Ecological Engineering 158 (2020) 106071

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Ecological Engineering

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Coupling breakwalls with oyster restoration structures enhances living T shoreline performance along energetic shorelines ⁎ I. Safaka, , P.L. Norbyb, N. Dixc, R.E. Grizzled, M. Southwelle, J.J. Veenstrae, A. Acevedoe, T. Cooper-Kolbe, L. Masseye, A. Sheremetf, C. Angelinib,f a Department of Civil Engineering, Faculty of Engineering and Natural Sciences, Istanbul Bilgi University, Eski Silahtaraga Elektrik Santrali, 34060 Eyupsultan, Istanbul, Turkey b Department of Environmental Engineering Sciences, Engineering School of Sustainable Infrastructure & Environment, University of Florida, 365 Weil Hall, P.O. Box 116590, Gainesville, FL 32611, USA c Guana Tolomato Matanzas National Estuarine Research Reserve, 505 Guana River Road, Ponte Vedra Beach, FL 32082, USA d Department of Biological Sciences, University of New Hampshire Durham, NH 03824, USA e Department of Natural Sciences, Flagler College St. Augustine, FL 32084, USA f Civil and Department, Engineering School of Sustainable Infrastructure & Environment, University of Florida, 365 Weil Hall, P.O. Box 116590, Gainesville, FL 32611, USA

ARTICLE INFO ABSTRACT

Keywords: Interest and investment in constructing living shorelines rather than harder engineering structures are on the rise Ecosystem worldwide. However, the performance of these interventions in rejuvenating coastal habitats, depositing fine Restoration sediments with elevated organic content, and reducing varies widely and is often low along energetic shorelines. In this study, we test the efficacy of a living shoreline design that couples breakwalls and oyster restoration structures, in protecting coastal estuarine ecosystems and their services along energetic shorelines. A Sediment field experiment was conducted between 2015 and 2019 along a section of the Atlantic Intracoastal Waterway in Erosion northeast Florida, which experiences commercial and recreational vessel traffic. We discovered that organic matter, silt and clay content all increased in sediments collected in the living shorelines compared to paired control treatments. In addition, oysters established and developed into robust reefs on the – wire cages filled with seasoned oyster shells - that were used to facilitate oyster recovery within this living shorelines design, although oyster growth was highest where the gabions were placed at lower intertidal elevations. Additionally, salt marsh cordgrass along shoreline margins protected by the living shoreline structures remained stable or began advancing toward the Intracoastal Waterway channel at rates of ~1 m per year, whereas cordgrass in control treatments retreated at rates approaching 2 m per year. This study provides powerful evidence that vessel wake stress is indeed driving ecosystem loss and that simple nature-based living shoreline structures designed to dissipate this energy can slow or reverse ecosystem decline. More research is needed to optimize these nature-based solutions for shoreline protection in coastal and estuarine settings, and to improve their durability.

1. Introduction eutrophication), development, and vessel traffic, as well as the artificial wave climate, and altered bathymetry and channelization for naviga­ Coastal estuarine ecosystems, including salt marshes, tional purposes associated with this traffic (e.g., Houser, 2011; Safak forests, meadows, and both coral and bivalve reefs, are some of et al., 2020). These factors can both drive direct coastal ecosystem loss the most productive and economically valuable systems on Earth via erosion and cause ecosystem degradation (Silliman et al., 2012; (Barbier et al., 2011; Costanza et al., 2014). However, the spatial extent Deegan et al., 2012; Crotty et al., 2020). Given that about 65% of and health of these systems are declining globally due to hydrodynamic coastal , 27% of coral reefs, 29% of , and 85% of stressors, such as tide- and wind-induced currents and waves, and an­ oyster reefs have been lost or degraded within the last century (Waycott thropogenic stressors including sea level rise, pollution (sucs as oiling, et al., 2009; Beck et al., 2011; Davidson, 2014), there remains a critical

