Phragmites australis dieback in the Mississippi River Delta, USA: Vegetation restoration

doi:10.21203/rs.3.rs-2489228/v1

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Phragmites australis dieback in the Mississippi River Delta, USA: Vegetation restoration

https://doi.org/10.21203/rs.3.rs-2489228/v1

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published 23 Oct, 2023

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The Mississippi River Delta (MRD) is one of the largest wetlands in the world and is experiencing widespread dieback of its most prevalent plant species, Phragmites australis. The potential for revegetation of dieback sites was assessed by planting three genetically distinct lineages of P. australis at nine sites in the MRD that varied in water depth. These lineages include Delta, the most prevalent in the MRD, EU, a highly invasive lineage, and Gulf, a common lineage of the Gulf Coast and southern United States. Four additional wetland species were planted at a subset of these sites. Plantings were monitored for 32 months during which time, we surveyed plantings for survival, growth and infestation by a scale insect that has been implicated in P. australis dieback. Survivorship differed considerably among plant types and by 6 months, Delta, Gulf and California bulrush (Schoenoplectus californicus) had twice the survivorship of the other plant types. However, by the end of the study, Delta, Gulf and bulrush were the only plant types alive, with Delta persisting in more than twice as many sites and having > 6 times more area of coverage as Gulf and bulrush. For all plant types, water depth was negatively related to survivorship. Finally, although scale insects achieved twice the abundance on Delta than the other lineages, we found no evidence that scales were inhibiting the growth and survival of P. australis. Despite potential negative legacy effects of dieback, Delta can be used revegetate shallow-water sites within a year following dieback.

Invasive species

legacy effects

Nipponaclerda biwakoensis

roseau cane scale

Schoenoplectus californicus

water depth

Globally, the Mississippi River has the seventh largest discharge of freshwater, nutrients and sediments in the world (Milliman & Meade, 1983; Allison, Demas, Ebersole, Kleiss, Little, Meselhe, Powell, Pratt, & Vosburg, 2012). Currently, about two-thirds of the river is drained via the Holocene-age deltaic plain known as the Mississippi River Deltaic Plain (MRD) (Saucier, 1994). This coastal wetland is amongst the most productive ecosystems in the world, providing invaluable economic and ecological services including flood control, habitat for wildlife, maintenance of water quality, support of natural resource-based industries and recreational and commercial fisheries (Day, Martin, Cardoch, & Templet, 1997; Day, Ko, Rybczyk, Sabins, Bean, Berthelot, Brantley, Cardoch, Conner, Day, Englande, Feagley, Hyfield, Lane, Lindsey, Mistich, Reyes, & Twilley, 2004). Despite its economic value, the MRD has experienced amongst the highest rates of coastal wetland loss worldwide (Day, Britsch, Hawes, Shaffer, Reed, & Cahoon, 2000; Barras, Beville, Britsch, Hartley, Hawes, Johnston, Kemp, Kinler, Martucci, Porthouse, Reed, Roy, Sapkota, & Suhayda, 2004; Allison & Meselhe, 2010).

An important contributor of wetland loss in the MRD has been vegetation dieback (Webb & Mendelssohn, 1996; Alber, Swenson, Adamowicz, & Mendelssohn, 2008; Suir, Saltus, & Reif, 2022). In 2000, over 10,000 ha of Spartina alterniflora Loisel (= Sporobolus alterniflora; Poaceae) dominated salt marshes experienced a sudden and acute dieback event (termed “brown marsh”). More recently, the dominant emergent vegetation in the Lower MRD, a perennial marsh grass, Phragmites australis (Cav.) Trin. ex Steud has experienced extensive dieback (Ramsey III & Rangoonwala, 2017; Knight, Wilson, Gill, Aviles, Cronin, Nyman, Schneider, & Diaz, 2018; Suir, Saltus, & Reif, 2018; Suir et al., 2022). Large-scale dieback events have been reported from other regions of the world including P. australis throughout Europe (van der Putten, 1997; Brix, 1999) and in China (Li, Shao, Qiu, & Li, 2013), seepweed (Suaeda salsa L.) in China (He, Silliman, Liu, & Cui, 2017), and mangrove forests in Australia, South Africa and the Caribbean (Breen & Hill, 1969; Duke, Kovacs, Griffiths, Preece, Hill, van Oosterzee, Mackenzie, Morning, & Burrows, 2017; Lovelock, Feller, Reef, Hickey, & Ball, 2017; Rossi, Archer, Giri, & Layman, 2020). Vegetation dieback events are often precursors to land loss because they result in the conversion of marshes into bare mudflats or open water (Turner, 1990; van der Putten, 1997; Brix, 1999).

In the MRD, P. australis inhabits freshwater and brackish wetland systems and grows in dense monospecific stands. Symptoms of P. australis dieback include retreat from deep waters, clumped distribution of stems, stunted stem and root growth, premature senescence, and increased occurrences of dead or decaying rhizomes (Armstrong, Armstrong, & van der Putten, 1996; van der Putten, 1997; Brix, 1999). The cause of dieback in P. australis is complex and currently under investigation. However, one important contributing factor appears to be a nonnative scale insect, Nipponaclerda biwakoensis (Kuwana) (Hemiptera: Aclerdidae), whose discovery coincided with the appearance of widespread P. australis dieback in the MRD (Knight et al., 2018; Cronin, Johnston, & Diaz, 2020; Knight, Cronin, Gill, Nyman, Wilson, & Diaz, 2020). The scales reach mean abundances in the summer of 150 per meter of stem and reduce P. australis growth and biomass (Knight et al., 2020). Moreover, scale insect abundance across stands of P. australis in the MRD is negatively correlated with the normalized difference vegetation index (NDVI); a measure of standing plant biomass (I. A. Knight, G. Suir and R. Diaz, unpublished data). Abiotic factors associated with climate change and human activities such as extended flooding, elevated soil salinity or nutrient pollution have also been linked to P. australis dieback (Cronin et al., 2020; Knight et al., 2020; Lee, Diaz, & Cronin, 2023).

