Directly plant non-woody plants: freshwater wetlands
Overall effectiveness category Awaiting assessment
Number of studies: 24
Background information and definitions
This action involves planting whole emergent plants, directly into soil or sediment, to restore/create marshes or swamps. These plants might be individual seedlings, rooted cuttings or mature plants. Plants may be 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 was vegetation planted? 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); planting facultative wetland plants in upland sites; or planting for commercial purposes (e.g. mangrove plantations; Kaly & Jones 1998). 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: Introduce vegetation fragments; Introduce seeds or propagules; Transplant or replace blocks of vegetation; Transplant or replace wetland soil; Introduce organisms to control problematic plants; Introduce nurse plants; 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.
Kaly U.L. & Jones G.P. (1998) Mangrove restoration: a potential tool for coastal management in tropical developing countries. Ambio, 27, 656–661.
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 controlled, site comparison study in 1978–1980 of four freshwater marshes in Florida, USA (Swanson & Shuey 1980) reported that an excavated marsh planted with wetland herbs contained a similar number of plant species to both unplanted and natural marshes. Statistical significance was not assessed. After two years, the planted marsh contained 76 vascular plant species (vs 70 in the unplanted marsh and 76–88 in natural marshes). The planted marsh was dominated by the three planted species whereas the unplanted marsh was dominated by broadleaf cattail Typha latifolia (cover was not quantified). Methods: In summer 1978, two 0.16-ha depressions were excavated in rangeland. One was then planted with three herb species collected from nearby marshes: maidencane Panicum hemitomon, pickerelweed Pontederia lanceolata and common rush Juncus effusus. The other depression was left unplanted. The whole site was seeded with pioneer herbs before planting (to prevent erosion) and limed and fertilized after. In summer 1980, plant species were recorded in each excavated marsh and two natural marshes, along a transect extending from the centre to the edge of each.Study and other actions tested
A before-and-after study in 1992–1993 on a tourist resort in Guam (Ritter & Sweet 1993) reported that a freshwater pool created by excavation, lining with wetland soil and planting herb species contained two of the four planted species after one year, and four additional species. The two planted species present after one year were spikerush Eleocharis dulcis (60% cover) and rusty flatsedge Cyperus oderatus (<1% cover). All planted taro Colocasioa esculenta died; the study suggests it was “excessively flooded”. Planted water lettuce Pistia stratioides was deliberately removed after five months, when it had reached 20% cover. Four additional species were present after one year: two rushes, one grass and one forb (<1–10% cover). Methods: In January 1992, a 600-m2 wetland was excavated on a natural valley slope, lined with wetland soil (30 cm deep) and planted with four herbaceous species (120 spikerush, an unclear number of rusty flatsedge, 20 taro, 5% cover of water lettuce). The study does not distinguish between the effects of these interventions on non-planted vegetation. The wetland was fed by ground and surface water, and had a stable 20–60 cm water depth. Final vegetation cover was estimated in January 1993.Study and other actions tested
A replicated study in 1993–1994 in a freshwater marsh in Ontario, Canada (Chow-Fraser & Lukasik 1995) reported 0–100% survival of planted emergent herbs, depending on water depth and use of silt screens. Vegetation was surveyed approximately one year after planting. Across three plots in shallow water (15–20 cm at planting), approximately 24% of arrowheads Sagittaria latifolia and 10% of broadleaf cattails Typha latifolia survived. Cattail survival was also reported, but not quantified, in two other shallow-water plots. In seven of eight plots in deep water (≥30 cm at planting), no arrowheads survived. In the other plot, surrounded by a fine-mesh silt screen, all planted arrowheads had survived and spread. Methods: In August 1993, volunteers planted 90 plants into each of thirteen 6-m2 plots in Cootes Paradise Marsh: 30 arrowhead, 30 cattails and 30 submerged plants. All plots were fenced in an attempt to exclude muskrats Ondatra zibethicus, but they entered at least eight plots and ate the above-ground vegetation. Vegetation was surveyed in July and August 1994.Study and other actions tested
