Strategies for Successful Mangrove Living Shoreline Stabilizations in Shallow Water Subtropical Estuaries

By combatting erosion and increasing habitat, mangrove living shorelines are an effective alternative to hard-armoring in tropical and subtropical areas. An experimental red mangrove living shoreline was deployed within Mosquito Lagoon, Florida, using a factorial design to test the impact of mangrove age, breakwater presence, and mangrove placement on mangrove survival within the first year of deployment. Mixed mangrove age treatments were included to identify if seedling (11-month-old) survival could be enhanced by the presence of transitional (23-month-old) and adult (35 to 47-month-old) mangroves. Environmental factors were monitored to detect possible causes of mangrove mortalities. Approximately half (50.6%) of mangroves died, and of those, 90.7% occurred within the annual high-water season, and 88.9% showed signs of flooding stress. Planting seedlings haphazardly among older mangroves did not attenuate enough wave energy to significantly increase seedling survival. Breakwaters alleviated stress through a reduction in water velocity and wave height, increasing the odds of survival by 197% and 437% when mangroves were planted in the landward and seaward rows, respectively. Compared to seedlings, deployment of adult mangroves increased survival odds by 1087%. Collectively, our results indicate that sites with a high-water season should utilize a breakwater structure and mangroves with a woody stem.


CHAPTER 1: INTRODUCTION
Historical efforts to stabilize shorelines have focused on the hard-armoring of extensive portions of coastlines with artificial structures such as seawalls, jetties, and breakwaters (e.g., Dugan et al., 2011;Gittman et al., 2016;Peterson et al., 2019). As a result, 14% of shorelines in the United States are hard-armored, with 64% of that total occurring in estuaries and lagoons (Gittman et al., 2016). Unfortunately, these methods have caused loss of natural habitats through direct removal of native plants, increased scouring adjacent to the structures, shading, and competition with exotic species (e.g., Bozek and Burdick, 2005;Beck and Airoldi, 2007;Bulleri and Chapman, 2010;Heerhartz et al., 2015). Overall, seawalls support 23% less biodiversity and 45% fewer organisms when compared to natural shorelines (Gittman et al., 2016). Additionally, seawalls reduce the ability of plant communities to migrate landward as sea levels increase, resulting in further habitat loss over time (Doody, 2004;Pontee, 2013;Phan et al., 2015).
"Living shoreline" is the term for shoreline stabilizations that use natural materials such as native vegetation to reduce erosion while also providing habitat (Currin, 2019). This method of restoration allows for habitat migration over time, wildlife movement between terrestrial and marine habitats, and increased wave attenuation as the vegetation grows larger (Bilkovic et al., 2016). For areas of high wave energy, structural components such as breakwaters are often placed in front of the planted vegetation to aid their survival (e.g., Bilkovic et al., 2016;Moosavi, 2017;Hardy and Wu, 2020). The presence of a breakwater has an impact on wave energy by lowering wave height and reducing incoming velocity (Losada et al., 2005;Spiering et al., 2018).
For example, an oyster shell bag breakwater was reported to decrease near-bed velocity by 62% and to reduce wave height by 42% when the water level was 5 cm above the structure (Spiering et al., 2018).
In tropical and subtropical areas, mangroves are frequently used in living shoreline stabilization efforts, are considered a foundational taxon, and are used by over 1300 animal species for shelter, foraging, and nesting (e.g., Baran and Hambrey, 1999;Ellison et al., 2005;Rusnak, 2016). Increasing areas of mangrove habitat, therefore, has the potential to increase local biodiversity and provide multiple ecosystem services, including fisheries production, carbon sequestration, and ecotourism (e.g., Carlton, 1974;Faunce and Serafy, 2006;Estrada et al., 2014;Gorman and Turra, 2016;Spalding and Parrett, 2019). Many mangrove species include complex, above-ground root structures that slow water movement, capture suspended sediments, and provide microhabitats for invertebrates and fish (Carlton, 1974;Zhang et al., 2019). Moreover, McClenachan et al. (2020) demonstrated that combining the results of multiple smallscale mangrove living shoreline projects reversed system-wide erosion patterns. In their example, 14 mangrove living shoreline deployments ranging from 104 to 327 m in length and 2-7 years in age resulted in a net shoreline gain of 347.62 m 2 yr -1 in a Florida estuary.
Rhizophora mangle (red mangrove) is frequently used in living shoreline and mangrove restoration efforts in the Southeastern United States, Caribbean, and Central America. (e.g., Teas, 1997;Winterwerp et al., 2013;Peters et al., 2015;Bayraktarov et al., 2016;Donnelly et al., 2017). Successful planting of R. mangle on Florida's east coast has traditionally had a northern limit of approximately 40 km south of Fort George Inlet, where the natural expansion of R. mangle populations have been restricted by the frequency of freeze events that drop below -4° C (Kangas and Lugo, 1990;Cavanaugh et al., 2014;Cavanaugh et al., 2019). Compared to the other mangrove species native to Florida [Avicennia germinans (black mangrove), Laguncularia racemosa (white mangrove)], R. mangle is able to settle and survive amid greater magnitudes of flooding due to larger propagules and interspecific differences in root aeration (McKee, 1993(McKee, , 1996Elster, 2000). According to data collected from Tampa Bay, FL, R. mangle occupy elevations ranging from +0.06 to +0.49 m, where the base of the mangroves are flooded on average 30% each day (Lewis, 2005). In order to imitate the observed hydrology of naturallyrecruited fringe mangrove stands, R. mangle used for living shorelines are planted in the middle to high intertidal zone (Primavera and Esteban, 2008;Samson and Rollo, 2008;Donnelly et al., 2017). Frequent inundation alleviates high pore-water salinity and increases phosphorus abundance, but extended periods of submersion can deplete a mangrove's stored oxygen, negatively impacting survival and growth through the accumulation of ethanol (Ball, 1988;Ball, 1998;Krauss et al., 2006;Lara and Cohen, 2006). Rhizophora mangle can withstand a greater range of flooded conditions as they get older due to larger stems and the growth of prop roots above the sediment surface, both of which have lenticels and aerenchyma for the intake and storage of oxygen (Tomlinson, 1986;Ball, 1988).
Wave energy contacting the portion of a mangrove submerged in water is a primary source of seedling mortality; individuals can be lost through dislodgement or failure at the stem (e.g. Balke et al., 2011;Boizard and Mitchell, 2010). Wind wave energy is produced based on wind speed and direction, bathymetry, and fetch. It can be enhanced by nearby boating activity, with resulting boat wakes contacting the shorelines (e.g. Gorman and Neilson, 1999;Bilkovic et al., 2017;Walters et al. 2021). Boizard and Mitchell (2010) found that the probability of dislodgement of seedling R. mangle by wave energy was inversely related to grain size; in their study, R. mangle anchored 3.5 times better in coral rubble than sand. Areas that are solely made up of small grains can vary in soil strength based on cohesiveness, and this impacts plant anchorage (Schutten et al., 2005). The shear strength of the sediment inhabited by mangroves spans from 2.5 to 46 kPa (Cahoon et al., 2003). Sediment accretion and erosion, which is influenced by the amount of wave energy at a site, can have an impact on young mangrove survival. Pilato (2019) showed that removal force of R. mangle seedlings increased by 0.20 N for every gram increase in root biomass. As the sediment around a mangrove erodes, the buried root biomass decreases, and less wave energy is required to displace the plant (Bywater-Reyes et al., 2015). Previous research, however, also indicates that accretion of sediment can lead to hypoxic conditions that result in mangrove mortality (Craighead and Gilbert, 1962;Terrados et al., 1997). For example, survival of 6month-old planted Avicennia marina (grey mangrove) seedlings was significantly impacted by sediment accretion once burial reached 14 cm above the original sediment level at time of planting (Kamali and Hashim, 2011). Additionally, seedling Rhizophora apiculata (tall-stilt mangrove) experienced a 3% increase in mortality rate for every cm of sediment added, and there was 0% survival for the 32 cm of additional sediment treatment at the 321-day mark (Terrados et al., 1997).
Coexisting with other vegetation has proven to have both negative (competition) and positive (facilitation) impacts on mangrove recruitment, survival, and growth (McKee et al., 1988;Farnsworth and Ellison, 1996;Donnelly and Walters, 2014;Teutli-Hernandez et al., 2019). Surrounding vegetation can influence young mangroves negatively by reducing light availability . Rhizophora mangle was once considered a shadetolerant mangrove species due to their ability to establish as propagules and grow to the seedling stage under shaded conditions (Sousa et al. 2003). However, further research revealed that R. mangle required canopy openings resulting in at least 20% light availability to proceed from a seedling to a juvenile (López-Hoffman et al., 2007). The presence of mature mangroves encourages propagule recruitment through increased surface complexity and decreased wave energy (Donnelly et al., 2017), and seedling mangroves can benefit from establishing near A. germinans and R. mangle secondary roots since they increase soil redox potential and lower sulfide concentrations (McKee et al., 1988). The presence of vegetation such as Batis maritima (saltwort) and Sarcocornia perennis (glasswort) was shown to have a positive impact on R. mangle establishment by increasing propagule retention time, reducing interstitial salinity, and increasing nutrients such as carbon, nitrogen, and phosphorus (e.g. Donnelly and Walters, 2014;Teutli-Hernandez et al., 2019). Wrack (collections of decaying organic matter) have also been observed to retain propagules of the 3 Florida mangrove species (Pinzón et al., 2003;Ruiz-Delgado et al., 2014;Breithaupt et al., 2019;Smith et al., 2020). Moreover, if wrack abundance is not great enough to smother mangroves, wrack presence can lead to increased growth due to the nutrient additions (Chapman and Roberts, 2004;Breithaupt et al., 2019).
Mangrove survival is crucial for reversing patterns of shoreline erosion and providing natural habitat (Faunce and Serafy, 2006;McClenachan et al., 2020), but many mangrove living shoreline projects have reported low levels of survival even with a breakwater present (Riley and Kent, 1999;Primavera and Esteban, 2008;Hashim et al., 2010;Tamin et al., 2011;Motamedi et al., 2014;Cuong et al., 2015;and Jayarathne et al., 2020). For example, a living shoreline in Malaysia costing $175,000 per 0.01 km 2 , that utilized a breakwater and planted 1030 A. marina and R. apiculata seedling mangroves (height: ~20 cm), reported a survivability index of 5% (Motamedi et al., 2014). A separate living shoreline in Malaysia, costing a total of $85,000, utilized a breakwater and planted A. marina saplings (~40 cm) in coir logs; the restoration had 30% survival after 8 months (Hashim et al., 2010). A review paper by Kodikara et al. (2017) revealed that out of 67 mangrove plantings in Sri Lanka, 97% of which were Rhizophora spp., 37 of the deployments resulted in 100% mortality. The reported reasons for these mortalities included drought, flooding, smothering by wrack, browsing and trampling by vertebrates, and infestation by insects and barnacles (Kodikara et al., 2017).
As demonstrated above, living shorelines can be expensive to deploy, and few studies start with pilot experiments to test different living shoreline designs at each deployment site and monitor them closely enough to identify the reason(s) for failures (Myszewski and Alber, 2016;Morris et al., 2018). In order to fill this gap and explore mangrove success when used in living shoreline stabilization in a shallow, subtropical estuary, I asked: 1) How does initial mangrove age, breakwater presence, and mangrove placement impact mangrove survival and growth? 2) Which structural characteristics of mangroves were most influential for survival? 3) What was the source of observed mangrove mortalities? 4) Is seedling survival enhanced by being planted with older mangroves? 5) How did the living shoreline impact local wrack and mangrove propagule abundance?

