Variability in nutrient and carbon cycling of constructed marshes

Salt marshes provide valuable ecosystem services, including filtering nutrients and trapping sediments from the water column, but are increasingly threatened by sea-level rise and coastal development. Constructed marshes offer a potential pathway to offsetting marsh loss, but the extent to which they can replace the ecosystem services provided by mature natural marshes is not well understood. Using a chronosequence of constructed marshes (1, 21, and 44 years old) and a natural reference marsh (>100y old) in Beaufort, NC, we assessed the development of plant communities, sediment characteristics, and nutrient and carbon cycling over time. We compared aboveground biomass, surface sediment texture and bulk density, soil carbon (C) and nitrogen (N) stocks, sediment respiration, and porewater nutrient and sulfide concentrations to assess trajectories toward functional equivalence. Vegetation structure, sediment texture, bulk density, porewater nutrients, and soil respiration converged with the natural reference marsh within approximately two decades of construction. In contrast, surface sediment C and N stocks remained lower than reference levels even after ~20 years, and patterns appeared to be influenced by geomorphic setting and episodic sediment deposition rather than marsh age alone. Methane fluxes were negligible across all sites, and carbon dioxide respiration exhibited seasonal patterns but limited age-related differences. Our results demonstrate that while constructed marshes in sandy back barrier environments can achieve vegetation and physical equivalence with natural marshes within a few decades, long-term C and N storage develop more slowly. Further, the trajectory of development can be particularly complicated in shoreline marshes as a result of episodic deposition events. These findings highlight the importance of sediment texture, geomorphic context, and long-term monitoring when evaluating restoration success and estimating carbon benefits of estuarine living shoreline projects.

Wetlands are highly valued for the ecosystem services they provide, including improving water quality, trapping suspended sediment, and their disproportionate ability to sequester carbon (Bansal et al., 2023; Mason et al., 2023). Despite only covering 3-8% of land surface (Mcleod et al., 2011; Lehner and Doll, 2004; Bansal et al., 2023), wetlands store up to half of all terrestrial soil organic carbon (Mitsch and Gosselink, 2015; Nichols and Peteet, 2019; Bansal et al., 2023). Over the past century, large expanses of salt marsh were historically filled for agriculture and coastal development, resulting in a global decrease of marsh area (Crooks et al., 2011; Bu et al., 2015; Mason et al., 2023). Loss of saltmarsh area is expected to continue as a consequence of growing coastal populations, rising sea level, and increased intensity of hurricanes (Culliton, 1998; Craft et al., 2016; Saintilan et al., 2022; Mason et al., 2023). In response, federal strategies (The White House, 2024) emphasize the need to support restoration methods that mitigate wetland loss and increase coastal resilience.

One such mitigation strategy is the use of natural infrastructure: a restoration method that utilizes natural elements sometimes in conjunction with engineered structures to offset and reduce habitat loss and increase coastal resilience (Arkema et al., 2013; Barbier et al., 2013; Sutton-Grier et al., 2015; Davis et al., 2015). Living shorelines are a form of natural infrastructure that incorporate natural shoreline vegetation either alone, or in combination with structural components like sills (Currin et al., 2010; Davis et al., 2015). Living shorelines have gained popularity over the past decade (Smith et al., 2020; Mason et al., 2023), aiming to protect coastlines while maintaining the ecosystem services provided by natural marshes. Technically, “living shorelines” describes marshes that have been modified or constructed, ranging from simple supplemental plantings in native sediment to the creation of new marshes using imported fill. This distinguishes them from “natural” marshes, which are defined in this context as being unmodified. However, these restoration efforts are undertaken with the assumption that constructed marshes will follow a predictable development ultimately mirroring the biotic and abiotic functions in nearby natural marshes. This trajectory and more precisely, the point at which constructed marshes function like natural ones (e.g., similar nutrient uptake, sediment accretion and stabilization, carbon storage) remains uncertain, with studies showing varied results (Craft et al., 1999; Mossman et al., 2012; Moreno-Mateos et al., 2012; Osland et al., 2012; Zhao et al., 2016; Burden et al., 2019; Mason et al., 2023).

