Introduce fragments of non-woody plants: freshwater wetlands
Overall effectiveness category Awaiting assessment
Number of studies: 5
Background information and definitions
This action involves introducing fragments of non-woody emergent plants to restore/create marshes or swamps. This includes unrooted cuttings, roots, tubers/bulbs/corms (underground storage organs), rhizomes (underground horizontal stems) or stolons/runners (above-ground horizontal stems). Vegetation fragments may be planted directly into the soil, or spread on the soil surface. Fragments may be obtained from plants raised in greenhouses/laboratories, or collected from natural sites (with potential damage to donor site; Laegdsgaard 2002).
Introduction of target vegetation might be useful in severely degraded or bare sites – which may lack remnant plants or seed banks to kick start revegetation with desirable species, and may be at risk of being taken over by undesirable species (Brown & Bedford 1997). It might also be useful in isolated wetlands, far from sources of marsh or swamp plant propagules. However, note that up-front costs can be high.
The effects of planting may be highly dependent on the environmental conditions in each study. Questions you might ask when interpreting the evidence include: Is the study site degraded? Where and when were fragments introduced? Was there any intervention to improve conditions before planting? What were the environmental conditions over the duration of the study?
The scope of this action does not include planting nurse plants; planting submerged or floating plants; planting to restore bogs, fens, fen meadows or peat swamp forests (see Taylor et al. 2018); or planting facultative wetland plants in upland sites. In contrast, the scope does include planting non-native species to conserve marshes or swamps – whilst acknowledging that this is often considered ethically unacceptable due to the risk of invasion (e.g. Ren et al. 2009).
Related actions: Directly plant whole plants; Introduce seeds or propagules; Transplant or replace blocks of vegetation; Transplant or replace wetland soil; Restore/create marshes or swamps using multiple interventions, often including planting.
Brown S.C. & Bedford B.L. (1997) Restoration of wetland vegetation with transplanted wetland soil: an experimental study. Wetlands, 17, 424–437.
Laegdsgaard P. (2002) Recovery of small denuded patches of the dominant NSW coastal saltmarsh species (Sporobolus virginicus and Sarcocornia quinqueflora) and implications for restoration using donor sites. Ecological Management & Restoration, 3, 202–206.
Ren H., Lu H., Shen W., Huang C., Guo Q., Li Z. & Jian S. (2009) Sonneratia apetala Buch.Ham in the mangrove ecosystems of China: an invasive species or restoration species? Ecological Engineering, 35, 1243–1248.
Taylor N.G., Grillas P. & Sutherland W.J. (2018) Peatland Conservation: Global Evidence for the Effects of Interventions to Conserve Peatland Vegetation. Synopses of Conservation Evidence Series. University of Cambridge, Cambridge.
Supporting evidence from individual studies
A replicated study in 1991–1992 in an excavated freshwater wetland in Pennsylvania, USA (Stauffer & Brooks 1997) reported that 38–79% of planted lurid sedge Carex lurida tubers survived over one growing season. Survival was 79% in plots with added leaf litter, but only 38% in plots without added leaf litter (see Action: Add below-ground organic matter before/after planting). Methods: In October 1991, lurid sedge tubers (number not reported) were transplanted from one wetland into a nearby recently excavated wetland (formerly cropland). The tubers were planted 10 cm deep into eight 6 x 6 m plots, then watered. Leaf litter was mixed into the surface of four plots before planting. Survival was last recorded in August 1992.Study and other actions tested
A replicated study in 1994–1996 in three experimental freshwater wetlands in Minnesota, USA (Yetka & Galatowitsch 1999) reported 0–73% survival of planted sedge Carex spp. rhizomes over 1–9 months, and that the abundance of one species increased over two growing seasons. Statistical significance was not assessed. Overall survival rates were 27% for lake sedge Carex lacustris and 4% for tussock sedge Carex stricta. However, for each species, survival varied with planting season, water regime and elevation. For example, 73% of lake sedge rhizomes were alive in June after planting in spring under a rising water regime. This dropped to 38% for spring-planted rhizomes under a falling water regime, and <2% for autumn-planted rhizomes under any water regime. The study also monitored the abundance of lake sedge in plots planted with that species. After one growing season, there were 14 shoots/m2 and 80 g/m2 above-ground biomass. After two growing seasons, there were 36–39 shoots/m2 and 236–497 g/m2 above-ground biomass (averaged across implementation options). Methods: Field-collected sedge rhizomes were trimmed (to 10 cm length; roots removed) and planted (2–4 cm deep) into three adjacent wetlands. There were 56 rhizomes for each combination of species, season (autumn 1994 or spring 1995), water regime (stable, low over winter/rising through growing season, high over winter/falling through growing season) and elevation (six levels). Survival (presence of living shoots) was monitored in June 1995. Shoots were counted in October 1995 and 1996. Biomass was cut, dried and weighed in August 1995 and 1996.Study and other actions tested