⁎ Corresponding author. E-mail address: [email protected] (I. Safak). https://doi.org/10.1016/j.ecoleng.2020.106071 Received 18 April 2020; Received in revised form 6 September 2020; Accepted 25 September 2020 Available online 05 October 2020 0925-8574/ © 2020 Published by Elsevier B.V. I. Safak, et al. Ecological Engineering 158 (2020) 106071 need for interventions to slow coastal ecosystem loss and rehabilitate traffic channels. Gabions are wire cages filled with seasoned oyster degraded systems (e.g., Myers et al., 2019). Although restoration ac­ shells that are commonly used to stabilize intertidal sediments, facil­ tivities are being widely pursued to counteract coastal ecosystem de­ itate oyster recruitment and growth, and stimulate the recovery of salt gradation, their performance under high-wave energy conditions that marsh cordgrass (Walles et al., 2016). We conducted the experiment typify the many coastal environments where natural hydrodynamics along the Intracoastal Waterway (ICW) in northeast Florida. Due to have been altered via dredging, channelization or intensive vessel high levels of commercial and recreational vessel traffic, higher eleva­ traffic has been limited. tion salt marsh and lower elevation oyster reef habitats (that fringe the To date, structures, such as , , and lower margins of the salt marsh) have been degraded over recent dec­ bulkheads, have largely been the default approach for protecting ades along tens of kilometers of the ICW in the region (Grizzle et al., coastal infrastructure and sustaining coastal habitats under high levels 2002; Wall et al., 2005; Garvis et al., 2015; Silliman et al., 2019). We of wave stress. Despite their ability to withstand substantial wave designed a relatively low-cost living shoreline technique that would loading, these structures often come at a significant cost to ecological serve two purposes: i) to dissipate boat wake energy enough to facilitate functioning (e.g., Gittman et al., 2016a; Gittman and Scyphers, 2017; the accretion of finer sediments and organic material that are vital for Mayer-Pinto et al., 2018). In particular, hard engineering structures supporting benthic infaunal communities and saltmarsh establishment, often disrupt ecological connectivity by preventing organisms from and ii) to support the recovery of both the foundational bivalve, the migrating naturally across intertidal landscapes, and can further alter Eastern oyster (Crassostrea virginica) and dominant salt marsh grass, hydrodynamic and sediment transport processes, such as through the smooth cordgrass (Spartina alterniflora), hence regaining key ecological scouring of adjacent habitats that intercept refracted wave energy and structure of intertidal coastlines in the region. We chose to use break­ dislodgement of the organisms that occur within these systems (e.g., walls given their low cost, and high success at supporting sediment Seitz et al., 2006; Bishop et al., 2019; Lee et al., 2018). As a result, accretion and salt marsh grass establishment along shorelines that ex­ ecological engineering and designing living shorelines that provide perience significant fetch-generated wave energy (Hofstede, 2003). shoreline protection (strikingly better than hard engineering methods Likewise, we implemented gabions in this living shoreline experiment in some cases; e.g., Gittman et al., 2014), rejuvenate habitat-forming for two main reasons: (1) the wire cages were intended to contain the organisms, stimulate ecosystem services such as biodiversity enhance­ oyster shells for long enough to support oyster recruitment, growth, and ment, and stabilize sediment are sparking the interest of coastal man­ eventual reef formation, after which the wire itself would naturally agers and property owners around the world (e.g., Borsje et al., 2011; biodegrade (as opposed to plastic bags traditionally used in restoration Temmerman et al., 2013; Cuong et al., 2015; Davis et al., 2015; Gittman that degrade into micro-plastic debris), and (2) seasoned oyster shells et al., 2016b; Smith and Scyphers, 2019; Schotanus et al., 2020). Al­ are used as natural settlement structure for oyster spat and available though increasing in popularity, living shorelines are highly variable in through a local program focused on recycling oyster shell and returning their efficacy in reducing shoreline erosion and restoring ecosystem it to coastal waters. We hypothesized that the coupling of semi- structure and function (e.g., Morris et al., 2018). Moreover, living permeable breakwalls and gabions would be vital to jumpstarting shorelines have notoriously high rates of failure in settings where hu­ coastline recovery and that the relative performance of these structures mans have altered environmental conditions to such an extent that in improving several metrics of ecosystem structure (i.e., here defined habitat-forming organisms cannot persist, such as those where coastal as: fine sediment content, sediment organic content, oyster establish­ waters are too turbid, nutrient-polluted, or flow too fast or slow to ment, oyster reef invertebrate diversity, and salt marsh and vegetation support habitat-forming plant or bivalve growth (e.g., Gittman et al., recovery) would vary across sites depending on local conditions. 2016a; Gittman and Scyphers, 2017; Mayer-Pinto et al., 2018). Typically, the objectives of implementing living shorelines are two- 2. Methods fold: to stabilize coastal sediments and to support the persistence or recovery of habitat-forming foundation species, including salt marsh 2.1. Study sites grasses, mangrove trees, and reef-building oysters. In urbanized estu­ aries where persistent (and often escalating) vessel traffic is altering The field experiment was conducted between November 2015 and what would otherwise be low-energy conditions, achieving these ob­ June 2019 at three sites along the ICW within the Guana Tolomato jectives is especially challenging because the energy climate is no Matanzas National Estuarine Research Reserve (GTM) in northeast longer conducive to sustaining plants or reef-building species – Florida, United States (Fig. 1). The northernmost site, the middle site, organisms that can withstand some, but not excessive levels of, hy­ and the southernmost site along the Tolomato River are hereafter re­ drodynamic energy (Grizzle et al., 2002; Wall et al., 2005). In these ferred to as A, B, and C, respectively. A, B, and C were located 22.5 km, areas, additional breakwater-type interventions to dissipate wake en­ 19 km, and 17.5 km north-northeast of the St. Augustine Inlet, re­ ergy may be essential to support living shorelines success. While prior spectively. Tides are semi-diurnal and have a range between 1 m and work has investigated the use of quarried rocks and concrete oyster 1.5 m; due to small fetch distances, wind waves are negligible (Safak castles and domes in functioning as breakwaters in living shorelines et al., 2020, under review). Wave climate is controlled by boat wakes designs (e.g., Burt et al., 2011; Scyphers et al., 2015; Morris et al., from vessel traffic within the ICW. Hydrodynamic observations at this 2019), these structures can induce additional scour due to their low section of the ICW showed that waves from ~25–50 prominent vessel porosity and are often expensive to build (Bersoza Hernandez et al., wake events per day (with wave heights occasionally exceeding 0.5 m 2018). More porous bagged shell and biodegradable fiber logs have at 1 m depth) reached the shorelines of the three sites (Safak et al., been popular in low- to moderate-energy environments, but they fall 2020). Vegetation at these sites is predominantly smooth cordgrass, apart in high-energy settings and do not stay in place on steep slopes oyster reefs are dominated by the Eastern oyster, and surficial sedi­ (SCDNR, 2019). Therefore, there remains a need to explore the poten­ ments are fine sand (median grain size of ~200 μm). Along the inter­ tial efficacy of other low-cost semi-permeable structures that can dis­ tidal margins of the ICW, prior field surveys indicate that expansive sipate wave/wake energy but allow water to flow through them to re­ ‘oyster rakes’ (piles of dead shells) are common and that there are duce scour in supporting living shorelines in high-energy contexts. virtually no live reefs in the area, despite oyster reefs being extremely In this study, we explore the potential for a new hybrid living common in nearby tidal creeks and flats not exposed to persistent vessel shoreline design comprised of brush-filled breakwalls, also referred to traffic (FWC, 2019). Additionally, prior research at a site about 47 km as groins, and gabions. Brush-filled breakwalls have been widely used south of our study area indicates that cordgrass, and its underlying in northern Europe to protect high-energy coastlines from wave stress organic sediments are eroding laterally at a rate of ~1 m/yr along the (Hofstede, 2003) but never to rejuvenate salt marshes along high vessel ICW (Silliman et al., 2019). For the first 22 months of the experiment,

