Dynamics, threats and management of salt marshes

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It is recommended this article together with the article Salt marshes.


Processes and mechanisms driving natural dynamics and ecosystem development

Coastal areas, like estuaries, are high-energetic environments where organisms are exposed to hydrodynamic forces from waves and tidal currents. Ecosystem engineering species (Jones et al., 1997) play an important role in shaping the intertidal landscape (Temmerman et al., 2007[1]; Weerman et al., 2010). Coastal vegetation, like salt marsh vegetation, are ecosystem engineers in that they can strongly attenuate hydrodynamic energy from tidal current and waves (Bouma et al., 2005[2], 2007[3], 2010). This has a positive effect on sediment accretion rates, and hence results in increased sediment elevation. In turn, increased sediment elevation stimulates plant growth because the inundation duration for the vegetation is shortened. This results in positive feedbacks between plant growth and sediment accretion. Implications of this feedback can be observed in the field in the form of dome shaped hummocks of cord‐grass (Spartina spp.). They can be found on the mud flats seaward of the salt marsh edge (Figure 1), where the salt marsh is developing.

Feedbacks between hydrodynamic forces, sediment accretion and vegetation are key processes in shaping salt marshes (Temmerman et al., 2007[1]; van Wesenbeeck et al., 2008). Locally the canopy of a vegetation stand can attenuate currents and waves which result in a net sedimentation. However, the same canopy also obstructs the flow, thereby diverting it and increasing flow velocities in the areas adjacent to the canopy because of conservation of mass and energy (Bouma et al., 2009). This biomechanical stress diversion can result in negative feedbacks on vegetation settlement and growth at some distance from the canopy (van Wesenbeeck et al., 2008). However, the outcome of these feedbacks may be dependent on the local context, seeing as these kinds of feedbacks are density‐dependent (Bouma et al., 2009). In other words, the strength of these negative feedbacks may vary with vegetation age, composition, or even the sediment type it is growing in (van Hulzen et al., 2006[4]). Overall these feedbacks cause complex patterns of gullies and hummocks until eventually a mature marsh arises, dissected by a complex drainage system (Kirwan and Murray, 2007; Temmerman et al., 2007[1]).

Figure 1: (A) Eroding cliff. (B) Patchy vegetation at pioneer zone of mudflat and saltmarsh interface. A dome-shaped patch is seen in front. (C) Salt marsh with eroding cliff separating the low marsh and pioneer zone. Pioneer vegetation (Spartina) has colonized the area below the eroding cliff (see also fig. 1.A). Photographs by J. van Belzen.

Many marshes are characterized by a cyclic nature, where marsh formation is followed by destruction (Figure 2). After a period of lateral extension, large scale lateral erosion of salt marshes can set in when the marsh edge becomes disturbed, a phenomenon often referred to as cliff erosion (see Figure 1.a, Figure 2.B.b; Allen, 2000[5]; Adam 2002[6]). For example, a disturbance from a storm surge can initialize this erosion process by forming a steep slope. At the disturbed edge, sediment is more vulnerable to wave action and currents. So once a cliff starts to erode, this process will not easily be stopped. Thus the steep slope remains particularly vulnerable for waves and currents until it is protected by new marsh vegetation emerging in front of the cliff. The initiation of cliff erosion is intrinsic to natural temporal salt marsh dynamics (Allen, 2000[5]; van de Koppel et al., 2005[7]). However, human activities can contribute significantly to the severity of the cliff erosion (Allen, 2000[5]; Adam, 2002[6]).

An example is eutrophication from agricultural effluents (especially N-enrichment) that favors the development of marsh plants with fewer and less dense roots and increases microbial decomposition of organic matter during denitrification, leading to less accumulation of organic material belowground and smaller particle sizes[8]. By disrupting organic matter storage processes, eutrophication may also turn salt marshes from a carbon sink into a source of C to the atmosphere (see Blue carbon sequestration).

Another example is shipping traffic and dredging that can increase exposure to currents and waves, thereby increasing the pace at which lateral erosion proceeds. Moreover, human induced activities may also take away the space for natural marsh recovery in front of the eroding cliff. The latter would result in the permanent loss of a marsh.

