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. Author manuscript; available in PMC: 2021 Jun 1.
Published in final edited form as: Ecol Appl. 2020 Mar 2;30(4):e02085. doi: 10.1002/eap.2085

Rapid peat development beneath created, maturing mangrove forests: ecosystem changes across a 25-year chronosequence

Michael J Osland 1,*, Laura C Feher 1, Amanda C Spivak 2, Janet A Nestlerode 3, Alejandro E Almario 3, Nicole Cormier 4, Andrew S From 1, Ken W Krauss 1, Marc J Russell 3, Federico Alvarez 3, Darrin D Dantin 3, James E Harvey 3, Camille L Stagg 1
PMCID: PMC7423248  NIHMSID: NIHMS1611098  PMID: 31991504

Abstract

Mangrove forests are among the world’s most productive and carbon-rich ecosystems. Despite growing understanding of factors controlling mangrove forest soil carbon stocks, there is a need to advance understanding of the speed of peat development beneath maturing mangrove forests—especially in created and restored mangrove forests that are intended to compensate for ecosystem functions lost during mangrove forest conversion to other land uses. To better quantify the rate of soil organic matter development beneath created, maturing mangrove forests, we measured ecosystem changes across a 25-year chronosequence. We compared ecosystem properties in created, maturing mangrove forests to adjacent natural mangrove forests. We also quantified site-specific changes that occurred between 2010 and 2016. Soil organic matter accumulated rapidly beneath maturing mangrove forests as sandy soils transitioned to organic-rich soils (peat). Within 25 years, a 20-cm deep peat layer developed. The time required for created mangrove forests to reach equivalency with natural mangrove forests was estimated as: (1) < 15 years for herbaceous and juvenile vegetation; (2) ~55 years for adult trees; (3) ~25 years for the upper soil layer (0–10 cm); and (4) ~45–80 years for the lower soil layer (10–30 cm). For soil elevation change, the created mangrove forests were equivalent to or surpassed natural mangrove forests within the first five years. A comparison to chronosequence studies from other ecosystems indicates that the rate of soil organic matter accumulation beneath maturing mangrove forests may be among the fastest globally. In most peatland ecosystems, soil organic matter formation occurs slowly (centuries, millennia); however, these results show that mangrove peat formation can occur within decades. Peat development, primarily due to sub-surface root accumulation, enables mangrove forests to sequester carbon, adjust their elevation relative to sea level, and adapt to changing conditions at the dynamic land-ocean interface. In the face of climate change and rising sea levels, coastal managers are increasingly concerned with the longevity and functionality of coastal restoration efforts. Our results advance understanding of the pace of ecosystem development in created, maturing mangrove forests, which can improve predictions of mangrove forest responses to global change and ecosystem restoration.

Keywords: chronosequence, coastal wetland, ecosystem development, functional equivalency, mangrove forest, peat, sea-level rise, soil organic matter

Introduction

In the face of climate change and global-scale ecosystem loss and degradation, ecologists and natural resource managers are increasingly challenged to identify conservation, restoration, and land management practices that can support critical ecosystem services, enhance the ability of species and ecosystems to adapt to climate change, and mitigate against the effects of climate change via carbon storage (Glick et al. 2011, Schlesinger and Bernhardt 2013, Stein et al. 2014, Jackson et al. 2017, Fargione et al. 2018, USGCRP 2018). Here, we examine ecosystem development and soil organic matter accumulation in created, maturing mangrove forests, which are among the most productive, valuable, and carbon-rich ecosystems on Earth (Alongi 2009, Donato et al. 2011, Costanza et al. 2014).