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Fig. 1. Locations of the field sites (marked with squares) A (orange), B (blue), and C (green) in GTMNERR along the ICW. Aerial views were obtained from the United States Geological Survey EarthExplorer database. Inset map (top left) shows the location of the GTMNERR in Florida, USA. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

we monitored ambient rates of salt marsh retreat/advance and oyster with an area of 0.9 m × 0.45 m; www.bese-products.com) were placed reef cover (which remained zero across all sites) before implementing between each section of breakwall and the vegetated zone, 3 m onshore the breakwall-gabion living shorelines in April 2017 and monitoring of the breakwalls (Fig. 2). Each gabion was composed of three sub- their performance for two years. units. Gabions and BESE were anchored into the sediment using L- shaped pieces of rebar. Immediately after installation, we monitored 2.2. Hybrid breakwall-gabion living shoreline implementation the latitude, longitude, and elevation relative to mean sea level using Trimble RTK GPS of the breakwalls and oyster restoration structures Each site included two 14-m-long segments of shoreline, separated deployed at each site (see values reported in Fig. 1). by at least 20 m. One segment had the hybrid breakwall-gabion treat­ Of note, our initial project design involved a third treatment in­ ment (hereafter treatment), and the other segment was unaltered to cluding gabions alone - a treatment that would in principle enable us to serve as a control. The semi-permeable breakwalls were built parallel to disentangle the relative importance of the breakwalls versus the ga­ the shoreline, 6 m offshore of the marsh edge, defined as the shoreward bions in dissipating wave energy and modifying environmental condi­ margin of salt marsh vegetation (Figs. 2 and 3). Three separate sections tions enough to facilitate saltmarsh recovery. However, in a six-month- of breakwall were built within each treatment segment. Each of these long pilot experiment conducted during the year preceding this focal sections were 4.3 m long and comprised of two parallel rows of seven, experiment intended to assess different gabion designs, we observed 12 cm diameter x 2.5 m long wooden pressure treated fence posts that gabions of even larger sizes and greater weight and positioned at driven into the ground to a depth of 0.75 m. Each post was placed similar elevations relative to the gabions deployed in our focal ex­ approximately 0.7 m from its neighboring posts; the two rows were periment that were not protected by a breakwall were consistently spaced 0.5 m apart. Crepe myrtle (Lagerstroemia indica) branches were uplifted and pushed back into the adjacent salt marsh by the high wave packed between the two rows of fence posts, compacted, and tied down energy in the ICW in this region. Thus, we decided to remove the ga­ using vinyl-coated wire nailed to the posts. From April 2018 onwards, bions-only treatment in our focal experimental design as we felt that breakwalls were maintained using bundles of branches bound by pallet placing large numbers of gabions at all three of our sites simply to prove strapping that were packed tightly between the two rows and tied down that the salt marsh protection was derived from the breakwalls as op­ using wire (Figs. 2 and 3). In addition to the breakwalls, two oyster posed to gabions, did not warrant the cost of materials and the potential shell gabions (0.5 m long, 0.2 m wide, 0.2 m high) and two Biode­ damage this treatment may cause to already eroding and hence de­ gradable Elements for Starting Ecosystems (BESE; interlocking sheets graded salt marshes at our field sites.