An additional factor complicating our understanding of dieback in the MRD is that there are three genetic and ecologically distinct lineages of P. australis that differ in their susceptibility to dieback. The dominant lineage in the MRD is Delta (haplotype M1; Hauber, Saltonstall, White, & Hood, 2011), accounting for > 95% of P. australis coverage (Hauber et al., 2011; Knight et al., 2018). It is of North African and Mediterranean origin (Hauber et al., 2011; Lambertini, Mendelssohn, Gustafsson, Olesen, Riis, Sorrell, & Brix, 2012). Another lineage found in the MRD is of Eurasian origin (haplotype M; hereafter, the EU lineage) is highly invasive of coastal and wetland marshes throughout the continental United States and southern provinces of Canada (Saltonstall, 2002). The EU lineage did not occur in the MRD until the 1990s when it was first noted in the Lower MRD (Hauber et al., 2011). Stands of the EU lineage are often found embedded within or contiguous with much larger Delta stands (Knight et al., 2018; Cronin et al., 2020). Finally, the Gulf lineage (haplotype I, Land-type or subspecies P. australis berlandieri) is widely distributed along the southernmost United States from Florida to California and south of the border into Central and South America (Saltonstall, 2002; Meyerson, Lambert, & Saltonstall, 2010; Hauber et al., 2011; Colin & Eguiarte, 2016). The Gulf lineage inhabits more upland and less saline areas of the marsh (Achenbach & Brix, 2014; Cronin et al., 2020). Based on field observations and experimental studies, the EU lineage has lower scale densities, is more tolerant of scale infestations and inundation (i.e., has proportionately greater biomass accumulation in the presence of scales), and is less prone to exhibiting symptoms of dieback than the Delta lineage (Cronin et al., 2020; Knight et al., 2020).

Restoration has become a key conservation strategy to combat coastal wetland degradation, promote biological diversity and enhance biogeochemical functions (Zhou, Qin, & Xie, 2003; Nyman, Walters, Delaune, & Patrick, 2006; Ogburn & Alber, 2006; Silliman, Schrack, He, Cope, Santoni, van der Heide, Jacobi, Jacobi, & van de Koppel, 2015). With the acceleration of coastal habitat loss due to climate change and anthropogenic activities, planting habitat-forming species has been widely adopted for wetland plant species (Wilsey, McKee, & Mendelssohn, 1992; Rolletschek, Rolletschek, Kuhl, & Kohl, 1999; Edwards, Travis, & Proffitt, 2005; Howard, Travis, & Sikes, 2008; Brix, Ye, Laws, Sun, Li, Ding, Yuan, Zhao, Wang, & Pei, 2014; Silliman et al., 2015). However, in dieback events, habitat restoration through natural revegetation or replanting efforts may be hindered by the presence of adverse edaphic, hydrological or biotic stressors that were the cause of consequence of dieback (i.e., legacy effects; Crawford & Stone, 2015). For example, dieback soils in wetland marshes are known to have toxic levels of sulfides (Armstrong et al., 1996; Webb & Mendelssohn, 1996; van der Putten, 1997; Brix, 1999) that could inhibit revegetation.

Given the unprecedented and substantial loss of P. australis in the Lower MRD (Suir et al., 2022), we conducted an in situ restoration field trial to assess the potential for revegetation of recent P. australis dieback sites. At nine dieback sites varying in water depth, we planted three lineages of P. australis (Delta, EU and Gulf). Although P. australis is our primary focus, at some of the sites, we also planted up to four other common wetland plant species: Schoenoplectus californicus (C.A. Meyer) Palla (California bulrush), Spartina alterniflora L. (tall cordgrass), Spartina patens (Aiton) Muhl. (saltmeadow cordgrass) and Zizaniopsis miliacea (Michx.) Döll. & Asch. (giant cutgrass). Plantings took place in 2018 and were monitored (through ground surveys and drone imagery analysis) over the next 32 months to assess the growth and survival of each plant lineage and species. For P. australis, we also monitored scale insect abundance as these insects may play a role in inhibiting successful re-establishment. We predicted that (1) because these sites were formerly occupied by P. australis, P. australis lineages would have higher survival and growth than the other plant species; (2) the EU lineage of P. australis would outperform the other lineages of P. australis because it more tolerant of scale insects and dieback conditions (Cronin et al., 2020; Knight et al., 2020; Lee et al., 2023); (3) shallow-water sites would have higher initial survivorship of plantings because wetland transplants are often limited by inundation (Lenssen, Ten, & Blom, 1998; Lenssen, Menting, van der Putten, & Blom, 1999; Konisky & Burdick, 2004; Howard & Rafferty, 2006); and (4) scale insects would be more abundant on the Delta and Gulf than EU lineages, potentially limiting restoration using these former lineages (Cronin et al., 2020; Knight et al., 2020). Based on our findings, we discuss recommendations for restoration of dieback sites in the MRD.

Study Area

The study was conducted in the Lower MRD, Plaquemines Parish, Louisiana, USA along two of three main channels stemming from the Head of Passes: Pass-a-Loutre Wildlife Management Area (WMA) and South Pass Campground. Dieback of P. australis has been widespread within these areas since 2016 (Suir et al., 2018; Suir et al., 2022). Pass-a-Loutre WMA and South Pass are composed of coastal freshwater marshes (average salinity < 1 ppt) and river channels and accompanying banks, natural bayous and manmade canals with surrounding beach and barrier island habitats. We chose nine sites for our plantings, seven within Pass-a-Loutre and two within South Pass (Table 1). All sites were coastal freshwater marsh habitats with mean salinity < 1 ppt for each site (range: 0.15 – 3.0 ppt). Water depth was variable over the course of the 32-month study, ranging from 0 cm (exposed mudflat) to 85 cm. Eight of nine sites had a previous history of P. australis dieback (within the past 1-2 years since trials were initiated). The one non-dieback site (Sawdust MC South) was a recently excavated dredge disposal area created by the U.S. Army Corps of Engineers (Park, 2018). All but one of those sites were, at the time of the restoration trial, devoid of any emergent P. australis. At Rattlesnake Bayou, the Delta lineage of P. australis was sparsely distributed amongst other marsh plant species (Colocasia esculenta (L.) (elephant ear), Iris hexagona Walker (Dixie iris), Typha latifolia (L.), Alternanthera philoxeroides (Mart.) (alligatorweed), Sagittaria lancifolia (L.) or Pontederia crassipes (Mart.) (water hyacinth) (Table 1). The minimum distance between sites was 0.3 km (Figure 1).

