CHAPTER TWO

Coastal Blue Carbon

INTRODUCTION

Overview of Coastal Blue Carbon—Definition and Motivation

Coastal carbon sequestration, in this report, refers to carbon dioxide (CO₂) removal from the atmosphere in conjunction with plant growth and the accumulation and burial of plant organic carbon (OC) residue in the soil of tidal wetland and seagrass ecosystems. Tidal wetlands, including salt marshes and mangroves, thrive in soft-sediment, shallow regions of estuaries between high and mean sea level, while seagrasses inhabit adjacent soft-sediment estuarine bottoms with adequate light penetration. Macroalgal systems such as kelp forests, while extremely productive, do not have root systems and soils to accumulate carbon (Howard et al., 2017). Little is known about the fate of macroalgal OC, a portion of which could potentially be sequestered by other means when exported from the macroalgal bed (Krause-Jensen et al., 2018). Further understanding of macroalgal transport processes or other farming approaches that could permanently remove CO₂ from the atmosphere is needed to assess the potential for globally significant levels of CO₂ sequestration. Given the state of research on this topic, which is summarized in Appendix C, the CO₂ removal capacity of macroalgal storage is not a focus of this chapter. Other ocean-based negative emissions technologies (NETs), such as sequestration in microbial or planktonic biomass (Zhang et al., 2017b) and ocean alkalization (Rau et al., 2013; Renforth and Henderson, 2017), may be conducted in the coastal environment but are essentially open ocean approaches. They remove carbon through very different mechanisms from coastal wetlands, which bear closer similarity to terrestrial-based NETs. As described in Chapter 1, these approaches are not included in this report and warrant further investigation because of the potentially large role that ocean sinks can play in CO₂ removal.

Tidal wetlands and seagrasses are among the most productive regions on Earth, sequestering CO₂ at a rate of 7.98 t/(ha y) CO₂ for tidal wetlands and 1.58 t/(ha y) CO₂ for seagrass meadows (EPA, 2017; IPCC, 2014a). Scaled to their current global areal extent, they are an important component of the global carbon cycle. Tidal wetlands grew to their current areal extent only in the past 4,000-6,000 years, once the rate of sea-level rise (SLR) declined to less than 2 millimeters per year. The OC that has

accumulated over this time is deep and vast, ranging from 2 to 25 Gt C (best-estimate of 7 Gt C; Bauer et al., 2013; Donato et al., 2011; Fourqurean et al., 2012; Pendleton et al., 2012; Regnier et al., 2013). The recent State of the Carbon Cycle Report (SOCCR-2) estimates that tidal marsh wetland soils and estuarine sediments of North America store roughly 1.323 Gt C as the top 1 m of soil and sediment. However, coastal soil profiles are known to reach greater depths, and thus carbon storage is higher than this estimate (Sanderman et al., 2018). Variation in soil and sediment depths have become an increasingly recognized source of uncertainty in tidal wetland coastal carbon stocks, in addition to other sources of uncertainty (Holmquist et al., 2018). Although seagrasses have lower OC burial rates than tidal wetlands per unit area, they potentially cover a much larger area and thus could have higher total carbon rates and sequestration capacity. However, substantial difficulties in large-scale estimation persist because there is large variability in measurements of seagrass OC stocks and burial rates. The mapping of seagrass areas in U.S. water is also limited; less than 60 percent of meadows are mapped, and existing maps have varying degrees of accuracy because of difficulties in remote sensing of underwater habitat (Oreska et al., 2018).