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Fig. 2. Diagram of the experimental setup showing treatment and control shoreline segments at each site. The dimensions are not to scale.

Fig. 3. Site C at 30.053o latitude north, 81.368o longitude west. Oyster gabions and BESE are seen onshore of the breakwalls. Inset (top left) shows oyster reef development on gabions at this site during the final sampling in June 2019. Accreted fine sediment can be seen behind the breakwalls and gabions.

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2.3. Experimental data collection and analysis the nearest mm the distance from the gabion surface to the outermost shell edge of the five tallest oysters on the top and sides of each sampled 2.3.1. Sediment and organic matter gabion. Three replicate gabion sub-units were measured at each site on To evaluate sediment particle size and organic matter content, three each sampling occasion. Areal percent coverage by live oysters was replicate surface sediment samples were obtained at each of three dif­ determined on three replicate gabion sub-units at each site using a ferent elevations within the at each site (Fig. 2). At the point-intercept method modified from Bergquist et al. (2006). Mea­ treatment sections of shoreline, samples from the ‘breakwall’ zone were surements were made only on the top of each gabion. A template made collected 40 cm behind the breakwalls (the 40-cm distance was ob­ from the gabion wire mesh material and containing 50 openings was served to be where the deposition was occurring), ‘gabion’ samples haphazardly placed on the surface of the gabion and sampled by sliding were collected 40 cm behind the gabions, and ‘marsh’ samples were a flag pin down from the intersection of the wire in one corner and collected at the reference poles delineating the vegetated edge of the identifying the object on the substrate as live oyster, dead oyster, cultch marsh (see Section 2.3.3). The control sections of shoreline at each site (shell), sediment, or ‘other.’ Only the live oyster cover data are pre­ were sampled in the same way, with three replicate samples collected sented herein. Percent cover was calculated using the following equa­ from elevations corresponding to the breakwall, gabion, and marsh tion: zones of the treatment shoreline. This generated 18 samples for each %cover by live oysters= (# of observations/50) x 100. (1) site (nine from the treatment section of shoreline and nine from the control section of shoreline). The top 1 cm of sediment was collected On the same gabion sub-units chosen for vertical height measure­ using one half of a Petri dish, following Southwell et al. (2010). Ad­ ments, and after vertical height and percent cover measurements had jacent control areas at each site were sampled in the same way, with been made, all material in the gabion was emptied into a 4-mm-mesh three replicate samples collected from elevations corresponding to the sieve and washed to remove sediment and other debris. All shell cultch breakwall, gabion, and marsh zones. Sediment sampling was conducted pieces were counted and the number recorded, and 30 were hapha­ in October 2018 for sites A and B, and July 2019 for site C. zardly chosen for further processing. All live oysters were counted on Samples were frozen (−20oC) at the laboratory, thawed, and their each of the 30 shells, and shell height (to nearest mm) was measured on wet weights recorded. They were then dried at 80oC for 48 h, ground the first 30 live oyster encountered in the counting process. Alllive with a mortar and pestle, and sieved to remove material coarser than organisms other than oysters (algae, invertebrates, and fish) captured in 2 mm. Particles coarser than 2 mm particles primarily consisted of the sieve after washing were removed and placed in a labeled bag and oyster shell and were not included in the organic matter content or returned to the laboratory where they were identified to lowest taxo­ particle size analysis. Sieved sediments were partitioned into two por­ nomic level practical and counted. Only data from the final sampling in tions, one for organic matter content, and one for particle size analysis. spring 2019 are presented. For organic matter analysis, we used the loss on ignition method of ANOVA was used to test effects of site, year, and site x year on Wang et al. (2011). Triplicate splits were included for 10% of samples, oyster growth metrics. Tukey's HSD was used to determine significant as well as reference sediment samples that had been previously char­ differences between treatments. Normal distribution of residuals was acterized. The standard deviation of sample splits was typically less tested using Shapiro Wilk W goodness of fit test. Live oyster density and than 0.3%; sample batches were rerun if split standard deviation or oyster vertical height were log transformed to meet homogeneity of reference samples differed by more than 1%. variance and normality of residuals assumptions. The second portion of sediment sample was used to determine particle size distribution using the modified methods from Erftemeijer 2.3.3. Salt marsh vegetation recovery and Koch (2001) and Kettler et al. (2001). To remove organic matter, Possible responses of S. alterniflorato the living shorelines treatment 15 mL of distilled