Constructed living shoreline marshes rapidly resemble natural marshes in terms of vegetative cover, yet ecosystem functions linked to soil biogeochemistry typically differ from natural marshes for decades, or even centuries (Craft et al., 2003; Hossler and Bouchard, 2010; Osland et al., 2012; Craft et al., 2016). Soil biogeochemistry governs sediment nutrient cycling, microbial activity, and vegetation dynamics, which influence primary production and organic matter turnover. For example, orthophosphate and ammonium, the bioavailable forms of phosphorus and nitrogen in salt marsh sediments and porewater, control plant growth and microbial metabolism, and their availability is tightly linked to sediment redox conditions and organic matter decomposition (Alongi, 1998; Sundareshwar et al., 2003; Guo et al., 2007; Neubauer, 2014; Mitsch and Gosselink, 2015). In natural marshes, tidal flushing and plant-mediated oxygen transport maintain redox gradients that facilitate nutrient cycling while limiting toxic hydrogen sulfide (H 2 S) accumulation and methane production (Mendelssohn et al., 1981; Howes et al., 1986; Laanbroek, 2010; Mueller et al., 2020; Capooci et al., 2023; Zhou et al., 2022). However, landscape modifications, like impoundment and restricted tidal exchange, can disrupt these processes by altering redox conditions, increasing sulfide accumulation and methane production, and changing nutrient concentrations relative to unaltered marshes (Chambers et al., 1998; Craft et al., 2003; Morse et al., 2012). Consequently, constructed marshes may differ from natural marshes in porewater nutrient concentrations and sediment chemistry long after vegetation has established.

Evaluating the development of these biogeochemical functions requires observations at longer timescales; yet, most research on living shorelines and constructed marshes has focused on projects less than five years old, providing insight primarily into short-term performance (Smith et al., 2020). Because many ecosystem services linked to sediment biogeochemistry develop slowly, long-term monitoring is essential for assessing whether constructed marshes achieve functional equivalence with natural systems. However, such monitoring is time- and resource-intensive, often making it impractical. We are not aware of any living shoreline funding programs that provide for long-term monitoring. As a result, there is currently only a rudimentary understanding of how created living shorelines evolve as they mature. A space for time approach offers one approach for filling this knowledge gap by allowing inference of restoration trajectories across multiple stages of marsh development without the need for extended temporal studies (Craft et al., 2003; Hutto and Belote, 2013; Onorevole et al., 2018; Palinkas and Staver, 2026). Chronosequences that extend past the average five years are key to determining restoration trajectories given findings that plant aboveground biomass and wildlife frequently match reference goals within a decade, while biogeochemical services, such as carbon storage, take much longer and may never be fully restored (Craft et al., 2016; La Peyre et al., 2014). Traditionally, research conducted on living shorelines focuses on their ability to provide protection and create habitat, with few studies analyzing the parameters affected by soil biogeochemistry like carbon storage and nutrient cycling (Smith et al., 2020). In a literature review conducted by Smith et al. (2020), only two papers analyzed nutrient cycling and carbon storage on living shorelines: Onorevole et al. (2018) and Davis et al. (2015). To calculate net carbon benefits of restoration, both greenhouse gas fluxes and carbon sequestration need to be accounted for (Mason et al., 2023). Hence, Mason et al. (2023) specifically calls for more data on greenhouse gas fluxes and carbon accumulation rates of salt marshes, which are required for constructing net carbon budgets needed to inform management and future restoration strategies. In spite of this, greenhouse gas dynamics are rarely quantified at constructed marshes, even with the potential difference in carbon dioxide and methane fluxes between constructed and natural sites (Bartolucci et al., 2024).

Utilizing a space for time approach to evaluate the trajectory of ecosystem services provided by constructed marshes in Beaufort, NC, this study aims to answer the question: How long does it take for constructed shoreline marshes to function like natural shoreline fringing marshes? To address this, we examined how aboveground biomass (AGB), sediment texture, soil nitrogen and carbon content, sediment respiration, and porewater nutrient concentrations differed between sites of varying restoration ages compared to a nearby reference marsh.

