolerance to extremes salinities compared to Zostera marina due to its intertidal habitat. Vermaat et al. (2000) investigating salinity tolerance in Zostera noltei found considerable mortalities of plants at a salinity of 35 ppt. These findings suggest that both Zostera species are ill-equipped to withstand high saline conditions. A review by 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 plants display a wide tolerance to a range of salinities, an increase from 35 to 38 units for the period of one year will cause some mortality in Zostera marina. The subtidal habitat makes the species more vulnerable to salinity extremes compared to the intertidal Zostera noltei resulting in a ‘Low’ resistance score. Zostera marina will thus be adversely affected by activities such as brine discharges from seawater desalination plant. Recovery, enabled by recolonization from surrounding communities, will be fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure 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 salinities as low as 5 ppt. Hyposaline conditions (reduced salinity) can, however, affect seagrass performance as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007).
A study by Salo et al. (2014) found that hyposaline conditions can seriously impair plant performance and survival rates. The study determined that the severity of impact will be population specific as seagrass populations from different areas may substantially differ in their salinity tolerance range with population naturally occurring in low saline areas having greater resistance to this pressure.
Salo & Petersen (2014) experimentally tested the effects of different combinations of salinity and temperature on the physiological performance of Zostera marina. The study found that the combination of high temperature and low salinity resulted in high mortality rates, highlighting negative synergistic effects when seagrasses are exposed to multiple pressures.
A review by d’Avack et al. (2014) determined 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. Zostera marina has a wide salinity tolerance. Reduced salinity will, however, impact performance causing some mortality. Resistance is therefore considered ‘Medium’. Effects can be exasperated when the seagrass is exposed to multiple stressors at the same time, highlighting the importance to consider negative synergistic effects when conduction assessments. Recovery is considered fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure at the pressure benchmark.
A complex interaction exists between seagrass beds and water flow. Water flow determines the upper distribution of plants on the shore whilst plants mediate the velocity of the flow by extracting momentum from the moving water. Reducing the flow increases water transparency (see ‘changes in suspended sediments’ pressure) and causes the deposition and retention of fine sediments. Increased flow rates, on the other hand, are 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 re-suspension 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 sea bed 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. Fonseca et al. (1983) found a lower maximum for Zostera marina and estimated the highest current velocity at approximately 120–150 cm/s.
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. Aquaculture installations can also change water flow and have shown to directly impact seagrass habitats. Everett et al. (1995) experimentally altered water flow to investigate the effects of the commercial culture of the oyster Magallana gigas on Zostera marina, using both stake and rack methods. The study found that both culture methods caused a sharp decline in Zostera marina plants with cover being less than 25% compared to control plots after one year of culture due to changes in local hydrological regime. Both culture methods produced strong, although dissimilar, changes in local hydrological conditions, which had clear effects on sediment characteristics. In general, stakes resulted in local sediment deposition while racks produced local erosion, both leading to the reduction and eventual death of nearby seagrass beds.
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 demonstrate higher resistance to desiccation than Zostera marina which occurs more frequently in the subtidal. To understand the differences in desiccation tolerance between the two 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 Zostera 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.
Tolerances vary not only between species but also within species. For instance, annual and perennial forms of Zostera marina were observed to tolerate desiccation to different extents. Van Katwijk & Hermus (2000) noted that in intertidal areas of the Wadden Sea, annual Zostera marina plants tended to lie flat on the moist sediment when exposed at low tide. Perennial plants, on the other hand, had stiffer stems inhibiting contact with the sediment. These upright sheaths desiccate more rapidly when exposed. Morphology is, therefore, a factor partly determining tolerance to desiccation. The same phenomenon was observed by Boese et al. (2003) on Zostera marina in Aquinas Bay, USA.
The overall low tolerance of seagrass species to aerial exposure means that an increase in tidal amplitudes could force seagrass to grow deeper where there was less chance of exposure to the air. As the depth limit of seagrasses is set by light penetration, this change is likely to reduce the extent of suitable habitat. Changes in seagrass distribution along a depth gradient will have an impact further down the food chain.