Annual rates of coastal carbon sequestration are also high. Globally, the total carbon sequestration rates are estimated at 31-34 Mt/y C for mangrove, 5-87 Mt/y C for salt marshes, and 48-112 Mt/y C for seagrass beds, or 226 ± 39 g/(m² y)C for mangrove, 218 ± 24 g/(m² y)C for salt marshes, 138 ± 38 g/(m² y)C for seagrasses, summing up to a global annual rate of 0.84 Gt/y CO₂ (Mcleod et al., 2011). According to the U.S. Inventory of Greenhouse Gas Emissions and Sinks, U.S. tidal wetlands (marshes and mangroves) sequester more than12 Mt/y CO₂ and a net of about 8 Mt/y CO₂e when both CO₂ and CH₄ (methane) fluxes are considered (EPA, 2017). The CH₄ fluxes introduce significant uncertainty in large-scale estimates because of difficulties with detecting salinity boundaries that determine CH₄ flux rates (Poffenbarger et al., 2011). For North America, soil OC accretion rate (sediment burial) was 5 ± 2 Mt/y for tidal marshes, 2 ± 1 Mt/y for mangroves and 1 ± 1 Mt/y for seagrass meadows (30 Mt CO₂ in total; SOCCR-2). The estimate produced by the U.S. Environmental Protection Agency (EPA) was only for the continental United States and included fewer wetland areas and tidal wetland types than the North American SOCCR-2 report. The EPA report also did not include seagrass meadows. Tidal wetlands, particularly mangroves, also sequester CO₂ in aboveground biomass via long-term storage in wood and woody stems. Although this contributes to negative CO₂ emissions, this CO₂ removal flux is omitted from estimates presented in this report because most recent U.S. estimates focus on soils only as it is the proportionally larger sink for tidal marsh ecosystems that represent the largest area of coastal carbon ecosystems in the United States.

The motivation for including coastal blue carbon as a potential NET is the potential to more than double the current rate of CO₂ removal through several approaches that restore and create coastal wetlands. Further, there is concern that the current rate of sequestration will drop substantially because of expected changes in factors that contribute to sequestration, especially those that affect the current areal extent and carbon burial rate per unit area. Sequestration is vulnerable to impacts from climate change, including increasing rates of SLR (Figure 2.1) and rising temperatures, and human activities in the coastal zone. Although conversion of U.S. coastal wetlands has slowed, it is estimated that global drainage and excavation of mangrove, tidal marsh, and seagrass soils release 450 million tons of CO₂ annually (range 150-1,005 Mt/y CO₂; Pendleton et al., 2012). In addition, declining sediment supplies and groundwater/oil and gas extraction pose indirect threats to wetlands (Megonigal et al., 2016). Reversing historic loss and degradation through restoration, incorporating wetland creation into coastal adaptation projects, and managing wetland area and carbon

accumulation rates provide an opportunity for increased carbon removal and storage through the 21st century.

Coastal wetlands and seagrasses are already the targets of restoration and management for the broad range of ecosystem services they provide beyond CO₂ removal, including coastal storm protection and wave attenuation, water quality improvement, wildlife habitat, and support of fisheries (Alongi, 2011; Barbier et al., 2011; Lee et al., 2014; Nagelkerken et al., 2008; Zhang et al., 2012). These activities and investments, which are not included in this report as NET costs, can be leveraged to provide CO₂ removal advantages at marginal costs.

COASTAL BLUE CARBON PROCESSES

Coastal carbon sequestration is calculated for each ecosystem as the product of its areal extent (horizontal dimension) and its vertical OC accumulation rate (vertical dimension) (Hopkinson et al., 2012). During the past 4,000-6,000 years, the rate of SLR slowed enough to allow tidal wetlands to maintain or expand their areal extent and their vertical elevation relative to SLR (Figure 2.2). The survival of existing tidal wetlands requires that vertical elevation gains at least match the rate of SLR. Tidal wetlands form with salt-tolerant plants in soft-sediment intertidal regions above mean sea level (MSL). Once established, the elevation of wetlands increases relative to sea level (SL) through the accretion of particles trapped by wetland vegetation from tidal waters (watershed, oceanic, or local sources) and the accumulation of undecomposed wetland plant organic matter. The rate of accretion and carbon burial varies hyperbolically with flooding frequency and depth and thus the rate of SLR (see Figure 2.3; Morris, 2016). With adequate sediment supply relative to the rate of SLR and local subsidence (or rise), sedimentation leads to the shallowing of estuaries. Tidal wetlands prograde into formerly open water regions, once depths reach MSL, thereby increasing their horizontal extent. Tidal wetlands will also expand horizontally through transgression when rising seas flood adjacent upland areas and wetland plants invade.