Within recent decades, coastal scientists have increasingly demonstrated that, in addition to providing many valuable ecosystem goods and services (Barbier et al. 2011, Costanza et al. 2014), mangrove forests and other vegetated coastal ecosystems (e.g., salt marsh and seagrass ecosystems) have the potential to support very high rates of soil carbon sequestration and burial (Chmura et al. 2003, Nellemann et al. 2009, Mcleod et al. 2011, Pendleton et al. 2012). High rates of soil carbon burial in mangrove forests are due primarily to the production, accumulation, and burial of roots under rising seas, which can lead to the development of deep, organic-rich soils (i.e., peat) (McKee 2011, Breithaupt et al. 2012, Krauss et al. 2014). Despite growing understanding of the factors controlling mangrove soil carbon stocks (Alongi 2014, Sanders et al. 2016, Atwood et al. 2017, Hayes et al. 2017, Hinson et al. 2017, Holmquist et al. 2018, Osland et al. 2018b, Rovai et al. 2018, Sanderman et al. 2018, Twilley et al. 2018), there is a need to advance understanding of the speed of peat development in mangrove forests, especially in created and restored mangroves.

The rate of global mangrove loss in the last century has been very high, due primarily to direct conversion of mangrove forests to other land uses (Duke et al. 2007, Hamilton and Casey 2016). To compensate for the ecosystem goods and services that are lost during mangrove conversion, there has been an increasing emphasis on mangrove restoration in many parts of the world. Although mangrove restoration efforts can be highly successful if natural hydrologic regimes are established, many poorly-planned restoration efforts fail (Lewis 2009, López-Portillo et al. 2017). Moreover, post-restoration funding for monitoring and research is often limited. Thus, there are knowledge gaps regarding the potential trajectories of mangrove ecosystem development following restoration, including the amount of time it takes for restored and created wetland ecosystems to become functionally equivalent to their natural counterparts. Chronosequence-based investigations of restored or created mangrove wetlands can be used to advance understanding of ecosystem development in maturing mangrove forests (McKee and Faulkner 2000, Proffitt and Devlin 2005), particularly when balanced with results from studies of ecosystem development in mangrove forest plantations (Chen et al. 2012, Salmo et al. 2013, Lunstrum and Chen 2014) and natural mangrove forest chronosequences (Lovelock et al. 2010, Kelleway et al. 2016, Walcker et al. 2018, Soper et al. 2019).

In this study, we investigated ecosystem development across a 25-year chronosequence of created, maturing mangrove forests in Tampa Bay, Florida, USA. We compared plant and soil data from created mangroves to nearby natural reference mangroves. We also quantified site-specific changes that occurred across a 5.5-year period. We investigated the following questions: (1) how quickly does peat development occur in created, maturing mangrove forests; (2) what are the trajectories of soil and plant community change following mangrove wetland creation; (3) how long does it take created mangrove forests to become equivalent to their natural reference counterparts; and (4) how does the rate of soil organic matter development beneath created, maturing mangrove forests compare to other ecosystems?

Methods

Study area

This study was conducted in natural and created mangrove forests along the eastern coast of Tampa Bay, which is the largest open-water estuary in the state of Florida, USA (27.519°–27.835°N; 82.673°–82.391°W; Appendix S1: Figure S1). Created tidal wetlands in the region are typically excavated to a target tidal elevation and planted with salt marsh grasses to stabilize substrates, create immediate structural habitat, and jump-start ecosystem development (Lewis and Dunstan 1975, Yando et al. 2019). Then, tides and currents transport mangrove propagules from adjacent mangrove forests to the created wetlands, where natural recruitment and forest development begins within the first few years after wetland creation.

Study design

Our study design incorporates two approaches. First, we used a chronosequence approach (that is, a space for time substitution) (Pickett 1989, Walker et al. 2010), which included nine created mangrove forest sites of different ages and nine natural reference mangrove forest sites. The nine created sites collectively span a 25-year period. The second approach complements the space-for-time approach with site-specific temporal comparisons, where we directly measured the plant and soil change that occurred at each of the nine created sites between 2010 to 2016 (i.e., a 5.5-year period). See Appendix S1: Figure S1 for a map of the 18 sites and Appendix S1: Table S1 for a table with site names and ages. Osland et al. (2012) contains more background information about the sites including photos and prior land use histories.