Sources and propagation of plants

In June of 2018, approximately four months prior to transplantation, plant material was acquired from various sources and propagated by a private nursery, Stream Wetland Services, L.L.C. (Lake Charles, Louisiana; https://streamwetlands.com/). For the EU and Delta lineages of P. australis, rhizome material was excavated from healthy stands along the north bank of Pass-A-Loutre WMA. Because of lower survivorship of Delta during propagation, a second collection of Delta was made one month later from Rockefeller Refuge in southwestern Louisiana. Gulf rhizomes were also collected from Rockefeller Refuge. For S. californicus, (Louisiana ecotype) and Zizaniopsis miliaea (Louisiana ecotype), Stream Wetland Services sourced the plant materials through their nursery. Lastly, S. patens (Gulf Coast ecotype) and Spartina alterniflora were obtained from the USDA Natural Resources Conservation Services, Golden Meadow Plant Materials Center (Golden Meadow, Louisiana). Coordinates for the location of source populations is provided in Table S1. With the exception of S. alterniflora, rhizomes of all plant species were grown in peat-based garden soil in trade-gallon pots (2.8 l). Spartina alterniflora was not potted individually but was derived from small plugs (≈ 5 cm diameter).

Plants for use in transplantation had to meet specific growth criteria: a minimum number of live and actively growing stems with a minimum stem height (see Table S1). For the three lineages of P. australis, the criteria were the same. For the other plant species, owing to their different growth forms, different heights and numbers of live stems were required (Table S1). Within ten days prior to planting at the nine sites, stems were pruned to a maximum height of

122 cm to facilitate transportation and planting.

Experimental setup

On October 18-19, 2018, four months after initial planting, personnel from Stream Wetland Services transported the potted plants and plugs to our sites in Pass-a-Loutre and South Pass. Separately for each plant species and P. australis lineage, plantings were laid out in a gridded plot consisting of four columns and 13 rows (only two plantings in row 13) and a total of 50 plantings (Figure 2, Figure S1). One exception was Rattlesnake Bayou (see below). A team of assistants walked each grid and sunk the plants (sans plastic pot) flush with the soil surface. Individual plantings within a plot were placed 1.5 m apart and the overall area of each plot was 83.6 m² (18.3 m x 4.6 m). The first row was oriented closest to the interior marsh side and the 13th row was closest to the channel, bayou, or open water. Plots for each species and lineage were separated by 4.6 m and aligned in parallel to each other (Figure S1). At all nine sites, plots for each of the three P. australis lineages were established. Owing to the size of some dieback sites and limited plant material, the other four plant species were not represented at every site (see Table S2). For S. californicus, we had sufficient material and space to establish two plots at Chenier Crevasse and Willow Tree Site. Finally, although it wasn’t the ideal situation, for logistical reasons, the arrangement of species/lineage plots was the same at every site (e.g., EU P. australis was always planted in the middle).

Rattlesnake Bayou was the only dieback site with emergent vegetation present. To fit our experimental grid within this space, and avoid overlap with denser emergent vegetation, we modified the layout of plantings. Plantings for a species/lineage were spaced 0.9 m apart in either a 9 × 5 grid (26.8 m²) or 6 × 5 grid (16.7 m²). Plots per species/lineage were separated by 1.5 m. For both grid arrangements, the last row contained only 3 plantings, resulting in either 43 or 28 plantings per plot (Figure S1; Table S2). Overall, a total of 443 plants per P. australis lineage, 428 of S. californicus, 143 of S. patens, 128 of Z. miliacea, and 43 S. alterniflora were transplanted for the restoration field trials for a grand total of 2,071 plants (Table S2).

Field surveys

We conducted field surveys of the plants within each restoration plot and within the nearest-naturally occurring P. australis patch (hereafter within-plot and nearest-patch) at periodic intervals over a 32-month period. The first survey was conducted approximately two months post-planting on 17 December 2018. Subsequent surveys were conducted on 9 April 2019 (6-months), 29-30 July 2019 (9-months), 23 September 2019 (11 months), 15-16 June 2020 (20-months), 10-11 August 2020 (22 months), and 2 June 2021 (32 months). Data collection differed among sites and is described below (see also Table 2).

Plant Survival (2, 6 and 9 months). At each site and for each plant species and lineage, we recorded whether each planting was alive or dead. Plants were marked as “alive” if green plant tissue was visible.Plant survival was also converted to a proportion alive (total # alive/total planted). At two edges and center of each plot per site, we measured water depth and salinity (ppt; using a YSI Professional Plus multimeter; YSI Inc., Yellow Springs, OH, USA). Finally, aerial photographs using a drone (DJI Phantom 3 Professional with a built-in 4K-color camera; DJI Corporation, Shenzhen, Guangdong, China) were taken at each site from a height of 15 m above ground level (for example, see Figure 2).

P. australis scale abundance and plant condition (11 months). This survey was conducted in September (2019), a time period when scale insects are usually near peak abundance (Knight et al., 2020). During this survey, we focused only on the P. australis plantings and the factors that influenced scale insect abundance per planting. At each site and for each P. australis lineage, we randomly selected five plantings (using the RAND function in Microsoft Excel 365; Microsoft Corp, Redmund, Washington, USA). For each planting, we inserted a 0.25 m² ring over its center and counted the number of live stems within that ring. We converted this count to the number of stems per m². Water depth was measured at the perimeter of each growing planting. Three live stems were haphazardly selected, clipped at the soil surface and transported to the laboratory. In the laboratory, stem heights were measured, number of live scales per stem counted and the mean number of scales per meter of stem for each planting was estimated.

Using similar methods (see also Knight et al., 2020), we also estimated scale per meter of stem from the nearest natural P. australis stand (in all cases, the Delta lineage). Live stem density was measured from five randomly chosen locations within the stand (number of live stems per 0.25 m²circular quadrat). Scale incidence (presence/absence) per stem was assessed from visually inspecting 30 stems. Ten scale-infested stems were clipped at the soil surface and returned to the laboratory where we measured stem heights and live scales per stem. The mean number of scale per meter of stem was computed as the number of live scale per stem stem height in m × proportion of stems infested with scales.