Study Site
Mosquito Lagoon is located on the east coast of central Florida and makes up the northernmost portion of the Indian River Lagoon (IRL) system. The IRL is classified as one of the most biodiverse estuaries in the continental United States, which supports over 4,000 species of plants and animals (Dybas, 2002). This area experiences an annual high water season each fall, and water movement is primarily wind-driven (Smith, 1987(Smith, , 1993Brockmeyer et al., 1996).
An experimental living shoreline was deployed in Mosquito Lagoon within the boundaries of Canaveral National Seashore (CANA) (Fig. 1). The 30 m of shoreline between the 2 sections was not stabilized because this stretch had no obvious erosion due to cover of mature A. germinans. The experimental living shoreline was planted along a shell-dominated shoreline once occupied by the Timucuan people (800 to 1400 CE) (National Park Service, 2020). Tribes harvested large amounts of oysters (Crassostrea virginica) and clams (Mercenaria mercenaria), discarding the empty shells in large piles (middens) along shorelines in Mosquito Lagoon (Donnelly et al., 2017). These shell middens contain culturally significant items, including broken pottery and animal bones (National Park Service, 2020). The US National Park Service is dedicated to protecting these historic sites with as little disturbance as possible, and stabilizing this area using living shoreline techniques directly supports this goal. Due to the shelly substrate, it is difficult for mangroves to naturally recruit to the shoreline; a total of 4 R. mangle developed into mature trees along the 650 m of adjacent shoreline (Donnelly et al, 2017).