It is critical to evaluate these biogeochemical trajectories within their specific geomorphic context. While much of the existing literature on salt marsh restoration focuses on fine-grained, silt/clay-dominated tidal wetlands with high organic inputs, created living shorelines in back-barrier environments are frequently constructed on coarse sandy substrates subject to wind and wave energy. Consequently, functional recovery rates in sandy estuarine shorelines may diverge significantly from fine-grained systems. In this study, we constrain our conclusions to sand, high-energy back barrier marsh systems, providing a necessary comparison to established silt/clay restoration settings.

Study site

Data was collected from three constructed salt marshes of known ages and one natural shoreline fringing salt marsh surrounding Pivers Island in the Newport River estuary near Beaufort, NC (Figure 1). Located within 0.1 km 2 of each other, the proximity of these sites ensured similar tidal inundation, salinity, and environmental conditions. The three constructed marshes varied in age: the Pilot Project (Pilot) was 1 year old at the time of sampling, Pivers Island West (PIW) was 21, and Kirby Smith Island (KS) was 44. The Pilot site was constructed in 2022 along the Northeast end of Pivers Island. The site features a tiered design with a QuickReef ® structure delineating the high and low marsh areas, both of which are filled with coarse sand, and coarse graded gravel aggregate. A rock sill delineates the waterward edge. PIW was established in 2002 and consists of a narrow, fringing marsh constructed with coarse sand fill graded to the appropriate elevation for vegetation and protected by a stone sill along the shoreward edge to reduce wave energy. Following the construction of both Pilot and PIW sites, Spartina alterniflora and Spartina patens were planted at elevations appropriate for the species. KS was established in 1979 by planting a few S. alterniflora (<20) on an exposed mud flat, and has since expanded to a ~2600 m 2 patch marsh island. The reference marsh, located north of Pivers Island, is referred to as Pivers Island North/Natural (PIN) and given an arbitrary age of >100 years. The naming scheme for PIN, KS, and PIW was established in previously published work (Davis et al., 2015) and kept the same here for continuity. Pilot is the newest living shoreline in the area and reported on for the first time in this manuscript. All of the created sites in this study involved the planting of marsh vegetation in pure sand. As a result, downcore profiles of sediment organic matter illustrate the full depth of the marsh sediment column; areas below the root zone remain pure sand. As of 2015, the depth of carbon-enriched marsh sediments at both PIW and KS was < 15cm. For this study, we focused on shallow surface sediments (< 12 cm) since it is at these shallow depths that we would expect to see differences among sites that are attributable to living shoreline establishment.

At each of these sites, three monitoring plots were established for measuring vegetation, porewater, and sediment characteristics (Figure 2). Plots occupied similar elevations, both within and among sites, to ensure analogous tidal inundation (Figure 3; 0.2-0.6 ± 0.02 m NAVD88). All plots were above local mean sea level (-0.03 m, NAVD88) and below the mean high water level (0.55 m, NAVD88), subjecting the plots to regular daily tidal inundation throughout the study period from March 2023 to May 2024 as measured by the DUML (station #8656483) NOAA COOPS NWLON station which is ~ 0.5 km from the sites (Figure 3). Elevation was determined by optically leveling (Dewalt ® DW096) to a benchmark of known elevation for plots at Pilot and PIW. At KS and PIN, elevation was measured with Real Time Kinematic GPS (Trimble ® R12i) with the base station located and referenced to benchmark 865–6483 E TIDAL (PID DE7961).

Vegetation surveys

Vegetation was surveyed quarterly at the established monitoring plots (Figure 2). These quarterly surveys recorded species presence and visually estimated percent cover within the marked 1 m² plot, followed by counting S. alterniflora stems in a 0.25 m² sub plot (Figure 2). Within the same 0.25 m² sub plot, the height of 10 randomly selected S. alterniflora stems were measured. During the last survey in May 2023, vegetation was clipped at PIW, KS, and PIN to determine above ground biomass (AGB). AGB was harvested from three 0.0625 m² areas per site, located near the vegetation monitoring plots. Cut stems were washed to remove all epiphytes and sediment before being dried at 60 °C. The weight of the dried material was recorded as standing AGB. The one-year-old site (Pilot) was excluded from clipping to protect new plantings.