Plant Area Coverage (20-22 and 32 months). By twenty months after the start of this restoration trial, many plantings had died, and others coalesced making it impossible to accurately assess survival of individual plantings. For the remainder of the survey dates, we measured area of coverage for plant species/lineages that were still present. We used the DJI Phantom 3 drone and the Drone Deploy mission planning software (Drone Deploy, San Francisco, CA, USA) to create linear transects from 15 m above ground level with 80% frontal and 75% lateral overlap to ensure optimal speed, direction, and alignment of the images. The aerial images were processed by Pix4D Cloud (Pix4D, Lausanne, Switzerland) to generate 2D orthomosiac and 3D maps. Maps were imported into ArcGIS Pro 3.0 (ESRI, 2022). An outline of the perimeter of each growing planting was traced in ArcGIS Pro, and the area computed. Total area per plant species or lineage was computed as the sum for all plantings for that species or lineage. Weather conditions prevented us from obtaining drone images from all sites in during the 20-month survey. The remaining sites were surveyed two months later. We refer to this as the 20–22-month survey.

Data Analysis

Plant Survival

Plant survival was measured at 2, 6 and 9 months. Although these are repeated measures on the same plants, and therefore non independent, we opted to treat each date as a separate analysis so that survival differences within a time period could be more easily understood. At each date, we conducted a plant-level analysis that focused on whether species or lineage type influenced each planting’s status: alive (= 1) or dead (= 0). We used a generalized linear mixed model (GLMM) with plant status (live, dead) as our response variable and species/lineage as a fixed factor. To account for nonindependence within sites and for differences among sites (e.g., water depth), site was included in the model as a random effect. A separate analysis for each survey period was conducted using R version 4.2.2 (R Core Team, 2022) and package glmer with family=binomial and a logit link function. Overdispersion was deemed not to be an issue using package blmeco and dispersion_glmer function. Differences in log-odds ratios between species/lineage pairs were evaluated with Tukey’s method and package emmeans. Goodness-of-fit was calculated using the Delta method in package MuMin which reports the marginal R² (variance explained by the model’s fixed effects) and the conditional R², (the combined variance explained by the entire model with fixed and random effects) (Nakagawa & Schielzeth, 2013). Finally, because S. alterniflora was only present at one site, it was omitted from this and all subsequent analyses.

We conducted a second set of analyses at the plot level to assess the effects of mean water depth and species/lineage on the proportion of plantings surviving at each survey period. Separate GLM analyses were conducted for each time period. It was not possible to include a lineage/species × water depth interaction in the model because mean water depth was measured at the site-level and was the same for each lineage and species. As a check for whether lineages or species were exhibiting similar relationships with water depth, we examined the slopes of the water depth-proportion surviving relationship for each lineage/species and time period. Analyses were conducted using the lm function in R. Diagnostics for the GLM were assessed using residual plots and quantile-quantile plots (for normality of residuals). Presence of outliers was assessed by examining studentized residuals.

P. australis growth, water depth and scale abundance

We conducted separate GLMMs to determine whether two measures of P. australis growth, mean stem height and stem density (numbers/m²) were related to P. australis lineage, water depth, lineage × water depth interaction and mean number of live scales per meter of stem within each planting. As in the analysis of plant survival, site was included in the model as a random effect. To normalize data distributions, scale abundances and stem densities were ln transformed. We used Akaike’s information criteria corrected for finite sample size (AICc) to select the most informative model (Burnham & Anderson, 2010). Candidate models were constructed using all possible combinations of predictor variables. There were three restrictions to the possible combinations of variables. (1) The basic mixed-effects framework was retained in all models to account for the nonindependence within sites, (2) Interaction terms could only be present in the model if their main effects were also present in the model, and (3) only two-way interactions were considered. Candidate models were ranked by AICc from lowest to highest value and AICcs with a Δ_i value (= AICc_i - AICc_min) of ≤2 were considered to have substantial support (Burnham & Anderson, 2010). AICc weights (w_i) were reported which indicate the weight of evidence (as a proportion) in favor of model i being the best model given the set of candidate models. Model selection was performed in R using package AICcmodavg. We also computed the Pearson’s product moment correlation between our two response variables using package cor in R.

We note that in the analyses above, we cannot ascribe cause and effect to the relationship between scale abundance and plant growth metrics. Because we were also interested what influences scale abundance among plants, we conducted another analysis in which scale abundance was the response variable. P. australis lineage, mean stem height, stem density, water depth and all two-way interactions were included in the full model. Because scale abundance within plantings may be related to background scale densities, we also included the mean scale abundance of the nearest neighboring P. australis stand (ln-transformed). Site was again included as a random effect and model selection was used to choose the model that best explained the data.

Plant area coverage at 20-22 and 32 months

By 20 months, area covered by each lineage or species was quite variable, with some sites having complete mortality of a lineage or species and others exhibiting strong evidence of clonal growth. For each site, we summed the areas of coverage (m²) for all surviving out-planted material for each lineage and species. Among all sites, the summed areas per lineage/species were neither normally distributed nor were the variances among lineages/species homogeneous (i.e., zeros were common, some lineages/species were absent from all site and others were limited to one or a couple of sites). Considering this issue, we simplified the analysis to reporting means and nonparametric 95% CIs for each lineage and species. Confidence intervals were computed via bootstrapping using package boot and boot.ci function with 1,000 bootstrapped estimates and basic bootstrap method. Means with non-overlapping 95% CIs were deemed significantly different.

Site conditions and fates

Sites varied considerably in water depth but not salinity over the 32 months of this study. Four of the sites (Boundary Line Ponds 1 and 2, Cheniere Crevasse and Willow Tree Site) averaged 0.51 cm deep over the course of the study, while the other four sites averaged ca. 1/3 as much water (Figure 3). Salinities registered within the margin of error of our YSI meter as pure freshwater at all survey dates.

Following the six-month survey (April, 2019), Sawdust MC South had dredge sediment deposited over the experimental plots. Also, at this time, background vegetations at Rattlesnake Bayou had grown so dense that distinguishing between the naturally occurring vegetation and our experimental plantings became unreliable. Sawdust MC South and Rattlesnake Bayou were omitted from all analyses conducted after the six-month mark. Table 2 shows the status of our plantings (fraction of sites with at least one live planting).