Each of the nine created and nine natural sites contained three 100-m2 plots, and each 100-m2 plot contained two 4-m2 subplots, three 1-m2 subplots, and three 0.25-m2 subplots. The three subplot sizes were each designated for a specific combination of plant and/or soil metrics, and all subplots were randomly positioned within 100-m2 plots. Plant and soil data were first collected within the nine created and nine natural sites in 2010 (Osland et al. 2012). Here, we incorporate additional plant and soil data that were collected from the nine created sites in 2016. The natural sites were not resampled in 2016 because the focus of this manuscript is on peat development in the maturing created mangrove forest sites. We assumed that the soils beneath the mature natural reference mangrove forest sites have already approached the upper limits of their soil organic matter concentrations.

Vegetation

Plant data were collected for the following three strata: herbaceous layer, juvenile mangrove tree layer, and adult mangrove tree layer. In 2010, plant data were collected from the nine natural and nine created sites. In 2016, plant data were collected from the nine created sites. For the herbaceous layer, we measured marsh percent cover within the 1-m2 subplots and Spartina alterniflora stem density (hereafter, Spartina stem density) within the 0.25-m2 subplots (Appendix S1: Table S2). For the juvenile mangrove tree layer, we measured the maximum juvenile tree height within the 4-m2 subplots (Appendix S1: Table S2). Juvenile trees were defined as those individuals with a height greater than 0.3 m but smaller than 3 m. To characterize the adult tree layer, we measured the maximum tree diameter [diameter at breast height (DBH), measured at 1.4 m height above the soil surface] within the 100-m2 plots (Appendix S1: Table S2).

Soil

Soil data were collected within the following two categories of soil samples: (1) coarse-increment soil samples, which represent the 0–10 cm and 10–30 cm depth increments; and (2) fine-increment soil samples, which represent the ten 2-cm depth increments that extend from the soil surface to a depth of 20 cm. Soil samples were collected using customized split-coring cylinders (Appendix S1: Figure S3; Osland et al. 2012) of differing diameters: 4.7-cm diameter cylinder for the coarse-increment samples and 6.9-cm diameter cylinder for the fine-increment samples. To make these coring devices, we worked with a machine shop to split each stainless steel cylinder into two semicylinders, which were then reconnected on one side via a welded piano hinge. The piano hinge allows the coring cylinders to easily open and close. During coring, a cap is placed on top of the device to keep the cylinder closed.

For the coarse-increment soil samples, we collected three soil cores to 30-cm depth from within each of the three 100-m2 plots at each site (Appendix S1: Table S3). In 2010, coarse-increment soil data were collected from the nine natural and nine created sites. In 2016, coarse-increment soil data were collected from the nine created sites. Within each 100-m2 plot, three cores were collected from within a 1-m buffer surrounding each of the three 1-m2 subplots. In the field, these three plot-specific cores were divided into two depth increments (0–10 cm and 10–30 cm) and composited to produce one coarse-increment soil sample per plot for each of the two depth increments (Appendix S1: Table S4). In 2010, an additional core was collected from each 100-m2 plot for bulk density (BD; i.e., dry bulk density) determination (Appendix S1: Tables S3 and S4) (Osland et al. 2012). In 2016, we used the composite coarse-increment samples for BD measurements. For the fine-increment soil samples, we collected one fine-increment soil core to 20-cm depth from within one of the three 100-m2 plots at each of the nine created sites (Appendix S1: Tables S3 and S4). The fine-increment cores were divided into 2-cm increments while in the field.

All soil samples were kept cool (~4°C) until processing. Soil BD was determined as a simple dry weight to volume ratio (Blake and Hartge 1986). SOM was determined via loss on ignition in a muffle furnace at 475°C for 16 h (Wang et al. 2011). Prior to SOM analyses, soils were dried, sieved (4-mm mesh), and homogenized with an analytical (2010) or planetary mill (2016).

Surface elevation change

Each of the 18 study sites has surface elevation table-marker horizon (SET-MH) stations (Cahoon et al. 2002), which were installed in January 2011 and have been used to quantify cumulative surface elevation change, shallow surface elevation change, vertical accretion, and sub-surface elevation change in the created and natural reference mangrove forests (Krauss et al. 2017). Here, we used these data to evaluate their relationships to time and the rate of soil change.