A study in 1991–1993 of an excavated and planted freshwater wetland in Ohio, USA (Niswander & Mitsch 1995) reported that it developed vegetation cover, including 13 of 17 planted herb species, after 18 months. Eighteen months after planting, 50 herbaceous plant species were recorded in the marsh and wet meadow zones (vs 35 after six months and 44 after 15 months). Of these, 13 were planted species (12 emergent marsh and wet meadow species, plus one cover crop). The other 37 species had colonized spontaneously. No submerged vegetation was recorded within pools in the wetland. Methods: In spring 1992, seventeen wetland herb species (including three intended as cover crops) were planted into flooded and saturated areas of an excavated wetland (two connected basins; 6.1 ha total area; excavated from former farmland in autumn 1991). In autumn 1992, summer 1993 and autumn 1993, herbaceous plant species were recorded along six transects spanning the wetland. The study does not distinguish between the effects of planting and excavation on non-planted vegetation.Study and other actions tested
A replicated study in 1987–1989 of lakeshores planted with bulrushes Scirpus spp. in the Netherlands (Clevering & van Gulik 1997) reported that where bulrushes persisted over three growing seasons, their density, biomass and extent increased. Statistical significance was not assessed. After three growing seasons, lakeshore bulrush Scirpus lacustris ssp. lacustris was present in all three sites where it was planted. There were 370–390 shoots/m2 with 1,730–2,360 g/m2 biomass (vs only 70–130 shoots/m2 and 90–430 g/m2 biomass after one growing season). Saltmarsh bulrush Scirpus maritimus was present in two of three sites where it was planted. There were 70–220 shoots/m2 with 310–1,070 g/m2 biomass (vs only 20–40 shoots/m2 and 30–60 g/m2 biomass after one growing season). In the other site, plants were uprooted by muskrats Ondatra zibeticus. Finally, in four of four cases with data, bulrush had spread outside planted plots (by 27–372 cm, on average). Methods: In spring 1987, each bulrush species was transplanted (12 plants/m2) into 24 plots (6–25 m2) across three sites. All sites were at the margins of freshwater lakes and two were tidal. Half of the plots were fertilized at planting and all were fenced to exclude waterfowl. Bulrush shoots were counted, and above-ground dry biomass estimated from length-mass relationships, in spring and summer until August 1989. Lateral spread was recorded in July 1989.Study and other actions tested
A study in the mid-1990s in a reflooded freshwater wetland in Israel (Kaplan et al. 1998) reported variable survival of five planted emergent species. Three species were planted in saturated soils, without protection from herbivores. Survival rates were low for dwarf waterclover Marsilea minuta (0 of 12 alive eight months after planting) and flowering rush Butomus umbellatus (1 of 19 alive four months after planting). Cover of water purslane Ludwigia palustris increased from 0.2 m2 when planted to 1.8 m2 after 26 months. For the other two species survival rates depended on soil type, water depth and/or herbivore protection. For example, the highest survival of yellow flag iris Iris pseudacorus occurred in peat soils, in 30 cm of water and with herbivore protection (75% of 16 plants alive after 18 months). The highest survival of papyrus Cyperus papyrus occurred in peat soils, without standing water and without herbivore protection (86–89% of 36 plants alive after 12 months). Methods: Plants native to the study area were planted into recently rewetted cropland. The plants were sourced from nearby natural wetlands, botanical gardens or private collections. Some plants were protected from herbivores with wire mesh (5 x 5 cm holes) and plastic netting. The study does not report precise planting and survey dates.Study and other actions tested
A replicated, paired, controlled study in 1992 in a wet meadow restoration site in Iowa, USA (van der Valk et al. 1999) reported that the number of shoots on tussock sedge Carex stricta seedlings increased over two months after planting. Statistical significance was not assessed. Two weeks after planting, there were 4.7–5.8 shoots/plant. Two months after planting, there were 11.5–15.5 shoots/plant. Adding compost or fertilizer sometimes increased the number of shoots/plant, but adding topsoil never had a clear effect (see Actions: Add upland topsoil before/after planting, Add inorganic fertilizer before/after planting and Add below-ground organic matter before/after planting). Methods: In June 1992, tussock sedge seedlings were planted into twelve sets of eight 1-m2 plots of mineral soil (topsoil had been removed). The number of seedlings/plot was not clearly reported. Seven plots/set were amended with topsoil, Scott’s® Starter Fertilizer and/or composted garden waste.Study and other actions tested