water and 30 mL of 30% hydrogen peroxide were were monitored by surveying three intertidal zones at both the break­ added to 30 g sediment sub-samples in 1 L high-density polyethylene wall and control sections of shoreline (Fig. 2). The middle of these three bottles. If the sample was extremely reactive after 45 min, an additional zones was located at the initial edge of the established marsh vegetation 15 mL of 30% hydrogen peroxide was added. Samples were left to react (zone 0 m). The other two zones were located 1.5 m onshore of the under a fume hood for 72 h, then dried at 80oC for 48 h. To disperse marsh edge (zone −1.5 m) and 1.5 m offshore of the marsh edge (zone aggregates, 100 mL of 3% sodium hexametaphosphate solution was 1.5 m). The initial marsh edge was delineated by driving 15 2-m-long added and samples were shaken on an orbital shaker for two hours. Silt polyvinyl chloride (PVC) poles 1 m into the ground every 1 m along the and clay were removed by wet sieving through a 63 μm sieve. The re­ vegetated marsh edge. These semi-permanent reference poles were maining sand was dried at 80oC for 24–48 h, then mechanically shaken placed prior to the installation of the living shorelines treatments to through the following sieve sizes for 15 min: coarse sand (2 mm - 500 capture ambient variation in cordgrass retreat/advance and remained μm), medium sand (250–500 μm), fine sand (125–250 μm), very fine in place throughout the duration of the study. During each monitoring sand (63–125 μm), and silt and clay (< 63 μm). Statistical differences event, these three zones were surveyed at every third reference pole between control and treatment samples were assessed with a linear (i.e., poles 3, 6, 9, 12, 15) by centering a 1 m2 quadrat over the pole model analysis of variance (ANOVA) with sampling zone (breakwall, (zone 0 m), centering a quadrat 1.5 m offshore of the reference pole gabion, or marsh) and the interaction between treatment and site as (zone 1.5 m), and centering a quadrat 1.5 m onshore of the reference explanatory factors. Organic matter and silt and clay content were log pole (zone −1.5 m). If a quadrat appeared to contain fewer than 100 S. transformed to meet the assumption of normality. alterniflora stems, each stem was counted individually. For quadrats estimated to contain more than 100 stems, stems were counted in either 2.3.2. Oysters 1/2 or 1/10 of the quadrat and the total number calculated by extra­ Performance of oyster establishment was evaluated by measuring polation. three universal live oyster metrics (Baggett et al., 2015): vertical height Statistical differences were assessed using repeated measures above the experimental substrate surface, oyster density, and oyster ANOVA with a Poisson generalized linear model with site, intertidal areal percent coverage. Because no oysters beyond a few loose, dead zone (1.5 m, 0 m, or − 1.5 m), treatment, and date as explanatory shells were ever observed along the control shorelines and fewer than variables. 100 live oysters were observed over time on the BESE elements, data To evaluate potential shifts in the location of the marsh edge in were only collected from oysters on the gabions placed in the living response to the presence of the living shorelines structures, the distance shorelines treatments. Vertical height of live oysters projecting above from the edge of the vegetation to each of the 15 aforementioned re­ the surface of the gabions was determined by measuring with a ruler to ference poles was measured at the control and treatment sections of

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Fig. 4. Variations of percent organic matter (top panels) and percent silt and clay (bottom panels) at control and treatment shorelines from the study sites. shoreline at each site. Distances were measured to the nearest cm using Mean vertical height of live oysters extending above the surface of the a meter stick or transect tape. A value of zero was recorded if the ve­ gabions was 2 cm, 2.7 cm, and 3.4 cm at sites A, B, and C, respectively getated edge of the marsh had neither advanced nor retreated with (Fig. 5d). As measured by these two metrics (percent cover of live oy­ respect to the reference pole. Positive values were recorded if the ve­ sters and vertical height above gabion), site A showed the least reef getated edge (i.e. the location of the S. alterniflora tiller located closest development, and site C the most. to the channel) was offshore of the reference pole (i.e., advance): ne­ The experimental gabions also provided habitat for invertebrate gative values were recorded if the vegetated edge was onshore of the species other than the oyster. Density and taxonomic richness of in­ reference poles (i.e., retreat). vertebrates that colonized the experimental oyster reefs also varied Statistical differences between control and treatment shorelines among the sites, with site C being significantly and substantially greater over the course of the experiment were assessed using a repeated for both metrics than sites A and B, which did not differ (Fig. 5e and f). measures ANOVA with site and the interaction between treatment and The details of the statistical analysis of the oyster data are presented in sampling date as explanatory variables. Table S3 in the supplementary material.