Plant survival at 2, 6 and 9 months

The transplantation of three lineages of P. australis and four additional wetland species was successful – among plots (the grid of plantings for a single lineage or species at a site) 80.0% ± 4.5% (mean ± se, n = 41) of the plantings survived to two months (Figure 4). At the individual plant scale, there were significant differences in the odds of being alive between species pairs but the lineage/species effect explained only 4.2% of the variance in the logistic-regression model (based on the marginal R² value; Table S3A). Differences in survival of plantings became stronger with increasing time since the initiation of the study. At six months, based on logistic regression analysis (Table S3B), the Delta and Gulf lineages and C. californicus had significantly higher survival – twice the mean survival – of the EU lineage, Z. miliacea and S. patens (Figure 4). At nine months, C. californicus had highest survival among all species/lineages (67%), 2.6 times higher than all other lineages/species combined (Figure 4, Table S3B). Within P. australis, Delta and Gulf had similar survival (35%); 1.9 times greater than for EU. By nine months, S. patens was absent from all plots. Overall, at six and nine months, the proportion of variation in the likelihood of being alive was 16% and 18%, respectively (Table S3A).

Through the first two survey periods, mean water depth (at the time of the survey) had no significant effect on the proportion of plantings that were alive; whether we focused on all lineages and plant species combined or just the P. australis lineages (Table S4). However, at the nine-month survey, there was a significant negative relationship between mean water depth and survival (Figure 5, Table S4; after accounting for the species effect, R² = 0.30 n = 28, P = 0.003). Separate regression analysis had low power (n ≤ 7) but slopes for all P. australis lineages and Z. miliacea were negative; S. californicus had a very slight positive relationship.

P. australis growth and scale abundance and at 11 months

Our P. australis growth measures, ln mean stem height and ln stem density per planting were significantly positively correlated (r = 0.41, n = 72, P < 0.001) but there was enough unexplained variance between these two metrics that we chose to treat them as separate measures of growth in our subsequent analyses. For stem density, the AICc-best model included only the intercept (Table S5). There was only one other supported model, which included only one fixed factor (scale abundance), and it had half as much support as the best model (based on AICc weights (Table S5). In comparison, the AICc-best model for P. australis mean stem height was included lineage, water depth, scale abundance and lineage × scale abundance interaction. This model had an AICc weight of 0.86 and there were no other supported models among the candidate list (Table S5). The proportion of variation in mean stem height explained by the fixed effects in the model was 0.58, whereas the random effect site explained almost no variation was explained by the random effects (Table 3). The Delta lineage averaged 188.5 ± 7.8 cm tall, 44% taller on average than the EU or Gulf lineages (Figure 6A). There was no evidence that scale abundance was negatively associated with stem height. In fact, for EU and Delta, taller stems were associated with more scales (Figure 6A). The Delta lineage exhibited no relationship between scale abundance and stem height. Finally, stems were generally taller as water depth increased but the relationship was weak (r = 0.12, n = 72, P = 0.32; after accounting for all other factors in the GLMM).

During the first summer following the initiation of the restoration trials, scale insects had infested many of our P. australis plantings. However, two of the sites, South Pass Campground and Willow Tree Site, had no scale insects. We excluded these two sites from the following analyses. The AICc-best model to explain scale abundance was one that included P. australis lineage, ln stem density and the lineage × ln stem density interaction (Table 3, Table S5). This model had an AICc weight of 0.87 and was 29 times more likely than the next best model to explain the data. Fixed effects in the model explained 22% of the variation in scale abundance. The Delta lineage had the highest mean abundance of scale insects, averaging 4.2 times higher than scale abundances on EU and Gulf (Figure 6B). The lineages also differed in their responses to stem density of the plantings (lineage × ln stem density interaction). Whereas scale abundance declined with increasing stem density for Delta and EU, it increased with density for Gulf (Figure 6B).

Plant area cover

By 20-22 months post planting, there were no surviving S. patens and S. alterniflora, and Z. miliacea was present at only one of three sites (total area: 8.3 m²; Table 2). Schoenoplectus californicus was present in four of six sites and had the highest mean area of coverage per site (110.0 m²; Figure 7). At Willow tree site, it existed in two large patches (there were two sets of plantings) with a total area of 355.7 m². Among the P. australis lineages, Delta, EU and Gulf were present in four, one and three sites, respectively (Table 2). Area of coverage for Delta averaged 48.7 m² and was 31.4 times greater than for EU and 9.7 times greater than for Gulf (Figure 7). Based on bootstrapped 95% confidence intervals, only the areas of coverage for Delta and EU were significantly different.

Between 20-22 and 32 months, the last remaining planting of EU (Chenier Pass Splay) disappeared. Also, Gulf disappeared from one site (Chenier Pass Splay) and S. californicus disappeared from two sites (Chenier Crevasse, Rattlesnake Flat). Interestingly, at all remaining sites with S. californicus, area of coverage shrunk to an average of only 7.9 m²; a 93% reduction from 20-22 months (Figure 7). Delta increased by 36% during that time and had the highest mean area of coverage at 66.3 m²; 6.1 times higher than Gulf (10.9 m²) and 8.4 times higher than S. californicus. Based on 95% CIs, Delta had significantly higher coverage than both Gulf and S. californicus (Figure 7).

Almost seven years since the first reports of widespread and pervasive dieback of Phragmites in the MRD (Knight et al., 2018; Suir et al., 2018; Cronin et al., 2020; Knight et al., 2020; Bumby & Farrer, 2022; Suir et al., 2022), there has been relatively little sign of recovery of the previously impenetrable and tall stands of P. australis (unpublished data). Ours is the first study to investigate the potential for restoration of dieback areas in the MRD. Moreover, this study is one of just a few to consider restoration trials with different lineages or genotypes of P. australis (but see e.g., Kuehl & Zemlin, 2000; Howard et al., 2008). Our main findings were that 1) P. australis outperformed the other four plant species in terms of survival and area of coverage at 32 months; 2) contrary to expectations, the EU lineage performed worst among the three P. australis lineages, barely hanging on in only one of the seven sites within which it was planted; 3) in general, for all plant species, survival during the first year following planting was greatest in the shallow-water sites; and finally 4) although scale insects quickly colonized the planted P. australis, we found no evidence in the short term (11 months) that they hindered plant establishment. Encouragingly, replanting dieback sites with Delta (the original vegetation in these sites) had the most promise and, despite legacy effects of dieback, replanting could be successful within a year after the dieback event.