Loss of salt marsh habitat due to lateral erosion is a major problem across the world, especially in those locations where the marsh does not seem to recover. For example, the marshes in the Venice Lagoon (Italy) laterally erode with 1.2‐2.2 [math]m yr^{-1}[/math] at their seaward edges (Day et al., 1998[9]) .The estuaries of South‐East England lose about 4,000 [math]m^2 yr^{-1}[/math] of tidal marsh area due to erosion at the seaward edges and channel widening of creeks dissecting the marsh (Hughes and Paramor, 2004[10]). However, the main drivers of salt marsh erosion are still subject of debate (Wolters et al., 2005[11]). Generally, it is believed that human activities are responsible for increasing erosion (Allen, 2000[5]; Adam, 2002[6]; Wolters et al., 2005[11]). Pollution, shipping and dredging are some of the proposed anthropogenic causes. In addition, climate change and sea level rise receive much attention as a cause of salt marsh disappearance. In addition to these extrinsic forcing factors, intrinsic biological processes are also proposed (Allen, 2000[5]; Wolters et al., 2005[11]). For example, vegetation‐sediment feedbacks (Allen, 2000[5]) and sediment destabilization by bioturbation and herbivory by worms (Hughes and Paramor, 2004[10]; van der Wal and Pye, 2004[12]) and geese (Dionne, 1985[13]) can also result in erosion of salt marshes. A fundamental understanding of the mechanisms that control cliff initiation and salt marsh re‐establishment in front of a cliff is needed in order to protect and manage these highly dynamic salt marsh ecosystems.

Figure 2: (A) Overview of saltmarsh by aerial photograph (RWS), comparable to situation (B.b). Cross-shoreprofile of salt‐marsh dynamics of conceptual ecomorphological model, which mimics the development of the marsh in (A). Here, (―) is the initial bare mud--‐flat profile, (―) is vegetated marsh profile at beginning, and (---) is the final profile. First, (a) saltmarsh generation due to the positive feedback between vegetation and sedimentation.Second, (b) cliff erosion of old marsh and subsequent growth at the pioneer zone (after van de Koppel, et al., 2005[7]).


Vulnerability & threats to salt marshes

Short-term effects of flooding and storms

Short-term flooding: Vulnerability of marshes to saltwater flooding

The salt‐marsh community is well adapted to salinity due to regular tidal exposure to seawater. The vast majority of salt-marshes are well drained and therefore at less risk to the endured flooding. In comparison, the community of grazing-marshes is adapted to very dilute seawater and the habitat drainage is often slow. The potential impact of saltwater flooding is therefore more severe for grazing marshes than for salt marshes. Much of the evidence regarding the effect of seawater on coastal vegetation therefore relates to oligohaline/grazing marshes.


Short-term flooding: Effect of flooding by saline water on salt marshes

A flooding event that originates from increased freshwater discharge, for instance due to heavy rainfall or ice melt in a catchment area, will result in a fresh‐water pulse through downstream marshes. Unless freshwater flooding lingers for extensive periods, the impact on the vegetation of salt-and grazing- marshes will be short lived (Flynn et al., 1995[14]; Grace and Ford, 1996[15]; Howard and Mendelssohn, 2000[16]). The ‘halophytic’ (salt-tolerant) species that dominate salt-and grazing-marsh communities will not be harmed by short‐term fresh water exposure. However, their physiological and biochemical adaptations to cope with salinity stress make them poorly competitive under fresh water conditions (Crain et al., 2004[17]). Their halophytic traits enable them to colonise saline environments (Pennings and Callaway, 1992[18]). Salinity exposure in the salt marsh, and consequently the inherent salt tolerance of the inhabitant community, does not necessarily decline linearly with shore level. Summer evaporation of seawater pools can leave concentrated deposits of salt on the high marsh where the habitat is infrequently flushed, leading to high levels of sediment salinity that exclude less halophytic species (Watson and Burne, 2009). Paradoxically, increased frequency of seawater flushing by storms might dilute the accrued sediment salinity of such high marsh environments and alter the zonation of species. For instance, increased tidal flooding of an elevated marsh plain can cause the normally very halophytic high marsh species to be replaced by salt‐intolerant lower shore species (Watson and Burne, 2009).

The severity of impact of saltwater flowing is likely to depend on the natural salinity occurring at a specific location. Relatively brackish marshes, dominated by halophytic plants, will see less changes to community composition than fresh water dominated grazing marshes and coastal flood plains (Brown et al., 1994). Increased flooding by salt water is most likely to have the greatest effect on the fresh‐water adapted members of the marsh vegetation (Crain et al., 2004[17]), which increases in dominance in the transitional and grazing‐marsh above the tidal marks. Stormy conditions that result in a temporary increase in sea level and which bring in salt water pulses to coastal marsh systems therefore should have a greater effect on the grazing‐marsh community than on the salt marsh community. For example, seawater flooding of a diked grazing-marsh, following a dike breaching, prevented most of the freshwater vegetation from developing in the following spring (Klein and Bateman, 1998[19]). Vegetation cover, species richness, recovery and re-establishment of an oligohaline marsh decreased during one month of experimental exposures to increased salinity (from 0.5‐5.0 salinity up to 12) (Howard and Mendelssohn, 2000[16]). In the longer term, the space left by dead vegetation is likely to be colonised by more salinity tolerant species, and thus grazing-marsh communities might come to resemble those of salt marshes (Doody, 1982[20]; Howard and Mendelssohn, 2000[16]).

