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. Author manuscript; available in PMC: 2024 Jun 1.
Published in final edited form as: Ecol Restor. 2023 Jun;41(2-3):84–98. doi: 10.3368/er.41.2-3.84

Exploring the Use of Living Shorelines for Stabilization and Nutrient Mitigation in New England

Mary Schoell 1,*, Suzanne Ayvazian 2,^, Donald Cobb 2, David Grunden 3, Marty Chintala 2, Anna Gerber-Williams 4, Adam Pimenta 2, Charles Strobel 2, Kenneth Rocha 2
PMCID: PMC10659082  NIHMSID: NIHMS1931330  PMID: 37990651

Abstract

As salt marsh habitats face challenges due to sea level rise, storm events, and coastal development, there is an effort to use nature-based approaches such as living shorelines to enhance salt marshes and provide coastal protection. A living shoreline restoration and seasonal monitoring was conducted between July 2016 and October 2018 at an eroding salt marsh on Martha’s Vineyard, Massachusetts, Northeastern USA to assess changes in two essential ecosystem services: shoreline stabilization and nitrogen removal. Neither the living shoreline nor unaltered sites demonstrated significant sediment deposition at the marsh edge or on the marsh platform between 2017 and 2018. While we expected nitrogen removal via denitrification to improve at the living shoreline sites over time as abiotic and biotic conditions became more favorable, we found limited support for this hypothesis. We found higher rates of denitrification enzyme activity (DEA) at the living shoreline sites when compared to unaltered sites, but these rates did not increase over time. This study also provides a qualitative assessment of our living shoreline structural integrity through the years, particularly following storm events that greatly challenged our restoration efforts. We demonstrate that living shorelines fortified solely with natural materials may not be the most effective approach to maintain these ecosystem services for Northeastern USA salt marshes exposed to intense northeasterly storms. We suggest the restoration of salt marshes to improve major functions be a priority among managers and restoration practitioners. Initiatives promoting the use of nature-based restoration solution where environmental conditions permit should be encouraged.

Keywords: denitrification potential, shoreline stabilization, salt marsh, shoreline restoration

Introduction

Coastal wetlands provide unique ecological, social, and economic benefits that extend well beyond shoreline communities (Barbier et al. 2011). These dynamic, productive habitats at the land-sea interface are valuable sources of both recreation and natural resources, yet natural and anthropogenic stressors threaten their extent and stability. While wind-driven waves and storms have been shifting and eroding coastlines for centuries (Schwimmer and Pizzuto 2000), human-driven factors such as changes in land use, altered hydrology and sedimentation rates, increased boat traffic, delivery of excess nutrients to water bodies, and climate change at large have further intensified shoreline change and degradation.

Widespread salt marsh loss along North America between 2005 and 2009 (282.6 km2; Campbell et al. 2022) was greater than in any other region globally and can be attributed to many co-occurring drivers (Bromberg and Bertness 2005, Deegan et al. 2012, Watson et al. 2016). In Northeastern USA, sea-level rise rates are increasing considerably faster than the global average and coastal development has altered sediment supply and transport, causing marshes to “drown” in place as their natural sediment accretion rates cannot keep pace with these stressors (Watson et al. 2017). Marsh submergence has led to vegetative dieback throughout the marsh platform, which has resulted in ponding, subsidence, and marsh slumping at the seaward edge (Fagherazzi et al. 2013, Watson et al. 2016). At their seaward edge, marshes are inherently unstable and are retreating laterally primarily due to repetitive, wind-driven waves and boat wakes (Leonardi et al. 2016). A change in sediment supply also controls marsh lateral erosion, which is largely driven by riverine and other terrestrial inputs. Edge erosion is one of the largest threats to salt marsh extent and stability (Fagherazzi et al. 2013).

Among the many ramifications of salt marsh loss is the effect on nitrogen (N) removal. Excess nutrient loading has impacted coastal embayments as global N inputs have doubled since the preindustrial era (Vitousek et al. 1997, Bertness et al. 2002, Alldred et al. 2017) due to over-fertilization and sewage waste, leading to more reactive N and resultant harmful algal blooms in coastal waters. Salt marshes of Northeastern USA are typically N-limited, yet they play an essential role in N uptake and removal due to their sediment microbial communities’ ability to transform, fix, and denitrify nutrients. Denitrification, a process by which microbes convert nitrate (NO-3) to inert N gas (N2) permanently removes N from coastal systems, is a vital mechanism through which eutrophication can be mitigated. Salt marshes provide ideal conditions for denitrification due to their emergent vegetation, labile carbon source, and the intertidal flooding regime (Velinsky et al. 2017). Protecting, enhancing, and expanding salt marsh habitat is clearly important for mitigating nutrient loading, and living shorelines can provide a holistic solution by enhancing existing marsh functions and creating new marsh habitat (Chambers et al. 2021). Shoreline stabilization techniques have long been in practice yet continue to evolve (Charlier et al. 2005, Dugan et al. 2011). Traditional coastal protection techniques have employed strong materials such as wood, rock, or cement to armor shorelines in the form of a seawall, groin, bulkhead, revetment, or riprap. While these static, hardened structures are designed to protect infrastructure from flooding and storm surge they disconnect the marsh from the intertidal zone, causing a loss of nursery habitat for nekton, interrupting sediment transport, nutrient uptake, and causing scouring and erosion in nearby areas by deflecting waves to adjacent shoreline (Peterson and Lowe 2009, Dugan et al 2011, Gittman et al. 2015, 2016a, Dugan et al. 2018). A meta-analysis of shoreline hardening by Gittman et al. (2016a) found that seawalls (walls to the high intertidal) supported 23% less biodiversity and 45% fewer organisms than natural shorelines.