Plant species/lineage performance

Spartina alterniflora inhabits saline, low-elevation sites in coastal marshes whereas S. patens is found at higher elevations and more brackish habitats (Bertness, 1991; Hester, Mendelssohn, & Mckee, 1996; 2001). In contrast, S. californicus and Z. miliacea occupy freshwater habitats and co-occur with P. australis in the area where our study was conducted (Chabreck, 1970; Howard & Rafferty, 2006; Li, Hopkinson, Schubauer-Berigan, & Pennings, 2018). At two months post planting, all species/lineages had high survivorship (mean of 80%), despite considerable variability in water depths. We attribute this to our selection process for transplantation – plantings were chosen that had a minimum height of 61 cm (Table S1) whereas water depths were, in most cases, below this level (see Figure 3). However, by six months post-planting, the two Spartina species disappeared from all sites (Table 2).

By nine months post-planting, survivorship of plantings from all species and sites was just 32%. One possible explanation for this generally low performance is legacy effects from the dieback. In studies of P. australis in Europe, eutrophic conditions, coupled with elevated sulfide concentrations and organic matter accumulation were evident following dieback events (van der Putten, 1997; Brix, 1999). Similar results were found in areas of S. alterniflora dieback in coastal Louisiana (Alber et al., 2008; Elmer, Useman, Schneider, Marra, LaMondia, Mendelssohn, Jiménez-Gasco, & Caruso, 2013; Crawford & Stone, 2015) and for the recent P. australis dieback in the MRD (Lee et al., 2023). There is also the possibility that dieback was brought about by soil-borne pathogens that may persist in the soil for some time after a dieback event (Elmer et al., 2013). Specialist soil pathogens may inhibit reestablishment of pre-dieback vegetation (e.g., P. australis in the MRD), whereas generalist pathogens may prevent successful establishment of new plant species. Interestingly, lineages of P. australis have distinctly different microbial communities (Bowen, Kearns, Byrnes, Wigginton, Allen, Greenwood, Tran, Yu, Cronin, & Meyerson, 2017; Schroeder, Halbrook, Birnbaum, Waryszak, Wilber, & Farrer, 2020) such that dieback-associated pathogens could be lineage specific. At this point, we do not know how long the legacy effects persist, although Lee et al. (2023) found that dieback soils differed markedly from healthy (non-dieback soils) 1-2 years following a dieback event and that these differences translated into reduced growth rates for all three lineages of P. australis. Although there have been other attempts to restore marsh habitats following dieback events (Ostendorp, Iseli, Krauss, Krumscheid-Plankert, Moret, Rollier, & Schanz, 1995; Kuehl & Zemlin, 2000; Ogburn & Alber, 2006; Tang, Cui, & Zhao, 2006; Bakker, Veen, Ter Heerdt, Huig, & Sarneel, 2018), research is needed to understand how soil legacies may affect restoration success and how long those legacy effects persist.

Another explanation for low plant survival was variation in water depth among restoration sites. We purposely selected sites that varied in water depth and by nine months post-planting, there was a significant negative relationship between planting survivorship and water depth (Figure 5). Thus, we found support for prediction three that deeper-water sites would have lower restoration potential. Our experimental design did not allow us to assess whether species or lineages responded differently to water depth but with the exception of S. californicus, all plant types exhibited a negative relationship with water depth. Even though our focal plants thrive in wetland environments, biomass and survival can be negatively affected by prolonged waterlogging (Mendelssohn & Mckee, 1988; McKee, Mendelssohn , & Burdick, 1989; Latham, Pearlstine, & Kitchens, 1994; Weisner, 1996; Sloey, Willis, & Hester, 2015; Snedden, Cretini, & Patton, 2015; Li et al., 2018). For example, P. australis is regarded as tolerant of high-water levels because of its ability to aerate flooded tissues by transporting oxygen through the aerenchyma, creating an extensive network of internal airspaces (Armstrong & Armstrong, 1991; Jackson & Armstrong, 1999; Eller, Sorrell, Lambertini, Whigham, Hazelton, Skalova, Brix, Cronin, Caplan, Kettenring, Meyerson, McCormick, Pysek, Mozdzer, Guo, Guo, & Bhattarai, 2017). Yet, prolonged flooding and deeper waters are known to negatively affect P. australis (Hellings & Gallagher, 1992; Cronin et al., 2020). Consequently, P. australis usually first establishes upslope and then, through rhizomatous spread, expands into deeper water (Kuehl & Zemlin, 2000; Chambers, Osgood, Bart, & Montalto, 2003).

It came as no surprise that the two Spartina species failed to persist at our experimental sites given their proclivities for more saline conditions. Although we expected them to be eventually outcompeted by freshwater species, particularly P. australis (Bertness, 1991; Silliman & Bertness, 2004; Yuan, Wang, Li, Pan, Lv, Zhao, & Gao, 2013), they disappeared well before other potential competing plant species expanded into their grid plots. Consequently, other unknown abiotic or biotic factors were involved in the disappearance of these species. Similarly, Z. miliacea was gone by 20 months from 2 of the 3 sites in which it was planted. This species has been shown to do best in shallow water (Latham et al., 1994; Li et al., 2018) which may have limited its success in these sites.

Our results supported our first prediction that P. australis would have higher survival and greater clonal growth than the other species used in this study. However, this conclusion was only true for the Delta plantings which, by the end of the study, were still present in five of seven sites. Phragmites australis, as with the other plant species used in this study, spread through the production of rhizomes. Compared to other wetland plant species, P. australis has a very large root/rhizome system that occupies a greater soil volume (Lissner & Schierup, 1997a) which has several advantages: greater ability to compete for belowground nutrients (Burdick & Konisky, 2003), the ability to penetrate into deep, permeable soil layers that contain lower salinity water (Burdick, Buchsbaum , & Holt, 2001) and resists wave action and storm surges (Coops & van der Velde, 1996; Coleman, Cassalho, Miesse, & Ferreira, 2022).

In comparison to Delta, plantings of S. californicus were present in only two of six sites and the area of coverage of this species averaged 8.4 times less than that of Delta. At first glance, our findings are at odds with the study by Howard, Rafferty, and Johnson (2020) which found similarly high survivorship of S. californicus and P. australis in a restored marsh in Louisiana. However, there are two main differences between our studies that make comparisons unwarranted. First, P. australis in the Howard et al. (2020) study was determined to be primarily of European origin but the genetic methods used (see Saltonstall, 2002) were not capable of distinguishing between Delta and EU. Also, up to 1/3 of the plantings in Howard et al. (2020) reportedly could have been the Gulf lineage. As our current study and past work on P. australis clearly demonstrate, lineages of P. australis are functionally very different (Cronin et al., 2020; Knight et al., 2020). Second, high rates of survival among species in Howard et al. (2020) could be due to the fact that fencing was erected around plantings to prevent herbivory from nutria (Myocastor coypus), an important and common non-native inhabitant of Louisiana coastal marshes (Sasser, Holm, Evers-Hebert, & Shaffer, 2018).