Sediment – elemental and grain size analysis

Surface sediment cores were collected from within the monitoring plots at all sites in March 2023 for analysis of bulk density, grain size, and carbon content (Figure 2). A total of 24 cores were collected, with two cores (one duplicate at each location; 3.5 cm in diameter) taken from each of the three monitoring plots across all four sites. Cores were 12 cm deep unless the total sediment column was < 12 cm; the shortest core collected here was 9 cm. One of each duplicate core was used for bulk density and elemental analysis, and the second core was used for grain size analysis. Once in the lab, all cores were extruded and horizontally divided into equal top and bottom sections (approximately 4.5–6 cm each).

The elemental analysis cores were dried at 60 °C for 48 hours, weighed to determine bulk density, and then homogenized by hand with a mortar and pestle. A subsample of the dried and homogenized sediment was ground into a fine powder (8000D Mixer/Mill ® - Dual High-Energy Ball Mill), weighed into silver capsules, and acidified dropwise with 1N HCl to remove calcium carbonate. The samples were then analyzed with a Perkin Elmer 2400 series II CHNS/O analyzer to determine organic carbon (C) and total nitrogen (N) content. Bulk carbon and nitrogen density were calculated for each depth interval. The organic carbon and nitrogen content was subsequently converted to carbon and nitrogen stock (stock = % C (or N) content * bulk density * 1000 * interval depth).

To characterize surface sediment grain size, cores were sectioned into top and bottom, and rinsed through a series of stacked sieves to sort the sediments into four categories: gravel and macro-organic matter (both > 2.0 mm), sand (0.063–2 mm), and silt (<0.063 mm). The gravel and macro-organic matter (MOM) fraction were manually sorted to separate inorganic and organic material. Following sorting, each fraction was dried and weighed to calculate its percentage contribution to the total weight of the core.

Sediment respiration

Soil respiration was measured monthly from May 2023 to February 2024. Sampling took place over one to two consecutive days during falling tides. Triplicate PVC collars (32 cm diameter, 16 cm height) were driven 5–10 cm into the ground at similar elevations (0.277-0.380 m NAVD88) to the monitoring plots at all four sites (Figure 2). Sediment respiration was measured monthly with an ultra-portable greenhouse gas analyzer (GGA; Los Gatos Research, Inc., CA, United States) by attaching an opaque, 19.15 L PVC chamber retro-fitted with bike tire tubes to create an air tight seal between the collar and the lid. The limit of detection for the GGA is 0.01 ppm for CH 4 and 1 ppm for CO 2. The chamber was equipped with a fan to ensure that the sampled air was well-mixed (Figure 4). Sampling ports on top of the chamber were connected to the GGA with gas-impermeable tubing (Figure 4). At each monthly sampling event, dry CO 2, dry CH 4, and temperature concentrations were simultaneously measured for 5 minutes with part-per-billion sensitivity via off-axis integrated cavity output spectroscopy. Chamber closure was restricted to 5 minutes per measurement to minimize pressure buildup and eliminate the need for mathematical pressure corrections. Temperature and atmospheric pressure were accounted for when solving for mol of gas in the chamber (n = PV/RT). Production rates were calculated from changes in concentration over time.

Porewater analysis

Between March 2023 and May 2024, porewater was collected monthly from within each of the monitoring plots from a depth of 10 cm (the active root zone) using a peristaltic pump connected to a push point sampler with gas-tight tubing. Samples were collected when water levels were below plot elevation. Porewater was filtered through a 0.45 μm Polyether Sulfone (PES) filter, and stored frozen until analysis (within 30 days of sampling), then thawed overnight in a refrigerator. All samples were processed independently. Orthophosphate and ammonium concentrations were determined by standard colorimetric methods, as described in Karl and Tien (1992) and Parsons (1984).

A second porewater sample for analysis of hydrogen sulfide was collected, filtered through a PES filter and complexed with zinc acetate (ZnAc) in the field. Hydrogen sulfide concentrations were determined by Cline’s method (Cline, 1969).

Statistical analysis

Statistical analysis was performed with R (version 4.5.1) using the tidyverse, lme4, ARTool, and emmeans packages. To test for normality and homogeneity, the Shapiro-Wilk test, Q-Q Plots, and Levene’s Test were utilized. Once normality and homogeneity were confirmed, one-way ANOVA and Tukey HSD post-hoc tests were run for inter-site comparison. For non-normal data, an ART ANOVA was used as a non-parametric test specifically to overcome non-normality in 2-way designs. Welch’s ANOVA was used if the assumption of homogeneity was violated.