Many coastal marsh plants are able to recover temporary increases in salinity (Flynn et al., 1995[14]; Grace and Ford, 1996[15]; Howard and Mendelssohn, 2000[16]). However, the potential for lasting changes to communities increases with the duration of flooding (Flynn et al., 1995[14]; Howard and Mendelssohn 2000[16]). Elevated salinity (from natural, 0.5‐5 to 15) slowed vegetation recovery more in flooded than in drained soils (Flynn et al., 1995[14]). The naturally slow drainage of grazing marshes, that follows temporary sea water flood, causes this habitat to remain immersed for longer periods than salt marshes. Grazing marshes are therefore at greater risk to the endured flooding. However, there are indications that these marshes are relatively resilient to exposure; if the water is brackish enough, it may require months of immersion before significant impacts to vegetation cover occurs (e.g. Howard and Mendelssohn, 2000[16]). Brewer and Grace (1990)[21] hypothesized that occasional storm‐generated pulses of salt water moving into an oligohaline marsh would generate short-lived salinity gradients that, along with biotic interactions, would regulate species distributions over longer terms. Sharpe and Balwin (2009)[22] proposed that an unexpected peak in vegetation species richness in the transitional marsh arose because pulsed variation in salinity (0‐5) prevented domination by fresh water or salt water species. Thus, pulsed salinity exposure might not necessarily diminish vegetation diversity. Nevertheless, increased salt water flooding of grazing marshes is likely to drive the succession towards more salt-tolerant vegetation, and increase the resemblance with salt marsh assemblages. Note that the empirical evidence for the rate of this transition is lacking (Nicholls and Wilson, 2001[23]). The consequence of increased coastal flooding might therefore be a gradual loss of grazing-marsh communities, in exchange for gain in area cover of salt marsh communities (Doody, 1982[20]; Klein and Bateman, 1998[19]; Nicholls and Wilson, 2001[23]).


Short-term flooding: Interactions of salinity with other disturbances

It is important to caution against a general interpretation that seawater flooding is a minimal risk to coastal marshes in general. The severity of seawater influence on grazing‐marshes might depend much on whether the salinity is paralleled with other plant stressors and disturbances. Sharpe and Balwin (2009)[22] sampled plant diversity in a marsh in the United States, across a fresh (salinity 0.5) to mesohaline (5-18) salinity gradient. In an undisturbed marsh, richness in transition zone oligohaline marshes was as high as or higher than in tidal fresh water-marshes. In an anthropogenically disturbed estuary, however, plant species richness declined linearly with an increase in salinity. Experimental flooding by brackish (6‐14) water had a greater effect on grazing‐marsh community structure and biomass when the vegetation was also disturbed by leaf clipping (Baldwin and Mendelssohn, 1998) or grazing (Gough and Grace, 1998). In comparison, flooding did not affect species richness in the absence of such additional disturbances (Baldwin and Mendelssohn, 1998). If the salinity and water regimes are permanently altered and/or the vegetation is destroyed by a combination of factors, the substrate might eventually subside. Substrate subsidence and associated increased water depth might prevent seed dispersal and germination of more flooding tolerant species, and thus hamper system recovery (McKee and Mendelssohn, 1989[24]). It is not known beyond which threshold the frequency of flooding will have permanent effects.


Short-term: Vulnerability of marshes to storm damage

As mentioned above, flooding can induce some disturbances to the vegetation composition of salt marshes. However, the threat storms impose on salt marshes is more likely to result from storm-associated damage than from flooding. Important lateral erosion of salt marshes can occur during storms. Wind-induced waves can destabilize sediments, initiate and propagate lateral cliff erosion, tear off plant material, and deposit wrack and debris in marshes. It has long been assumed that the vulnerability of salt marshes is largely related to the effects of waves on sediment stability and on lateral erosion. However, recent studies suggest that other factors can be more important. Wave erosion is strongly reduced in marshes with pioneer vegetation[25]. Observations in the Great Marsh (Massachusetts, USA) show that soil properties and wind waves are not strongly correlated with the presence and distribution of marsh erosion[26]. Lateral erosion can also result from the meandering of tidal channels, see Tidal channel meandering and marsh erosion. Severe salt marsh erosion will undoubtedly lead to an increased risk of sea water flooding and storm-associated damage for adjoined grazing‐marshes, and might eventually drive a transition of grazing‐marsh communities into salt marsh habitats.


Short-term storm: Sediment destabilization and lateral erosion of salt marshes

The main driving factors initiating marsh erosion are not always clear (Wolters et al., 2005[11]). Human activities can be, in part, responsible for increasing erosion rates through pollutant-driven diminishing of vegetation cover and/or by enhancing hydrodynamic energy reaching the marsh via ship waves and channel dredging (Allen, 2000[5]; Adam, 2002[6]; Wolters et al., 2005[11]). Thus, there is a strong need for a fundamental understanding of the cyclic functioning of salt marsh ecosystems in order to understand when disturbances by storms will start a natural cycle of rejuvenation versus when they cause the irreversible loss of a marsh and thus would benefit from protective measures.