Sensitivity assessment. Sensitivity to changes in emergence regimes varies between species and habitats. Species growing in intertidal habitats have greater tolerance to exposure to air than species inhabiting subtidal beds. The resistance of Zostera marina to this pressure is therefore assessed as ‘Low’. 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 score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure 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 the plants in several ways. Seagrasses are not robust. Strong waves can cause mechanical damage to leaves and rhizomes. 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 re-suspension 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 for the sensitivity of the beds 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 and is considered to be fairly rapid scoring a ‘Medium’ resilience. The biotope, therefore, scores a ‘Medium’ sensitivity to changes in wave exposure at the pressure benchmark.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available.
Zostera marina is known to accumulate TBT but no detrimental effects were observed in the field (Williams et al., 1994). Naphthalene, Pentachlorophenol, Aldicarb and Kepone reduce nitrogen fixation and may affect Zostera viability. TBT contamination is likely to adversely affect grazing gastropods resulting in increased algal growth, reduced primary productivity and potential smothering of the biotope.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available.
Zostera marina may be partially protected from direct contact with oil due to its subtidal habitat. Healthy populations of Zostera can occur in the presence of long-term, low level, hydrocarbon effluent, for example in Milford Haven, Wales (Hiscock, 1987). The Amoco Cadiz oil spill off Roscoff caused Zostera marina leaves to blacken for 1-2 weeks but had little effect on growth, production or reproduction after the leaves were covered in oil for six hours (Jacobs, 1980). The Amoco Cadiz oil spill did, however, result in the virtual disappearance of Amphipods, Tanaidacea and Echinodermata from Zostera marina beds and caused a decrease in numbers of Gastropoda, sedentary Polychaeta and Bivalvia. The numbers of most groups returned to normal within a year except Echinoderms which recovered more slowly and amphipods which did not show any signs of recovery (Jacobs, 1980). Removal of oil intolerant gastropod grazers may result in smothering of seagrasses by epiphytes (Davison & Hughes, 1998). Jacobs (1980) noted a larger algal bloom than in previous years after the Amoco Cadiz spill in Roscoff, probably as a result in increased nutrients (from dead organisms and breakdown of oil) and the reduction of algal grazers. However, herbivores recolonized and the situation returned to 'normal' within a few months.
Experimental treatment of Zostera sp. with crude oil and dispersants halted growth but had little effect on cover whereas pre-mixed oil and dispersant caused rapid death and significant decline in cover within 1 week suggesting that dispersant treatments should be avoided (Davison & Hughes, 1998).
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available.
Little information on the impacts of synthetic compounds on Zostera species is present in the literature. Triazine herbicides (e.g. Irgarol) inhibit photosynthesis and sublethal effects have been detected. Terrestrial herbicides may also damage seagrass beds. For example, the herbicide Atrazine is reported to cause growth inhibition and 50 % mortality in Zostera marina exposed to 100 ppb (ng/ l) Atrazine for 21 days (Davison & Hughes, 1998). Bester (2000) noted a correlation between raised concentrations of 4 triazine herbicides and areas where Zostera plants had been lost.
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
No evidence found.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed.
The effects of oxygen concentration on the growth and survivability of Zostera marina are not reported in the literature. Zostera sp. leaves contain air spaces (lacunae). Oxygen is transported to the roots where it permeates into the sediment, resulting in an oxygenated microzone, enhancing 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 deoxygenation, especially if combined with low light penetration and hence reduced photosynthesis will have an adverse 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 (benchmark). Loss of grazers will result in an unchecked growth of epiphytes and other algae which may smother Zostera marina.
Sensitivity assessment. Overall de-oxygenation is not likely to adversely affect seagrass beds. The loss of grazing gastropods could result in smothering and potential reduction in the extent of the seagrass. At the level of the benchmark, both resistance and resilience are assessed as 'High' (no impact to recover from). Overall the biotope is therefore 'Not Sensitive' to de-oxygenation at the pressure benchmark.