In seagrass meadows, lateral expansion occurs through asexual clonal growth. Sea-grasses also disperse seeds to colonize new areas. Within seagrass meadows, OC is accreted vertically when wave attenuation from the plant canopy causes sediment to settle from the water column and bury plant detritus as well as allochthonous carbon. About one-half of the carbon buried in seagrass beds is produced by non-seagrass sources (Oreska et al., 2018). Carbon burial varies spatially within a meadow; wave attenuation is more effective in inner parts of the meadow, with greater erosion occurring at the edges (Oreska et al., 2017). Seagrass bed productivity is controlled by

different factors including nutrient and light availability (Apostolaki et al., 2011; Hendriks et al., 2017). Light availability is controlled by a variety of factors including water depth and turbidity. Turbidity is related to local geomorphic drivers as well as eutrophication (which controls phytoplankton density) and growth of epiphytes on leaves (a factor contributing to reduced photosynthesis of submerged macrophytes).

Critical to coastal carbon sequestration is the rate of vertical elevation gain and specifically the relative contribution of undecomposed plant OC to mineral sediments. In tidal wetlands, most of the OC accumulating is that produced in situ (autochthonous), while in seagrass meadows trapping of OC from external sources (allochthonous) can be important as well. The fate of wetland plant biomass production varies tremendously from system to system depending on the rates of net primary production (NPP), respiration of microbes and larger animals living on and in tidal wetlands (Re) (i.e., decomposition and consumption of OC), and the tidal export of undecomposed plant material (Hopkinson, 1988), which is mostly above-ground plant material. The balance between NPP, Re, and export is the organic matter that is buried and preserved. The continued surface deposition acts to increase over time the depth of OC within the sediment, where conditions reduce organic matter decomposition, thus preserving it for longer and longer periods of time (Redfield, 1972).

The dynamic exchange and export of OC to adjacent systems is a unique characteristic of tidal and estuarine systems compared with terrestrial ecosystems. Carbon that is exported laterally may exit the coastal ocean through CO₂ outgassing or dissolved OC and particle carbon export to the open ocean, with carbon import across these interfaces also possible. The balance of these processes results in annual OC storage at higher rates than long-term OC sequestration (Breithaupt et al., 2012). To simplify the Committee’s treatment of negative carbon emissions in this study, the focus in this chapter is on long-term (50-100 years), buried soil carbon.

COASTAL BLUE CARBON IN THE FUTURE—THE IMPACTS OF CHANGING ECOSYSTEM DRIVERS

The baseline carbon sequestration capacity for coastal wetlands and seagrass meadows is the predicted changes in areal extent and OC burial rates of these ecosystems in the absence of human intervention. Without any human intervention, this baseline is possibly decreasing over time, as compared with the current burial rate, under the stress of climate change and human activity. It is upon this expected baseline that CO₂ removal approaches to increase future CO₂ removal trajectories can be evaluated. Extrapolating past and current rates of CO₂ removal to predict future rates of

CO₂ removal likely will not provide accurate estimates, because drivers are changing rapidly in many coastal areas. The heart of a research agenda is to fill knowledge gaps on the response to these changing drivers, and to constrain uncertainties in coastal blue carbon to better predict and manage future trajectories and accelerate new opportunities for CO₂ removal.

The drivers of concern are those most likely to change as a result of climate change or other anthropogenic impacts over the next hundred years, including

• relative SLR (which affects extent, depth, and duration of tidal flooding),
• sediment availability (from watershed inputs, tidal flooding and/or storms),
• temperature (and growing season length),
• light availability,
• salinity (related to river flow, local climate, and sea level),
• inorganic nitrogen and phosphorous availability and enrichment, and
• development of wetland area or uplands adjacent to wetlands.