While living shoreline is a broad term for a variety of shoreline restoration approaches, the NOAA Guidance for Considering the Use of Living Shoreline (2015) describes the living shoreline approach as incorporating ‘soft’ materials such as native vegetation and / or living, natural materials (coconut coir logs), alone or with harder shoreline structures such as oyster reefs or rock sills to enhance stability. These nature-based materials are used to attenuate waves, trap sediment, and reduce erosion along the edge of the salt marsh. By stabilizing the shoreline, these nature-based solutions have been shown to promote salt marsh vegetation (Smith et al. 2018), sequester carbon (Davis et al. 2015), and enhance nursery habitat for estuarine organisms (Gittman et al. 2015, 2016a, Bilkovic et al. 2016). Suitability and material choice depends heavily on local site parameters such as topography, bathymetry, fetch, and wave direction and intensity (Miller et al. 2015) as well as permitting requirements.

Living shoreline designs are gaining in popularity as the preferred method for protecting against marsh erosion and in some states are even mandated to be installed in place of hardened structures where appropriate (e.g., the Maryland Living Shoreline Act of 2008). However, little experimental research has examined the effectiveness of living shoreline designs in the Northeastern US, where in recent decades salt marshes have been exposed to an increase in the intensity and frequency of storms (Frumhoff et al. 2007) and accelerating sea-level rise rates (Watson et al. 2016). Living shoreline projects are typically installed with shoreline protection as the main goal, and successes and failures are assessed by observing the physical integrity of the installation over time. Additionally, few studies have evaluated the role of living shorelines in nitrogen cycling, particularly how they might alter the function of salt marshes to remove nitrogen over time. Understanding the potential use of a living shoreline as a nitrogen removal tool is of great importance to restoration practitioners and coastal managers as they consider different strategies to enhance the resilience of salt marshes.

Here we report on a three-year study in which we partnered with several local stakeholder groups to install and monitor three replicate living shoreline sites and three replicate unaltered control sites along the fringing salt marsh at the Massachusetts Audubon Felix Neck Wildlife Sanctuary. We hypothesized that addition of the living shoreline using coconut coir log biodegradable materials would promote shoreline stabilization and enhance nitrogen removal at these sites compared to unaltered sites (see Figure 1 for a conceptual diagram). We assessed shoreline stabilization by monitoring differences in sediment deposition, shoreline position, elevation, and flooding frequency over time and between living shoreline vs unaltered sites. We measured sediment deposition in living shoreline sites within and outside the coir log design to understand if the coir logs would increase sediment capture through their design and compared these to rates of unaltered sites in order to examine the influence of the living shoreline structures on their sedimentary environment. Additionally, we evaluated nitrogen removal by measuring spatial and temporal differences and changes in potential denitrification rates and examined relationships between abiotic and biotic environmental parameters that could influence these rates. We expected that enhanced marsh vegetation post-living shoreline installation would positively influence potential denitrification rates and assessed this by measuring aboveground biomass as well as carbon (C) and nitrogen (N) concentrations in marsh vegetation. While not a research objective, per se, we qualitatively assessed the integrity of the coir log restoration during this project. All aspects of this work will improve our understanding of the effectiveness of living shoreline restoration techniques in a Northeastern US salt marsh exposed to severe winds and winter storms, thus expanding our view of this restoration technique as both an erosion control method and an innovative nitrogen removal tool.

Figure 1.

Figure 1.

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Conceptual diagram of hypothesized changes in marsh ecosystem services (shoreline stabilization and nitrogen removal) resulting from the installation of a living shoreline. Ecosystem services were measured via metrics listed within the shaded boxes. Italicized terms represent the specific parameters monitored during the study.

Methods

Study Site

The experiment was conducted between June 2016 and October 2018 along a continuous strip of salt marsh owned by MA Audubon’s Felix Neck Wildlife Sanctuary in Edgartown, MA (−70° 34’ 1.2”, 41° 25’ 37.9”) (Figure 2). Fringing salt marshes represent the dominant habitat type along Sengekontacket Pond, with the short form of Spartina alterniflora lining the edge of the low marsh and Spartina patens occurring in the low-high marsh. The tall form of S. alterniflora was absent due to the degraded condition of the low marsh. Sengekontacket Pond is a shallow 745-acre coastal lagoon that undergoes a semi-diurnal tidal exchange with Vineyard Sound through two inlets located at the north and south ends of the pond. The average water depth in the pond is 0.91 meters with a 0.2 m tidal range (Howes et al 2011).

Figure 2.

Figure 2.

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Map of the study site on Felix Neck Audubon Sanctuary salt marsh, including positioning of the living shoreline and unaltered sites.

Severe marsh edge erosion at the Audubon Sanctuary caused by Hurricane Sandy and the subsequent intense winter storms of 2012 had impaired marsh habitat and caused concerns over loss of habitat for birds, juvenile fish, and shellfish. Sengekontacket Pond also suffers from nutrient over-enrichment, which has largely been attributed to wastewater entering the pond from private septic systems (Howes et al. 2011). In 2009, a Total Maximum Daily Load (TMDL) was issued for the Sengekontacket Pond embayment, and excessive total nitrogen was found to be the primary pollutant of concern (Howes et al. 2011). These considerations prompted the need for stabilization of the marsh.