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 stressed 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 pressure on ‘changes in suspended solids’).
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 caused 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 and Neverauskas (2004) determined that 8 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 marina 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 resulting in 'Very Low' resilience. 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. Coarser sediments reduce the vegetative spreading of seagrasses and inhibit seedling colonization (Gray & Elliott, 2009). Changes in sediment type can, therefore, have wider implications on the distribution of seagrass beds. Hence, 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.
Sensitivity assessment. 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 ‘abrasion’ and ‘penetration and/or disturbance of the substratum below the surface of the seabed’ pressures). Resistance is therefore assessed as ‘None’ and resilience is considered ‘Very Low’ resulting in a ‘High’ sensitivity score. The confidence assessment for this pressure is high as it is based on the characteristics of the pressure i.e. complete removal of the feature within the pressure footprint.
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. Evidence for abrasion impacts is summarised below for activities that give rise to this pressure.
Trampling: 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 the 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.
Boating activities: boats passing in close proximity to seagrass beds can create waves. Turbulence 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 can 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. Hall 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 widgeon (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 1 cm 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 a 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. 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. 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 in winter likely to have less impact than the damage that occurs 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. The leaves and stems of seagrass plants rise above the surface and the roots are shallowly buried. 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. Abrasion to the substratum to a depth greater than 5 cm will directly impact seagrass habitats and all biomass (leaves, rhizomes) will be completely removed leading to the death of the plant in the area impacted. 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. The extent of the damage on seagrass beds depends on activity.
Clam digging and clam raking: 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 causing 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 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: bottom trawling and dragging are industrial fishing methods which scour the seabed to collect target species. 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 marina 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 Zostera marina was in its 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 soft bottom. However, an increase in dredging effort (30 dredges) led to a significant reduction in Zostera marina 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 but, 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 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) determined that Zostera noltei was 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, plants are likely to be more intolerant to chronic increases in turbidity. 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 due to reduced photosynthesis caused by 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. Similar response to reduced light availability for Zostera marina was observed by Moore & Wetzel (2000).
Increases in turbidity over a prolonged period of time are therefore highly likely to impact seagrass species. Sensitivity 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). The growth of both Zostera marina and its associated epiphytes are reduced by increased shading due to turbidity (reduction of light penetration by 42, 28 and 9%). Backman & Barilotti (1976) further established that intensive shading (reduction of light penetration by 63%) inhibited flowering in Zostera marina plants.
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 marina 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. Vermaat et al. (1997) found that adult Zostera marina in the Dutch Wadden Sea was able to cope with sedimentation rates between 2 and 13 cm per year as the plant has the capacity to elongate vertical stems enabling it to raise the leaf canopy above the sediment load. A study in the USA, however, observed a mortality of over 50% of plants of Zostera marina in field burial treatments of 4 cm (corresponding to 25% of plant height) for 24 days (Mills & Fonseca, 2003). Plants buried 75% or more of their height (16 cm) experienced 100% mortality indicating a low resistance of Zostera marina to burial. The differences observed between these two studies were probably caused by different phenotypes adapted to local conditions.
The timing of the siltation event also plays a role in particular for intertidal beds. At low tide, 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.
Sensitivity assessment. Above studies suggest that Zostera marina 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 'Low' 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 marina is intolerant of smothering by excessive siltation (see above). Seagrasses can cope with small rates of sedimentation by relocating their rhizomes closer to the sediment surface (Vermaat et al., 1997). Mills & Fonseca (2003) however observed 100% mortality in Zostera marina plants buried at a depth of 16 cm. Resistance to sedimentation at the pressure benchmark (30 cm of added material) is therefore assessed as ‘None’ as all individuals exposed to siltation are predicted to die and consequent resilience as ‘Low’ to ‘Very Low’. 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. 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
An increase in light might be beneficial while shading by artificial structures will decrease incident light and hence reduce photosynthesis and growth rates. For example, in mesocosm experiments, Frederick et al. (1995) noted that shading (at 11, 21, 41, 61, and 94% of incident surface light for one week) resulted in a reduction in shoot density and an increase in shoot height. But shading alone did not cause mortality in the experimental time frame. Holmer & Laursen (2002) noted that shading affected Zostera marina from a low-light, organic rich sediment population more than light saturated, low-organic sediment population. However, the effects were significant in spring but not in autumn, and were also related to the plant's ability to tolerant anoxic and sulfidic conditions.