Partially or fully driven by the above abiotic factors, biotic factors include:

• plant productivity and species composition,
• plant migration rates, and
• organic matter decomposition rates.

Expansion of tidal wetlands into open water areas of estuaries has slowed and, in some places, has reversed during the past 100 years or so, primarily because of decreased sediment input due to watershed management activities compounded by increased rates of relative SLR. Increased rates of SLR have also increased transgression (the expansion of wetlands into terrestrial uplands). SLR interacts with anthropogenic stressors and can result in accelerated erosion and subsidence in certain conditions. Several models have been developed to provide a predictive understanding of both the areal extent and OC burial components of tidal wetlands (French et al., 2008; Kirwan et al., 2010, 2016b; Morris et al., 2002; Morris, 2016; Mudd, 2011). Models show that a simple bell-shaped relationship between elevation and productivity characterizes tidal biomass production and a linear elevation–decomposition relationship characterizes belowground organic carbon degradation (Figure 2.3). Current models suggest the existence of a threshold level of sea level where the vertical elevation and lateral migration of tidal wetlands may not keep pace with the water level, resulting in wetland drowning and a sudden decrease in OC burial (Figure 2.4). In addition to driving the vertical response, SLR can drive the migration of tidal wetlands inland (transgression) and either into open water (progradation) or into interior wetlands (redeposition of eroded carbon) (Figure 2.5). Without adequate sediment supply to maintain

a critical depth of tidal flats adjacent to tidal wetlands, erosion of the edge of existing wetlands can result in decreased areal extent (Mariotti and Fagherazzi, 2010, 2013; Mariotti and Carr, 2014). Thus, the future expanse of tidal wetlands will reflect the balance between positive or negative progradation/erosion and upland transgression.

Despite the value of these predictive models, knowledge gaps remain that affect the ability to predict future CO₂ removal capacity. One key gap is the fate of eroded OC. When a wetland erodes or is drowned, the fate (i.e., whether it will decompose and contribute to carbon emissions, be deposited and buried long term, or be redeposited on the marsh platform) depends on geomorphological processes controlling erosion, deposition, and resuspension (Hopkinson et al., 2018). Better understanding of what could happen to eroded carbon will require better understanding of OC preservation and its refractory nature and transport depositional fate. In addition, although there is apparent agreement between current models of tidal wetland OC burial and selected field observations, reliance on old paradigms of OC preservation in soils (Lehmann and Kleber, 2015; Schmidt et al., 2011) and on microcosm results as evidence for a hyperbolic response of tidal wetland platform vegetation to flooding do not lend confidence to predictions of marsh survival and OC burial if sea level increases by 1-2 m by century end (Morris, 2016). Regional and local drivers and recent changes, including human impacts, can limit the broad predictive ability of these models. Nevertheless, experimental manipulations coupled with hierarchical approaches to scaling, and better integration of field-validated remote sensing, have greatly improved integration of plot-based drivers of OC accumulation rates and landscape-scale estimation of coastal wetland CO₂ removal (Byrd et al., 2018; EPA, 2017; Holmquist et al., 2018).

The future CO₂ removal capacity of tidal wetlands depends on their ability to transgress into upland areas as SLR increases. “Coastal squeeze” and vegetation shifts may reduce the lateral space for wetland transgression into uplands. Coastal squeeze occurs when there is no more lateral space for upland migration (DOE, 2017a; Doody, 2004). The decline in available lateral space results from upland barriers, when uplands are occupied by other land uses (e.g., agriculture, urban lands) or when the slope does not support the migration of wetlands into upland areas (Doody, 2004). Prediction of the trajectories of available lands is another key knowledge gap, including what factors may reduce barriers to wetland transgression into uplands that are occupied by other land uses. Vegetation shifts associated with wetland transgression and change in carbon uptake capacity, as in shifts to woody species or loss of vegetated wetland from inland subsidence due to marsh dieback, can result in overall changes in carbon burial rates.