Vulnerability of the saltmarsh to the initiation of cliff erosion largely depends on the age of the marsh. Cliff erosion is hypothesized to be an inevitable and intrinsic consequence of the ecomorphological dynamics of saltmarshes (van de Koppel et al., 2005[7]). That is, the capture of sediment by the vegetation leads to vertical salt marsh growth, which in the long term makes the salt marsh susceptible to lateral erosion (Allen, 2000[5]; van de Koppel et al., 2005[7]). Both conceptual modelling and empirical evidence showed that positive feedback between vegetation growth and sediment capture generates an increasingly steeper bank at the seaward edge of the marsh (van de Koppel et al., 2005[7]). As a consequence, salt‐marsh edges become more vulnerable to disturbance as they mature (see Figure 1.B). This means that in the end, relatively small disturbances like from minor storm events or ship waves may induce the erosion. Data on sedimentation in saltmarshes, obtained from sediment core transects and spatiotemporal analysis of aerial photographs, support this conceptual model (van de Koppel et al., 2005[7]; van der Wal et al., 2008[27]), suggesting that, in a wide range of circumstances, lateral retreat due to cliff erosion will happen sooner or later.

Resilience of the marsh edge to erosion will depend on the interplay between vegetation composition and sediment dynamics. For instance, Spartina plants reduce cliff erosion more than Limonium plants do due to differences in their root system. As a consequence, a Spartina dominated marsh edge is likely less vulnerable to storm events (van Eerdt, 1985). However, consequences of plant community on the vulnerability and resilience of the marsh can be more complex. The above ground plant traits can have other effects on the vegetation-sediment interaction. For example, stiffness and density of the plant may affect sedimentation rates (Bouma et al., 2005[2], 2009, 2010). The overall effect of above and below plant traits on salt‐marsh resilience/vulnerability remains largely unknown and subject to ongoing research.

Alongside cliff erosion, re‐growth of pioneer vegetation on the cleared mudflat in front of the saltmarsh cliff ideally occurs, thereby rejuvenating the mature marsh (see Figure.1.B; van de Koppel, et al., 2005[7]; van der Wal et al., 2008[27]). This pioneer vegetation determines the conditions for lateral retreat and gradually slows down erosion of the salt-marsh edge.

The establishment of pioneer vegetation is therefore of vital importance to the development and recovery of salt marshes (van de Koppel et al., 2005[7]; van der Wal et al., 2008[27]; Callaghan et al., 2010[28]). This emphasizes the importance of initial conditions for the establishment and growth of vegetation, and the physical conditions that may constrain these intrinsic processes for salt marsh development (van der Wal et al., 2008[27]). However, the factors limiting seedling establishment remain poorly understood (Bouma et al., 2009). We hypothesize that sediment destabilization plays a critical role in the ability for pioneer establishment on a mudflat and, therefore, in the ability of the salt marsh to recover from lateral cliff erosion (see Tidal channel meandering and marsh erosion).


Short-term storm: sediment and wrack deposits on salt marshes

Depositions of wrack, debris or large amounts of sediment, associated with extreme flooding events, can have significant effects on salt marsh vegetation. Wrack depositions may smother less hardy vegetation, leaving bare patches and opportunity for new colonization by neighbouring species, or from dispersed seeds (Bertness and Ellison, 1987[29]; Tolley and Christian, 1999[30]). Large algal mats can have residence times of 3‐4 months (Valiela and Rietsma, 1995[31]). While wrack depositions may not be as significant in cover (Valiela and Rietsma, 1995[31]), small-scale alteration in species cover by algal wrack depositions does have the potential for wider effects on community diversity if the same spots are regularly covered by seaweed (van Hulzen et al., 2006[4]). The deposition of sediments following storm flooding may be significant. Experimental flooding and sedimentation of seedbanks of an oligohaline marsh community showed that an addition of 2 cm of sediment decreased plant density and germination of seedlings, suggesting that an increase in sedimentation and relative sea level may reduce plant biodiversity (Peterson and Baldwin, 2004[32]]). The deposition of sediments might be positive for subducting and nutrient starved marshes. For example, within a year after 3-8 cm of sediment deposited by Hurricane Katrina, the vegetation of a high marsh had fully recovered and below‐ground root growth had increased 10‐fold (McKee and Cherry, 2009[33]).

While single factors may have limited effects on marshes, a collective of concurrent stressors is likely to generate significant impacts on marsh communities. The deposition of wrack and sediments is often concurrent with other habitat stressors that might jointly influence marsh vegetation cover. Thus, while Tolley and Christian (1999)[30] found little effect of sea water flooding on vegetation biomass, the simultaneous deposition of algal wrack greatly repressed plant cover and biomass, in some species irreversibly so.