Overall, there is little evidence of seagrass mortality resulting from shading directly but the effects of shading and smothering from epiphytes and macroalgae are discussed under nutrient enrichment and the effects of light attenuation under 'water clarity' above. However, 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 'Low' and sensitivity of 'High' 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|
Translocation of seagrass seeds, rhizomes and seedlings is a common practice globally to counter the trend of decline of seagrass beds. However, Williams & Davis (1996) found that levels of genetic diversity of restored Zostera marina beds in Baja California, USA, were significantly lower than in natural populations. A subsequent study by Williams (2001) determined that the observed genetic bottleneck was a consequence of the collection protocol of source material (i.e. founder effect). Founder effects are likely to occur if seeds used to revegetate restoration sites are collected from a limited number of sources. Similar to episodes of colonization, the ‘founding’ propagules can represent only a portion of the genetic diversity present in the source populations, and they might hybridize with local genotypes (Hufford & Mazer, 2003). 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) affirms that genetic variation is 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 a markedly reduced community resilience than natural populations (Hughes & Stachowicz, 2004).
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, resistance and resilience to this pressure are thus considered to be ‘High’ (no impact to recover from). Overall the biotope is therefore 'Not Sensitive' to this pressure. However, if hybridization occurred, recovery would not be considered possible unless the population is eradicated and replaced. In this case, resilience is thus deemed ‘Very Low’ resulting in an overall ‘Low’ sensitivity score. As there is no direct evidence to support assessments, these are based on expert judgement.
The effects on native species on seagrass species were 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. The (potential) impact of each invasive non-indigenous species (INIS) is reported below.
Non-native invasive plants: among the INIS currently present in the UK, the large brown seaweed Sargassum muticum has the most direct impact on Zostera species. Druehl (1973) was the first to raise concern about the potential negative effects of S. muticum on Zostera beds in British waters. Zostera and Sargassum muticum were thought to be spatially separated due to their preferred habitat. Zostera species grow on sand and muddy bottoms, whereas Sargassum muticum attaches to solid substratum. However, when the seabed consists of a mixed substratum of sand, gravel and stones both species may occur together. Even though there are no indications of direct competition between the two species (Den Hartog, 1997), Sargassum muticum establishes itself within seagrass habitats where beds are retreating due to natural or anthropogenic causes. The invasive seaweed almost immediately occupies the empty spaces thereby interfering with the natural regeneration cycle of the bed. In addition, a study in Salcombe, SW England by Tweedley et al. (2008) demonstrated that the presence of Zostera marina may help the attachment of Sargassum muticum on soft substrata by trapping drifting fragments thereby allowing viable algae spores to settle on the seagrass matrix in an otherwise unfavourable environment. Once the invasive seaweed establishes itself, Zostera marina is unable to regain the lost territory indicating that eventually, Sargassum muticum is able to replace seagrass beds particularly on mixed substratum (Den Hartog, 1997).
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) colonises 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.
The invasive green algae Codium fragile ssp. tomentosoides, now found throughout Britain has been reported to occur in habitats dominated by Zostera marina (Gabary et al.,1997). It was initially thought that Zostera out-competes Codium at high Zostera densities (Malinowski & Ramus, 1973). But a study by Gabary et al. (2004) in Canada found that the invasive alga has morphological adaptations that allow it to compete with Zostera even in healthy eelgrass beds. Codium fragile ssp. tomentosoides have a wide salinity tolerance 12 to 40 ppt and are thus a concern to biotopes in full as well as in reduced salinity. However, direct ecological impacts remain unknown and no quantitative evidence is available to assess resistance at the benchmark.