Warming of the air and sea water will impact the coastal carbon cycling and sequestration driven by complex interaction of plants, microbes, and physical processes (Megonigal et al., 2016). Both primary production and OC decomposition could increase with warming. While theory and mesocosm studies of estuarine water show that warming will differentially increase Re relative to NPP, which would thus decrease net ecosystem productivity (NEP) and the potential OC available for either burial or export, data are insufficient to extrapolate these results to tidal wetland and seagrass ecosystems (Yvon-Durocher et al., 2010). Warming could also drive the displacement of salt marsh by mangrove (Megonigal et al., 2016). The final carbon burial rate is determined by the rate and sensitivity of these processes to warming.

APPROACHES FOR COASTAL BLUE CARBON

The overall goal for a research agenda on coastal blue carbon is to be able to quantitatively evaluate the enhanced OC burial for a variety of management and engineering approaches under a changing suite of social barriers, human activities, and climate scenarios. These approaches build on our current understanding of the baseline of annual carbon burial rates and estimation of incremental carbon burial induced from these projects. The knowledge gaps for each of these approaches relate to both research and technology. The committee focuses on five approaches by which coastal blue carbon can be accelerated through 2060 and potentially for the remainder of the century. They vary in terms of potential cost, degree of human intervention required, technological readiness, and social barriers (likelihood of enactment).

1. Actively manage coastal wetlands and seagrass meadows to increase CO₂ removal against the decreasing baseline of carbon burial rates.
2. Restore coastal wetlands and seagrass meadows where they have been degraded or lost.
3. Convert hardened and eroding shorelines to natural shorelines consisting of wetland area as part of coastal adaptation.
4. Manage wetland transgression into uplands with change in SLR and human drivers/impacts.
5. Increase the carbon storage capacity of coastal wetlands and shorelines by augmentation with carbon-rich materials.

Active Ecosystem Management

As a result of increasing rates of SLR and human disturbance, the baseline carbon burial rate is declining over time (Figure 2.4), meaning the current capacity of natural

negative emissions in coastal wetlands is shrinking. With active management, this trend could be reversed and the carbon burial rate could become equal to the current rate or increase over time. The field has reasonable estimates of OC burial in coastal macrophyte systems in the United States, albeit there are large uncertainties in current areal extent of seagrasses and medium confidence in their appropriate OC burial rates.

Key to maintaining existing areas of natural tidal wetlands and seagrass meadows is management practices to reduce the impacts of human drivers that cause coastal change. As described earlier, progress has been made in developing models to predict the future areal extent and OC burial rate for extant wetlands of the United States. Some of the largest uncertainties are the (1) interplay between sediment availability and OC accumulation, (2) effects of climate drivers and SLR on NEP and OC burial, (3) factors controlling edge erosion and the importance of released sediment on wetland platform survival, and (4) effects of other human activities, such as pollutant runoff, on NEP and OC burial. A Habitat Evolution Model that incorporated future SLR scenarios applied to the Tampa Bay Estuary indicated that, if managed, coastal habitats would remove about 74 Mt CO₂ from the atmosphere by 2100 (ESA, 2016). That model can be used to prioritize coastal wetland areas for active management.

Given the projected decline in the baseline of carbon burial, any management practices and projects that can reverse the trend is a coastal NET. For example, enhancing coastal nitrogen management by building sewage treatment systems could significantly reduce nitrogen leaching to salt marshes, enhance rooting depth and marsh productivity (Deegan et al., 2012). Sediments are also required to maintain elevation of tidal flats proximal to wetlands to prevent shoreline erosion (Bilkovic et al., 2017; Fagherazzi et al., 2012). Edge erosion may be controlled by either direct (e.g., dredging) or indirect (e.g., river diversions or removal of dams on rivers) addition of sediments. It can also be controlled directly by living shorelines or breakwaters immediately adjacent to wetland shorelines. The potential problem with preventing shoreline erosion via breakwaters is that it prevents the liberation of sediments internal to the estuarine system, which may be a critical source of sediment contributing to elevation gain of wetlands interior to their edge. Management strategies to conserve carbon at the edge of eroding shorelines or creating breakwaters that protect the CO₂ sequestration of inland coastal marshes before barrier islands degrade could maintain the negative CO₂ emissions achieved by these coastal wetlands (Bilkovic et al., 2017). Some states are protecting shorelines from erosion as an active management effort with associated policy and a regulatory framework (Bridges et al., 2015). Further, protecting inland coastal marshes not only sustains areas for OC burial but also protects significant stores of peat. Initial saltwater intrusion may influence CH₄ emissions (Neubauer et al.,