A replicated, before-and-after study in 1995–1997 in three recently excavated wet basins in Minnesota, USA (Budelsky & Galatowitsch 2000) reported >37% survival of lake sedge Carex lacustris in each of the three years after planting, and increases in sedge biomass, density and height over time. In the first growing season after planting, the sedge survival rate ranged from 37% in the wettest plots to 95% in the driest. In the next two growing seasons, the average survival rate across all plots was ≥95%. Amongst variation related to planting density, water regime, elevation and weeding treatments (see Actions: Actively manage water level after planting and Remove vegetation that could compete with planted vegetation), there were significant increases over time in sedge biomass (from 12–81 g/m2 after one growing season to 272–1,160 g/m2 after three growing seasons), density (from <20 stems/m2 when planted to 143–219 stems/m2 after three growing seasons) and average height (from <35 cm when planted to 88–102 cm after three growing seasons). Maximum height also increased, but this was not statistically tested (<80 cm when planted; 139–158 cm after three growing seasons). Methods: In May 1995, nursery-reared lake sedge was planted into 48 bare, 5-m2 plots (10 or 45 plants/plot) across three wet basins (same as in Study 10). Each basin was managed with a different water regime: falling, stable or rising throughout the growing season. The plots were situated at four different elevations, and half were weeded (colonizing plants removed) throughout the study. Vegetation was surveyed through the 1995, 1996 and 1997 growing seasons.Study and other actions tested
A study in 1996–1997 in a mine tailing pool in Ireland (McCabe & Otte 2000) reported that planted floating sweetgrass Glyceria fluitans grew. Over 6–14 months after planting, the sweetgrass plants grew 0.5–2.9 shoots/month. Live above-ground biomass increased by 0.1–0.6 g/month. A batch of sweetgrass planted in deeper water in spring grew faster than a batch planted in shallower water the previous autumn (see original paper). Methods: A total of 21 wild-collected sweetgrass plants (runners with two shoots; 5.5–7.0 g fresh mass) were planted into a pool of mining waste. Fourteen plants were planted in July 1996 (in 20–30 cm deep water, but water table dropped below surface in summer). Seven plants were planted in March 1997 (in 30–50 cm deep water, with sediment always waterlogged). In September 1997, all sweetgrass shoots were counted then harvested, dried and weighed.Study and other actions tested
A replicated study in 1995–1997 in three recently excavated wet basins in Minnesota, USA (Budelsky & Galatowitsch 2004) reported >48% survival of planted tussock sedge Carex stricta in each of the three years after planting, and increases in sedge biomass and stem number over time. In the first growing season after planting, the sedge survival rate ranged from 48% in the wettest plots to 99% in the driest. In the next two growing seasons, the survival rates in all plots were ≥90%. Amongst variation related to water regime, elevation and weeding treatments (see original paper), there were significant increases over time in sedge biomass (from 2–16 g/plant after one growing season to 49–234 g/plant after three growing seasons) and stem number (from <5 stems/plant when planted to 50–310 stems/plant after three growing seasons). Methods: In May 1995, nursery-reared tussock sedge was planted into 48 bare, 5-m2 plots (10 or 45 plants/plot) across three wet basins (same as in Study 8). Each basin was managed with a different water regime: falling, stable or rising throughout the growing season. The plots were situated at four different elevations, and half were weeded (colonizing plants removed) throughout the study. Vegetation was surveyed through the 1995, 1996 and 1997 growing seasons.Study and other actions tested
A replicated, paired, controlled study in 2001–2004 in 12 ephemeral freshwater wetlands undergoing restoration in South Carolina, USA (De Steven & Sharitz 2007) found that plots planted with southern cutgrass Leersia hexandra and maidencane Panicum hemitomon had greater cover of wetland-characteristic vegetation than unplanted plots, but similar overall vegetation cover and species richness. After approximately three years, planted plots had greater cover of wetland-characteristic vegetation (overall: 65–79%; cutgrass and maidencane: 41–66%) than unplanted plots (overall: 54%; cutgrass and maidencane: 0%). However, total vegetation cover did not significantly differ between treatments (planted: 87–100%; unplanted: 90%). The same was true for plant species richness: both for wetland-characteristic