Non-native invasive invertebrates: benthic macroinvertebrates can have a significant impact on seagrass beds, by either influencing abundance through seed herbivory (Fishman & Orth, 1996) or by influencing seed germination and seedling development by affecting vertical distribution of seeds. Some species have a positive effect by burying seeds to shallow depths and thereby reducing seed predation and facilitating seed germination whilst other species bury seeds too deep to allow germination. The invasive polychaete Marenzelleria viridis, a species naturally occurring on the east coast of North America but introduced Europe via transport in ballast waters, was recorded to directly impact seed banks of Zostera marina beds in its new territory (Delefosse & Kristensen, 2012). The study carried out on the island of Fyn, Denmark, determined that the impact of Marenzelleria viridis on seagrass beds depended on the abundance of worms within a bed. Negative effects were only observed at high abundances (1600 individual per m2) causing seeds to be buried too deep to germinate. However, the study by Delefosse & Kristensen (2012) is the only publication on the impact of this particular invasive species on seagrass beds, and more evidence is needed in order to determine the ecological implications of this introduced polychaete in UK waters.
The invasive tunicate Didemnum vexillum has been reported growing on stalks and blades of Zostera marina plants in New England, USA (Carman & Grunden, 2010). The ecological effects of invasive tunicates introduced to seagrass beds remain unassessed, but in general terms, introduced epibionts have been shown to have negative effects on marine flora (Williams, 2007). Their considerable weight combined with their rapid asexual and sexual reproduction and an absence of predators (Carman et al., 2009) make them a considerable threat to marine plant communities as they increase the risk of smothering. The absence of predators could be related to anti-fouling microbial compounds present in Didemnum vexillum (Tait et al., 2007). Although the direct effect of invasive tunicates on seagrass remains unknown and no records of Didemnum vexillum growing on Zostera plants in the UK exist yet, there are concerns about possible negative interactions. No quantitative evidence regarding the level of impact has been found to assess this pressure.
Other invasive species could affect seagrass beds via indirect pathways. For instance, the Atlantic oyster drill Urosalpinx cinerea, a small predatory sea snail is unlikely to have a direct effect on seagrass beds but by preying on mussels and other bivalves, the sea snail could be responsible for a drop in water clarity which in turn will affect Zostera species (see sections below on changes in suspended solids). The invasive Pacific oyster Magallana gigas can also have negative effects. Oysters physically alter their environment by increasing habitat complexity and altering water flow and causing sulphide to accumulate in the sediment. Sulphide is toxic to eelgrass and a decline in Zostera marina as a consequence of invasive oyster growth was observed in British Columbia, Canada (Kelly & Volpe, 2007). The authors did not state the level of effect quantitatively and therefore the level of impact in terms of the resistance benchmarks used in this study is not clear.
Sensitivity assessment. Invasive species are affecting seagrass habitats around the UK with invasive flora having the greatest impact on seagrass beds so far recorded. However, there are extensive knowledge gaps on how invasive species influence the health of Zostera beds in UK waters. More research is 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 INIS 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 the eelgrass Zostera marina 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 continues to affect Zostera marina in temperate regions with variable degrees of losses but not to the extent of an epidemic (Short et al., 1988). The exact conditions responsible for an outbreak are still unknown but it has been shown that already weakened plants are more susceptible to infection (Tutin, 1938; Rasmussen, 1977) and that salinity plays a role the pathogen activity (Muehlstein et al., 1988).
Sensitivity assessment. Zostera marina is highly susceptible to microbial pathogens, which were in the past responsible for important reductions in seagrass populations. A sensitivity of ‘High’ has been recorded (‘Low’ resistance, ‘Low’ resilience)
Seagrass is not a species targeted by 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. However, in the USA, a mechanical seed harvesting technique was invented and put into practice (Orth & Marion, 2007). The mechanised harvester is 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’.