Experimental Design and Restoration
To test the efficacy of different living shoreline designs, an experimental living shoreline was deployed between 14 and 21 June 2019. In total, 1,050 oyster shell bags were deployed as breakwaters and 640 R. mangle planted with the help of 51 volunteers (324 volunteer hours).
Oyster shell bags were constructed from DelStar Technologies Naltex nylon mesh filled ~3/4 full with recycled oyster shells (18.9 L) collected from restaurants and quarantined outdoors at Protection permitting requirements, with each unit no more than 6.6 m in length and a minimum of 6.6 m stretches between each unit to enable wildlife movement. Shell bags were never placed closer than 1 meter to any existing seagrass (Halodule wrightti) or other wetland plants. Each breakwater unit consisted of 2 stacked rows of 21 oyster shell bags (total shells bags/unit = 42), attached together with cable ties (304 x 7 mm).
Rhizophora mangle used for the stabilization were collected as propagules from over 100 trees within the boundaries of CANA and grown at the University of Central Florida greenhouse in Orlando. Propagules were planted in 3.7 L pots with topsoil for approximately 1 year and then transferred to 11.3 L pots with additional topsoil. These pots were kept in shallow, plastic pools filled approximately to 14 cm with freshwater.
Rhizophora mangle used in the experimental living shoreline were separated into 3 developmental stages based on known plant ages and observations of the mangrove stems at the time of deployment. Mangroves were either seedlings at 11 months-old, transitional plants at 23 months-old (hereafter referred to as "transitionals"), and adults that ranged in age from 35 to 47 months-old. These developmental stages were identified by the percentage of woody tissue on the stem. Seedlings had 0%, transitionals had between 25 and 75%, and adults had 100% woody tissue.
A factorial design was used to test all combinations of mangrove developmental stages with the presence or absence of a wave break for a total of 10 treatments along the experimental living shoreline: seedlings only, seedlings with a breakwater, transitionals only, transitionals with a breakwater, adults only, adults with a breakwater, mixture of the developmental stages, mixture of the developmental stages with a breakwater, no mangroves (control), and no mangroves (control) with a breakwater. Each treatment with a mixture of developmental stages had between 5 and 7 seedlings, a minimum of 3 adults, and a minimum of 3 transitionals; however, the exact ratio of seedlings to transitionals to adults and the placement of each developmental stage within the treatment replicate was haphazard (Table 1). This planting scheme was intended to imitate a restoration strategy that uses a haphazardly deployed mixture of developmental stages with the goal of increasing seedling survival.
Each treatment was replicated 5 times along the shoreline. The placement of each treatment along the shoreline was randomly determined prior to restoration using a random number generator (random.org) ( Table 1). For treatments with a breakwater, shell bags were placed 1 meter seaward of the planted R. mangle. Treatments with R. mangle included 16 plants in 2 staggered rows of 8 at an elevation inhabited by the closest naturally recruited adult R.
mangle. Within each treatment replicate, mangroves were centered with approximately 0.7 m distance to adjacent mangroves. One week after the deployment, all mangroves were checked to ensure the root balls and topsoil from the pots were completely buried by sediment. Three R. mangle (0.4% of total) did not meet this standard and were replaced.

Mangrove Survival
A numbered Haglöf log tag (length: 43.0 mm, width: 27.0 mm, weight: 3.9 g), which did not cause bending or damage to any of the mangroves in pilot trials, was attached to each R.
mangle with flagging tape halfway up the stem for identification. Flagging tape was not tightened and did not visibly restrict growth. All tags and tape were removed at the end of monitoring.
Survival was monitored monthly from 28 June 2019 through 28 June 2020, plus on 19 September 2019, 1 week after Hurricane Dorian (Category 2, wind speeds 56 -96 kmh -1 ), to isolate any impacts of the storm (Cappucci, 2019). The eye of the storm was approximately 160 km east of the living shoreline (Butler, 2019). Categories included: 1) "alive" if the mangrove remained in place and had foliage, 2) "standing dead" if the mangrove remained in place with no foliage, 3) "dead" if the stem was bent or partially snapped at the base to the point that the entire mangrove was lying flush on the sediment and had no foliage, and 4) "missing" if the mangrove was no longer in the planted location. A category was not created for mangroves that were bent or partially snapped to the point where the entire mangrove was lying flush on the sediment but still had foliage due to lack of occurrence. "Missing" included loss from uprooting or stem breakage. Stem breakage encompassed individuals in which the root and a small stub of the mangrove stem remained in the sediment. Mangroves that were "missing" due to stem breakage, "dead", or "standing dead" were monitored throughout the year to account for the possibility of regrowth and new leaf production (Anderson and Lee, 1995;Feller, 1995;Imbert et al., 2000;Duke, 2002).
Environmental factors that potentially varied along the shoreline at the start of the trial and could have impacted survival results included slope, distance to other established vegetation, fetch, and direction of the shoreline. The slope was calculated 1 week after the living shoreline was implemented . A level and laser were used to find the change in elevation between the shoreline 1 m seaward of the shell bags and the shoreline 3 m landward of the planted mangroves. The distance to adjacent shoreline vegetation to the left (northward) and right (southward) of the planted mangroves was determined using a transect tape, with the maximum distance being the start of another replicate, not including controls. Shoreline vegetation included naturally-recruited, mature R. mangle, A. germinans, and Conocarpus erectus (buttonwood). ArcMap 10.6 software was used to find the fetch value for each treatment replicate from the S, SW, W, and NW directions using aerial imagery from 2017. Other directions were excluded because they all had a fetch of 0. The experimental shoreline was curved in nature ( Fig. 1), so shoreline direction (in degrees) was determined for each treatment replicate by pointing a compass towards the water, perpendicular to the water line. This value accounted for possible variation in mangrove protection from wave energy that could arise from the shoreline orientation.
Factors that could vary along the shoreline throughout the monitoring period included sediment transport and shading. To measure local erosion or accretion for each treatment replicate, a polyvinyl chloride (PVC) pipe (length: 0.6 m, diameter: 12.7 mm) was placed in the center of the intertidal zone where plants were deployed or the comparable area for control treatments (Rick et al., 2006). Each piece of PVC pipe was secured until 50% of the PVC pipe was belowground and secure. The height of each PVC pipe above the sediment was measured in mm with a meter stick at the beginning of the experiment and every 3 months thereafter for 12 months. None of the planted R. mangle were placed directly beneath a canopy created by other plants at the start of the trial. To account for any change over the 12-month period, shading from all plants was recorded as a binary variable (presence or absence) for each deployed R. mangle at the end of each month. Shading was visually classified as "present" if the R. mangle had another deployed R. mangle or naturally-recruited plants growing directly above it.