species, including cutgrass and maidencane (planted: 3.8–4.1 species/4 m2; unplanted: 4.7 species/4 m2) and all species (planted: 6.8–7.4 species/4 m2; unplanted: 8.7 species/4 m2). The study also reported data from one year after intervention, during a drought (see original paper). Methods: Twenty-four plots (each 80–150 m2) were established across 12 wetlands undergoing restoration (drainage ditches plugged and trees cleared in 2000–2001). In April–May 2001, southern cutgrass and maidencane were each transplanted into 12 plots (1 plot/wetland; 2–3 plants/m2, giving 1–4% cover). In August 2002 and 2004, plant species and cover (excluding resprouting trees) were recorded in three 4-m2 quadrats/wetland: one quadrat/plot and one in the adjacent, unplanted area.Study and other actions tested
A before-and-after, site comparison study in 1996–2008 aiming to restore a reedbed on farmland in England, UK (Booth & Ausden 2009) found that an area planted with common reed Phragmites australis (after excavating wet basins) contained a greater density of reeds but fewer plant species than a nearby natural reedbed. The restored area was initially drained farmland. Five years after planting finished, the restored area contained a greater density of live reeds (96 stems/m2) than the natural reedbed (63 stems/m2). There was no significant difference in the density of dead reeds (restored: 52; natural: 48 stems/m2). Although the restored area contained fewer plant species than the natural reedbed at a large scale (restored: 5; natural: 9 species/30 m2), both sites had the same species richness at a small scale (3 species/2 m2). Statistical significance of these richness results was not assessed. Methods: Between 1996 and 2003, a quarter of a million common reed stems were planted into 300 ha of excavated wet basins. The study does not distinguish between the effects of reed planting and excavation on non-planted vegetation. In August 2008, reed stems and plant species and were recorded in thirty 2-m2 quadrats: 15 in the restoration area and 15 in a natural (never-farmed) reedbed.Study and other actions tested
A study in 1997–2006 of a levelled, irrigated and partially planted freshwater marsh in California, USA (Miller & Fujii 2010) reported that it developed vegetation dominated by emergent plants, including planted tule Schoenoplectus acutus – although vegetation cover and density depended on the water level. After 2–9 years, the shallower half of the site had 89–98% total vegetation cover. This included 77–81% cattail Typha spp., 11–19% tule and 0–5% submerged vegetation cover. Emergent vegetation density fluctuated between 49 and 76 stems/m2. The deeper half of the site had 77–100% total vegetation cover, including 38–58% cattail, 3–8% tule, and 10–46% submerged vegetation cover. Emergent vegetation density fluctuated between 44 and 59 stems/m2. Across the entire site, above-ground biomass of emergent vegetation was 1,630 g/m2 after 1–3 years (vs submerged, floating and algae combined: 389 g/m2) then fluctuated between 925 and 2,360 g/m2 for the following six years. Methods: In autumn 1997, tule was planted into 0.5 ha of a 6-ha site: 0.25 ha in the shallower half (25 cm water depth) and 0.25 ha in the deeper half (55 cm water depth). The site used to be farmland, but had been levelled before planting and was continuously irrigated after. The study does not distinguish between the effects of planting, levelling and irrigation on non-planted vegetation. All plants and algae were surveyed along transects, in summer/autumn, at least biennially between 1998 and 2006. Biomass was cut, dried and weighed (years 1–3) or estimated from plant height and diameter (years 4–9).Study and other actions tested
A replicated study in 2005–2006 of 22 lakeshore restoration sites in Minnesota, USA (Vanderbosch & Galatowitsch 2010) reported that 17–40% of planted/sown species reliably established across multiple sites, and that no planted/sown species established in some individual sites. In the seasonally flooded zone, only 22 of 128 planted/sown species reliably established (survived in >75% of sites where planted, or ≥25% cover in ≥1 site). Fifty-six species failed to establish at any site. However, some planted/sown species established at 100% of sites. In the permanently flooded zone, 10 of 25 planted/sown species reliably established. Six species failed to establish at any site. Planted/sown species completely failed to establish at 27% of sites. Methods: In summer 2005 and spring 2006, plant species and their cover were surveyed in 22 urban lakeshore restoration projects. Native plants had been introduced between 1999 and 2004. Species lists were obtained from project reports or interviews with staff. Almost all introduced plants were emergent herbs, and most (but not all) were wetland species. Some plants were directly planted (as plugs or on pre-vegetated coconut-fibre mats) and some were sown. The study does not distinguish between the effects of planting and sowing. Most sites were protected with fences and/or wave breaks, at least for the first growing season after planting/sowing.Study and other actions tested