Data analyses: Ecosystem development across the chronosequence

For the chronosequence regression analyses, we incorporated data from 2010 and 2016, which produced a dataset that contained the following 18 site ages: 1.8, 3.2, 3.6, 4.8, 5.9, 7.3, 8.7, 9.1, 10.3, 11.3, 11.4, 11.8, 14.4, 16.8, 17.3, 19.6, 19.9, and 25.1 years. We used simple linear regression analyses to characterize the relationships between time and the vegetation and soil variables (Appendix S1: Table S5). To determine chronosequence-based rates of soil BD change, we used the slope of the linear relationship between time and BD. To determine chronosequence-based rates of SOM change, we calculated the first derivative of the relationship between time and SOM.

Simple linear regression analyses were also used to evaluate the relationships between the surface elevation change variables and time and SOM change (Appendix S1: Table S5). For SOM change in these analyses, we used site-specific soil data (i.e., the site-specific change that occurred between 2010 and 2016); hence, the site-specific SET-MH data were paired with the corresponding site-specific soil change data. For site age in these rate-focused regression analyses, we used the age of each site at the mid-point of the 5.5-year period separating the two sampling events.

To quantify a natural reference-based target zone for response variables, we used plant and soil data (Osland et al. 2012) and surface elevation change data (Krauss et al. 2017) that were collected from the nine natural sites. We quantified natural reference-based target zones from the natural reference mangrove forest means and standard errors. We used the natural mangrove forest means and the created mangrove regression equations to calculate times to equivalence (teq in figures), which represent the amount of time it takes created mangrove forests to become equivalent to natural reference mangrove forests. All regression analyses were conducted in Sigma Plot (Systat Software, San Jose, CA, USA). Derivative equations were determined in R (R Core Team 2017) using the Ryacas package, an R interface to the Yacas computer algebra system (Goedman et al. 2016). For all analyses, statistically significant relationships were defined as those with a P value less than 0.05. Non-significant relationships are identified in figures with the NS abbreviation.

Data analyses: Fine-increment soil data

We used the fine-increment SOM data to illustrate the site-specific fine-increment soil change that occurred between 2010 and 2016. The fine-increment data were also used to identify the relationships between time and the depth to which SOM reached and maintained concentrations greater than 5, 10, or 30%. We used linear and sigmoidal regression analyses to identify the relationships between time and soil depth for each of these three SOM levels.

Data analyses: Ecosystem comparisons

We compared the rate of SOM change determined in this study to rates determined from chronosequence studies in other ecosystems. First, we conducted a literature search and obtained data from representative salt marsh, freshwater wetland, Mediterranean forest, temperate grassland, tropical forest, temperate forest, and boreal forest ecosystems (Insam and Domsch 1988, Chapin et al. 1994, Edwards and Proffitt 2003, Banning et al. 2008, Ballantine and Schneider 2009, Mukhopadhyay et al. 2014, Čížková et al. 2018) (Appendix S1: Tables S6–S8). If an equation for soil organic matter change was provided, we used it to determine the 15-year, 20-year, and maximum rates of SOM change. If an equation was not provided, we used simple linear regression analyses to identify an equation for the data provided (Appendix S1: Table S7).

Results

Vegetation

Across the 25-year chronosequence, salt marsh grasses were replaced by growing mangrove forests (Figure 1; Appendix S1: Figure S2). Salt marsh grasses (primarily S. alterniflora) were initially planted at created wetland sites to stabilize sandy sediments and facilitate natural mangrove recruitment via tidally-transported propagules from adjacent forests. Hence, salt marshes dominated the plant community during the first 3–5 years; however, once mangrove propagules established via natural recruitment and developed into forests, salt marsh dominance declined (Appendix S1: Figure S2). In Florida, there are three common mangrove species (Rhizophora mangle, Avicennia germinans, and Laguncularia racemosa), and all three species were present and common at the created and natural mangrove sites. Across the 25-year chronosequence, there was: (1) a decrease in salt marsh cover (Appendix S1: Figure S2a); (2) a decrease in Spartina stem density (Appendix S1: Figure S2b); (3) an increase in juvenile tree height (Figure 1a); and (4) an increase in adult tree maximum diameter (dbh) (Figure 1b). The time to equivalence for marsh cover, Spartina stem density, juvenile tree height, and adult tree dbh were estimated to be 10, 12, 4, and 53 years, respectively (Figure 1 and Appendix S1: Figure S2).