Harvesting of seagrasses as craft material is a small but growing, industry. 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). Bivalves also contribute pseudofaeces to fertilize seagrass sediments (Bradley & Heck Jr, 1999).
Seagrass plant may be directly removed or damaged by static or mobile gears that target 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 ‘None’. Studies of the effects of wildfowl grazing (see resilience and recovery above) suggest that recovery from the removal of target species will be rapid resulting in 'Medium' 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 ‘Medium’ 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 pressure ‘penetration and/or disturbance of the substratum’. Seagrass plants and the sedimentary habitat 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.
Incidental removal of the key characterizing seagrass species and associated species would alter the character of the biotope. The biotope is characterized by the presence of beds of seagrass, these provide habitat structure and attachment surfaces for epiphytic species. These may also modify local habitats through changes in water flow and the trapping of sediments. The loss of the turf due to incidental removal as by-catch would, therefore, alter the character of the habitat and result in the loss of habitat structure and species richness. The ecological services such as primary and secondary production and habitat engineering provided by seagrass and the associated species would also be lost.
Sensitivity assessment. Incidental removal of seagrass as by-catch would be detrimental, altering the character of the biotope and removing the habitat structure, and could lead to reclassification of the biotope where extensive removal occurs. 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.
Anonymous, 1999p. Seagrass beds. Habitat Action Plan. In UK Biodiversity Group. Tranche 2 Action Plans. English Nature for the UK Biodiversity Group, Peterborough., English Nature for the UK Biodiversity Group, Peterborough.
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.
Backman, T. & Barilotti, D., 1976. Irradiance reduction: effects on standing crops of the eelgrass Zostera marina in a coastal lagoon. Marine Biology, 34 (1), 33-40.
Bamber, R.N., 1993. Changes in the infauna of a sandy beach. Journal of Experimental Marine Biology and Ecology, 172, 93-107.
Bester, K., 2000. The effects of pesticides on seagrass beds. Helgoland Marine Research, 54, 95-98.
Biebl, R. & McRoy, C., 1971. Plasmatic resistance and rate of respiration and photosynthesis of Zostera marina at different salinities and temperatures. Marine Biology, 8 (1), 48-56.
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., Alayan, K.E., Gooch, E.F. & Robbins, B.D., 2003. Desiccation index: a measure of damage caused by adverse aerial exposure on intertidal eelgrass (Zostera marina) in an Oregon (USA) estuary Aquatic Botany, 76, 329-337
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.
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.
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.
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.
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.
Carman, M.R. & Grunden, D.W., 2010. First occurrence of the invasive tunicate Didemnum vexillum in eelgrass habitat. Aquatic Invasions, 5 (1), 23-29.
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.
Connor, D.W., Dalkin, M.J., Hill, T.O., Holt, R.H.F. & Sanderson, W.G., 1997a. Marine biotope classification for Britain and Ireland. Vol. 2. Sublittoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06.
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.
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.
Davies, C.E. & Moss, D., 1998. European Union Nature Information System (EUNIS) Habitat Classification. Report to European Topic Centre on Nature Conservation from the Institute of Terrestrial Ecology, Monks Wood, Cambridgeshire. [Final draft with further revisions to marine habitats.], Brussels: European Environment Agency.
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
Delefosse, M. & Kristensen, E., 2012. Burial of Zostera marina seeds in sediment inhabited by three polychaetes: Laboratory and field studies. Journal of Sea Research, 71, 41-49.
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., 1997. Is Sargassum muticum a threat to eelgrass beds? Aquatic Botany, 58 (1), 37-41.
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.
Druehl, L.D., 1973. Marine transplantations. Science, 179 (4068), 12.
Dyrynda, P.E.J., 1997. Seasonal monitoring of the Fleet Lagoon aquatic ecosystem (Dorset UK): 1995-1996. Report to the World Wildlife Fund UK from Marine Environmental Research Group, University of Wales, Swansea.