A replicated study in 2006–2007 along the shores of five freshwater lakes in Minnesota, USA (Vanderbosch & Galatowitsch 2011) reported 15% overall survival of planted softstem bulrush Schoenoplectus tabernaemontani plugs after approximately one year. The survival rate per lake ranged from 4% to 31%. It was significantly affected by month of planting and plug type, but not water depth. For example, plugs grown in pots and planted in June had the highest survival rate (39%) whilst plugs grown in pots and planted in August or grown in mats and planted in September had the lowest survival rate (3%). Methods: Between May and September 2006, a total of 3,750 bulrush plugs were planted along degraded shorelines within five urban lakes (150 plugs/month/lake, spaced 45 cm apart). Each greenhouse-reared plug contained 3–5 individual plants. Biomass at planting varied between months. Half of the plugs had been grown in pots and half on coconut fibre mats. Half were planted in shallow water (maximum depth 0–30 cm) and half in deep water (maximum depth 31–60 cm). All planted areas were fenced to exclude muskrats Ondatra zibethicus. Plugs containing ≥1 live plant were recorded until May–June 2007.Study and other actions tested
A site comparison study in 2008 of five sedge meadows in Illinois and Wisconsin, USA (Lawrence & Zedler 2013) found that a restored meadow – planted with plugs of tussock sedge Carex stricta, after removing trees and excess sediment – contained more but smaller sedge tussocks than nearby natural meadows after 11–14 years. In four of four comparisons, the restored meadow contained a greater density of sedge tussocks (8.4 tussocks/m2) than natural meadows (4.5–5.6 tussocks/m2). Sedge tussocks were also smaller in the restored meadow than in the natural meadows. This was true in four of four comparisons for height (restored: 5 cm; natural: 11–18 cm), perimeter (restored: 39 cm; natural: 51–82 cm) and volume (restored: 560 cm3; natural: 2,342–6,604 cm3). The basal area of tussocks in the restored meadow was only 0.07 m2/m2, compared to 0.12–0.23 m2/m2 in the natural meadows (statistical significance not assessed). Methods: In 2008, sedge tussocks were surveyed in one restored and four natural sedge meadows (15–30 quadrats/meadow, each 1 m2). The restored meadow was formerly a wooded floodplain. Trees and accumulated sediment were removed, then plugs of tussock sedge planted 30 cm apart, between 1994 and 1997. The study does not distinguish between the effects of these interventions on any non-planted sedges.Study and other actions tested
A replicated, before-and-after study in 2010–2013 in a seasonally flooded depression on farmland in South Dakota, USA (Zilverberg et al. 2014) reported that planted prairie cordgrass Spartina pectinata occurred in 65–86% of sampled quadrats after two growing seasons and 90–100% of quadrats after four, depending on elevation. Two growing seasons after planting, 65% of quadrats at low elevations (≤10 cm from wetland bottom) and 86% of quadrats at higher elevations (>10 cm from wetland bottom) contained at least one cordgrass stem. Four growing seasons after planting, cordgrass plants had spread and there was at least one stem in 100% of quadrats at low elevations and 90% of quadrats at higher elevations. Methods: Four plots were established in a historically cultivated ephemeral wetland. Each plot ran perpendicular to the slope of the wetland, so included a range of elevations. Spring floodwaters were typically 50 cm deep. In spring 2010, each plot was planted with >760 greenhouse-reared cordgrass plugs (90 cm apart). All plots were mown once in 2011 to control weeds. Each autumn from 2011 to 2013, cordgrass presence was surveyed in 1-m2 quadrats along the length of each plot. This study used the same farm as (18), but used a different experimental set-up.Study and other actions tested