Figure 1.

Figure 1.

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The relationships between time and mangrove (a) juvenile tree maximum height and (b) adult tree maximum diameter (dbh; diameter at breast height). The gray boxes depict target zones based upon data from natural reference sites. Teq represents time to equivalence.

Soil

Across the 25-year chronosequence, there was a linear decrease in soil BD (Figure 2a) and an exponential increase in SOM (Figure 2b) within the upper soil layer (0–10 cm). Within the lower soil layer (10–30 cm), the rate of change was much smaller; nevertheless, there was also a linear decrease in soil BD (Figure 2c) and an exponential increase in SOM (Figure 2d) with time. In the upper soil layer, the time to equivalence for soil BD and SOM were 24 and 23 years, respectively (Figures 2a,b). For the upper soil layer, the rate of soil BD decrease was −0.05 g cm−3 yr−1 (slope of linear relationship in Figure 2a) and the rate of SOM increase ranged from 0.2 to 6.0% yr−1 (Appendix S1: Figure S4; determined from first derivative of relationship identified in Figure 2b). Within the lower soil layer, the time to equivalence for soil BD and SOM were estimated to be 78 and 47 years, respectively (Figures 2c,d).

Figure 2.

Figure 2.

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The relationships between time and soil (a) bulk density in the upper layer, (b) organic matter in the upper layer, (c) bulk density in the lower layer, and (d) organic matter in the lower layer. Upper soil layer: 0–10 cm. Lower soil layer: 10–30 cm. The gray boxes depict target zones based upon data from natural reference sites. Teq represents time to equivalence.

The temporal analyses within individual sites provide a complementary approach for characterizing SOM accumulation. The fine-scale soil data depict, at 2-cm vertical resolution, the SOM increases that occurred between 2010 and 2016 (compare open and closed circles in Figure 3). These data show that the depth of the peat layer has increased with time (Figures 3–4 and see photos in Appendix S1: Figure S3). There were positive relationships between time and the depth of soil with organic matter levels above 5, 10, and 30% (Figure 4a–c, respectively). Within 25 years, a 20-cm deep peat layer developed (i.e., SOM greater than 30% to 20-cm depth) (Figure 4c).

Figure 3.

Figure 3.

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For nine created mangrove sites, the change in soil organic matter with soil depth in 2010 (open circles) versus 2016 (closed circles). The legends within each panel denote the age of the site during sampling. Sites are ordered according to age (top left: youngest; bottom right: oldest).

Figure 4.

Figure 4.

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The relationship between time and the depth to which soil organic matter concentrations reached and maintained concentrations greater than 5, 10, or 30% (a, b, c, respectively).

Surface elevation change

Across the 25-year chronosequence, there were no relationships between time and surface elevation change, shallow surface elevation change, vertical accretion, or sub-surface elevation change (Figure 5a–d, respectively) (Krauss et al. 2017). The time to equivalence for surface elevation change, shallow surface elevation change, vertical accretion, and sub-surface elevation change were all less than 5 years. Whereas the rates of surface elevation change, shallow surface elevation change, and sub-surface elevation change were generally higher in created mangroves than in natural mangroves (Figures 5a, b, d, respectively), the rates of vertical accretion in created mangroves were similar to natural mangroves (Figure 5c). At the created mangrove sites, surface elevation change ranged from 4.1 to 11.2 mm year−1 (mean ± se = 7.1 ± 1.0) (Figure 5a), shallow surface elevation change ranged from 4.1 to 13.5 mm year−1 (mean ± se = 7.1 ± 1.1) (Figure 5b), vertical accretion ranged from 4.3 to 7.8 mm year−1 (mean ± se = 5.9 ± 0.4) (Figure 5c), and sub-surface elevation change ranged from −1.9 to 4.9 mm year−1 (mean ± se = 1.2 ± 0.8) (Figure 5d) (Krauss et al. 2017). There were no relationships between SOM change and surface elevation change, shallow surface elevation change, or vertical accretion (Figures 5e–g, respectively). However, there was a negative relationship between SOM change and sub-surface elevation change (Figure 5h).