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.
Erftemeijer, P.L. & Robin, L.R.R., 2006. Environmental impacts of dredging on seagrasses: A review. Marine Pollution Bulletin, 52 (12), 1553-1572.
Everett, R.A., Ruiz, G.M. & Carlton, J., 1995. Effect of oyster mariculture on submerged aquatic vegetation: an experimental test in a Pacific Northwest estuary. Marine Ecology Progress Series, 125 (1), 205-217.
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. & Bell, S.S., 1998. Influence of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Marine Ecology Progress Series, 171, 109.
Fonseca, M.S., 1992. Restoring seagrass systems in the United States. In Restoring the Nation's Marine Environment (ed. G.W. Thayer), pp. 79 -110. Maryland: Maryland Sea Grant College.
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.
Fonseca, M.S., Zieman, J.C., Thayer, G.W. & Fisher, J.S., 1983. The role of current velocity in structuring eelgrass (Zostera marina L.) meadows. Estuarine, Coastal and Shelf Science, 17 (4), 367-380.
Garbary, D., Vandermeulen, H. & Kim, K., 1997. Codium fragile ssp. tomentosoides (Chlorophyta) invades the Gulf of St Lawrence, Atlantic Canada. Botanica Marina, 40 (1-6), 537-540.
Garbary, D.J., Fraser, S.J., Hubbard, C. & Kim, K.Y., 2004. Codium fragile: rhizomatous growth in the Zostera thief of eastern Canada. Helgoland Marine Research, 58 (3), 141-146.
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.
Hiscock, K., 1984. Rocky shore surveys of the Isles of Scilly. March 27th to April 1st and July 7th to 15th 1983. Peterborough: Nature Conservancy Council, CSD Report, No. 509.
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.
Holmer, M. & Laursen, L., 2002. Effect of shading of Zostera marina (eelgrass) on sulfur cycling in sediments with contrasting organic matter and sulfide pools. Journal of Experimental Marine Biology and Ecology, 270 (1), 25-37.
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
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.
Hufford, K.M. & Mazer, S.J., 2003. Plant ecotypes: genetic differentiation in the age of ecological restoration. Trends in Ecology & Evolution, 18 (3), 147-155.
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.
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
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, J., Young, J., Haynes, G., Moss, B., Eaton, J. & Hardwick, K., 1999. Do submerged aquatic plants influence their periphyton to enhance the growth and reproduction of invertebrate mutualists? Oecologia, 120 (3), 463-474.
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.
Kelly, J.R. & Volpe, J.P., 2007. Native eelgrass (Zostera marina L.) survival and growth adjacent to non-native oysters (Crassostrea gigas Thunberg) in the Strait of Georgia, British Columbia. Botanica Marina, 50 (3), 143-150.
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., 1999. Sediment resuspension in a shallow Thalassia testudinum banks ex König bed. Aquatic Botany, 65 (1), 269-280.
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.
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.
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.
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.
Mathieson, S., Cattrijsse, A., Costa, M., Drake, P., Elliott, M.J., Gardner, J. & Marchand, J., 2000. Fish assemblages of European tidal marshes: a comparison based on species, families and functional guilds. Marine Ecology Progress Series, 204, 225-242.
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.
Mills, K.E. & Fonseca, M.S., 2003. Mortality and productivity of eelgrass Zostera marina under conditions of experimental burial with two sediment types. Marine Ecology Progress Series, 255, 127-134.
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.
Moore, K.A. & Jarvis, J.C., 2008. Environmental factors affecting recent summertime eelgrass diebacks in the lower Chesapeake Bay: implications for long-term persistence. Journal of Coastal Research, 135-147.
Moore, K.A. & Wetzel, R.L., 2000. Seasonal variations in eelgrass (Zostera marina L.) responses to nutrient enrichment and reduced light availability in experimental ecosystems Journal of Experimental Marine Biology and Ecology, 244, 1-28
Moore, K.A., Shields, E.C. & Parrish, D.B., 2014. Impacts of varying estuarine temperature and light conditions on Zostera marina (eelgrass) and its interactions with Ruppia maritima (widgeongrass). Estuaries and coasts, 37 (1), 20-30.