Aboveground biomass

All sampled plots were dominated by Spartina alterniflora. Despite the young age of the Pilot (1 y), other plant species, including Borrichia frutescens and Limonium carolinianum, were already recruiting to the site (Table 1). At KS (44 y) plots, only S. alterniflora was present. At PIW (21 y) and PIN (>100 y), S. alterniflora, L. carolinianum, B. frutescens, and Salicornia spp. were observed (Table 1). Mean dry AGB was between 235–257 g m -2 for PIW (21 y), KS (44 y), and PIN (>100 y). (Pilot (1 y) was excluded due to recent restoration at the site) (Figure 5). Though within-site variance contracted with marsh age (Figure 5), AGB was similar across site ages (Welch’s F-test, F (2.00,5.24) = 0.069, p = 0.935).

Sediment characteristics

Average bulk density ranged between 1.36-1.54 g cm -3 at all four sites within standard deviation of each other between sites (Table 1). Grain size measurements showed all sites were dominated by sand (Figure 6; >68% for all cores). The surface sediments at Pilot (1 y) contained lower proportions of sand and higher proportions of silt compared to all other sites (Tukey’s HSD, p < 0.001 for both), along with greater gravel fractions than at KS (44 y; p = 0.005) and the natural site (p = 0.014). These trends persisted throughout the core: Pilot bottom intervals exhibited lower sand and higher silt fractions relative to all mature sites (p ≤ 0.001). In contrast, grain sizes were statistically indistinguishable among PIW (21 y), KS (44 y), and the natural reference (p ≥ 0.058 across all pairwise comparisons). The larger percentage of gravel and silt at Pilot (1 y) compared to the other sites were attributed to the composition of the placement material.

Surface organic carbon stocks differed among sites, ranging from 101 to 1,086 g C m −2 (Figure 7), but showed no significant overall difference among sites (F 3,8 = 2.471, p = 0.136). The unusually high carbon stock measured in one Pilot (1 y) core resulted from a pocket of carbon-rich clay in the placed material. Excluding this anomalous core, organic carbon was highest at the natural reference site and lowest at KS (44 y). Surface nitrogen content differed significantly by site (F 3,8 = 6.896, p = 0.013). The natural reference contained greater nitrogen stocks than Pilot (1 y; p = 0.020) and KS (44 y; p = 0.017), whereas PIW (21 y) did not differ significantly from the reference or the other marsh sites (p ≥ 0.05).

Sediment respiration

During the study period, methane respiration was continuously at or near zero with little variation over time or between sites. Carbon dioxide respiration fluctuated seasonally at all sites (Figure 8). Kirby Smith (KS) consistently had the highest rate of respiration relative to all other sites (p < 0.023). The max respiration rate was recorded in August 2023 at KS (44 y: 43.9 µmol CO 2 m -2 min -1). The minimum rate was recorded in February 2024 at Pilot (1 y: 0.68 µmol CO 2 m -2 min -1).

Porewater analysis

Porewater orthophosphate concentrations across all four sites remained consistently low throughout the year (0-1 µM PO 4), with no overall difference among sites (p = 0.9287). Ammonium concentrations were similarly low across all study sites (0-74 µM NH 4). Although non-parametric ART ANOVA detected statistically significant main and interaction effects (p < 0.001) driven by subtle rank order shifts at low concentrations (e.g., persistent near-zero baselines at PIN versus minor seasonal pulses at KS), absolute concentration differences among sites were minimal. In contrast, hydrogen sulfide varied significantly by site (F 3,49.59 = 9.36, p < 0.001) and season (F 11,66.43 = 16.15, p < 0.001); site x date interaction, (F 25,66.27 = 10.06, p < 0.001) (Figure 9). This interaction was driven by the oldest natural reference site, PIN (>100 y), which accumulated significantly higher porewater sulfide than all constructed sites (p ≤ 0.0017) during the summer, peaking in mid-August.