Long-term effects due to climate change and sea level rise

Coastal squeeze, due to sea level rise, and erosion are primary threats to salt marshes across Europe. They can result in reduced coastal defence value and in an increased risk of flooding. Although sea level rise may pose serious threats to the survival of salt marshes, there is growing evidence that as long as sediment supply is sufficient, the vegetation-sedimentation feedback of marshes enables marshes to accrete vertically at the rate of the rising sea-level (Kirwan and Temmerman, 2009[34]). However, if the suspended matter load is reduced by climate change or by significant human alteration in a catchment area, vegetation- sedimentation feedbacks can become limited, affecting the potential of marshes to accrete (Kirwan and Temmerman, 2009[34]). As explained in the previous sections, lateral marsh erosion can become a serious threat to salt marshes over time if seedling establishment in front of the marsh is not possible so that re- growth of the marsh is prevented. Many aspects that affect the cyclic dynamics of marshes are still not well understood. Important in maintaining the vegetation-sedimentation feedback is that the sedimentary conditions remain more or less the same.

Several managerial aspects are likely to compromise the capacity for marshes to persist and to protect the coast. Reduction in area by coastal squeeze will reduce the wave attenuation capacity, as the efficiency of energy reduction is strongly dependent on the depth of the marsh (Möller, 2006[35]). Whether this might have negative feedback on marsh accretion and accelerate area loss is not known. Effects of grazing might also reduce the vegetation‐sedimentation feedback by reducing vegetation cover and height, thereby hampering the development of salt marshes (Kiehl et al., 2007[36]). Finally, very little is known about the implications on salt marsh resilience from interactions between different environmental, climatic and managerial variables. Interactions between climate stressors (e.g. desiccation, irradiation), physical forcing (extreme flooding events, increased storminess) and environmental management (eutrophication, grazing, and managed retreat) will be a likely reality for many marshes. This is important because interactive stresses can be synergistic and cause shifts in the stable states of ecosystems, which can compromise the naturally delivered services (Scheffer et al., 2001[37]; Scheffer et al., 2009[38]). Our evaluation of the resilience of salt marshes to disturbance, including climate change, might for the time being still be somewhat naïve and based on limited current research.


Case study: forecasting the effects of sea‐level rise

Figure 3: Impact of sea-level rise on tidal flat and tidal marsh complex by 2050 (a and b) and 2100 (c and d) year at Chongming Dongtan nature reserve.

Located at the mouth of the Yangtze Estuary, the Chongming Dongtan nature reserve is extremely vulnerable to climate change and especially to an accelerated sea-level rise. We use a variety of data from remote sensing, an in situ global positioning system (GPS), tidal gauges, nautical charts, geographic spatial analysis modelling and IPCC sea-level rise scenarios to forecast the potential impacts of increased sea level on the coastal wetland habitat of the Chongming Dongtan Nature Reserve (Figure 3). The results indicate that around 40% of the intertidal zone of the nature reserve will be inundated by the year 2100 due to an estimated 0.88 m increase in sea level (Figure 3.c and 3.d). In particular, the Scirpus mariqueter communities and bare tidal flats are more vulnerable to sea‐level rise. The identification, mapping and statistical summary of environmental impacts of the projected sea-level rise at Chongming Dongtan Nature Reserve represent an important initial step for decision makers concerned with mitigation of the adverse impacts of sea-level rise. In this study, the inundation‐based assessment was developed to inform policymakers, managers and the public about the amount and the spatial distribution of tidal wetland change as a result of sea‐level rise. The results indicate that the zones most vulnerable to sea-level rise at the Chongming Dongtan Nature Reserve is the S. Mariqueter zone, the bare tidal flat zone and the tidal creeks, which are the most suitable habitats for migratory birds. A ~30% loss of the S. Mariqueter marsh community by the year 2100 would eliminate a rich invertebrate food source and cause deterioration in the estuarine food web for migrating birds; such a loss could arise from human-induced stressors such as land reclamation, seawall constructions, overfishing and local pollution. As tidal marshes and flats submerge and decline in size and productivity, increased crowding in the remaining areas could lead to reductions in and eventually even exclusion of some local shorebird populations (Tian et al., 2010[39]).


Key processes to focus on for maintaining ecosystems integrity

Effects of single disturbance events on marsh responses to long-term change

Single events, such as violent storms, normally have short-lived effects on the species composition and on the ecological functioning of saltmarshes (Flynn et al. 1995[14], Howard and Mendelssohn, 2000[16]; McKee and Cherry, 2009[33]), and are thus of less importance compared to long term persistent changes in environmental condition. Long-term processes of coastal squeeze with sea level rise and lateral erosion with increased storminess are the primary threats to salt- and grazing-marshes across Europe (Nicholls and Wilson, 2001[23]). A single storm can push a marsh over the tipping point, shifting it from laterally expanding towards laterally eroding. If erosion persists, and the marsh cannot re-establish in front of the cliff, in time this will result in reduced coastal defence value and an increased risk of flooding of adjacent terrestrial environment (e.g. grazing- marshes) (Klein and Bateman, 2007).