Muehlstein, L., Porter, D. & Short, F., 1988. Labyrinthula sp., a marine slime mold producing the symptoms of wasting disease in eelgrass, Zostera marina. Marine Biology, 99 (4), 465-472.
Muehlstein, L.K., Porter, D. & Short, F.T., 1991. Labyrinthula zosterae sp. nov., the causative agent of wasting disease of eelgrass, Zostera marina. Mycologia, 83 (2), 180-191.
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.
Nelson, T.A., 1997. Epiphytic grazer interactions on Zostera marina (Anthophyta monocotyledons): effects of density on community structure. Journal of Phycology, 33, 740-753.
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.
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.
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., 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.
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, 1995a. Effect of periphyton grazing by Hydrobia ulvae on the growth of Zostera noltii on a tidal flat in the Dutch Wadden Sea. Marine Biology, 122, 431-437.
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.
Proctor, C., 1999. Torbay Zostera mapping project. Report to English Nature, World Wide Fund for Nature UK, and Torbay Council
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., Ehlers A., Hämmerli, A. & Worm, B., 2005. Ecosystem recovery after climatic extremes enhanced by genotypic diversity. Proceedings of the National Academy of Sciences of the United States of America, 102 (8), 2826-2831.
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.
Salo, T. & Pedersen, M.F., 2014. Synergistic effects of altered salinity and temperature on estuarine eelgrass (Zostera marina) seedlings and clonal shoots. Journal of Experimental Marine Biology and Ecology, 457, 143-150.
Salo, T., Pedersen, M.F. & Boström, C., 2014. Population specific salinity tolerance in eelgrass (Zostera marina). Journal of Experimental Marine Biology and Ecology, 461, 425-429.
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., Ibelings, B.W. & Den Hartog, C., 1988. Comparison of a current eelgrass disease to the wasting disease in the 1930s. Aquatic Botany, 30 (4), 295-304.
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., Burdick, D.M. & Kaldy III, J.E., 1995. Mesocosm experiments quantify the effects of eutrophication on eelgrass, Zostera marina. Limnology and Oceanography, 40 (4), 740-749.
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.
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.
Tait, E., Carman, M. & Sievert, S.M., 2007. Phylogenetic diversity of bacteria associated with ascidians in Eel Pond (Woods Hole, Massachusetts, USA). Journal of Experimental Marine Biology and Ecology, 342 (1), 138-146.
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.
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.
Turner, S.J. & Kendall, M.A., 1999. A comparison of vegetated and unvegetated soft sediment macrobenthic communities in the River Yealm, south western Britain. Journal of the Marine Biological Association of the United Kingdom, 79, 741-743.
Tutin, T., 1938. The autecology of Zostera marina in relation to its wasting disease. New Phytologist, 37 (1), 50-71.
Tweedley, J.R., Jackson, E.L. & Attrill, M.J. 2008. Zostera marina seagrass beds enhance the attachment of the invasive alga Sargassum muticum in soft sediments Marine Ecology Progress Series, 354, 305–309
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.
Van Katwijk, M. & Hermus, D., 2000. Effects of water dynamics on Zostera marina: transplantation experiments in the intertidal Dutch Wadden Sea. Marine Ecology Progress Series, 208, 107-118.
van Lent, F. & Verschuure, J.M., 1994. Intraspecific variability of Zostera marina L.(eelgrass) in the estuaries and lagoons of the southwestern Netherlands. I. Population dynamics. Aquatic Botany, 48 (1), 31-58.
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., 2007. Introduced species in seagrass ecosystems: status and concerns. Journal of Experimental Marine Biology and Ecology, 350 (1), 89-110.
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.
This review can be cited as:
Last Updated: 14/08/2015
Tags: seagrass sea grass eeelgrass eel grass