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- Materials and methods
- Supporting Information
Global losses of coastal marsh have been and will continue to be extensive. Approximately 50% of saltmarsh area worldwide has already been lost or degraded (Adam 2002; Barbier et al. 2011). In recognition of the scale of habitat loss and the value of saltmarshes, there is considerable interest in reestablishing or creating saltmarsh covering the full range of habitats that have been lost. In the USA, section 404 of the Clean Water Act affords a high degree of protection to wetlands and has been interpreted as requiring ‘no net loss’ of both wetland area and function (Zedler 2004). Legislation in Europe seeks to maintain saltmarsh (and other natural habitats) in ‘favourable status’, including the maintenance of marsh area and the provision of compensatory habitat if areas are lost to development (European Commission 2000). Subsequent guidance from the EU has clarified that this means that ‘a wetland should normally not be drained before a new wetland, with equivalent biological characteristics is available’ and that ‘the compensatory measures proposed for a project should address, in comparable proportions, the habitats and species negatively affected’ (European Commission 2000).
There are various approaches to the creation of saltmarsh habitat, including plantings of ecosystem engineer species and the placement of dredged material. However, in the face of rising sea levels, managed coastal realignment has become an increasingly important option (French 2006). Despite numerous examples of such restoration, the extent to which they provide species and habitats in ‘comparable proportions’ to natural marshes is not clear. Nor do we know how long in advance of habitat loss the pre-emptive replacements need to be initiated. Many of the wetlands created to date in mitigation have not been equivalent to those destroyed (Race & Fonseca 1996; Turner, Redmond & Zedler 2001), and the speed with which natural saltmarsh structures and functions develop on restoration schemes has been highly variable. Halophytic species may colonize some newly restored saltmarshes quickly, and species richness may parallel reference areas within 7 years (Morgan & Short 2002; Mossman et al. 2012). At other sites, however, colonization has been slower (Onaindia, Albizu & Amezaga 2001; Wolters, Garbutt & Bakker 2005), and subsequent succession of plant communities may also be slow (Havens, Varnell & Watts 2002; Williams & Orr 2002). Differences in vegetation composition between natural and restored marshes have been detected after many decades (Burd, Clifton & Murphy 1994; Bakker 2002; Garbutt & Wolters 2008). Vegetation composition and physical structure inevitably have significant influences on saltmarsh ecosystem functions (Zedler, Callaway & Sullivan 2001; Doherty, Callaway & Zedler 2011), and large differences between the vegetation on restored and reference sites are likely to result in considerable disparity in ecosystem functioning.
Most of the deliberate reactivation (managed realignment) of saltmarsh sites has occurred within the last 20 years and therefore does not provide information to assess their equivalence in the long term. On the other hand, there are areas of saltmarsh that have developed after sea defences were accidentally breached during storm surges and remained unrepaired (Burd 1992; Burd, Clifton & Murphy 1994). In most cases, it has now been more than 50 years since tidal flow was restored to these accidentally realigned (AR) sites, and they can provide useful space-for-time analogues (Gray et al. 2002) for marsh development. On the basis of a survey of four deliberately restored marshes and 14 older accidentally flooded (AR) sites in the UK, each paired with one reference site, Garbutt & Wolters (2008) concluded that plant species richness was lower on restored sites and their plant communities different. This study, however, sampled only a small number of quadrats at each site and was focused on a restricted elevation range within the local tidal frame. Elevation is a major determinant of vegetation zonation and, consequently, of many ecosystem functions (e.g. Stagg & Mendelssohn 2010; Shepard, Crain & Beck 2011). In order to provide an adequate assessment of how well restored saltmarshes in the UK replicate reference marshes, we need to examine the whole elevation range of a site. The need to compare the attributes of restored sites to multiple reference sites is widely acknowledged (e.g. SER 2004), but few studies have done so. Furthermore, few studies of intertidal restorations have measured both biological and environmental attributes and, those that have, tend to be intensive studies of a single site (e.g. Spencer et al. 2008; Mossman et al. 2012).
In this study, we examine the vegetation that has developed on 18 deliberately realigned saltmarsh sites and 17 accidentally realigned sites, together ranging from 1 to 131 years since the restoration of tidal flow, and compare it with that of 34 reference saltmarshes. The aim was to assess the extent to which these realigned marshes have equivalent plant communities to natural reference marshes, and the time-scale for these to be established. It was expected that the vegetation of older realigned sites (25–131 years) would be more similar to the vegetation on reference sites than that on more recently realigned sites. Specific objectives were to (i) assess the colonization of halophytes with time since restoration, using a space-for-time substitution; (ii) compare species abundance and plant communities on realigned sites with those from reference marshes; and (iii) investigate environmental characteristics that may influence plant colonization and compare them with equivalent reference measurements.