Living Shoreline Installation

Installation of the living shorelines began in June 2016 and extended over 424 meters (m) of available marsh edge. The selection of coir logs and other natural materials to be used in the project were predetermined based on local permitting regulations. The restoration design was chosen after on-site discussion with other practitioners and the local stakeholders, including the Audubon Sanctuary, the Shellfish Constables of Edgartown and Oak Bluffs, and the Friends of the Sengekontacket Pond. The siting of the restoration on the marsh edge was prescribed by the Audubon Sanctuary. The design consisted of two treatments with three living shoreline sites (experimental) and three unaltered marsh sites (control), which were subject to the same sampling regime. The sites were labelled by treatment type (control, C; experimental, E) and a site number (1, 2, 3) denoting their position from north (site 1) to south (site 3) along the marsh. All sites were an average length of 27 m (± 2.5 m standard deviation or SD), with 20 m (± 1.5 m SD) buffer zones separating experimental and control sites (Figure 2). Living shoreline and unaltered sites 1 and 2 have a northeastern exposure while living shoreline and unaltered site 3 are positioned along a southeastern facing shoreline. The positioning of the paired living shoreline and unaltered sites was determined randomly for each of the three sites.

Construction of the living shoreline sites involved installation of coconut fiber coir logs (0.4 m diameter x 3 m length). Coir logs were oriented to face the predominant wave direction and energy to enhance the robustness of the existing marsh and facilitate the seaward growth of new salt marsh area. The logs at each site cupped the marsh edge in a “scallop” shape. With this design, each coir log had three points of contact with another log, which allowed for a more reinforced structure where potential failures or losses could be localized. Outer logs were intended to serve as wave breaks, while logs closest to the marsh edge supported vegetation (Figure 3). Each of the three living shoreline sites was constructed with 15 to 22 coir logs, depending on the shoreline configuration. Twelve 1.2 m oak stakes were hammered into the sediment along the sides of each log and hemp rope (later nylon rope) was used to secure the log in place. The coir logs were reinforced with 900 hand sewn coconut fiber bags (0.3 m width x 0.6 m length) containing local recycled clam and oyster shell. During each sampling we carefully made note of the condition of the coir log physical structure and made decisions regarding the maintenance of the coir log structure, including replacement, in order to maintain the rigor of the experimental design.

Figure 3.

Figure 3.

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Photograph of the living shoreline restoration site (site E3) with coir logs, sandbags to the seaward edge, stakes and rope supporting the coir logs, coir fiber shell bags, sand amendments, and planted Spartina alterniflora.

Sampling Design

Site Characterization

Water temperature (oC), salinity (PSU), and dissolved oxygen (DO, mg l−1) were measured with a Hach Probe (HQD portable meter with 4-pole conductivity probe and luminescent dissolved oxygen probe) at a depth of 0.5 m at each site during each seasonal visit.

Shoreline Stabilization

Sediment deposition rates were measured using sediment traps positioned 1 m (nearshore) and 3 m offshore of each of the six sites (See Figure 4 for sampling design). Nearshore sediment traps at living shoreline sites were positioned within the coir log design (n = 2) and the offshore sediment traps (n = 2) were just outside of the coir logs. At unaltered sites there was one nearshore and one offshore sediment trap. Sediment traps were deployed for two weeks in the fall of 2016, and two weeks during the spring, summer, and fall of 2017. They were constructed of PVC slabs with triplicate cup holders attached. Each cup contained a honeycomb filter to intercept suspended sediment. Traps were secured on the sediment using a U-shaped piece of steel rod (rebar). Upon retrieval, lids were placed on the cups to prevent sediment loss during transport and the traps were returned to the laboratory. Care was taken with the deployment and retrieval of the traps to ensure accurate results. The sediment within each cup was rinsed into pre-weighed aluminum pans and placed into a drying oven at 60 °C until the weight (grams dry weight = g dw) was stable. Sediment deposition rates were calculated as the grams dry weight/days deployed.

Figure 4.

Figure 4.

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Generalized diagram of the sampling design for one living shoreline site. Sediment trap placement is 1 m and 3 m offshore as shown with gray boxes. Elevation plots (0.5 m x 0.5 m quadrats) are shown in blue, and adjacent vegetation biomass plots are shown in red. DEA core locations were collected at four random points along each transect and are shown in yellow.

Yearly elevation surveys of the marsh edge and of permanent quadrats located on the marsh platform allowed us to calculate the change in vertical (elevation) and lateral (erosional edge) position of the marsh at each site (see Figure 4 for sampling design). Elevation points were recorded every 3 m along transects that ran parallel to the vegetated edge of the marsh. and at permanently established quadrats on the marsh platform that were located 0.5 and 5 m from the marsh edge at each site (4 per site). Points were taken at the corners and center of each quadrat. On July 15, 2016, an RTK-GPS was used to establish benchmarks and perform the elevation surveying after the installation of the coir logs. No pre-installation surveying was conducted. In 2017 and 2018, all elevation surveying was conducted using a Northwest Model NTSO2B total station (angular measurement accuracy 5mm; laser plummet accuracy +/− 0.8 mm/1.5 m) (Northwest Instruments and Controls, Timnath, CO, USA).

For each site, the elevation data from 2017 and 2018 were used to create an interpolated surface. For each transect of elevation locations, additional locations were added at approximately 0.3 m intervals along the transects and an elevation value for the added location was derived from the interpolated surface. The nearest location from each 2017 location to each 2018 location was identified and the elevation value for each pair of locations was used to calculate an estimate of elevation change.

Hydrologic data were paired with elevation data from the permanent quadrats on the marsh platform to calculate the flooding regime at each site. A water level logger was deployed from mid-March through July 2016, recording water levels every 6 minutes over 257 tidal cycles. A strong correlation between our recorded data and Nantucket tide data from NOAA’s Nantucket Island station (no. 8449130) allowed us to model local high tides for any time period and evaluate and eliminate bias in estimates of flood frequency from our 4.5-month long dataset. Flooding frequency was calculated as the fraction of high tides in this dataset that were greater than or equal to the known marsh edge elevation.