Current management practices

Making space for water

Currently, saltmarshes are managed extensively because of their acknowledged role in coastal protection. Many countries like e.g. the UK, the Netherlands, etc, have developed management schemes in order to make space for water along river flood plains, estuarine and coastal areas (Bakker, et al., 2005; DEFRA, 2004). This way, river run‐off and occasional high sea water levels can be attenuated by the natural buffer and retention capacity of the landscape. For example, restoring the water storage volume in an estuary can reduce the tidal prism, smoothing the tidal amplitude, which reduces the risk of flooding in up-stream estuarine areas. Salt‐marshes play an important part in this contemporary policy, because creating new marsh‐land both increases tidal water storage in up‐stream estuarine areas and wave attenuation of storm surges along exposed coast lines (Bakker et al., 2005; Kiehl, et al., 2007[36]).


Grazing management and coastal protection

There is evidence to suggest that grazing management could be of particular importance to the capacity of marshes for protecting the coast, although there has been little quantitative research on this subject (Bakker, et al., 2005). The vegetation is of key importance to coastal protection by marshes, through consolidation of the soil and by representing a structural hindrance to wash-over waves. Evidently, livestock has large potential for altering the vegetation structure directly through feeding and indirectly by altering the conditions for vegetation growth (Bakker, et al., 2005; Kiehl, et al., 2007[36]). Feeding and defecation moderate vegetation structure‐composition and above- and below-ground biomass production. Trampling and hoof holes lead to soil compaction and can cause saltpan formation (Vera, 2000[40]). The potential of management of grazing regime to influence the salt marsh coastal protection potential is therefore high. Intense grazing modifies zonation patterns and transforms complex communities with woody species into homogenous lawns dominated by short flexible grass (Andresen, et al., 1990[41]; Kiehl, et al., 2007[36]), with an associated likely reduction in wave attenuation (Möller, 2006[35]) and sedimentation rates (Andresen, et al., 1990[41]). Grazing at low intensity increases vegetation patchiness and biodiversity due to selective grazing of palatable species (Bakker, 1985[42], 1998; Kiehl et al., 1996[43]; Adler et al., 2001[44]; Bouchard et al., 2003[45]; Marriot et al. 2005). Patchiness may cause specific spatial patterns in turbulence and sedimentation (Boorman, 1999; van Wesenbeeck, et al., 2007), so that the sum effect of patchiness on marsh coastal protection is not known. Conversely, grazing pressure can lead to greater resource allocation of below‐ground biomass (Pucheta, et al., 2004[46]), thus reducing surface erosion and below-ground contributions to an increase in marsh surface elevation.


A separate article deals with Restoration of estuarine and coastal ecosystems.