Mangrove Growth
To track individual mangrove growth, initial measurements of each plant were recorded 1-week post-deployment on 28 June 2019 and every 3 months afterwards for 1 year. To ensure that mangrove growth results were not confounded by survival, only mangroves classified as "alive" at 12 months were included in the analyses of mangrove growth (N=316). Measurements included height, diameter, and number of branches, leaves, anchored prop roots, free-hanging prop roots, flowers, flower buds, and propagules. Without manipulation of the mangrove, height was recorded to the nearest cm from the base of the stem to the highest point with a meter stick.
To account for any impact that change in sediment level had on height measurements at the end of the monitoring period, the value of accretion or erosion based on the month 12 erosion stake measurement was added or subtracted from the month 12 mangrove heights. Diameter was measured with calipers to the nearest mm at the thickest portion of the stem. Branches were classified as an extension at least 2 cm long with a minimum of 1 attached leaf. After the initial 1-week measurement, leaf counts were not recorded above 100 leaves due to reduced accuracy.
Free-hanging prop roots were defined as secondary roots originating from the stem and at least 2 cm long, but not touching the sediment. Anchored prop roots originated from the stem and contacted the surface of the sediment. Prop roots were not included if they were shriveled and black in color. Prop roots were not tagged for the experiment; therefore, when analyzing change in number of prop roots from month 0 to month 12, free-hanging and anchored prop roots were combined. This method was used to account for any free-hanging prop roots that grew into anchored prop roots over the 12-month period.

Temporal Environmental Factors
Temporal factors that could collectively impact the deployed mangroves included water level, wind speed, precipitation, and minimum temperature. To determine the mean water level experienced by the plants, 5 PVC pipes (length: 0.6 m, diameter: 12.7 mm), with colored zip-ties attached 2 cm apart, were secured into the sediment at the same elevation as the planted R.
mangle. The spacing between zip-ties was re-calibrated each month. These PVC pipes were placed along the restoration site at ( Within the same 5 quadrats placed for wrack quantification, the abundance and diversity of mangrove propagules were also recorded. To further explore propagule abundance and location in the area, a transect tape was extended from 3 m seaward of the planting zone to the terrestrial ecotone within each treatment replicate. In order to capture the same location each monitoring period, the transect tape was placed to the right of the erosion stake at a 90 o angle to the water line. The bottom left corner of a 0.25 m 2 quadrat was placed at each meter of the transect tape starting at 0. In addition to propagule species and count, the percent cover of substrate (shell, sand, wrack, woody debris, vegetation) within each quadrat was calculated using the point-intercept method (Jonasson, 1983). Species of vegetation were also recorded. Notes: "Subject of Measurement" indicates whether the measurement was taken for each individual mangrove, each treatment replicate, or for the entire restoration site. *indicates the measurement was a dependent variable.

Site Characteristics
Grain size, wave height, and boating pressure at a restoration site can all impact erosion rate of a shoreline; these measurements were therefore reported to provide context for future living shoreline endeavors. To find the mean grain size on this shell midden shoreline, 100 fossil oyster or clam shells were sampled, from the intertidal portion of the shoreline at the elevation where the mangroves were deployed, using the Wolman pebble count method (Wolman, 1954).
Shell midden shorelines, such as this one, have 100% cover of recent and historic oyster and

Statistical Analyses
A logistic regression was used to determine the impact mangrove treatment, breakwater presence, and mangrove placement had on mangrove survival at 12 months (Table 3, Test # 1).
Survival indicated the mangrove was classified as "alive" at the end of the monitoring period.
Mangrove placement was a binary covariate used to separate mangroves planted in the row closest to the water (seaward row) from those planted in the row furthest from the water (landward row). A likelihood-ratio test was used to identify the overall impact of developmental stage on survival. To account for any variation along the shoreline that may influence individual mangrove survival, distance to other vegetation (left and right of treatment), fetch (S, SW, W, NW), shoreline direction, wrack thickness, wrack cover, and total erosion or accretion at 12 months were considered as possible predictors of mangrove survival and growth. Shading was removed from consideration since there were only 4 occurrences. Scatter plots of the possible influential covariates were analyzed to create multiple plausible models. Each model was tested for multicollinearity and covariates with a variable inflation factor exceeding 10 were removed (Hair et al., 1995). The best model was identified using weighted Akaike information criterion (AIC), data visualization, and covariate p-values. All statistical analyses were conducted using R, version 3.5.1 (R Core Team, 2018).
A logistic regression was used to identify the impact of temporal environmental factors on overall mangrove survival (Table 3, Test # 2). Each mangrove was marked as a "1" for the month in which they changed from "alive" to "standing dead", "dead", or "missing", and as a "0" if there was no change. This time of death marker was used as the binary response variable.
Plots were analyzed comparing mean salinity, wind speed, minimum temperature, precipitation, and water level to time of death to create a plausible model.
A Welch's t-test was utilized to detect if erosion and accretion patterns along the shoreline were being impacted by breakwater presence (

Mangrove Survival
At the end of the 12-month monitoring period, 49.4% of mangroves survived. The best model chosen by AIC, explaining mangrove survival after 12 months, included breakwater, mangrove treatment, and row with an interaction between row and mangrove treatment (Table   4). However, a significant interaction was only present between row and mangrove treatments with mixed developmental stages. Since the placement of developmental stages within mixed treatments was haphazard, by chance, a greater amount of younger mangroves were placed in the seaward row (58.5% of seedlings and transitionals). The significant interaction between row and mangrove treatment was therefore an artifact of experimental design as opposed to mangrove age. Consequently, the second model chosen by AIC was selected as the best model; this model included breakwater, mangrove treatment, and row with an interaction between breakwater and row. (Table 4). The overall influence of mangrove treatment was substantial (p<0.001), with an increase in mangrove survival from seedlings to transitionals (p<0.001) and from transitionals to adults (p<0.001) (Fig. 2, Table 5). The total survival for each developmental stage was 28.2% for seedlings, 45.5% for transitionals, and 75.6% for adults. Increased survival was also associated with presence of a breakwater (Fig. 2 presence of a breakwater, and mangroves planted in the landward row. Notes: Colon represents an interaction between two covariates.  At the end of the 12-month monitoring period, 50.6% of the deployed mangroves did not survive. Of these mangroves, 54.0% were classified as "standing dead", 42.3% "missing", and 3.7% "dead". Greater percentages of missing mangroves were associated with younger developmental stages, absence of a breakwater, and being in the seaward row for mangroves in both mixed and single developmental stage treatments (Figs. 3, 4). Compared to adults, the number of mangroves classified as missing increased by 255.6% for transitionals and 966.7% for seedlings.