A replicated study in 2008–2013 in two seasonally flooded depressions on farmland in South Dakota, USA (Zilverberg et al. 2014) reported 91% survival of transplanted prairie cordgrass Spartina pectinata after one growing season. The study also measured wet, above-ground vegetation biomass. After two growing seasons, cordgrass biomass was greater in plots with closely spaced transplants (0.9 m apart: 4 Mg/ha) than loosely spaced transplants (1.5 m apart: 2 Mg/ha). After 3–6 growing seasons, total above-ground biomass (including plants other than cordgrass) did not significantly differ between transplant density treatments (closely spaced: 10–16 Mg/ha; loosely spaced: 8–14 Mg/ha). Methods: In May–July 2008, greenhouse-reared cordgrass plugs were transplanted into two historically cultivated, ephemeral wetlands (corn and soybean fields in the years before planting). Half of each wetland was planted at each transplant spacing. The wetlands were sprayed with herbicide before planting, and individual invasive plants were sprayed or pulled up after planting. Cordgrass survival was monitored in October 2008. Between 2009 and 2013, vegetation was cut from an average of twenty-four 1-m2 quadrats/wetland/year, then weighed in the field. This study used the same farm as (17), but used a different experimental set-up.Study and other actions tested
A replicated study in 2012–2013 in a freshwater wetland in Wisconsin, USA (Doherty & Zedler 2015) reported 27–100% survival of planted tussock sedge Carex stricta over 1–2 growing seasons, and that survivors grew. Survival rates depended on how wet plots were, and how sedges were planted. Survival was lowest for sedges planted into 16-cm-tall soil mounds in a drier area (27% after two growing seasons) and highest into 8-cm-tall mounds, peat pots or flat ground in a wetter area (100% after two growing seasons). Surviving plants grew, on average – although not in all cases during the first growing season (2012), when there was a drought (see original paper for data). After 1–2 growing seasons, planted plots contained 3–70% tussock sedge cover. As for survival, variation was related to plot wetness and planting method. Methods: Across spring 2012 and 2013, a total of 300 nursery-reared tussock sedges were planted into 60 plots in a wetland undergoing restoration (five sedges/plot). An invasive shrub that had colonized the site was cut down in January each year. Half of the plots were in a wetter area and half in a drier area. The sedges were planted into mounds, hollows or peat pots in 48 of the plots, and into flat ground in the other 12 plots. Sedges planted in 2012 were regularly watered and weeded. Survival and cover were surveyed in June–August 2013. Growth rates were calculated from leaf lengths measured in 2012 and 2013.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 planted sedges and reeds survived for three months in 10 of 12 cases, but were present after two years in only 2 of 12 cases. Three species were planted in each of four areas. After three months, all four areas contained planted California bulrush Schoenoplectus californicus (44–94% of individual plants alive) and hardstem bulrush Schoenoplectus acutus (25–75% of individual plants alive). However, only two of four areas contained planted broadleaf cattail Typha latifolia (where 25–56% of individual plants were alive). After 24 months, the bulrush species were each present in only one of four areas (area covered: 0.1–2.4 m2) and broadleaf cattail was not present in any area. For all species, initial survival was statistically similar in open water areas (0–94%) and on the marsh fringe (0–75%). Methods: In June 2010, one hundred and ninety two nursery-grown plants were planted into four areas within the marsh (16 plants/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
A study in 2005–2013 of an excavated, planted and harvested water treatment marsh in Sardinia, Italy (De Martis et al. 2016) reported that it supported 275 plant taxa. This included 201 plant species in 161 genera. Approximately 63% of the taxa were Mediterranean (found predominantly or solely in this region) and approximately 16% were known non-natives in Italy. As expected in the study area, 56% of the taxa were annual plants that complete their life cycle rapidly in favourable conditions (“thereophytes”). Only 2% of taxa had underwater resting buds (“hydrophytes”). Methods: Between 2005 and 2013, plant taxa were recorded in the 37-ha EcoSistema Filtro marsh, which had been constructed with the dual aims of habitat creation and water treatment. There were monthly surveys (a) across the whole site, including banks and upland areas, and (b) in three 16-m2 plots, each April–July and September–December. The wetland had been constructed by excavating basins of varying salinity and levees (including removal of all existing vegetation; beginning 1990) and planting bundles of 2-m-tall common reed Phragmites australis (2004). Some “plant biomass” was mechanically removed between 2005 and 2007. Note that this study evaluates the combined effect of these interventions, and does not separate results from fresh, brackish and saline areas.Study and other actions tested