The results from this study suggest that in sandy, high-energy shoreline marsh systems, sediment structure and vegetation follow a rapid trajectory toward equivalence with natural marshes, whereas surface sediment carbon (C) and nitrogen (N) stocks may require longer time scales, likely exceeding two decades. Across the shoreline sites (Pilot and PIW), vegetation structure, soil texture, respiration rates, and porewater nutrients mirrored the reference marsh. In contrast, surface C and N stocks were lower than the natural marsh even 20 years post construction, confirming that biogeochemical equivalence lags behind sediment structure and vegetative metrics. These results contrast with findings from silt- and clay-dominated marshes (Craft et al., 1991, 2003; Onorevole et al., 2018), where sediment structure develops more slowly due to dependence on the accumulation of fine-grained sediments. These findings highlight that sediment texture strongly influences restoration trajectories, and that even in sandy marshes, it may be multiple decades before full development of soil carbon and nitrogen in restored marshes.

Aboveground biomass

Measurements of vegetative biomass, plant diversity, and plant density are commonly used as indicators of marsh restoration success because vegetation is foundational to marsh system function and relatively easy to quantify. Previous studies document vegetative communities reaching equivalence within 3–5 years (Ebbets et al., 2020; Craft, 2022). Our chronosequence does not include sites from the early post-construction period; however, earlier monitoring of PIW indicated that vegetation (plant species composition and the percentage cover, stem density and stem height) resembled the natural fringing marsh within three growing seasons (Currin et al., 2008). These observations support the conclusion that plant community structure likely reaches equivalence within 5 years and before our second time step of 21 years. Vegetative equivalence should result in similar abilities to trap sediment and therefore similar accretion rates across sites of different ages. The vegetative community creates a uniform physical baffle that reduces tidal flow velocity and attenuates wave energy, facilitating the settling of suspended particles regardless of the marsh age (Cahoon et al., 2020; Moffett et al., 2015); however, vegetative similarity is less of a controlling factor on carbon sequestration (Craft, 2022).

Sediment characterization

There was little variability in sediment bulk density across our study sites, a finding that diverges from the established restoration literature. Typically, constructed marshes exhibit significantly higher bulk densities than natural reference marshes, with a predictable decline over multi-decadal timescales as vegetation traps fine silt/clay particles and organic matter accumulates (Craft et al., 1991, 2003; Onorevole et al., 2018). While achieving physical parity can take decades in fine-grained systems, our natural reference marsh (PIN) was predominantly sandy. As a shoreline fringing marsh in a low-estuary environment, PIN experiences wave conditions and tidal flushing that restrict the retention of fine silts (Grabowski et al., 2011; Stoorvogel et al., 2024).

Our data suggests that when sand fill in a constructed marsh closely matches the coarse substrate of a reference marsh as it happens in these sandy, shoreline adjacent systems, the physical sediment structure (bulk density and grain size) achieves equivalence rapidly, bypassing the 30–100 year physical maturation period typical of silty marshes (Burden et al., 2019; Craft, 2022). This broadly applies to thin-layer placement as well, as adding sediment aims to bring physical and chemical properties into equivalence with background conditions (La Peyre et al., 2009). It is important to note that if the initial fill or planting material is inconsistent across the chronosequence, the estimated trajectory will be skewed, leading to inaccurate estimates of time to equivalence, a risk when using space for time. Notably, both observed and modeled nearshore wave energy at these sites are on the lower end for the region, characterized by short fetch (<500 m) and negligible boat traffic (Currin et al., 2017). Consequently, the sandy substrate across these sites is not driven by high wave energy, but rather by narrow marsh widths that limit fine particle retention.

Rapid physical stabilization did not serve as a proxy for overall sediment equivalence. Unlike findings where reference-like fill accelerated total ecosystem recovery (Wolters et al., 2005), our sites demonstrate a clear decoupling between physical sediment structure and carbon and nitrogen accumulation. Even with reference-like soil structure, surface C and N stocks lagged behind natural reference levels. While surface C and N increased from the 1 year old site (Pilot) and the 21-year-old site (PIW), levels remained well below the natural reference marsh, demonstrating that biogeochemical equivalence requires more than two decades.