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see also


References

  1. ↑ 1.0 1.1 1.2 TEMMERMAN, S.; BOUMA, T.J.; VAN DE KOPPEL, J.; VAN DER WAL, D.; DE VRIES, M.B.; HERMAN, P.M.J.(2007). Vegetation causes channel erosion in a tidal landscape. Geology. 35(7), 631-634. Available from: www.vliz.be/imis
  2. ↑ 2.0 2.1 BOUMA T.J.,DE VRIES M.B., LOW E., PERALTA G., TNCZOSI.C.,VANDEKOPPELJ., HERMAN P. M. J., 2005. Trade‐offs Related to Ecosystem Engineering: A Case Study on Stiffness of Emerging Macrophytes. Ecology. 86, 2187‐2199.
  3. ↑ BOUMA, T.J.; VAN DUREN, L.A.; TEMMERMAN, S.; CLAVERIE, T.; BLANCO-GARCIA, A.; YSEBAERT, T.J.; HERMAN, P.M.J. (2007). Spatial flow and sedimentation patterns within patches of epibenthic structures. Cont. Shelf Res.. 27(8): 1020-1045. dx.doi.org/10.1016/j.csr.2005.12.019 Available from:www.vliz.be/imis
  4. ↑ 4.0 4.1 VAN HULZEN J.,VAN SOELEN J.,HERMAN P.M.J., BOUMA T.J., 2006.The significance of spatial andtemporal patterns of algal mat deposition in structuring salt marsh vegetation. J Veget Sci.. 17, 291‐298.
  5. ↑ 5.0 5.1 5.2 5.3 5.4 5.5 5.6 5.7 ALLEN J.R.L., 2000. Morphodynamics of Holocene salt marshes: a review sketch from the Atlantic and Southern North Sea coasts of Europe. Quaternary Science Reviews. 19, 1155-1231.
  6. ↑ 6.0 6.1 6.2 6.3 ADAM P., 2002. Salt marshes in a time of change. Environmental Conservation. 29, 39‐61
  7. ↑ 7.0 7.1 7.2 7.3 7.4 7.5 7.6 7.7 VAN DE KOPPEL J.,VAN DER WAL D., BAKKER J.P.,HERMAN P.M.J., 2005. Self‐Organization and Vegetation Collapse in Salt Marsh Ecosystems. The American Naturalist. 165, E1-12.
  8. ↑ Fagherazzi, S., FitzGerald, D.M., Fulweiler, R.W., Hughes, Z., Wiberg, P.L., McGlathery, K.J., Morris, J.T., Tolhurst, T.J., Deegan, L.A., Johnson, D.S., Lesser, J.S. and Nelson, J.A. 2022. Ecogeomorphology of Salt Marshes. Ch. 8.14 of Treatise on Geomorphology (ed. Shroder, J.F.), Elsevier
  9. ↑ DAY J.W., SCARTON F., RISMONDO A., ARET D., 1998. Rapid Deterioration of a Salt Marsh in Venice Lagoon, Italy. Journal of Coastal Research. 14, 583‐590.
  10. ↑ 10.0 10.1 HUGHES R.G., PARAMOR O.A.L., 2004. On the loss of saltmarshes in south‐east England: methods for their restoration. Journal of Applied Ecology. 41, 440‐448.
  11. ↑ 11.0 11.1 11.2 11.3 11.4 WOLTERS M., BAKKER J.P., BERTNESS M.D., JEFFERIES R.L., MÖLLER I., 2005. Saltmarsh erosion and restoration in south-east England: squeezing the evidence requires realignment. Journal of Applied Ecology. 42, 844‐851.
  12. ↑ VAN DER WAL D., PYE K., 2004. Patterns, rates and possible causes of saltmarsh erosion in the Greater Thames area (UK). Geomorphology. 61, 373‐391.
  13. ↑ DIONNE, J.-C., 1985. Tidal marsh erosion by Geese, St. Lawrence estuary, Québec. Géographie physique et Quaternaire. 39, 99‐105.
  14. ↑ 14.0 14.1 14.2 14.3 14.4 FLYNN, K.M.; MCKEE, .KL.; MENDELSSOHN, I.A, 1995.RECOVERY OF FRESH-WATER MARSH VEGETATION AFTER A SALTWATER INTRUSION EVENT. OECOLOGIA. 103(1), 63-72 DOI:10.1007/BF00328426.
  15. ↑ 15.0 15.1 GRACE J.B., FORD M.A., 1996. The potential impact of herbivores on the susceptibility of the marsh plant Sagittaria lancifolia to saltwater intrusion in coastal wetlands. Estuaries. 19, 13–20.
  16. ↑ 16.0 16.1 16.2 16.3 16.4 16.5 16.6 HOWARD R.J., MENDELSSOHN I.A., 2000. Structure and composition of oligohaline marsh plant communities exposed to salinity pulses. Aquat Bot. 68, 143–164.
  17. ↑ 17.0 17.1 CRAIN, C.M.; SILLIMAN B.R.; BERTNESS, S.L. ; BERTNESS, M.D., 2004. Physical and biotic drivers of plant distribution across estuarine salinity gradients. ECOLOGY. 85(9), 2539-2549.DOI: 10.1890/03-0745
  18. ↑ CRAIN, C.M.; SILLIMAN B.R.; BERTNESS, S.L. ; BERTNESS, M.D., 2004. Physical and biotic drivers of plant distribution across estuarine salinity gradients. ECOLOGY. 85(9), 2539-2549. DOI: 10.1890/03-0745
  19. ↑ 19.0 19.1 KLEIN R.J.T., BATEMAN I.J., 1998. The recreational value of Cley marshes nature reserve: An argument against managed Retreat? Water and Environment Journal. 12, 280-285.
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  24. ↑ MCKEE K.L.; MENDELSSOHN I.A.,1989. Response of a fresh-water marsh plant community to increased salinity and increased water level. AQUATIC BOTANY. 34(4), 301-316. DOI: 10.1016/0304-3770(89)90074-0.