Nitrogen Removal & Soil Characteristics

Soil cores for denitrification enzyme activity (DEA) were collected on the marsh platform during the spring, summer, and fall seasons at living shoreline and unaltered sites from July 2016 to October 2018 for a total of eight sampling events per site. At each site, two transects running parallel with the edge of the marsh were established at 0.5 m and 5 m from the marsh edge (Figure 4). Cores (4.7 cm diameter x 5 cm depth) were collected at four random points along each transect (n = 4 per transect; 8 per site) from the marsh platform using a slide hammer coring device with a plastic insert. Cores were capped in their plastic sleeves, kept on ice during transport, refrigerated (4°C) at the laboratory, and processed within two days of collection. Soil moisture (VWC) and conductivity (dS/m) were recorded in the field at the same location of each core using a TEROS 12 soil moisture and conductivity probe (Meter Environment, Pullman, WA, USA). Shear strength measurements (kPa) were taken at 5 cm depth at each sampling point using an AMS Field Vane Shear Tester (AMS, Inc, American Falls, ID, USA).

Cores were weighed and cut in half longitudinally. One half of each core was processed for bulk density, % C, and % N, and the other half was used for DEA analysis. Bulk density was calculated by recording the mass of each core before and after being dried at 60 °C for a minimum of 48 hours. Dried cores were then sieved through a 2 mm mesh to remove large rhizomes and rocks, ground to a fine powder, and used for elemental analysis of % C and % N on a ThermoFinnigan Flash EA 112 (Thermo Fisher Scientific, Waltham, MA, USA) (Katz et al. 2013). Peach leaves were used as the reference standard. Bulk density values were corrected for gravel (anything larger than 2 mm), and calculated as:

Bulkdensity(g/cm3)=Dryweightofsediment(g)/Volumeofsediment(cm3)

All sample preparation, processing, and analysis was conducted at the US EPA laboratory in Narragansett, RI.

The DEA assay is a laboratory-based procedure that measures the potential denitrifying enzyme activity in a soil sample under anaerobic environmental conditions with unlimited C and nitrate (NO3-). DEA assays were performed on soil slurries using the acetylene (C2H2) inhibition technique described by Smith & Tiedje (1979) and Groffman et al. (2006). After rock and rhizome removal and homogenization, 6.5 g subsamples were weighed into 60 ml serum bottles. Each sample received 12 ml of nutrient-rich, filtered seawater solution containing nitrate (7 mmol N l−1), glucose (17 mmol C l−1), phosphate (6 mmol P l−1), and chloramphenicol (0.125 g l−1); an antibiotic that prevents the production of new enzymes. Bottles were capped with rubber septa, then evacuated and flushed with nitrogen gas (N2) three times to create anaerobic conditions. Five ml of C2H2 was injected into each bottle at the start of the timed assay to inhibit the reduction of nitrous oxide (N2O) to N2. Bottles were then promptly placed on an orbital shaker table at 125 rpm. At 30, 60, and 90 minutes, headspace samples (5 ml) were extracted from each bottle and injected into 12 ml evacuated glass exetainers (Labco, UK). Ten ml of N2 was also added to each exetainer. The lost headspace in each sample bottle was replaced by 4 ml of N2 and 1 ml of C2H2. Concentrations of N2O in each exetainer were analyzed via gas chromatography with an electron capture detector (ECD) (GC, Shimadzu GC2014, Shimadzu Scientific Instruments, Columbia, MD, USA) at Yale University (New Haven, CT, USA).

Vegetation Characteristics

To assess the condition of the marsh vegetation and examine differences between the living shoreline versus unaltered sites, we used 0.25 × 0.25 m quadrats to harvest aboveground vegetation at each sampling location in July 2017 and 2018. The vegetation quadrats were placed directly adjacent to each permanent elevation quadrat on the marsh platform (4 per site) (Figure 4). Vegetation within the quadrat was clipped at soil level, placed into labelled plastic bags, kept on ice, and returned to the laboratory. Harvested vegetation was washed to remove salt and sediment and sorted by species. Vegetation was dried at 60 °C for at least 24 hours and dry weight recorded for each quadrat. Subsamples from vegetation collected in summer 2017 were ground in a Wiley mill (Thomas Scientific, Logan Township, NJ USA) using a size 40 screen and weighed for elemental analysis. Total C and N was determined by dry combustion at 950 °C on a Thermo Finnegan FLASH EA 1112 Series CHN analyzer (Katz et al. 2013). Peach leaves were used as the reference standard (National Institute of Standards and Technology).

Statistical Analysis

Data analyses were conducted using R version 0.98 (2012). We tested the effects of treatment (experimental or control) and position (distance from the marsh edge) on potential denitrification rates and all other environmental parameters (soil characteristics, vegetation characteristics, elevation characteristics) using 2-way ANOVA. Statistical significance was accepted at p = 0.05. Tukey post-hoc tests were used to examine the living shoreline application (treatment) effects, distance effects, and cross-factor differences. Yearly differences for each metric were assessed using a one-way ANOVA. A linear mixed-effects model was used specifically to test for significant differences and interactions in denitrification potential rates with treatment, position, and time since the installation of the living shorelines. Through this model, we were focused on understanding 1) whether there was a change in denitrification potential rates over time; 2) if change over time was greater for areas with living shorelines (experimental) than unaltered (control) treatment; and 3) if the position on the marsh had a significant interaction with time. This meant testing the interaction of treatment and time, as well as the 3-way interaction among treatment, time, and position. Site was included as a random effect to account for variability among our sites, whereas treatment (living shoreline or unaltered shoreline), time since installation (a continuous explanatory variable), and position on the marsh platform (0.5 m vs 5 m) were included as fixed effects. By including a random effect, we were able to test the average differences across our treatment groups over time, rather than comparing means like we would in a 3-way ANOVA. Pearson correlation analyses were used to examine relationships between DEA rates and environmental parameters. Prior to analysis, all variables were tested for normality and homogeneity of variance. Shoreline position and elevation change analyses were performed using ArcGIS version 10.6.1. The sediment deposition rate for each of the six sites was calculated by averaging across duplicate sediment traps and across four seasons (fall 2016 and spring, summer, and fall of 2017), and the difference was calculated by subtracting the deposition rate at a living shoreline site from its associated unaltered site.