3. Results 3.3. Salt marsh vegetation recovery

3.1. Sediment and organic matter Stem density (stems per m2; Fig. 6) varied significantly by site, date, and intertidal zone (1.5 m, 0 m. -1.5 m); p≪0.001 for each term (Table Sediment at the control sites was predominantly fine sand (76% on S4). Over the course of the experiment, stem density across all sites average) with silt and clay contents ranging between 2% and 11% increased at both the control and treatment shorelines (p≪0.001), but (Fig. 4). Sediments from the treatment shorelines had significantly the treatment effect of the breakwalls - gabions was not significant greater amounts of silt and clay (~3–58%) than those of control shor­ (p = 0.763). elines (p≪0.001), but the magnitude of this difference varied by site Compared to their initial positions, treatment marsh edges ad­ (p = 0.01) (Fig. 4; Table S1 in the supplementary material). The or­ vanced a mean of 0.66 m and control marsh edges retreated −0.64 m ganic matter content of sediments in control areas ranged between 1% (Fig. 7). The marsh edge of the treatment shorelines at sites A and B and 3%. Sediments from the treatment shorelines had significantly showed an average channel-ward advance of 1.00 m and 1.36 m, re­ greater amounts of organic matter content than those of control areas spectively (Fig. 7). At site B, the marsh edge at the control shoreline (p≪0.001; Table S2), with treatment sediments ranging from 10 to advanced an average of 0.14 m, whereas the marsh edge at the control 500% more organic matter content than those of control sediments. The shoreline at site A retreated an average of −0.18 m (Fig. 7). At site C, effect of the living shoreline treatment varied by site (p≪0.001; Table the marsh edge at both the treatment and control shorelines retreated, S2), with the largest effect measured at Site A, and the smallest at Site C but the retreat was far less for the treatment shoreline than for the (Fig. 4). control shoreline (−0.36 m compared to −1.87 m; Fig. 7). The effect of the breakwalls-gabions on the channelward advance of the marsh edge 3.2. Oysters was significant (p≪0.001; Table S5). During the first two monitoring events following construction of the breakwalls-gabions, the treatment No oysters were observed along the control shorelines during the did not significantly influence the position of the marsh edge (May sampling period. However, substantial oyster populations developed on 2017: p = 0.95; September 2017: p = 0.28). During the five final the living shoreline experimental substrates (gabions) at all three study monitoring events, however, the treatment effect became significant sites over the duration of the study. Live oyster density approached (January 2018 – June 2019: p≤0.005; Table S6). 4000/m2 at all sites by the final sampling period in spring 2019 (Fig. 5a). In other metrics, there were differences among the study sites, 4. Discussion suggesting differences in related environmental conditions. Mean live oyster size (shell height) was about 3 cm at sites A and B, and 2 cm at Overall, this study provides experimental evidence that breakwalls Site C (Fig. 5b) but sites A, B, and C had 15%, 24%, and 66%, respec­ coupled with oyster shell-filled gabions can facilitate the recovery of tively, cover of live oyster reef development on the gabions (Fig. 5c). coastal ecosystem structure along otherwise rapidly eroding shorelines