Figure 5.

Figure 5.

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The left panels show the relationships between time and (a) surface elevation change, (b) shallow surface elevation change, (c) vertical accretion, and (d) sub-surface elevation change. The right panels show the relationships between soil organic matter change and (e) surface elevation change, (f) shallow surface elevation change, (g) vertical accretion, and (h) sub-surface elevation change. The gray boxes depict target zones based upon data from natural reference sites. Teq represents time to equivalence. Soil elevation data are from Krauss et al. 2017.

Ecosystem comparisons

The equations for the ecosystem comparisons and the maximum, 20-year, and 15-year rates of SOM change are provided in Appendix S1: Table S8. The rate of SOM development beneath created, maturing mangrove forests (i.e., this study) was greater than all other ecosystems that we evaluated (Figure 6; Appendix S1: Table S8). Whereas the maximum rate, 20-yr rate, and 15-yr rates of SOM development beneath created, maturing mangrove forests were determined to be 6.0, 2.9, and 1.4% SOM yr−1, respectively, the corresponding values for other ecosystems were all less than 0.8% SOM yr−1, respectively (Figure 6; Appendix S1: Table S8).

Figure 6.

Figure 6.

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Comparison of soil organic matter development in mangrove forests to representative studies from other ecosystems. (a) The trajectory of soil organic matter development in a mangrove forest, freshwater wetland, and mineral salt marsh. (b) The maximum, 20-year, and 15-year rates of soil organic matter change in mangrove forests compared to other ecosystems. See Methods section and Tables S6–S8 for description and sources for the non-mangrove data. The mangrove forest data are from this study.

Discussion

Rapid peat development beneath created, maturing mangrove forests

In most freshwater peatland ecosystems, peat formation occurs slowly (i.e., centuries, millennia) (Moore and Bellamy 1974, Cameron et al. 1989); however, our results indicate that peat development beneath created, maturing mangrove forests can occur within decades. Our study sites began as blank canvases with sandy, dense soils with comparatively little organic matter (~2%). These starting soil conditions provided a valuable opportunity for characterizing the rate of soil organic matter change during mangrove ecosystem development. Across the chronosequence, the sandy soils transitioned to less dense, organic-rich soils (i.e., peats) as soil organic matter accumulated beneath the maturing mangrove forests. Within 25 years, a 20-cm deep peat layer developed. Rapid peat development in maturing mangrove forests has been observed in other mangrove forests across the world, including mangrove forest plantations (Chen et al. 2012, Salmo et al. 2013, Lunstrum and Chen 2014) and natural mangrove forest chronosequences (Lovelock et al. 2010, Kelleway et al. 2016, Walcker et al. 2018). The production and accumulation of refractory roots is the primary mechanism leading to peat development in mangrove forests, but benthic mat formation can also play a role— turf algae, microbial communities, and accumulated litter and detritus can also contribute to peat formation (McKee 2011).

How does the rate of soil organic matter increase beneath developing mangrove forests compare to other ecosystems? Our analyses indicate that the rate of peat development beneath created, maturing mangrove forests may be among the highest documented for any ecosystem globally. There are many terrestrial ecosystems that have soil organic matter concentrations that are lower than the higher annual concentration increases quantified for the maturing mangrove forests in this study. For example, our results indicate that soil organic matter concentration increases in mangrove forests can be greater than 2% per year. In contrast, soil organic matter concentrations in many ecosystems are less than 2% (Brady and Weil 2002). In addition to mangrove forests, peat development can also be rapid in other highly-productive tidal saline wetland ecosystems (e.g., salt marsh ecosystems dominated by productive grasses, sedges, and/or rushes) (Craft et al. 1999, Edwards and Proffitt 2003, Herbert et al. 2016, Abbott et al. 2019). The mangrove forests in our study area are biogenic, meaning that they receive comparatively little terrigenous sediment inputs that would dilute soil organic matter additions. The rate of soil organic matter change as a percentage has the potential to be much higher in biogenic mangrove forests (Rovai et al. 2018, Twilley et al. 2018) compared to minerogenic tidal saline wetlands that receive larger terrigenous sediment inputs [e.g., the Louisiana and North Carolina salt marsh chronosequences studied by Edwards and Proffit (2003) and Craft et al. (1999), respectively]. We expect that a biogenic salt marsh chronosequence from a similar geomorphic setting would also have high rates of soil organic matter change comparable to the biogenic mangrove forests evaluated here (Osland et al. 2018b, Radabaugh et al. 2018). However, we are not aware of a study that evaluates soil organic matter change across a chronosequence of biogenic salt marshes. Conversely, we expect that the rate of soil organic matter change in minerogenic mangrove forests would be lower than at our sites due to the dilution effect of sediment inputs.