- Top of page
- Materials and methods
- Supporting Information
The establishment of saltmarsh through managed realignment has often been described as successful (e.g. Boorman, Hazelden & Boorman 2002; Spencer et al. 2008). Our work only agreed with this in so far as typical halophytic plant species quickly colonized MR sites. Garbutt & Wolters (2008) found that species richness was lower on 72% of restored sites compared to a paired reference marsh. However, species richness was highly variable between the reference marshes studied by Garbutt & Wolters (2008), and the species richness of 77% of the restored marshes was within the range of reference marshes. High variability can both mask differences between restored and natural sites (Neckles et al. 2002) and impede the identification of developmental trajectories (Simenstad & Thom 1996). This highlights the importance of comparing restored sites to at least two reference sites (Ruiz-Jaen & Aide 2005) and setting restoration targets that allow for between-site variability (Short et al. 2000).
While the species richness of whole realigned sites was similar to that of reference sites, plant communities in individual quadrats on MR sites were certainly not equivalent. There was more bare ground on MR sites, and species characteristics of pioneer and low saltmarsh, such as Salicornia europaea (Davy, Bishop & Costa 2001) and Puccinellia maritima (Gray & Scott 1977), were most abundant. In the low-lying areas of some MR marshes, limited vegetation cover and dominance of pioneer species could be a result of bioturbation by invertebrate infauna (Paramor & Hughes 2005), or it may simply indicate an early-successional state.
Our environmental data, however, indicate a more fundamental reason for these vegetational differences on the low marsh. Sediment conditions in lower-lying areas of realigned marshes were less oxygenated than those at corresponding elevations of reference marshes. Sediment redox potential has effects on plant abundance, independently from elevation (Davy et al. 2011), and low redox potential may inhibit vegetation colonization or limit colonization to species tolerant of waterlogged conditions (Mossman et al. 2012). The less oxygenated sediment conditions of low- and mid-marsh elevations on newly realigned sites thus shift the vegetation towards more inundation-tolerant, pioneer communities. Studies that assess the success of saltmarsh restoration by comparing the species composition of paired reference and restored samples of the same elevation may therefore exaggerate the differences in species abundance between restored and reference samples.
In contrast, at higher elevations, sediments of MR sites were better oxygenated but nevertheless remained significantly less vegetated. These sediments were drier and contained less organic matter than those from the same elevations on reference marshes. These areas may remain sparsely vegetated because patches of bare sediments on the high marsh are frequently hypersaline as a result of high surface evaporation and infrequent tidal inundation (Bertness 1991; Bertness, Gough & Shumway 1992). The initial surface elevation at an MR site depends, among other things, on the period of reclamation for agricultural use before restoration, during which there will have been shrinkage and consolidation of the sediments (Crooks et al. 2002).
Atriplex portulacoides, potentially the physiognomic dominant on marshes around much of the European coastline, was significantly more abundant on older realigned marshes than on reference sites, and it is rapidly increasing in dominance at several MR sites (e.g. Freiston Shore (H.L. Mossman, pers obs). Atriplex portulacoides is intolerant of waterlogging (Davy et al. 2011), particularly at the seedling stage (Chapman 1950) and is therefore not an early colonist of naturally accreting and establishing marshes. However, areas of some managed realignment sites are at elevations in the tidal frame that have suitable sediment conditions for its colonization immediately following reinstatement of tidal flow. In these areas, facilitated succession is not necessary. Species that are quick to colonize, and that are both fast growing and long-lived, may become dominant by inhibiting the invasion of subsequent colonists for as long as they persist (inhibition succession; Connell & Slatyer 1977). Atriplex portulacoides is a long-lived perennial shrub (Chapman 1950). While few individuals have been observed colonizing mature, dense saltmarsh swards (Mohamed 1998), individuals of A. portulacoides can produce very large numbers of fruits, and germination rates can be high (Mohamed 1998). The establishment of species such as Armeria maritima on to realigned sites may be inhibited by the rapid growth and subsequent dominance of Atriplex portulacoides, with opportunities for recruitment of rarer species limited to disturbance events. Our study sites were biased towards the southern UK in which species of predominantly southern distribution, such as Limonium vulgare and Atriplex portulacoides, are prominent, but the general principles should be widely applicable.