2013), while old peat deposits may be vulnerable to decomposition upon salinization (Wilson et al., 2018). Breakwater-type living shoreline projects can slow edge erosion, but their contribution to the sediment needed for platform elevation gain is uncertain, because edge erosion may be an important contributor of sediment to maintain elevation of the remaining marsh platform. For example, at the Plum Island Ecosystems LTER, marsh erosion liberates sufficient sediment to meet almost 30 percent of the annual rate of sediment accumulation, with rivers only contributing 9 percent (Hopkinson et al. 2018).

Restoration of Lost or Degraded Coastal Wetlands and Wetland Creation

Wetlands are drained, excavated, and tidally-restricted as a result of human activities that reduce their area or their capacity to sequester CO₂ and confer other ecosystem services (Kroeger et al., 2017a). The goal of restoration is to return or improve wetland functions and provision of ecosystem services. OC sequestration accompanies ongoing restoration activities (Kroeger et al., 2017a), but often is not the primary or even secondary objective. However, reversing the effects of anthropogenic activities on coastal wetlands to both reduce the greenhouse gas (GHG) burden on the atmosphere and reinitiate processes that promote CO₂ removal has been taken up as part of national and international policy actions—for example, recent International Governmental Panel on Climate Change GHG Inventory Guidance (IPCC, 2014a) and a new component of EPA’s National Greenhouse Gas Inventory (EPA, 2017)]. Tidal marsh restoration sites in the Snohomish Estuary have measured annual accumulation rates between 0.9 t/(ha y) C to 3.52 t/(ha y) C (3.3 t CO₂ to 12.9 t CO₂), depending on characteristics of the sites including project age and elevation (Crooks et al., 2014). Restoration sites in the Tampa Bay Estuary are estimated to have accumulated 217,000 t CO₂e¹ between 2006 and 2016 (ESA, 2016).

While restoration of tidal wetlands and seagrass meadows is occurring, the potential to increase the areas for restoration is significant (EPA, 2017). In the United States, approximately 1.3 million hectares (ha) of tidal wetlands and seagrasses have been converted to other land uses or otherwise lost and currently potentially available for restoration. Notably, the different land uses for these former coastal wetlands may pose significant barriers to their restoration. These land uses within the coastal boundary (upper limit mean higher high water [MHHW]) are characterized by the National Oceanic and Atmospheric Administration’s (NOAA’s) Coastal Change Analysis Program

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¹ For any concentration and type of greenhouse gas (e.g. methane, perfluorocarbons, and nitrous oxide) CO₂e signifies the concentration of CO₂ which would have the same amount of radiative forcing.

(C-CAP) as developed (low to medium-high intensity); cultivated (including pasture/hay and grassland), including some portion that are tidally restricted (Kroeger et al., 2017a); and open water/unconsolidated shore (and possibly older eroded areas) because of erosion of tidal wetlands or loss of seagrass meadows (Waycott et al., 2009).