A replicated, randomized, paired, controlled, before-and-after study in 1999–2000 in a floodplain marsh in the Northern Territory, Australia (Paynter 2004) reported that 50% of plots planted with wick grass Hymenachne acutigluma runners contained wick grass after one year, but found that planting had no significant effect on vegetation cover. After one year, wick grass was present in 5 of 10 planted plots (at approximately 1% cover). Presence in unplanted plots was not clearly reported. Planted and unplanted plots had statistically similar cover of vegetation overall (approximately 90%), sedges and grasses overall (approximately 12%) and invasive mimosa Mimosa pigra (approximately 10%). Before planting, plots destined for each treatment had statistically similar cover of vegetation (<1%), dead mimosa stumps (15%) and bare mud (85%). Methods: In November 1999 (at the end of the dry season), fifteen 5 x 5 m plots were established (in five sets of three) on a degraded floodplain marsh. Mimosa had recently been cleared from the marsh using herbicide, crushing and burning. Then, 10 plots (two random plots/set) were planted with locally-collected wick grass runners (36 or 121 runners/plot). The other five plots (one random plot/set) were not planted. Vegetation was surveyed immediately before planting and approximately one year after (October 2000). This study used the same marsh as (4), but a different experimental set-up.Study and other actions tested
A replicated, randomized, paired, controlled, before-and-after study in 2000–2003 in a floodplain marsh in the Northern Territory, Australia (Paynter 2004) found that plots planted with wick grass Hymenachne acutigluma runners contained more wick grass than unplanted plots over three years and had greater vegetation cover after three years, but supported similar mimosa germination rates. Immediately before planting, these plots had no vegetation cover. After one year, wick grass was more frequent and had greater cover in planted plots (present in 10 of 12 plots at 6% cover) than unplanted plots (present in 2 of 12 plots at <1% cover). Overall vegetation cover was statistically similar in planted plots (60%) and unplanted plots (66%). After three years, planted plots still had greater wick grass cover (24%) than unplanted plots (<2%) and now had greater overall vegetation cover (68%) than unplanted plots (50%). Finally, germination rates of invasive mimosa Mimosa pigra did not significantly differ between planted and unplanted plots in any year (see original paper). Methods: In July–September 2000 (at the end of the wet season), twelve pairs of 7.5 x 7.5 m plots were established on a degraded floodplain marsh. Mimosa had recently been cleared from the marsh using herbicide, crushing and burning. Then, one plot in each pair was planted with 16 locally-collected wick grass runners. The other plots were not planted. Vegetation was surveyed immediately before planting and in the following three dry seasons (July–October 2001–2003). This study used the same marsh as (3), but a different experimental set-up.Study and other actions tested
A replicated study in 2010–2012 in a tidal freshwater marsh in California, USA (Sloey et al. 2015) reported that all planted sedge and reed rhizomes died for two of three species, but that they survived and spread for the other species. Three months after planting, all rhizomes of hardstem bulrush Schoenoplectus acutus and broadleaf cattail Typha latifolia had died (i.e. none had produced shoots). In contrast, California bulrush Schoenoplectus californicus rhizomes were alive in all four areas where they were planted, with 6–31% of individual rhizomes having produced shoots. After 24 months, California bulrush was still present in all four areas and had spread to cover 4–23 m2/site. Methods: In June 2010, one hundred and ninety two rhizomes were planted into four areas within the marsh (16 rhizomes/species/area). Survival was quantified in September 2010. Cover was measured until June 2012. The study areas were flooded for 82–99% of each summer.Study and other actions tested