Unlike vegetative structure and porewater nutrients which converged with natural reference levels within approximately two decades, surface carbon and nitrogen stocks exhibited a delayed and non-linear trajectory. This contrasts with fine-grained, silt/clay restored marshes, where soil organic carbon stocks often achieve functional equivalence within 20–30 years due to higher mineral-organic particle binding and lower physical export (Craft et al., 2002, Craft et al., 2003).

While targeting shallow surface sediments (0–12 cm) effectively isolates the primary zone of recent organic matter input and root-zone biogeochemistry in these young sand-fill systems, it also introduces a key constraint when evaluating net blue carbon capacity. Because carbon stock calculations were restricted to the upper 12 cm, our estimates represent surface accumulation rates rather than total system-wide blue carbon pools. Given a regional sea-level rise rate of ~3 mm/y, the 44 year old site (KS) would suggest approximately 13 cm of accumulation which is almost entirely captured in surface cores collected at the KS site and is completely captured in the surface sediment cores of the Pilot (1 y) and PIW sites (21 y). Therefore, we feel confident that the surface sediment cores captured the comparable biogeochemical root-zones at all sites. However, limiting sampling to the top 12-cm omits deeper soil organic carbon pools that develop over decadal to centennial timescales. Future evaluations incorporating deep soil cores (>30 cm) are necessary to fully quantify deep carbon burial and total blue carbon capacity in aging constructed shorelines.

The lower C stock observed at KS (44 y) highlights that biogeochemical trajectories in shoreline marshes are strongly modulated by geomorphic setting and episodic sedimentation rather than marsh age alone. Current C densities at KS (0.001 g C cm −3) and PIW (0.005 g C cm −3) are markedly lower than the values reported by Davis et al. (2015) for the same sites (0.024 and 0.016 g C cm −3, respectively). This discrepancy likely stems from differences in sampling plot elevation between the two studies rather than a loss of organic matter. The Davis et al. (2015) PIW sampling plots were at 0.19 m (NAVD88) while PIW plots in this study (measurements in 2023) were all above 0.4 m (NAVD 88). At KS, Davis et al. (2015) specifically identified the highest elevation of KS at -0.14 m NAVD88. In contrast, the 2023 plots from this study were located between 0.28 and 0.35 m NAVD88, suggesting at least 0.42 m of elevation gain in just eight years. Such a rapid elevation gain far exceeds projected accretion rates based on regional sea-level rise (~3 mm y -1) and points to substantial inorganic sediment deposition. This is supported by a visible sandy deposit at KS, but the geomorphic context of the site suggests the dynamics may be even more complex. KS is a small island surrounded by extensive sand bars and shallow channels, forming a geomorphically enclosed system well-positioned to trap sediments during high-energy events. Several major storms have affected the region since 2015 (including Matthew, Irma, Florence, and Dorian), likely contributing pulses of mineral sediment from adjacent sand flats and channels. These storm-driven deposition events can quickly raise marsh surface elevation, particularly in island settings where sediment delivery is constrained under normal tidal conditions but becomes amplified during storm surge or overwash (Turner et al., 2006; Walters and Kirwan, 2016; Castagno et al., 2018; Fitzgerald et al., 2020; Pannozzo et al., 2021; Phillips, 2022). As a result, surface soils sampled in this study may primarily consist of recently deposited, carbon-poor sediments, rather than representing long-term organic accumulation. Consequently, the 44-year age assigned to the cores is likely inaccurate. This dynamic sedimentation regime complicates assumptions about the trajectories of carbon accumulation in restored systems. Assuming a consistent rate of elevation increase and belowground biomass production, natural marshes reach an equilibrium with respect to inputs of carbon and mineralization (Mudd et al., 2009; Davis et al., 2015). While newly constructed marshes are expected to increase in carbon content until an equilibrium is reached (Davis et al., 2015), episodic sedimentation can obscure these trajectories. If sedimentation outpaces belowground biomass production, surface soils may appear biogeochemically immature (contain less carbon) despite the advanced age of the established marsh. This highlights the importance of considering site-specific geomorphology and disturbance history when evaluating the long-term development and function of restored marshes.