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Fig. 5. Metrics of (a-d) live oyster, and (e-f) invertebrate on gabions at the study sites. For live oyster metrics (a-d), sampling intervals are arranged chronologically from left to right by site: fall 2017, fall 2018, and spring 2019. Note no data for oyster vertical height (d) for fall 2017. Invertebrate metrics (e-f) correspond to only the final sampling in spring 2019. Vertical bars indicate the standard error. Lowercase letters indicate levels of significance for ANOVA followed by Tukey's HSD tests: means with same letter were not significantly (p > 0.05) different, means with different letters were significantly (p < 0.05) different. Note there was no statistical difference among site means for invertebrate taxonomic richness (f). lining high vessel traffic channels. Compared to the control shorelines particularly narrow channel location where vessel wake energy may not where sediments were sandy, no live oyster reefs established, and have dissipated enough as it moved through the breakwalls to enable marsh cordgrass failed to advance toward the channel, we observed all fine sediments to settle and accrete on the bed. of these ecological metrics improving along our treatment shorelines Importantly, our observations of the intertidal bed being eroded over this experiment. along control shorelines highlight that vessel wakes not only damage Indeed, organic matter content increased in response to the treat­ salt marsh and oyster habitats but also greatly disturb intertidal mud­ ment at all three sites, especially at sites A and B (Fig. 4), an effect flats that can be important foraging habitats for birds, rays, and other correlated with the relative impact of treatments on the silt and clay wildlife and hotspots for denitrification within estuaries (e.g., content at these sites (Fig. 4). This correlation can be attributed to the Onorevole et al., 2018). Thus, interventions such as breakwalls that fact that fine sediments have greater surface area and, therefore, greater reduce wake disturbances and promote the re-establishment of fine organic matter content (Mayer, 1994). The organic matter content may sediments with elevated organic content have the potential to support be further augmented by biodeposits from oysters living in the gabions the recovery of multiple ecosystem types and services. (Haven and Morales-Alamo, 1972). These findings reveal that this In following, we found that oysters reefs successfully developed at living shoreline design dissipated vessel wake energy enough to enable all three sites (Fig. 5). By the end of the study, percent cover of live the deposition of fine sediments at otherwise high-energy, sand-domi­ oysters on the gabions had reached 35%, averaged across the study nated sites. These variations in contents of sediment and organic matter sites. This development of oyster gabions was compared to 20 in­ can be important for modulating ecosystem responses such as oxygen dividual natural reefs nearby in the Tolomato River (Fig. 8) assessed by availability and nutrient cycling as well as mediating the distribution of GTM staff in a large baseline survey of the entire estuary (for experi­ benthic infauna and epifauna, such as polychaetes, fiddler crabs, and mental details of the survey, the reader is referred to Marcum et al., snails (e.g., Onorevole et al., 2018). We suspect that the breakwall- 2018). The 35% cover of live oysters reached by the end of our study is gabion treatment had stronger effects on surficial sediments at sites A comparable to the average live cover estimated for those natural in­ and B relative to site C where its effect was negligible, possibly due to tertidal reefs in the Tolomato River (27.1%; see yellow dots in Fig. 8 for differences in channel width and slope. Specifically, site C occurred at a locations). Development of oyster reefs (as measured by live oyster

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Fig. 6. Variation of stem counts (per m2) at the 1.5 m, 0 m, and − 1.5 m intertidal zones at the study sites. Dashed and solid curves correspond to control and treatment, respectively. Vertical bars indicate the standard error. Vertical dotted lines indicate when construction of the breakwall-gabion living shorelines was completed.

density) on the gabions varied widely by site, as has been reported in (unpublished GTM data). other areas of the southeastern US for restored/constructed intertidal Our observation that stem densities varied by site (Fig. 6) suggests reefs. In the present study, oyster densities exceeded 12,000/m2 during that plant vegetative growth strategies may have also varied in response 2017 and 2018 in some samples, then on average declined to ~4000/ to hydrodynamic stress. Specifically, we found that cordgrass advanced m2 at all three sites by spring 2019, two years after initiation of the toward the channel fairly rapidly, resulting in increased stem densities, experiment (Fig. 5). Similarly, total oyster densities on restored reefs in at sites A and B where wave/wake stress was more strongly dissipated, Louisiana studied by La Peyre et al. (2014) averaged > 1200/m2 the compared to site C where we suspect wave/wake stress acting at the first year, then decreased to ~600/m2 for years 2 and 3. In contrast, marsh edge was still quite high (see details above). Similar observations Milbrandt et al. (2015) reported mean oyster densities ranging from of other Spartina species shifting their growth strategy in response to 220/m2 after eight months to 690/m2 after three years on two restored/ decreasing hydrodynamic stress have been reported by van Hulzen constructed reefs in southwest Florida, with an overall pattern of in­ et al. (2007). They attributed this shift in growth strategy to a repri­ creasing densities over time. In their study of five constructed reefs in oritization from the benefits of drag-reduction generated by high stem South Carolina over three years, Hadley et al. (2010) also reported densities under energetic conditions to those of resource acquisition via widely different temporal patterns in oyster density with some reefs at lateral expansion in less energetic contexts (which results in lower stem maximum densities in year 1 and others in the third year. This brief densities). review indicates there may be no typical pattern of development of We also found partial evidence that salt marsh vegetation responded oyster populations on restored reefs. It seems reasonable, however, to positively to our living shoreline design over the course of just a few expect that development patterns of restored reefs would reflect the years (Fig. 7). Although the number of stems per m2 increased sub­ temporal dynamics of nearby natural reefs. Although the present study stantially at the treatment sections of shorelines, it increased slightly did not include sampling of natural reefs, so no direct comparisons are more in the control than treatment sections. Given the limited ecolo­ possible, data collected as part of the GTMNERR monitoring program gical relevance of the difference between the control and treatment indicate that oyster densities on the experimental gabions were likely shorelines (i.e., the difference between the means for each treatment greater than those on natural reefs. Based on data collected between type was less than 2 stems per m2), we conclude that the presence of the 2014 and 2016, natural reefs in GTM averaged 1621 oysters/m2 (FWC, breakwall-gabion structures does not measurably influence this plant 2019) but some samples from the Tolomato Estuary exceeded 4000/m2 performance metric. In contrast to the lack of treatment effect on stem