Results

Living Shoreline Integrity

Over the course of our study, we observed weather-driven damage and degradation to the coir logs in our living shoreline designs. This necessitated decisions about rebuilding many of these areas to maintain the integrity of our study design. Tropical Storm Hermine in early September 2016 caused significant damage to many coir logs, in which coir fiber was lost from the log’s mesh casing and washed up on to the marsh surface. After this storm, the team replaced 7 total coir logs in the living shoreline sites. In the spring of 2017, extensive repairs were required after severe winter ice scouring and 27 coir logs were replaced. Adjustments were also made to the materials we used during this round of repairs as a response to the need for much greater durability. We purchased different coir logs with a tighter mesh weave casing and replaced the hemp rope used to secure the logs to the wooden stakes with thin nylon rope (the only non-natural material used). During the summer of 2017, 350 burlap bags were filled with sand and placed on the seaward side of the coir logs as a secondary buffer to wave action. 19 m3 of sand was added to the inner scalloped coir log areas and 1000 Spartina alterniflora plugs were planted into the sand amendment. In late February and early March 2018, three successive and massive storm fronts battered the coastline. During the spring and summer of 2018, we observed coir material debris on the top of the marsh which we removed. Some logs required repairs at sites 1 and 2, and other logs were buried partially with sand. The sand and plant amendments from summer 2017 were still intact. Our visit to the marsh in the spring 2019 revealed the degradation of some logs at site 1. Coir logs remained at site 2 and were buried partially, and the logs at site 3 remained in place with minimal damage. Given an estimated life span of five years, degradation of the coir logs from aging was expected; however, the severe weather events hastened the process.

Site characterization

Water temperature values measured seasonally at the six sites from 2016–2018, ranged from 13.0 ºC in the spring, 24.7 ºC in the summer, and 10.27 ºC in the fall. Salinity values were consistent ranging from 29.5–30.4 ppt. Dissolved oxygen ranged between 10.1–10.5 μg l-1.

Shoreline Stabilization

Our hypothesis of significantly greater sediment deposition at the living shoreline sites was only partially supported. The average sedimentation rate (± SE) across the four seasons from the nearshore living shoreline sediment traps (within coir logs) was 0.171 g dw day−1 (± 0.027) compared to the nearshore unaltered site sediment traps 0.262 g dw day−1 (± 0.035). At offshore (3 m) areas, the average sedimentation rate (± SE) across the four seasons at living shoreline sites was greater (0.330 ± 0.056 g dw day−1) compared to unaltered site (0.210 ± 0.030 g dw day−1) (Figure 5). The highest sedimentation rate observed during our study was at living shoreline site 3 (E3) in the fall of 2017. No strong patterns in sedimentation rates over time emerged, particularly at living shoreline sites, which were very variable between seasons and between individual sediment traps (as seen in panels A and B in Figure 5).

Figure 5.

Figure 5.

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Average deposition rates (g dw day−1) from sediment traps within living shoreline (E) (n = 2) and unaltered (C) (n = 1) sites positioned 1 m (nearshore) and approximately 3 m (offshore) from the shoreline for four sampling events during the study. The points for each site and sampling have been offset so the individual standard error bars can be easily visualized on panels A and B. At living shoreline sites, the nearshore sediment traps were positioned within the coir log design. A. Living shoreline sites at 1 m offshore (within coir logs); B. Living shoreline sites at 3 m offshore (just outside of coir logs); C. Unaltered sites at 1m offshore; D. Unaltered sites at 3 m offshore.

Detailed elevation surveys along the marsh edge at each site between 2017 and 2018 revealed that all sites, except E3 and C3, eroded laterally over the year as noted by negative average values (Figure 6). There were no patterns in vertical elevation loss or gain by treatment. Sites E2, E3, and C3 had an overall loss in elevation, while sites E1, C1, and C2, had minor gains in elevation. Flooding frequency was calculated as a function of elevation. Overall, flooding frequency ranged from 10 % to 90 % and was higher in unaltered areas (F(1,32) = 4.172, p = 0.049) and at the marsh edge (F(1,32) = 5.430, p = 0.026) and did not vary among years.

Figure 6.

Figure 6.

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Boxplots depicting the median, interquartile range, and outliers (defined as being either 1.5 x IQR (IQR = interquartile range) above the 75th percentile, or 1.5 x IQR below the 25th percentile) of A. shoreline position and B. elevation change data at living shoreline (E) and unaltered (C) sites from 2017 to 2018, where positive values indicate a gain in either shoreline position or elevation.

Nitrogen Removal & Soil Characteristics

A 2-way ANOVA indicated significantly higher denitrification enzyme activity (DEA) potential rates at the living shoreline sites than the unaltered sites (F(1,88) = 30.725, p < 0.001). There was a significant interaction between treatment and distance from the marsh edge (F(1,88) = 5.852, p = 0.017), where DEA rates (± SE) at 5 m landward of the marsh edge at living shoreline sites (0.630 ± 0.078 g N m−2 d−1) were significantly greater than rates at both 0.5 m (0.368 ± 0.048 g N m−2 d−1) and 5 m (0.176 ± 0.038 g N m−2 d−1) at unaltered sites. DEA rates summarized by treatment, distance from the marsh edge, and season/year revealed that the highest rates at living shoreline sites were at 5 m from the marsh edge during fall 2016 (1.27 ± 0.236 g N m−2 d−1). At unaltered sites, the highest rates were at 0.5 m from the marsh edge during spring 2018 (0.593 ± 0.037 g N m−2 d−1) (Figure 7). Overall, there was a trend of DEA increasing linearly with days since living shoreline installations (F(1,80) = 1.066, p = 0.078; Table 1) albeit non-significant. However, the interaction between treatment and time was not significantly correlated with DEA rates, meaning change in DEA rates over time was not different for living shoreline and unaltered treatments.