A replicated, site comparison study in 2013–2015 around two fresh/brackish lakes in South Australia (Jellinek et al. 2016) found that planted stands of river club-rush Schoenoplectus tabernaemontani became more similar to mature natural stands over time – in terms of structure and rush abundance – and supported similar near-shore vegetation to the natural stands within 8 years. Older planted rush stands were more similar to mature natural stands in terms of stand width (young planted: 1–3 cm; old planted: 5–12 cm; natural: 35 cm), maximum height (young: 60–142 cm; old: 131–152 cm; natural: 155 cm) and stem density (data not reported). All stands were a similar average height (data not reported). Near-shore vegetation (i.e. between the rush stands and the shoreline) behind older planted rush stands was similar to that behind mature natural stands, whereas young planted stands supported similar near-shore vegetation to areas without rush stands. This was true for overall community composition (data reported as graphical analyses; statistical significance of differences not assessed), plant species richness (no rushes: 30; young: 45; old: 150; natural: 330 species/site) and abundance (no rushes: 940; young: 1,370; old: 14,000; natural: 31,300 plants/site). Methods: In autumn 2013–2015, vegetation was surveyed at 21 sites on the margins of two connected fresh/brackish lakes. Ten sites had been planted with nursery-reared rushes (1 m apart): six sites ≤3 years ago (young plantings) and four sites 8–11 years ago (old plantings). Three sites had mature natural rush stands (≥20 years old) and eight had no rushes. All sites were fenced to exclude livestock. Rush stands were surveyed in five 1-m2 quadrats/site/year. Other near-shore vegetation was surveyed in approximately thirty-six 3-m2 quadrats/site/year.Study and other actions tested
A before-and-after study in 2008–2014 in two sites containing freshwater marshes and wet meadows in southern China (Liu et al. 2016) reported that after planting herbs into the wetlands (and a polluted river feeding them), plant species richness, diversity, cover, biomass and height all increased. Statistical significance was not assessed. The sites contained 13–14 plant species before planting but 26–42 plant species five years after. In the one site for which data were reported, marsh and wet meadow habitats experienced increases in plant diversity (data reported as diversity indices), total vegetation cover (from 22–64% to 64–93%), total vegetation biomass (from 520–638 g/m2 to 768–919 g/m2) and vegetation height (from 43–86 cm to 86–161 cm). Methods: In May 2009, two degraded wetland sites were planted with herbaceous plants (number of species not reported). In one site, a lakeshore marsh was planted with emergent and floating herbs. In the other site, slightly uphill from the lake, marshes and meadows were planted with forbs and grasses. The river feeding the lake was also planted with pollution-reducing vegetation. The study does not distinguish between the effects, on any non-planted vegetation, of planting directly in the lakeshore marsh and planting in the river. Vegetation (emergent, floating and submerged) was surveyed before (July 2008) and for approximately five years after (July 2009–2014) planting (details not fully reported).Study and other actions tested
A replicated, randomized, controlled study in 2013–2014 in a degraded floodplain swamp in Florida, USA (Smith et al. 2016) reported 2–57% survival of planted wetland herbs after one year, and found that planted plots had higher plant species richness than unplanted plots. Four herb species were planted. After one year, survival rates were 2% for purple bluestem Andropogon glomeratus and pine barren goldenrod Solidago fistulosa, 46% for common rush Juncus effusus, and 57% for red-top panic grass Panicum longifolium. The study reported that cattle damaged or completely removed some plants, especially purple bluestem and pine barren goldenrod. Over the year following planting, plant species richness was higher in planted plots (total: 5.2; native: 3.8 species/0.56 m2) than in unplanted plots (total: 1.8; native: 0.6 species/0.56 m2). Planting had no significant effect on Mexican petunia Ruellia simplex density, cover or biomass (see Action: Introduce organisms to control problematic plants). Methods: Fourteen 1.5 x 1.5 m plots were established in a floodplain swamp where invasive Mexican petunia had been controlled (but not eradicated) with herbicide. In November 2013, seven random plots were planted with greenhouse-reared herbs (four species; four plants/species/plot; individual plants 30 cm apart). The other seven plots were not planted. Vegetation was surveyed for one year after planting. During this period, surface water was present in 6 of 12 months and was up to 21 cm deep.Study and other actions tested