Although distance from the nearest potential propagule source was not related to species richness on realigned marshes in this study, low seed viability and long reproductive cycles of Limonium vulgare, Triglochin maritima and Plantago maritima (Boorman 1967; Hutchings & Russell 1989; Davy & Bishop 1991) may slow the ability of these species to colonize and spread within newly realigned marshes. Propagule addition or the transplantation of rarer species on to newly restored sites may increase their frequency in the longer term (Armitage et al. 2006; Varty & Zedler 2008). The creation of small-scale heterogeneity of edaphic conditions may also provide refuges for rarer species (Ewanchuk & Bertness 2004; Varty & Zedler 2008).
To what extent are the restored marshes likely to provide similar ecosystem functions to reference marshes? As we have not measured functions, we do not know how relatively small differences in plant communities might translate into important differences in ecosystem functions and services. A large meta-analysis of wetland restoration suggested that structural recovery may be necessary to achieve functional recovery (Moreno-Mateos et al. 2012). However, equivalent vegetation composition does not guarantee functional equivalency (Zedler & Lindig-Cisneros 2000), so it is unlikely that restored marshes with very different vegetation will be functionally equivalent. Floristic differences on saltmarshes have been associated with substantial functional differences. For example, the dominant plant species in a saltmarsh community has a major influence on metrics of productivity and nitrogen accumulation (Sullivan, Callaway & Zedler 2007; Doherty, Callaway & Zedler 2011). The plants that occur at a particular location are indicative of the redox status of the soil (Davy et al. 2011), and low redox is associated with increased greenhouse gas emissions (Ding, Zhang & Cai 2010; Adams, Andrews & Jickells 2012). Sparse vegetation is likely to result in lower wave energy attenuation (Möller et al. 2001) and lower productivity, whereas areas where the evergreen shrub Atriplex portulacoides is dominant may be more effective in protecting sea walls from wave action. Dominance of species such as Atriplex portulacoides may further affect functioning by inhibiting colonization by less common species that perform differing functions (Zedler, Callaway & Sullivan 2001). Lower abundance of plant species such as Limonium vulgare on realigned marshes will reduce faunal biodiversity, as many invertebrate species are exclusively dependent on these plants (Agassiz et al. 2000). As Limonium vulgare and Armeria maritima flower in summer, when visitor use of marshes is at its highest, their low abundance will reduce the aesthetic and recreational value of created marshes.
While the vegetation of realigned marshes was more similar to reference marshes after 50 or 100 years, some differences in species abundance and diversity remained. This may be related to the relatively high floristic diversity of these marshes; studies of much less diverse North American Atlantic saltmarshes suggest vegetation cover or species richness can reach equivalence within 10 years (LaSalle, Landin & Sims 1991; Morgan & Short 2002). However, vegetation structure expressed as stem density (Zedler 1993), and ecosystem functions, such as carbon burial (Craft et al. 1999; Morgan & Short 2002), may take decades to reach equivalence. We have only examined sites resulting from coastal realignment, but it is likely that marshes restored by other approaches would exhibit similar properties in areas of high floristic diversity. Even where an ecosystem engineer, such as Spartina maritima, is planted extensively, high-marsh development may still depend on subsequent successional processes (Castillo & Figueroa 2009).
It is clear from our work that marshes reactivated by managed realignment do not provide habitats and species in comparable proportions to natural marshes and do not have equivalent biological characteristics. They therefore do not satisfy the requirements of the EU Habitats Directive. It may be possible to improve adherence to the Directive in the future. Additional management interventions, such as the creation of topographic heterogeneity and the planting of mid- and upper-marsh species, could accelerate convergence. On the other hand, given the inherent variation in both natural saltmarshes and local responses to realignment, in the context of projected sea-level rise and climatic change, exact equivalence at individual sites may not be feasible. The requirements of the US Clean Water Act may be more achievable, if implemented by requiring minimum levels of a range of ecosystem functions, and no net loss on larger spatial scales. Nevertheless, a focus on ecosystem functions must not obscure the fact that we do not yet know how to create saltmarshes similar to the flower-rich ‘general saltmarsh’ characteristic of the upper parts of natural marshes and sites with these characteristics merit even higher protection. Planned MR schemes should include measures to encourage the development of these communities, so that we maintain the full range of habitats and species across catchments and regions, even if this is not achieved at every individual site.