Implications of rapid peat development for mangrove adaptation to global change

In addition to affecting important biogeochemical processes (McKee and Faulkner 2000), peat development in mangrove forests and other tidal wetlands also influences the stability of these dynamic ecosystems in the face of sea-level rise. To keep pace with rising seas, mangrove forests must adjust their vertical position in the landscape via positive soil surface elevation change (McKee et al. 2007, McKee 2011). In this study, the rates of surface elevation change within created and natural mangrove forests were positive (7.1 ± 1.0 and 3.9 ± 0.6 mm yr−1, respectively) and larger than current rates of sea-level rise in the region (2.6 mm yr−1) (Krauss et al. 2017).

For all the measured aspects of soil elevation change (i.e., net elevation change, accretion, shallow surface elevation change, sub-surface elevation change), the created mangrove forests were equivalent to or surpassed their natural mangrove forest counterparts within the first five years. However, the sub-surface layer was a zone where the created sites differed greatly from their natural counterparts. While most of the created sites had positive rates of sub-surface elevation change, the natural sites had negative rates of sub-surface elevation change. Whereas root zone expansion was a primary driver of the rapid elevation changes observed in the younger, created mangrove forests, shallow subsidence was a primary contributor to elevation change in the natural mangrove forests (Krauss et al. 2017). Shallow subsidence can be caused by compaction, root mortality, and decomposition, which are all processes that are more likely to occur in soils with high organic matter concentrations and high root biomass. Thus, we expect that the reduced rates of shallow subsidence at the created sites were associated with lower root and soil organic matter concentrations. Indeed, there was a positive relationship between soil organic matter change and sub-surface elevation change at the created sites; hence, we expect that as the rate of soil organic matter change increases in the younger, created sites, sub-surface contraction rates will also increase and ultimately be comparable to the rates measured within the natural mangrove forests. In the meantime, the rates of elevation change at the younger, created sites are very high due to rapid rates of root zone expansion.

Mangrove forests are dynamic ecosystems that are affected by many different aspects of global change. In response to moderate rates of sea-level rise, mangrove forests can adapt locally via vertical gains in elevation due to positive feedbacks between inundation, plant growth, and sedimentation (Morris et al. 2002, Krauss et al. 2014, Woodroffe et al. 2016). Mangrove forests can also adapt to sea-level rise by migrating landward at the expense of upslope and upriver ecosystems (Enwright et al. 2016, Borchert et al. 2018). In addition to sea-level rise, mangrove forests are sensitive to changing temperature and rainfall regimes (Duke et al. 2017, Lovelock et al. 2017, Osland et al. 2017, Osland et al. 2019c). In response to warming temperatures, mangrove forests are expected to expand poleward (Cavanaugh et al. 2014, Saintilan et al. 2014, Gabler et al. 2017, Osland et al. 2019a). Mangrove are also expected to expand and contract in response to fluctuations in precipitation and salinity (Eslami-Andargoli et al. 2009, Osland et al. 2018a). In response to these many different aspects of global change, some mangrove forests are expected to adapt in place and some are expected to move to new locations. We expect that the successful expansion and proliferation of mangrove forests in these new and existing locations will depend upon rapid rates of organic matter accumulation, like those observed in this study, which enable recently-established and maturing mangrove forests to modify their abiotic environment as they gain or lose elevation to attain the ideal vertical position relative to sea level (McKee 2011).