Surprisingly, we must reject our second prediction, that the EU lineage would outperform the other P. australis lineages in our restoration trials. In fact, EU performed worst among the three lineages; surviving to 32 months in none of seven sites in comparison to five of seven and two of seven for Delta and Gulf, respectively (Table 2). The one site in which EU persisted to 20-22 months, Cheniere Pass Splay, was one of our shallowest sites (Table 1) that receives fresh water and sediments from an adjacent channel. Regarding area of coverage, the differences among lineages were more apparent. Numerous studies have reported that the EU lineage – the cryptic and highly invasive haplotype M (Saltonstall, 2002) – exhibits higher rates of clonal growth (i.e., rhizome spread, tiller production, above and belowground biomass production), greater plasticity in response to different environmental conditions, and more efficient nutrient assimilation than the North American native lineage of P. australis (Vasquez, Glenn, Brown, Guntenspergen, & Nelson, 2005; League, Colbert, Seliskar, & Gallagher, 2006; Holdredge, Bertness, von Wettberg, & Silliman, 2010; Price, Fant, & Larkin, 2014; Bhattarai, Meyerson, Anderson, Cummings, Allen, & Cronin, 2017; Eller et al., 2017). Comparisons between EU and other P. australis lineages are less common (but see Eller, Lambertini, Nguyen, & Brix, 2014; Allen, 2016; Cronin et al., 2020; Knight et al., 2020). Most notably, we previously found that in both field surveys and common-garden experiments, the EU lineage was more resistant to scale insects (i.e., supported lower abundances) and appeared to be impervious to the dieback syndrome exhibited by co-occurring Delta (Knight et al., 2018; Cronin et al., 2020; Knight et al., 2020). One possible reason why Delta performed better than EU is that Delta is more tolerant of inundation (a water depth of 40 cm) than EU or Gulf (Cronin et al., 2020). In that study, Gulf was least tolerant of inundation and this lineage is generally found at higher elevations, usually on roadsides and river embankments on the margins of the MRD (Cronin et al., 2020). Consequently, we were surprised that it did better than EU in our restoration trials.

Relationship between scale abundance and plant performance

The roseau cane scale, N. biwakoensis, is an important pest of P. australis in its native habitat in eastern Asia (Kaneko, 2004; Brix et al., 2014; Schneider, Broadley, Andersen, Elkinton, Hwang, Liu, Noriyuki, Park, Dao, Lewis, Gould, Hoelmer, & Diaz, 2022) and has been implicated in the dieback of P. australis in the MRD (Cronin et al., 2020; Knight et al., 2020). Even at modest densities (20 adult females per stem), the scale can depress P. australis growth (Cronin et al., 2020) and therefore can impose a constraint on restoration success. The potential for herbivores to thwart restoration efforts has been reported for P. australis (Bakker et al., 2018; Temmink, van den Akker, van Leeuwen, Thöle, Olff, Reijers, Weideveld, Robroek, Lamers, & Bakker, 2022) and other wetland systems (e.g., Llewellyn & Shaffer, 1993; Qian, Chen, Zhang, Wu, Ma, Silliman, Wu, Li, & He, 2021; Wasson, Tanner, Woofolk, McCain, & Suraci, 2021).

In this study, we found that the scale insects can quickly colonize P. australis plantings in less than a year, spreading up to 40 m to our most isolated plantings. Active dispersal by this species occurs via first-instar crawlers; which probably limits its movement to a meter or less. However, passive dispersal by winds, water currents and avian vectors is possible and is probably responsible for most of the colonization of our experimental plantings. As has been reported among naturally occurring stands of P. australis in the MRD (Knight et al., 2020; Bumby & Farrer, 2022), plantings of the Delta lineage had more than four times the abundance of scale insects as the other two lineages. Interestingly, scale abundance (number per meter of stem) was related to stem density of the plantings, although the relationship differed among lineages. With passive dispersal, it would make sense that more dense patches would “capture” more colonists (Hambäck & Englund, 2005). Alternatively, colonists may simply have higher growth rates on more vigorously growing patches of plants (plant vigor hypothesis; Price, 1991). But, a positive relationship between scale abundance and stem density was only evident for Gulf. For both Delta and EU, the relationship was negative. We cannot easily explain this result, although a temporal separation between when scale colonization occurred and when we measured stem density at 11 months could have resulted in a disconnect or even reversal in the relationship between scale abundance and stem density. Unfortunately, the dispersal and colonization behavior of N. biwakoensis, as with most scale insects, is poorly understood (but see e.g., Gullan & Kosztarab, 1997; Magsig-Castillo, Morse, Walker, Bi, Rugman-Jones, & Stouthamer, 2010).

We have no evidence to suggest that N. biwakoensis negatively impacted survivorship of our plantings of P. australis. Neither stem density nor mean stem heights of the plantings exhibited a negative relationship with scale abundance. In fact, we actually found positive associations between stem height and scale abundance for the EU and Gulf lineages. Again, this finding could be due to taller plants being better traps for passive dispersers than shorter plants or that scales do better on vigorous. However, at 11 months, it may have been too early to expect to find measurable negative effects of scales on planting growth. We did not measure scale insect abundance on subsequent survey dates because of the rapidly declining survivorship of EU and Gulf. However, given that the Delta lineage has greatest susceptibility to scale attack (Cronin et al., 2020; Knight et al., 2020), and was most prevalent at the end of the study, it seems unlikely that the scales played a determining role in which lineage was best in these restoration trials. It remains to be seen if the scales eventually cause dieback of our surviving Delta plantings.