Mangrove Growth
Seedling vertical growth rate (11.9 cm yr 1 ) was 22.8% greater than transitional growth rate (9.7 cm yr -1 ) and 59.2% greater than adult growth rate (7.5 cm yr -1 ). Presence of a breakwater increased growth for seedlings, adults, and mixed treatments, but decreased growth for transitionals (Fig. 5). Growth rate for adults (5.6 cm yr -1 ) increased by 67.0% and growth for seedlings (9.9 cm yr -1 ) increased by 39.9% when a breakwater was present. For the transitional developmental stage, the absence of a breakwater increased vertical growth rate by 82.4% from 6.87 cm yr -1 . Diameter growth over the 12-month period was consistent for all treatments regardless of mangrove age or breakwater presence. Decreased ranges were observed for seedling treatments without a breakwater and mixed treatments without a breakwater due to the decreased sample size (Fig. 6).

Figure 6: Diameter growth (cm) from month 0 to month 12 based on mangrove and breakwater treatment. "N" represents sample size of each treatment.
Mean change in branch count over the 12-month period was 2.90 branches (44.2%) greater for mangroves with a breakwater compared to those without a breakwater (6.58 branches). Although mean change in branch count was similar among all mangrove treatments, range increased as mangrove age increased (Fig. 7) Figure 7: Change in branch count from month 0 to month 12 based on mangrove and breakwater treatment. "N" represents sample size of each treatment.
Of the 316 mangroves that were alive at the end of the experiment, 111 (35.2%) were flowering or had buds in June 2020. None of the mangroves had flowers or buds at month 0, making all counts positive. Increased number of flowers was associated with older developmental stages and breakwater presence (Fig. 8). By month 12, 2 adult mangroves each had a single propagule hanging from their branches. The buds of these 2 propagules were first observed 9 months after the restoration (March 2020). Neither anchored nor free-hanging prop roots ever developed on the seedling mangroves.
Out of the total 253 transitional and adult mangroves that were alive at the end of the monitoring period, 134 (53.0%) produced prop roots after they were deployed. Of these mangroves, 44.8% were free-hanging and 55.2% were anchored. Increased growth was associated with older mangroves (Fig. 9). Compared to the mean prop root growth over the 12-month period for transitionals (0.77 roots), prop root production increased by 185.4% for adult mangroves (2.20 roots). There were 3 instances on 3 different mangroves where a single free-hanging prop root was shriveled and black in color. These mangroves were in the standing dead category. Differences in branch growth based on row position were consistently observed among mangrove treatments (Fig. 10). For mangroves planted in the landward row, mean branch growth over the 12-month period was 57.1% (2.87 branches) greater than those planted in the seaward row (5.02 branches yr -1 ). Consistent changes in growth based on row placement were not observed for height, diameter, prop roots, and flowers.

Structural Characteristics
Mean starting size dimensions for each developmental stage are displayed in Table 6. All mangroves had leaves at the start of the trial; none had flowers, flower buds, or propagules.
According to ANOVA tests, starting size dimensions among each developmental stage were significantly different (p < 0.05) from one another with exception of free-hanging prop roots.
The category of free-hanging prop roots was therefore removed from consideration as a main driver of increased survival with increased mangrove age. Height, diameter, and anchored prop roots had the greatest variation between mangroves that survived and those that did not at month 12. Greater starting size measurements for these categories all had a positive impact on survival (Fig. 11). Odds of survival increased by 11.0% for every mm of diameter, 3.5% for every cm of height, and 26.3% for every anchored prop root. Figure 11: Size metric distribution for mangroves that survived versus mangroves that did not survive over the 12-month monitoring period.
For mangroves without a breakwater, a linear increase in mangrove survival was observed as height and diameter increased (Fig. 12). The same pattern was observed for mangroves with a breakwater until diameter reached 2.0 cm. After this point, height had a smaller overall influence on survival (Fig. 13). The required starting height and diameter for survival was lowered whenever a breakwater was present (Fig. 12, 13).

Source of Mangrove Mortalities
Water level (p<0.001) was the best model predicting total mangrove mortality. The majority of the 324 mangrove mortalities occurred 4 months after the restoration, approximately 2 months after the onset of the annual high water season. More specifically, these mortalities occurred in October (62.0%), November (16.0%), September (8.0%), and December (4.6%) when the mean water levels (cm) (± SE) above the sediment interface of the mangroves were 20.5 ± 0.2, 14.2 ± 0.8, 20.0 ± 0.2, and 15.4 ± 0.4, respectively (Fig. 14). Total percent mortality for the remaining months ranged from 0% to 2.2% when mean water level was between 0.0 and 5.6 cm above the sediment (Fig. 14). Post Hurricane Dorian monitoring was conducted on 19 September 2019, at which time only 3 (11.5%) of the 26 mortalities from the month of September occurred. This equates to less than 1.0% of total mortalities.  Growth measurements taken on 28 September 2019 revealed that at the beginning of the high water season, seedling mangroves were a mean (±SE) of 42.7 ± 0.5 cm tall, transitionals 50.9 ± 0.6 cm, and adults 64.6 ± 0.7 cm. During September and October survival monitoring, when water level was ~20 cm, 3 seedlings were completely submerged underwater, and 41 mangroves had only the top portion of their highest leaf bundle exposed (21 seedlings, 19 transitionals, 1 adult). Of these 44 mangroves that experienced extreme submersion, 81.8% experienced mortality by October and 88.6 by the end of the 12-month monitoring period. For the 5-day period that water level was monitored each month, video footage revealed that mangroves were completely exposed (water level = 0.0 cm) for part of every month except September and October. For the remainder of months, none of the mangroves experienced complete submersion.
Temperatures were never below freezing (0° C) during the 12-month monitoring period.
The lowest temperature reached was 1.6° C for 6 hours in January; during this month, total mortality for the month was 1.5%. Although small changes in precipitation and wind speed were not correlated with overall mangrove mortality, maximum monthly precipitation (8.3 mm) and wind speed (16.3 km h -1 ) values occurred in October and September, respectively (Tab. 7).
These factors most likely contributed to the overall stress of flooding and wave energy. Mean values for water level, minimum temperature, salinity, precipitation, and wind speed are provided in Table 7.   According to historical water gage data collected from Haulover Canal by USGS, between the years of 2008 and 2020, mean water level within the tidal zone was 23.7 cm. Mean water level for each year ranged from 18.6 cm (2008) to 30.4 (2020). For the experimental monitoring period at the restoration site (July 2019 to June 2020), the mean water level at the Haulover Canal station was 31.5 cm (Fig. 17). Out of the 137 R. mangle that were missing by 1 year post-restoration, 26.3% went straight from "alive" to "missing", and 73.7% were classified as "standing dead" or "dead" before progressing to "missing." For mangroves that progressed from standing dead or dead to Although the proportion of dead mangroves remained steady over the 12-month period, this was due to the transitory state of progressing from standing dead to missing.