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Fig. 7. Changes in the onshore-offshore position of the vegetated marsh edge relative to its location at the start of the experiment at our study sites. Positive values indicate cordgrass advanced toward the channel relative to the start of the experiment while negative values indicate it retreated. Vertical bars indicate the standard error. Vertical dotted lines indicate when construction of the breakwall-gabion living shorelines was completed. count densities, there was a clear positive treatment effect on the po­ along an energetic shoreline. Over a period of four years, observations sition of the marsh edge. The strength of this effect varied by site: the on sediment grain size and organic matter content, oyster development, treatment marsh edges at sites A and B advanced toward the channel and marsh vegetation change were collected. The results showed that 1 m and 1.3 m, respectively, whereas the marsh edges at their re­ this living shoreline design increased the proportion of fine sediments spective control shorelines remained near their original position. In and organic matter content, promoted development of oyster reefs, and contrast, the edge of the treatment marsh at site C retreated, although mitigated the erosion of the shoreline. Additionally, it was found that not nearly as much as its control shoreline. This inter-site variability in waves and hydrodynamic processes are among the major controls in­ treatment response corresponds to the inter-site variability in channel fluencing the performance of living shorelines (Safak et al., 2020). This width and cross-channel slope. Site B, with the shallowest cross-channel study highlights the need for living shorelines used for coastal protec­ slope (Fig. 1), experienced the greatest amount of marsh edge expan­ tion where vessel traffic is stressing coastal ecosystems and their ser­ sion at the treatment section of shoreline. The marsh edge along the vices, and that these nature-based methods are viable in such contexts. control section of shoreline also expanded here, albeit very little. In More research is needed to optimize the design of living shorelines in contrast, site C was located at the narrowest channel section among the different settings and for using living shorelines to address various study sites and had the steepest cross-channel slope (Fig. 1). The marsh challenges, including stabilizing sediments and edge at the control section of site C retreated far farther and faster than rejuvenating bivalve reefs. Although living shorelines show great pro­ at the other two sites. Although the marsh edge did not advance at this mise for the management of coastal ecosystems, we found that main­ site, the breakwall treatment here helped reduce the severity of retreat taining living shoreline breakwalls was labor intensive as branches re­ by 80%, highlighting that this living shoreline design may dampen the quired restocking every few months. Therefore, exploring how to make rate of salt marsh degradation even under particularly energetically- these nature-based structures more durable and self-sustaining will be stressful conditions. Therefore, this inter-site variability could be at­ essential for their widespread implementation. tributed to the fact that boat wakes and their shoreline-eroding effects would be dissipated by bottom friction more at sites with milder cross- channel slopes than at sites with steep cross-channel slopes, as briefly CRediT author statement mentioned above. These results provide evidence that the effectiveness of breakwalls in stabilizing shorelines and stimulating plant growth I. Safak: Writing, Original Draft Preparation, Field Experiment, Data could also depend on the hydrodynamic context in which they are Collection, Data Analysis placed. In general, our observation of salt marsh grasses expanding P. Norby: Field Experiment, Data Collection, Data Analysis channel-ward in response to our treatments, or at least retreating less N.Dix: Supervision, Data Collection, Data Analysis quickly, indicates this intervention has the potential to help maintain R.Grizzle: Data Collection, Data Analysis the functionality and areal coverage of this highly productive coastal M.Southwell: Data Collection, Data Analysis habitat in urbanized estuaries. Overall, this living shoreline design was J.Veenstra: Data Collection, Data Analysis shown to be effective in stabilizing and rejuvenating coastal wetlands A.Acevedo: Data Collection, Data Analysis and oyster reefs threatened by high vessel traffic. T.Cooper-Kolb: Data Collection, Data Analysis L.Massey: Data Collection, Data Analysis 5. Conclusions A. Sheremet: Supervision C. Angelini: Conceptualization, Methodology, Investigation, Field The performance of a living shoreline design that combines semi- experiment, Editing permeable breakwalls with oyster restoration structures was tested

9 I. Safak, et al. Ecological Engineering 158 (2020) 106071

Appendix A. Supplementary data

Supplementary data to this article can be found online at https:// doi.org/10.1016/j.ecoleng.2020.106071.

References

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