Figure 7.

Figure 7.

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Average DEA rates (± SE) by treatment, distance from the marsh edge, season, and year. Average rates from control areas are shown in the left panel and rates from living shoreline (experimental) areas are shown in the right panel. Black bars represent rates calculated from samples taken at the marsh edge, and gray bars represent rates calculated from samples taken at 5m from the marsh edge.

Table 1.

Linear mixed effects model estimates (± SE) for the effects of treatment, distance from the marsh edge, and days since living shoreline installation on rates of potential denitrification (DEA).

Dependent variable
DEA (g N m−1 d−1)
Treatment 0.266 (0.173)
Distance from marsh edge 0.067 (0.152)
Days since installation 0.0004* (0.0002)
Treatment x Distance from marsh edge 0.301 (0.215)
Treatment x Days since installation −0.0002 (0.0003)
Distance from edge x Days since installation −0.001*(0.0003)
Treatment x Distance from edge x Days since installation −0.0001 (0.0004)
Constant 0.196 (0.122)
Number of Observations 92
Groups 6
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Note

*

p < 0.1

Soil % N was found to be significantly higher at 5 m distances than along the marsh edge (F(1,76) = 13.551, p < 0.001), but there was no treatment effect or treatment by distance interaction. Similarly, soil % C was significantly higher at 5 m from the marsh edge (F(1,76) = 14.204, p < 0.001), with no significant differences between treatment and no significant treatment by distance interaction. In addition, DEA significantly decreased (r = −0.59, p < 0.001) with increasing flood frequency in our sampling areas (Figure 8).

Figure 8.

Figure 8.

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The relationship between flooding frequency and denitrification potential, summarized by treatment type. Black dots represent data from control sites, and triangles represent data from living shoreline (experimental) sites. Data from 2016 to 2018 are included. The solid trend line indicates a negative linear association.

Soil shear strength was significantly higher in unaltered sites along the marsh edge (0.5 m) than living shoreline sites at the marsh edge (F(1,20) = 6.907, p = 0.016). Soil moisture values collected during the study were overall significantly higher in living shoreline sites (F(1,64) = 5.485, p = 0.020) and significantly higher at 0.5 m distances (F(1,64) = 4.045, p = 0.04). Soil moisture was lowest in unaltered treatments at 5 m from the marsh edge (F(1,64) = 6.739, p = 0.012). Soil conductivity was the only metric found to be significantly different by year and was higher in 2018 (F(1,66) = 22.16, p < 0.001).

Vegetation Characteristics

Live aboveground biomass did not differ significantly between treatment or distance from the marsh edge but was generally higher in living shoreline treatments at 5 m (Table 2). Additionally, aboveground biomass values did not differ by year. Total C and N values for aboveground biomass samples from summer 2017 were not significantly different by treatment or position on the marsh platform. Total aboveground biomass was positively correlated with soil % C (r = 0.456, p < 0.05). There were no significant relationships between DEA and aboveground biomass.

Table 2.

Average values (± SE) of environmental parameters collected from the marsh platform (soil characteristics, vegetation characteristics, elevation characteristics) at the living shoreline (E) and unaltered sites (C). Superscripts show significant interactions between treatment and distance from the marsh edge for each metric. Means that share the same letter are not significantly different (p > 0.05) according to Tukey multiple comparisons of means tests.

Parameter Metric Living Shoreline Unaltered
0.5 m 5 m 0.5 m 5 m
Soil Characteristics Bulk Density (g cm−3) 0.80 ± 0.03a 0.63 ± 0.04a 0.74 ± 0.07a 0.67 ± 0.12a
Soil Nitrogen (%) 0.34 ± 0.03a 0.54 ± 0.05a 0.38 ± 0.07a 0.67 ± 0.11a
Soil Carbon (%) 4.18 ± 0.32a 7.29 ± 0.69a 5.05 ± 0.89a 8.73 ± 1.37a
Shear Strength at 5cm depth (kPa) 32.4 ± 2.7b 42.5 ± 5.5ab 59.2 ± 8.0a 37.2 ± 6.9ab
Soil Moisture (VWC) 0.71 ± 0.02ac 0.73 ± 0.02ac 0.72 ± 0.02a 0.58 ± 0.05b
Soil Conductivity (mS cm−2) 7.94 ± 0.72a 10.1 ± 1.2a 7.52 ± 0.83a 9.93 ± 1.8a
Vegetation Characteristics
Live aboveground biomass (g m−2) 231.84 ± 21.7a 404.20 ± 93.9a 290.90 ± 34.9a 279.82 ± 56.9a
Vegetation g N m−2 3.30 ± 0.45a 4.31 ± 0.67a 4.25 ± 0.67a 4.17 ± 1.04a
Vegetation g C m−2 87.18 ± 11.4a 137.7 ± 19.2a 131.41 ± 22.2a 149.92 ± 46.1a
Elevation Characteristics Flooding Frequency (%) 0.55 ± 0.09a 0.40 ± 0.07a 0.70 ± 0.04a 0.53 ± 0.07a
Elevation (m, NAVD88) 0.32 ± 0.04 0.39 ± 0.03 0.25 ± 0.02 0.36 ± 0.04
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Discussion

We sought to assess the ability of a living shoreline design consisting of coir logs and other natural materials to stabilize an eroding salt marsh edge and enhance nitrogen removal, as measured by an increase in sediment deposition, elevation, aboveground biomass, and DEA rates, over time. While this study provides multiple lessons learned about the functionality of using nature-based materials for restoration under Northeastern US climatic conditions, it falls short of confirming that coir log living shoreline restoration techniques can stabilize the marsh edge preventing further erosion and enhance DEA rates consistently in dynamic environments. The life expectancy of the coir logs is approximately five years, and we anticipated a substantiative improvement in the marsh condition during our three-year monitoring.