Seedling Survival
For treatments with and without a breakwater, the proportion of seedlings that survived did not vary between seedlings in mixed treatments and seedlings in seeding-only treatments (Fig. 20). Seedling-only treatments had 29.4% survival, and seedlings part of mixed treatments had 25.0% survival. The starting height for all seedling mangroves, regardless of treatments, ranged from 23.0 to 53.0 cm. For seedlings that survived the 12-month period, the mean starting height was 39.9 cm; for seedlings that did not survive, the mean starting height was 37.6 cm.

Wrack and Propagule Abundance
There was no significant difference in wrack thickness among mangrove treatments (p = 0.19) or between treatments with and without a breakwater (p = 0.16) (Fig. 20).
Additionally, there was no significant difference in wrack cover among mangrove treatments (p=0.79) or between treatments with and without a breakwater (p=0.41) (Fig. 21). The majority of wrack observed along the restoration site was Halodule wrightti (shoal grass) shoots, but algal and leaf debris from intertidal and terrestrial plants were also present (Brighthaupt et al. 2019).
Monitoring conducted at 3 months (5 October 2019) and 6 months (2 January 2020) after deployment of the living shoreline showed very minimal amounts of wrack cover (mean=0.0%).
Transects conducted perpendicular to each treatment replicate during the high water season placed the wrack line between 2 and 3 meters landward of where the R. mangle were deployed.
Outside of the high water season (month 9 and 12), the high water line varied between the breakwater and deployed mangroves, but wrack presence was minimal and did not to form a visible wrack line (Figs. 21, 22).  There were no naturally recruited mangrove propagules or seedlings recorded in wrack quadrats throughout the 12-month period. Sixty-two propagules were found in the transects run perpendicular to the shoreline and into the ecotone, averaging to 0.05 ± 0.02 propagules per 0.25m 2 for the year. These propagules were not marked, so it is unknown if any of these propagules were ever recounted or if they recruited to become seedlings. Only 1 R. mangle propagule was recorded within a quadrat that also had a deployed mangrove, and 1 A. germinans propagule was found caught in the mesh of a breakwater, the latter of which had started to germinate. Four mangrove propagules (2 R. mangle, 2 A. germinans) were found at the elevation of the deployed mangroves, whereas the remaining propagules were between 1 and 3 meters landward of this area.
A total of 33 naturally-recruited mangrove seedlings were recorded along the transects, placed perpendicular to each treatment replicate over the 12-month monitoring period. Seedlings were defined as recruited propagules. The observed seedlings were recently recruited, visible by their small size and the seed casings still attached to some of the A. germinans seedlings. All seedlings were found 1 to 3 meters behind the elevation at which the deployed mangroves were placed. There was only 1 quadrat that had mangrove seedlings in it consistently through months 3 to 12. This quadrat had 16 mangrove seedlings (1 R. mangle, 15 A. germinans) in month 3 and 3 mangrove seedlings (1 R. mangle, 2 A. germinans) in months 6, 9, and 12. The substrate of this quadrat included of A. germinans pneumataphores, shell, woody debris, and wrack.

A. germinans Seedlings
July 2019   Out of the 12 total boats that were captured by the wildlife cameras in the study, all had motors, but only 2 produced wakes. These 2 boats were far enough away from the shoreline that the waves originating from the boat dissipated offshore. The remaining 10 boats were either actively fishing or trolling with fishing rods visible in the boats.
Of the 300 shells sampled along the restoration site, 67.7% were clam shells and 32.3% were oyster shells. Clam shells ranged from 1.50 to 9.10 cm with a mean of 5.16 cm; oyster shells ranged from 0.70 to 6.90 cm with a mean of 3.07 cm.