Sediment respiration

Soil respiration in salt marshes reflects microbial and plant metabolic activity and is regulated by organic matter availability, oxygen, and temperature (Krauss et al., 2012; Forbrich and Giblin, 2015; Morris and Whiting, 2024; Bartolucci et al., 2024). Although restored wetlands may take more than 50 years to become net greenhouse gas sinks (Craft et al., 2016; Hemes et al., 2019), no significant differences in winter soil respiration were observed among sites in this study. The Pilot, PIW, and PIN sites exhibited respiration rates within one standard deviation across seasons, indicating that marsh age did not strongly influence respiration. This convergence suggests that, in sandy systems, microbial and plant respiration achieve functional equivalence within the first year of construction. In contrast, the KS site exhibited elevated summer respiration, likely due to the general boost in microbial metabolism from warmer summer temperatures, which could lead to the turnover of labile organic matter. This process would subsequently result in less macro-organic matter present (Figure 8) and lower soil C and N levels documented in the sediment samples collected from KS (Wigand et al., 2009; Figure 7). Additionally, the increased respiration at KS could result from depositional events on the island. The rapid or complete burial of active marsh vegetation triggers pulses of decomposition and microbial respiration as the newly buried organic matter is consumed. This scenario again suggests the sediment sampled when coring at KS is likely younger than marsh construction.

Porewater constituents

Porewater nutrient concentrations reflect organic matter availability and nutrient cycling processes (Craft et al., 1991). In contrast to studies showing prolonged nutrient deficits in constructed marshes (Craft et al., 1991), ammonium and orthophosphate concentrations in this study did not differ between constructed and natural sites and remained consistently low across the chronosequence. These results suggest that porewater nutrient dynamics in sandy, high-energy marshes converge relatively quickly, likely due to limited organic matter accumulation, rapid tidal flushing, and efficient plant uptake. The low ammonium and orthophosphate concentrations observed across all sites indicate that this aspect of biogeochemical function has reached equivalence one year after construction, despite persistent deficits in soil C and N stocks, and further highlights how sandy marshes can rapidly achieve functional similarity in some processes while remaining biogeochemically immature in others.

Hydrogen sulfide (H 2 S), a byproduct of organic matter decomposition in anoxic marsh porewaters, can inhibit plant growth, especially with organic-rich amendment soils (Koch et al., 1990; Craft et al., 2016). However, the high tidal elevation and sand content of this study’s sites promote aeration and low percent organic matter keep H 2 S concentrations low across all locations (Figure 9). Even the highest reported H 2 S concentration (365 µM), was below the millimolar phytotoxicity threshold for marsh macrophytes (500 µM, Lamers et al., 2013). Thus, H 2 S levels do not appear to limit vegetative success in these marshes.

It is essential to understand how restored wetlands function compared to natural wetlands, especially when they are used as a strategy to offset the loss of ecosystem services due to wetland degradation (Race and Fonseca 1996; Moreno-Mateos et al., 2012). Despite the value of restoration monitoring for tracking ecosystem services, time and budget constraints often limit long-term research (Onorevole et al., 2018). By utilizing a space for time approach, this study provides a proxy for the long-term trajectory toward functional equivalence in sandy, high-energy environments. Our findings indicate that the geomorphic setting significantly influences the pace of restoration. In sandy back-barrier systems, structural attributes, including vegetation density, sediment texture, and porewater nutrient dynamics, can reach equivalence with natural marshes in less than two decades. This rapid physical stabilization contrasts with silt-dominated systems and suggests that living shorelines in these environments can quickly provide habitat and coastal protection services comparable to natural marshes. However, a biogeochemical lag persists. This highlights that while a marsh may appear physically equivalent, its capacity for long-term carbon sequestration develops slowly.

Managers must distinguish between rapid structural equivalence and the much slower progression toward biogeochemical maturity, which may require decades. Project success is often evaluated under assumptions of zero post-construction management; however, the implementation of adaptive management may shorten the timeframe required to achieve functional equivalence. Overall, this study demonstrates that constructed sandy shoreline marshes can mature rapidly, while underscoring the need for site-specific monitoring to accurately quantify their contribution to carbon storage compared to unmodified natural marshes. Future work is necessary to better understand the nuances in biogeochemical maturation of constructed salt marshes across different settings and modification strategies. Expanding future research to include more time steps and a broader range of marsh ages, while excluding geomorphologically dynamic island sites like Kirby Smith, will help clarify these patterns.