Living Shoreline Effects on Shoreline Stabilization

During our three-year monitoring, results of shoreline accretion did not support our hypothesis that the living shoreline would attenuate waves, leading to decreased erosion, increased sediment deposition within the coir log design, and increased elevations on the marsh platform. The lateral face of the marsh at the four sites with a northeast exposure (two living shoreline and two unaltered sites) did not accrete sediment between the 2017 and 2018 surveys but eroded further, and there were no patterns in vertical elevation loss or gain. Over time, the third living shoreline and its unaltered control site showed minimal gains at the shoreline edge and a loss in average elevation. Sediment accretion in the nearshore area was less at the living shoreline sites than the unaltered sites in two of the three replicates, indicating that sediment deposition coming in from the seaward side may have been hindered by the living shoreline’s physical structure. This could also explain why there was greater sediment capture in the offshore sediment traps at living shoreline sites compared to the unaltered sites, where suspended sediment may be more likely to deflect off the coir log structure and build up in the offshore areas. The living shoreline coir logs did not reduce lateral or edge erosion, nor did they assist with sediment capture which we would have expected during our study period. As a result, there was little opportunity for the expansion of the marsh edge or any subsequent increase in vegetation biomass due to greater area at the living shoreline sites. Fringing salt marshes are noted for their ability to attenuate wave energy, trap sediment, and reduce shoreline erosion (Currin 2019 and references therein), however these attributes require an intact marsh and a conducive environmental regime to be fully realized. This fringing salt marsh experienced several severe weather events which accelerated the degradation and loss of function of the coir logs.

Living Shoreline Effects on Nitrogen Removal and Soil Characteristics

Unlike Onorevole’s et al. (2018) chronosequence study examining nitrogen removal over a 20 year time horizon, this study cannot speak to the denitrification potential of “new” marsh habitat created through installation of the living shoreline (i.e., the backfilled area between the eroding marsh edge and coir logs), but rather speaks to the impact that this shoreline restoration has on the nutrient removal capabilities of the pre-existing marsh. Overall, DEA rates were quite variable among study sites and over the course of our short-term monitoring. DEA rates were consistently higher at living shoreline sites throughout the study, yet they did not change significantly over time. Additionally, DEA rates were highest in the living shoreline sites at the 5 m from the marsh edge location, but in contrast were lowest at unaltered sites at the 5 m location. Overall, our reported DEA rates are comparable to previous studies of DEA in salt marshes along the Northeastern US coast (Wigand et al. 2004), Louisiana (Gardner and White 2010), and California (Cao et al. 2008).

Living shoreline sites at the 5m location reported consistently higher amounts of soil % C and % N compared to any other locations on the marsh and may explain why we saw highest DEA rates at the 5 m location in living shoreline site, since we would expect denitrification to be greater is areas where there is available organic carbon and nitrate. DEA rates in living shoreline areas didn’t differ significantly by position on the marsh, but did in unaltered sites, where DEA rates were higher at the marsh edge. These higher rates at the marsh edge might be explained by a complex set of interacting environmental variables not detected in this study.

DEA rates were higher in areas of lower flooding frequency and as expected, flooding frequency decreased with distance from the marsh edge. Unaltered sites were more frequently flooded than living shoreline sites, and the 0.5 m location more flooded than the 5 m location, meaning the flooding regime at our sites may partially explain the treatment differences in DEA rates. Relevant literature has competing views about how hydrologic conditions affect microbial activity in wetland soils. Multiple studies report that flooded, low marsh zones created anoxic conditions and produced the highest denitrification rates when compared to other parts of the marsh that were infrequently flooded (Johnston et al. 2001, Wigand et al. 2004, Bai et al. 2017). Tomsek et al. (2019) and Orr et al. (2007) highlight the importance of flood pulsing events in creating ideal oxidative-reductive conditions, or denitrification “hotspots”.

The combination of the high soil % C and % N and low frequency of flooding at the 5 m location at the living shoreline sites might have provided the needed substrates and oxidative-reductive conditions for the high DEA rates measured at the living shoreline sites. In coastal restoration work, the addition of labile C through soil amendments has been suggested as a method to facilitate productive, newly formed marsh area by stimulating nutrient cycling and reducing bulk density (Bruland et al. 2004, Ballantine et al. 2014). The coconut fiber logs at the living shoreline sites may have provided an indirect organic carbon subsidy to the adjacent salt marshes. This was an unplanned but possible effect of the installation of the coir logs in our study. As wave energy degraded coir logs over time, coconut fiber became embedded into the vegetation on the marsh platform which was apparent along the wrack line (about 5 m landward of the marsh edge). Despite a 20 m buffer distance between control and experimental sites, some coconut fiber was likely outsourced to adjacent control sites. The potential for living shorelines to export organic matter subsidies to adjacent unstructured control sites supports Onorevole et al. (2018) findings and suggests the impacts of a living shoreline extend well beyond what we delineate to be the habitat.

A consideration for future studies is the impact that storm events may have on coir log degradation, labile C supply, and ultimately denitrification rates. Our coir logs were subjected to several intense weather events throughout the study period, and while we did not directly measure this influence, we suggest that this could potentially be an explanation for the high variance in DEA rates in our study. These results suggest a need to better understand what the addition of coconut fiber may have on N cycling on the marsh platform, especially as coir logs degrade over time.