CHAPTER 4: DISCUSSION
Deploying 3 to 4 year-old adult mangroves and utilizing a breakwater were important strategies for the retention and thus success of the living shoreline stabilization project in Mosquito Lagoon, a shallow-water estuary in Florida with wind-driven circulation and an annual fall high water season (Provost, 1973;Smith, 1987Smith, , 1993Brockmeyer et al., 1996). Compared to planting seedlings, survival odds increased by 186.4% when transitional mangroves were used and 1,086.9% when adult mangroves were used. For mangroves in the seaward and landward row, respectively, survival odds increased 436.8% and 197.5% when a breakwater was present.
The timing of mangrove mortalities throughout the 12-month period indicates that flooding stress was the most important factor influencing success of the living shoreline (Fig. 14). Increased survival in older plants was linked to their larger starting diameter, height, and number of anchored prop roots at the beginning of the experiment (Fig. 11), which increased access and storage of oxygen needed for cellular respiration (Chen et al., 2006). Breakwater presence did not influence sediment patterns around the base of the mangroves within 1 year of the restoration ( Fig. 19) but increased survival significantly, indicating that wave energy was a source of stress, compounding the effects of flooding (Balke et al., 2013). Since initial size metrics were crucial for withstanding flooded conditions, deployment after the commencement of high-water season, presents an effective strategy to allow for maximum growth before flooding stress. Due to higher temperatures and greater exposure to sunlight, the majority of R. mangle shoot growth occurs during the summer months and the threat of freeze events are minimal (Gill and Tomlinson, 1971). Consequently, the beginning of summer offers optimal conditions for deploying a mangrove living shoreline in the IRL and other areas with similar conditions. The cause of mangrove mortalities was reinforced by the magnitude of standing dead mangroves, which is consistent with flooding stress as opposed to forceful removal of the plant by wave energy (Bouchon et al., 1994;Hoppe-Speer et al., 2011).  (Feller, 1995;Ellison and Farnsworth, 1996;Feller and Mathis, 1997;Duke, 2002;Burrows, 2006). Evidence of herbivory on the deployed mangroves was present on 4 plants in 4 different treatment replicates; 1 mangrove had a Phocides pigmalion (mangrove skipper) and insect bite marks present, whereas 3 mangroves had just insect bite marks. Wildlife cameras that were placed along this restoration site captured 1,419 observations of mammals (56% raccoons, 37% hogs, and 4% deer) (Rifenberg et al., 2021). Of those observations, 15 captured mammals contacted the deployed mangroves, but no dislodgement or consumption occurred. Standing dead trees can also be the result of freezing temperatures (Osland et al., 2019), but this threshold was never reached over the 12-month monitoring period.
If branches with a diameter > 2.5 cm are not broken and if root rot has not occurred, standing dead R. mangle have the possibility of recovering (Snedaker, 1995;Smith et al., 2009;Kodikara et al., 2017). However, the exact length of time R. mangle can be standing dead before recovering is not known (Snedaker, 1995;Barr et al., 2012;Goldberg and Hein, 2018).
Monitoring conducted 4 months, after the destruction of Typhoon Sudal in April 2004, revealed that 64% of the R. mangle refoliated after being classified as standing dead (Kauffman and Cole, 2010 (Lewis, 2005;Primavera and Esteban, 2008). Provost (1973), however, points out the difficulty that Florida is faced with defining the high-water mark due to its seasonal variation.
Data acquired from Haulover Canal showed that over 13 years, mean high water level varied by 12 cm, which can have large impacts on young mangroves survival (Figs. 14,17). For example, mangroves in the seaward row were placed ~ 0.7 m away from the landward row over a mean slope of 0.12; therefore, the seaward row underwent greater flood and wave energy. Being planted in the seaward row decreased survival probability by 241.1% and increased the need for a breakwater structure. Seasonal changes in water level are not restricted to the coasts of Florida, but are also observed in tropical areas of Asia, Australia, West Africa, and the Americas that experience a monsoon season (Webster et al., 1998). The negative impacts of monsoonal flooding on the survival of both planted and naturally recruited mangroves have been reported in Malaysia, Sri Lanka, and South Vietnam (Hashim et al., 2010;Nguyen, 2013;Motamedi et al., 2014;Jayarathne et al., 2020).
Choosing mangroves for a shoreline stabilization project based on development of a woody stem is a simple guideline that removes the need for extensive measurements. This method was an easy way to identify age and guarantee a plant with larger dimensions. Terms such as "seedling" and "sapling" are commonly used to describe young mangroves, but the exact definition can be based on age or various size measurements depending on the author and type of research being conducted. For example, Tamai and Iampa (1988) classified a sapling as greater than 1 year-old, Ashton and Macintosh (2002) considered saplings to be greater than 1 meter tall but less than 4 cm diameter, and Farnsworth and Ellison (1996) as having 1 to 2 aerial roots and between 1 and 3 growing shoot tips. Whether a restoration manager decides to grow their own mangroves or purchase them from a nursery, tracking age for a large number of mangroves can be difficult. Furthermore, mangroves that are the same age have been observed to grow at different rates and allocate growth to different areas based on environmental factors such as shading and soil nutrients Feller et al., 2003). Classifying mangroves based on the progression from a soft, herbaceous stem to a fully woody stem easy method for choosing mangroves. Moreover, for potential living shoreline restorations that parallel the conditions of my site, choosing mangroves with a woody stem is a proven method for increasing R. mangle survival.
Parallel with previous observations of mangrove growth, younger mangroves allocated the majority of growth to increasing height, whereas older mangroves spent energy on branch, prop root, and flower production (Kathiresan and Bingham, 2001;Nagarajan et al., 2008;Primavera and Esteban, 2008). Vertical growth rate for adults and seedlings increased with breakwater presence, consistent with attenuation of wave energy (Balke et al., 2013); but vertical growth rate for transitionals decreased with breakwater presence (Fig. 5). It is possible that the critical flooding pressure needed to trigger allocation of more resources to vertical growth was different for transitionals with and without a breakwater (Ellison and Farnsworth, 1997;He et al., 2007;Balke et al., 2013;Hoppe-Spear et al., 2011). The mean starting height required for overall mangrove survival was lowered by 5.1 cm when a breakwater was present, indicating the structures were effectively reducing wave height (Achab et al., 2014;Spiering et al., 2018).
During high water season, when water level was between 12 and 34 cm around the deployed mangroves, transitional mangroves were a mean of 50.9 cm tall. Transitional mangroves without a breakwater allocated more growth to increasing height (12.5 cm yr -1 ) compared to transitionals that had a breakwater (6.9 cm yr -1 ) (Fig. 5), indicating that transitional mangroves with a breakwater may have represented the threshold where allocation to vertical growth was no longer activated by the seasonal flooding.
Of the 157 living adults after 12 months, 51.6% produced flowers or flower buds, illustrating that flooding or wave energy stress were minimized. Propagule production from a biological standpoint indicates reproductive success and flower growth has been used as an indicator of long-term mangrove restoration success (Nagarajan et al., 2008;Primavera and Esteban, 2008). At my site, mangroves that were adults at the beginning of the experiment showed propagule growth, and compared to seedlings, odds of flower and flower bud production increased by 933.1% for transitional mangroves and 6724.5% for adult mangroves. Only 5 seedlings showed signs of flowering after 12 months for a total of 16 flower buds.
Forty-four mangroves (transitionals and adults) experienced branch loss over the monitoring period but never reached the standing dead category (Fig. 7). A branch was not counted if it did not have any leaves, making branch number a good representation of leaf number as well. Rate of branch loss was highly variable (-1 to -20 branches), and small changes could be attributed to natural leaf loss (Gill and Tomlinson, 1971). Of the mangroves that lost branches, 51.6% had > 50 leaves. Further monitoring would be required to see if these trends of branch loss reverse over time or eventually lead to the standing dead state. Seedlings only had 1 or 2 branches with between 5 and 18 leaves. Therefore, all seedlings that lost branches experienced mortality by the end of the monitoring period.
Stem diameter growth was not greater for younger mangroves as would be expected (Kathiresan and Bingham, 2001) (Fig 6.). Changes in sediment level could have reduced the accuracy of the diameter measurements since they were taken at the base of the mangroves (the thickest point). As sediment accreted, the point at which diameter was taken changed to a higher, possible thinner portion of the stem. Four mangroves had a negative reading for change in diameter from month month 0 to month 12.
Planting seedlings haphazardly among older mangroves was not an effective way to increase seedling survival in Mosquito Lagoon. Some studies have found positive effects of deploying mixed age plant communities in restoration (Cody, 1993;Ashton et al., 1997;Dulohery et al., 2000;Valenzuela et al., 2016), while others have not seen these facilitative effects (De Steven, 1991;Callaway and Walker, 1997;Coomes and Allen, 2007;Gomez-Aparico, 2009). In the case of this experiment, the reason for seedling failure was most likely two-fold. During flooding pressure in the fall, seedlings at a mean height of 42.7 cm were partially submerged for approximately 2 months with maximum wave heights up to 32 cm.
When the site was visited in October for survival monitoring, 192 (87.2%) of the seedlings had their leaves exposed. Secondly, wave energy attenuation was not sufficient to significantly increase survival.
The shoreline materials used for the restoration were not adequate for trapping wrack and mangrove propagules for time periods long enough to be registered by a quarterly monitoring scheme. Seagrass wrack contains nutrients that could potentially increase the growth of living shoreline restorations (Goforth and Thomas, 1980;Breithaupt et al., 2019) Propagule abundance at the site was surprisingly low considering that the shorelines of Mosquito Lagoon are dominated by mangroves (Dybas, 2002). A similar finding had been noted previously by Donnelly et al. (2017); they found lower mangrove propagule and seedling abundance at shell middens compared to shorelines with smaller grain sizes. Donnelly et al. (2017) additionally demonstrated in greenhouse experiments that R. mangle propagules were capable of penetrating shelly substrates and surviving rooted in the sediment. Thus, they hypothesized that the low natural seedling recruitment at shell middens was due to the unstable nature of the disarticulated shell. Video footage which captured boating activity indicated that the wave energy experienced by the restoration site was mostly natural. The ~150 m channel adjacent (west) to the restoration site ( Fig. 1) is bordered on the other side by an island. The channel progressively gets deeper as it approaches the island across from the restoration site. As a result, the only boats that were observed on plane were far enough away that the waves dissipated before reaching the mangroves. The analysis was limited by the range of the wildlife cameras (60 m) and could explain the minimal number of boats captured on plane. However, the indication of the footage that few boats came to the area and that boats were planing on the other side of the channel did line up with observations made in the field while monitoring (RF, pers. obs.). Because boating activity can increase local wave energy and limit mangrove recruitment (Donnelly et al., 2017), the low boating activity at the site indicates the low natural mangrove recruitment rate was likely the result of the shelly substrate.
Natural mangrove propagule dispersal was most abundant during fall monitoring (5 October 2019), when the area experienced a high-water season. The timing of the high water season most likely limited the number of propagules being placed among the planted R. mangle and indicated the complexity of the R. mangle and breakwater was not able to trap passing propagules. The majority of propagules discovered were landward of the R. mangle ecotone where there was more surface complexity from wrack, dead wood, and upland vegetation. Lack of complexity at the elevation suitable for mangrove settlement has been previously observed to result in propagules appearing at elevations too high for survival (Smith et al., 2020). A similar process could have been responsible for the seedling abundance along the restoration site that was observed to decrease from October to July. Thus, further research is needed on living shoreline methodology that increases propagule recruitment within the first year of a shell midden restoration. Possible solutions would be to plant a wider range of plant species to cover a greater range of elevations. Plants that have already been shown to trap Florida mangrove propagules species include A. germinans, Spartina alterniflora (marshgrass), Sesuvium portulacastrum (sea purslane), Distichlis spicata (salt grass), B. maritima and S. perennis (Lewis, 2005;McKee et al., 2007;Donnelly and Walters, 2014;Millan-Aguilar et al., 2016). Although my experimental living shoreline method was not effective at trapping propagules, it is possible that propagule retention will increase over time as the planted R. mangle grow more complex and produce their own propagules.
The substrate at the restoration site was very shelly (Fig. 23), which is characteristic of shell middens (Alvarez et al., 2011). Compared to finer sediments (i.e. sand, mud), sediment transport at the experimental shoreline should be reduced and anchoring strength of the planted mangrove increased (Schutten et al., 2005;Boizard and Mitchell, 2010;Peterson et al., 2014).
The shelly sediment provided an opportunity to isolate the impact of flooding and wave energy on mangrove survival and growth, separate from the threat of removal from erosion, regardless of breakwater presence. Many other restoration sites are made up of sandy and muddy sediments, therefore the results of this experimental living shoreline may not be directly comparable.
Detached breakwaters can effectively reduce incoming velocity, lower wave height, and lead to the buildup of sediment landward of the structure. The effectiveness of a breakwater, however, depends on many factors including breakwater design, sediment supply, land use, distance of the structure from the coast, tide level, sediment type, slope, and wave energy (Akbar et al., 2017;Palinkas et al., 2017;Fitri et al., 2019;Vona et al., 2020). There is evidence that R. apiculata seedlings planted in sediment with a mean grain size of 0.016 mm could be forcibly removed by high tides even with the presence of a breakwater (Motamedi et al., 2014). On the opposite spectrum, a study by Kamali and Hashim (2011) showed that breakwater presence can lead to accretion extreme enough to smother planted A. marina seedlings. Larger mangroves have the potential to combat both erosion and accretion through larger root and shoot systems, respectively (Terrados et al., 1997;Balke et al., 2011;Tamin et al., 2011;Pilato, 2019), but further research is needed to know how the experimental design would impact R. mangle living shoreline success at sites composed of different sediment types.