Ecological trajectories and functional equivalency

Restoration chronosequences have frequently been used to gauge functional equivalency and characterize post-restoration ecological trajectories (Zedler and Callaway 1999, Baer et al. 2002, Walker et al. 2010, Abbott et al. 2019). In many restored wetland ecosystems, vegetation development proceeds at a pace that is faster than soil development (Herbert et al. 2016, Richardson et al. 2016). Although our data indicate that it takes approximately 55 years for the adult mangrove trees at created wetland sites to become equivalent to their natural mangrove forest counterparts, equivalency in the herbaceous and juvenile vegetation strata can occur in less than 15 years. Our belowground results indicate that equivalency in the upper soil layer of mangrove forests can occur within approximately 25 years, which is rapid in comparison to many restored terrestrial and freshwater wetland ecosystems (Ballantine and Schneider 2009, Hossler and Bouchard 2010, Moreno-Mateos et al. 2012) but comparable to results from some salt marshes (Craft et al. 1999, Craft et al. 2003, Edwards and Proffitt 2003).

Soil development in the lower soil layers of restored and created wetlands often takes more time than in the surficial layers (Ballantine and Schneider 2009, Herbert et al. 2016, Richardson et al. 2016, Abbott et al. 2019). Indeed, in an earlier study at these same sites, we found no significant soil development in the lower soil layer (i.e., the 10–30 cm depth increment) (Osland et al. 2012). However, the data presented here — from a longer, 25-year chronosequence as well as from finer vertical resolution (i.e., 2-cm increment) analyses — show that soil development has extended into the lower layer. These more recent findings indicate that time to equivalency in the lower soil layer may be 2–3 times longer than the time for the upper soil layer (~45–80 years versus ~25 years for the lower and upper layers, respectively).

One key limitation of our study relates to the extrapolation required for a subset of the time to equivalency estimates. For three ecosystem properties (i.e., lower soil layer bulk density, lower soil layer organic matter, and the adult tree layer), we used relationships observed during the first 25 years of ecosystem development to estimate changes that will occur beyond 25 years. However, the trajectory of future change may be different, which would affect our estimate of the time it will take for these ecosystem properties to reach equivalency. For example, given the non-linear rate of organic matter change in the upper soil layer, it is likely that soil development in the lower layer will accelerate in the coming decade, meaning that equivalency in the lower soil layer could be on the faster end of our extrapolated estimate (i.e., closer to the 45-year side of the 45–80 year range). See Appendix S1: Figure S5 for an illustration and description of one potential hypothesis regarding the longer-term, non-linear relationships between time and soil organic matter change in the upper and lower soil layers. Incorporating older sites into our analyses would help refine these calculations.

Conclusions

Collectively, our analyses quantify the pace of ecosystem development and functional equivalency in created, maturing mangrove forests, which is information that can inform predictions of mangrove forest responses to global change and ecosystem restoration. Coastal wetland managers can use these data to estimate the amount of time needed for created mangrove wetlands to reach functional equivalency to their natural counterparts. Rapid peat development, primarily in the form of sub-surface root production and accumulation, enables mangrove forests to modify their abiotic environment, sequester carbon from the atmosphere, change their soil elevation relative to sea level, and adjust to dynamic abiotic conditions at the land-ocean interface. In the face of climate change and accelerated sea-level rise, coastal managers are increasingly concerned with the longevity and functional equivalency of coastal wetland conservation and restoration efforts. The rapid pace of ecosystem development documented here indicates that, if given the opportunity, created and restored mangrove forests have a high capacity to adapt to future change.

Supplementary Material

Sup1

Acknowledgments

This research was supported by the USGS Land Change Science Climate R&D Program, USGS Ecosystems Mission Area, USGS LandCarbon Program, US EPA Gulf Ecology Division, and the US EPA Office of Research and Development Post-Doctoral Research Program. The views expressed in this paper are those of the authors and the U.S. Geological Survey but do not necessarily reflect the views or policies of the U.S. Environmental Protection Agency. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government.

Footnotes

Data Availability

Data are available (Osland et al. 2019b) on USGS ScienceBase at: https://doi.org/10.5066/P9CW5VUC

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