Living Shoreline Effects on Vegetation

Salt marsh plants are understood to be an essential part of N cycling, either directly through their ability to uptake N and retain it in organic tissue (Schimel et al. 1989, Valiela et al. 2000), or indirectly by supplying labile C (Sherr and Payne 1978) and oxygen belowground via their root systems, which creates oxic-anoxic zones that are ideal for denitrification (Caffrey and Kemp 1990). We found a positive correlation between aboveground biomass and soil C, which supports the idea that vegetation is closely linked to accumulation of biomass. Though not statistically significant, we generally observed greater aboveground biomass in our living shoreline site, because there were not any significant differences in these values over time, we do not see that the living shoreline design had any influence on aboveground biomass. Given that unaltered sites overall had a greater flooding frequency, we expect that aboveground biomass is likely influenced by the tidal regime, meaning vegetation was always sitting at a more favorable flooding frequency in living shoreline areas. While we expected DEA rates to be positively influenced by vegetative biomass, we did not see a correlation between DEA and aboveground biomass in our study. We acknowledge that our sampling size and frequency of the aboveground biomass might need to be increased in order to better investigate connections between aboveground vegetation and DEA rates. Sampling belowground biomass may illuminate a relationship between vegetation, soil % C, soil % N, and rates of DEA.

Living Shoreline Integrity and Limitations to Measuring Ecosystem Service Outcomes

A fundamental aspect of this research was the approach taken to maintain the scientific rigor and integrity of the experiment in the face of constant environmental challenges. The deterioration of the living shoreline natural materials reduces the likelihood the restoration will continue to function in the long term. In future salt marsh restorations, it would be critical to design the restoration and the materials to be used in collaboration with a coastal engineer. The temporal trajectory of the many parameters measured during our study period is also an important consideration. A growing body of living shoreline research suggests that project age and maturity of nearby habitats is a major factor driving the variability in ecosystem services provided by living shorelines (Bilkovic and Mitchell 2013, Gittman et al. 2016b), and our three-year monitoring is likely only providing a snapshot of short term impacts when compared to unaltered reference sites, which are inherently difficult to compare given the natural spatial and temporal variabilities in unaltered marshes (Polk et al. 2022). We likely need longer term monitoring data to better understand changes in some parameters like DEA rates and the factors that influence these rates (edaphic conditions, vegetation growth), which may have a more delayed response to the living shoreline intervention. On the other hand, other living shoreline studies have shown an increase in some ecosystem services in the very short term. Findings by Safak et al. (2020), Manis et al. (2014), and Polk and Eulie (2018) showed that living shorelines provide wave attenuation benefits almost immediately, and these benefits were not influenced by the age of the living shoreline project. Other studies have observed that plant characteristics in living shoreline areas were comparable to natural marshes in under 5 years (Craft et al. 1999, Bilkovic and Mitchell 2013). However, what is not detailed here is the difference in project materials and overall living shoreline design among these studies. We suggest that project siting, materials, monitoring timeframe, and recurring storm events (and subsequently the deterioration of our living shoreline design) were major drivers in the variability of our results.

With predicted sea level rise and increased storm activity in the northeastern US, it is critical that we develop and improve living shoreline designs to support the resilience of salt marshes and to evaluate their success by monitoring ecosystem service outcomes such as shoreline stabilization and nutrient uptake (Whalen et al. 2012). Restoration practitioners and coastal managers might consider living shorelines using natural and hybrid materials to enhance the probability of success of restoring the salt marsh habitat.

Restoration Recap.

  • A living shoreline design consisting of coconut fiber coir logs has been proposed as a nature-based strategy to enhance marsh edge stabilization and the services the marsh provides.

  • This paper presents results of an experimental approach to quantify shoreline stabilization and nitrogen removal services following a living shoreline restoration at the Sengekontacket salt marsh on Martha’s Vineyard, Massachusetts, USA

  • Living shoreline sites did not demonstrate greater marsh edge stabilization over the study period compared to unaltered sites. Contributing factors included low sedimentation rates, severe weather events, our monitoring timeframe, and the lack of hardened materials such as rock or shell that could have provided extra support to the living shoreline structure.

  • Living shoreline sites demonstrated consistently higher rates of denitrification enzyme activity (DEA) than unaltered salt marsh sites. Under the right environmental setting living shoreline restoration has the potential to improve nitrogen removal from the marsh.

Acknowledgements

The authors sincerely thank Drs. Wayne Munns, Timothy Gleason, Richard McKinney, Autumn Oczkowski, Cathleen Wigand and Mr. Joseph Livolsi for their insightful and thorough reviews and constructive comments which greatly improved this manuscript. This research would not have been possible if not for the efforts of many people who assisted with the collection and processing of the samples for this study, for which we are grateful, including M. Brumsted, C. Gleason, T. Gleason, S. Grabbert, B. Hancock, A. Hanson, A. Hanian, S. Hanian, T. Hill, A. Humphries, L. Josephs, R. McKinney, L. O’Brien, L. Schissler, K. Tammi, C. Wigand, and A. Special thanks to Suzan Bellincampi, Philip Hunsaker and volunteers from the Felix Neck Audubon Wildlife Sanctuary, the Oak Bluffs and Edgartown Shellfish Department personnel, Paul Bagnall, the Edgartown Shellfish Constable, Friends of the Sengekontacket Pond summer interns, Mike Charpentier and Kenneth Miller. The research described in this article has been funded wholly by the U.S. Environmental Protection Agency. The views expressed in this article are those of the authors and do not necessarily represent the views or policies of the U.S. Environmental Protection Agency. Any mention of trade names, products, or services does not imply an endorsement by the U.S. Government or the U.S. Environmental Protection Agency. The EPA does not endorse any commercial products, services, or enterprises. This manuscript has tracking number ORD-044333.

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