CHAPTER 5: CONCLUSION
Testing different restoration strategies is an essential step for increasing future living Breakwater presence increased mangrove survival through the reduction of wave velocity and wave height. If seasonal flooding occurs, mangroves should be placed above the lowest high water level (HWL) to avoid flooding pressure and below the highest HWL to avoid smothering by wrack. Observing the elevation of nearby, naturally-recruited adult mangroves of the same species is a good first step for choosing planting elevation. To analyze how nearby elevation may have changed since the time when the mature mangroves recruited, historical aerial imagery can be evaluated using ArcMap 10.6 software as described in McClenachan et al. (2020). Planting younger mangroves haphazardly among older mangroves is not an effective method for increasing living shoreline success if the site experiences extensive seasonal flooding. Ensuring the success of planted mangroves is especially important along shell middens since natural mangrove propagule recruitment is severely limited by the shelly substrate.

CHAPTER 6: MANAGEMENT RECOMMENDATIONS
To establish that a shoreline is in need of stabilization, look for: 1) minimal natural mangrove recruitment, and 2) signs of erosion. Signs of erosion include the presence of scarps and a receding vegetation line, visualized through aerial imagery and field visits. Analyze the hydrology of the restoration site to choose the proper restoration materials and planting location.
High wave energy can be natural, driven by large fetches and high wind speeds, or it can be the byproduct of boating activity. If the site has high wave energy or a seasonal high water season, utilize a breakwater structure and select mangroves with a woody stem and anchored-prop roots.
If feasible, monitor water level of the site for a minimum of 1 year prior to restoration either directly or by accessing a nearby monitoring station that has historical data. Place the mangroves where they will be inundated ~30% of the year. To reach this optimal goal, reference nearby