Implications for Restoration

The fate of the Lower MRD depends on the stabilization of soil elevation by emergent vegetation which is under threat by eutrophication, salinity intrusion, elevated atmospheric CO₂ concentrations, increasing intensity and frequency of hurricanes and storms and flooding events (Turner, 1990; Cahoon, White, & Lynch, 2011). The Mississippi River Delta is of particular concern given its high rates of sea level rise (8.5-9.5 mm year^-1), high land subsidence rates that can exceed 1 cm/yr^-1 and coastal wetland degradation (Shea & Karen, 1990; Day, Conner, Costanza, Kemp, & Mendelssohn 1993; González & Tornqvist, 2006). As such, it is no surprise that the ongoing P. australis dieback in the Lower MRD is attributed to complex interactions of multiple abiotic and biotic stressors, making identification of the causal agent(s) difficult (Cronin et al., 2020; Knight et al., 2020). While natural recolonization following dieback may occur, the rate of growth may be insufficient to revegetate vast regions of the MRD before soil erosion makes that impossible. Moreover, some areas of dieback, have been colonized by less desirable invasive and shallow-rooting or floating aquatic plants; e.g., water hyacinth (Pontederia crassipes (Mart.) Solms or elephant ear (Colocasia esculenta (L.) Schott) (Cronin et al., 2020).

Restoration using vegetation plantings is often designed to promote vertical accretion for the maintenance or building of marsh elevation (Hatton, DeLaune, & Patrick Jr., 1983; Bricker-Urso, Nixon, Cochran, Hirschberg, & Hunt, 1989; Nyman, DeLaune, Roberts, & Patrick, 1993; Chmura & Hung, 2004). Vertical accretion via plant growth involves mechanisms that promote the accumulation of organic matter, expansion of belowground root structures and sediment trapping via surface litter (Wolaver, Dame, Spurrier, & Miller, 1988; Craft, Seneca, & Broome, 1993; Nyman et al., 2006). Phragmites australis is adept at accreting sediments (Rooth & Stevenson, 2000; Kiviat, Meyerson, Mozdzer, Allen, Baldwin, Bhattarai, Brix, Caplan, Kettenring, Lambertini, Pysek, Weis, Whigham, & Cronin, 2019). Additionally, the typically dense clonal stands of P. australis are connected through their extensive belowground rhizome network which facilitate nutrient acquisition and can mitigate against some, but obviously not all, potential biological and physiological stress (Lissner & Schierup, 1997b).

Although much is known about the ability of the EU lineage of P. australis to invade and spread under a wide variety of marsh conditions (Chambers, Meyerson, & Saltonstall, 1999; Clevering, 1999; Saltonstall, 2002), relatively little is known about the Delta lineage whose distribution is limited to the MRD and surrounding areas (but see e.g., Hauber et al., 2011; Achenbach & Brix, 2014; Cronin et al., 2020). Our study clearly suggests that the Delta lineage would be the best choice for revegetating areas of dieback. Revegetation can begin shortly after the dieback event, despite some modest negative legacy effects mediated through changes in soil chemistry (Lee et al., 2023). Planting would need to take place in relatively shallow water, preferably under 0.5 m in depth. Revegetation of deeper-water areas will likely have to take place through growth and expansion of the planted material from shallow- to deeper-water areas (e.g., down a channel embankment).

To offset wetland loss in the MRD, The Beneficial Use of Dredged Material Program was authorized by the Water Resources Development Act (WRDA) of 2007 - Section 7006(d). Maintenance dredging of navigation channels has been used frequently over the years, with dredge material being deployed to restore degraded marshes and to create new marsh habitat (CPRA, 2017). Although dredge soils are deficient in nutrients and some metals (Lee et al., 2023), our experimental work has shown that all three lineages can grow on those soils, with only modestly lower growth rates than on undisturbed marsh soils (Lee et al., 2023). Our field observations confirm the natural colonization of those sites by P. australis. In addition, (Howard et al., 2020) demonstrated that transplanted P. australis to dredge sites could rapidly spread within two-years. In our current study, we had included one dredge site, Sawdust MC South, but the unexpected addition of new dredge sediments, caused the destruction of all of our plantings. Consequently, it remains an open question as to which P. australis lineages would do best on dredge soils. Finally, efforts to restore marsh habitat by building elevation through the deposition of dredge material and/or by revegetation of open areas cannot possibly keep up with the scale at which dieback is occurring in the MRD. However, these methods remain viable for high-risk (e.g., areas along channels that have high water flow) or high-value (e.g., near residential or commercial structures) areas.

Acknowledgments We thank Todd Baker, Vaughan McDonald and Trebor Victoriano from the Louisiana Department of Wildlife and Fisheries for technical and logistical support. Cindy Steyer from the U.S.D.A National Resources Conservation Service was also instrumental in incorporating this research into the Coastwide Vegetation Planning Project. We also thank the following people for their assistance with field work and scale counts: Leslie Aviles, Matthew Berry, Joey Breaux, Wendy Gibson, Nate Harms, Ian Knight, Herie Lee, Giovana Matos, Tshililo Mphephu.

Authors’ contributions JT Cronin, JA Nyman and R Diaz conceived of and designed this study. JT Cronin conducted the statistical analyses and H Lee and J Johnston conducted the drone-image analysis and developed the GIS for this study. JT Cronin and H Lee wrote the manuscript and all coauthors participated in the editing of the manuscript.

Funding This project was funded by the LSU AgCenter, Louisiana Department of Wildlife and Fisheries, U.S.D.A. APHIS (award number LAB94095 to R.D, J.A.N. and J.T.C.) and the U.S.D.A (LA-39 CWVP Year 7A; Coastwide Vegetative Planting Project; https://lacoast.gov/new/Projects/Info.aspx?num=LA-39).

Data availability Data and drone images are available upon reasonable request from the corresponding author.

Code availability Not applicable.

Conflicts of interest The authors declare that they have no competing interests.

Consent for publication Not applicable.

Ethical approval and consent to participate Not applicable.

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Table 1. Sites from the Lower Mississippi River Delta that were included in the restoration trials.

* Habitat type, mean salinity and water depth range were assessed at 3, 6 and 22 months.

Table 2. Survey dates, main data collected during each survey period and the fraction of sites with plantings of that lineage or species present out of the number of sites. Presence of plant species and lineages of total sites during survey dates (Fractions indicate number of site present during survey/total planted sites).

** At this date, total number of remaining sites were reduced from nine to seven due to the removal of Rattlesnake Bayou and Sawdust MC South.

Table 3. AICc-best models for response variables (A) ln Stems per m², (B) ln Mean stem height and (C) ln Scales per meter of stem.

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Phragmites australis dieback in the Mississippi River Delta, USA: Vegetation restoration

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