  25. ↑ Wang, H., van der Wal, D., Li, X., van Belzen, J., Herman, P. M. J., Hu, Z., Ge, Z., Zhang, L. and Bouma, T. J. 2017. Zooming in and out: Scale dependence of extrinsic and intrinsic factors affecting salt marsh erosion, J. Geophys. Res. Earth Surf. 122: 1455–1470
  26. ↑ Houttuijn Bloemendaal, L.J., FitzGerald, D.M., Hughes, Z.J., Novak, A.B. and Phippen, P. 2021. What controls marsh edge erosion? Geomorphology 386: 107745
  27. ↑ 27.0 27.1 27.2 27.3 VAN DER WAL D., WIELEMAKER-VANDENDOOL A., HERMANP.M.J., 2008. Spatial patterns, rates and mechanisms of saltmarsh cycles (Westerschelde, The Netherlands). Estuarine, Coastal and Shelf Science. 76, 357‐368. Available from: www.vliz.be/imis.
  28. ↑ CALLAGHAN, D.P.; BOUMA, T.J.; KLAASSEN, P.; VAN DER WAL, D.; STIVE, M.J.F.; HERMAN, P.M.J., 2010. Hydrodynamic forcing on salt-marsh development: Distinguishing the relative importance of waves and tidal flows Est., Coast. and Shelf Sci.. 89(1), 73-88. dx.doi.org/10.1016/j.ecss.2010.05.013. Available from: www.vliz.be/imis.
  29. ↑ BERTNESS M.D., ELLISON A.M., 1987. Determination of pattern in a New England salt marsh plant community. Ecol Monogr. 57, 129‐14.
  30. ↑ 30.0 30.1 TOLLEY P.M.; CHRISTIAN R.R., 1999. Effects of increased inundation and wrack deposition on a high salt marsh plant community. ESTUARIES. 22(4), 944-954.DOI: 10.2307/1353074.
  31. ↑ 31.0 31.1 VALIELA I.; RIETSMA C.S.,1995. DISTURBANCE OF SALT-MARSH VEGETATION BY WRACK MATS IN GREAT-SIPPEWISSETT-MARSH. OECOLOGIA. 102(1), 106-112.
  32. ↑ PETERSON J.E., BALDWIN A.H., 2004. Seedling emergence from seed banks of tidal freshwater wetlands: response to inundation and sedimentation. Aquat Bot. 78, 243–254. Available from: www.vliz.be/imis.
  33. ↑ 33.0 33.1 MCKEE K.L., CHERRY J.A., 2009. Hurricane Katrina sediment slowed elevation loss in subsiding brackish marshes of the Mississippi river delta. Wetlands. 29, 2-15.
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  37. ↑ SCHEFFER M., CARPENTER S., FOLEY J.A., FOLKER C., WALKER B., 2001. Catastrophic shifts in ecosystems. Nature. 413: 591‐596. Available from: www.vliz.be/imis.
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  39. ↑ TIAN, B; ZHANG, LQ ; WANG, XR; ZHOU, YX ; ZHANG, W.; 2010. Forecasting the effects of sea-level rise at Chongming Dongtan Nature Reserve in the Yangtze Delta, Shanghai, China. ECOLOGICALENGINEERING, 36(10): 1383-1388.DOI: 10.1016/j.ecoleng.2010.06.016.
  40. ↑ VERA F.W.M., 2000. Grazing Ecology and Forest History. CABI Publishing, Wallingford, UK.
  41. ↑ 41.0 41.1 ANDRESEN H., BAKKER J.P., BRONGERS M., HEYDEMANN B., IRMLER U., 1990. Long‐term changes of salt-marsh communities by cattle grazing. Vegetatio. 89, 137–148.
  42. ↑ BAKKER J. P., DIJKSTRA M., RUSSCHEN P. T., 1985. Dispersa, germination and early establishment of halophytes and glycophytes on a grazed and abandoned salt‐marsh gradient. New Phytologist. 101, 291-308.
  43. ↑ KIEHL K., EISCHEID I., GETTNER S., WALTER J., 1996. Impact of different sheep grazing intensities on salt-marsh vegetation in northern Germany. Journal of Vegetation Science. 7, 99–106.
  44. ↑ ADLER P.B.; RAFF D.A.; LAUENROTH W.K.;, 2001.The effect of grazing on the spatial heterogeneity of vegetation. OECOLOGIA. 128(4): 465-479.
  45. ↑ BOUCHARD V.; TESSIER M.; DIGAIRE F.; VIVIER, J.P.; VALERY, L.; GLOAGUEN, J.C.; LEFEUVRE, J.C., 2003. Sheep grazing as management tool in Western European saltmarshes.InternationalCongress on Biodiversity Conservation and Management. COMPTES RENDUS BIOLOGIES. 326, S148-S157.DOI: 10.1016/S1631-0691(03)00052-0.
  46. ↑ PUCHETA, E.; BONAMICI, I.; CABIDO, M.; DIAZ, S.; 2004. Below-ground biomass and productivity of a grazed site and a neighbouring ungrazed exclosure in a grassland in central Argentina. AUSTRALECOLOGY. 29(2), 201-208.DOI: 10.1111/j.1442-9993.2004.01337.x.


The main authors of this article are van Belzen, Jim, Bouma, Tjeerd, Skov, Martin, Zhang, Liquan and Yuan, Lin
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