Within developed and cultivated lands in the coastal zone, restoration of hydrology, either through reversing drainage or removing tidal restrictions, is likely the most important activity to restore wetlands. Restoring saltwater flow to a tidally restricted wetland allows for both a reduction in methane emissions and a reconnection to the SLR processes that promote soil accumulation (Kroeger et al., 2017b). In submerged areas where elevation must be increased, two sediment management activities can be applied: (1) direct addition to the marsh surface (e.g., through thin layer deposition) and (2) indirect addition to the estuary and then tidal current conveyance to the wetland surface (e.g., river diversions and dam removal along rivers). OC burial can restart once intertidal elevations are reached and wetland vegetation is established (Osland et al., 2014). For example, reversing coastal wetland loss to subsidence and SLR is the primary focus of the Louisiana Master Plan (CPRA, 2017). Dredged material has proven to be a valuable sediment source for wetland creation—and hence raising sequestration above a baseline. Approximately 194 million cubic yards of sediment are dredged annually from the nation’s navigable waters (USACE, 2015). Mississippi River diversions are of a smaller scale and may be more influential in reversing wetland loss. However, significant uncertainty remains as to whether application of these approaches in all subsided or eroded coastal wetlands will achieve similar results given the interacting effects of other drivers of coastal change.

Only a fraction of coastal water has been surveyed for seagrass. Globally, the seagrass meadow is estimated to cover 30-60 million ha (Duarte et al., 2005; Fourqurean et al., 2012; Kennedy et al., 2010; Mcleod et al., 2011). Based on current extent and fraction of reported loss, the current area of U.S. seagrass meadows is estimated at 0.6 million ha (Waycott et al., 2009). For seagrass meadows, watershed management to improve water quality and clarity, control of sediment loads, and resuspension and replanting of degraded seagrass cover are well proven technologies to increase areal extent, productivity, and OC burial.

Although not widely applied in the United States, there has been a strong push to develop carbon markets for wetlands, particularly tidal wetlands. Despite the absence of a market, a methodology for crediting tidal wetland and seagrass restoration with carbon sequestration has been developed (Verified Carbon Standard VM0033).

Conversion of Hardened and Eroding Shorelines to Natural Shorelines

The conversion of hardened and eroding shorelines to natural and nature-based shorelines that can keep pace with SLR is a growing environmental risk reduction strategy (van Wesenbeeck et al., 2014). This strategy also serves as an effective approach to CO₂ removal when projects enhance wetland area or performance (Bilkovic and Mitchell, 2017; Bridges et al., 2015; Davis et al., 2015; Saleh and Weinstein, 2016). Risks from flooding and storm surge are increasing and will impact coastal inhabitants in the United States and globally. Flooding of upland areas due to SLR is expected to place between 2.2 and 13.1 million people at risk, depending on SLR and population projections (Hauer et al., 2016). Because the estimated cost for relocation is $1 million per resident (Huntington et al., 2012), extensive adaptation for coastal risk reduction is anticipated (Brody et al., 2007). Failure and expensive maintenance of infrastructure that extends to coastal areas portend a significant increase in new nature-based, coastal infrastructure as a means to avert risk and reduce the cost of coastal adaptation. Further, because shoreline armoring is prohibited or significantly restricted in several U.S. states (O’Connell, 2010), nature-based, living shoreline approaches will likely be more frequently used.

There is considerable activity around the world to employ natural and nature-based features (NBBF), which can enhance the CO₂ removal value of coastal adaptation projects (Bridges et al., 2015). Nature-based features mimic characteristics of natural features, but are created by human design, engineering, and construction to provide specific services such as coastal risk reduction. The built components of the system include nature-based and other structures that support a range of objectives, including erosion control and storm risk reduction (e.g., seawalls, levees), as well as infrastructure that provides economic and social functions (e.g., navigation channels, ports, harbors, residential housing). Case studies of demonstrated projects in the United States, Europe, Mexico, and China are emerging from this rapidly growing area of research (see Bilkovic et al., 2017; Bridges et al., 2015; Saleh and Weinstein, 2016; and Zanuttigh and Nicholls, 2015 for reviews). About 14 percent (22,000 km) of the U.S. shorelines have already been armored (Gittman et al., 2015) of the shoreline armored. Converting these armored shorelines to natural shorelines with plants, sediments, and tidal flooding features will provide significant benefits as a NET. However, although implementation of nature-based approaches to tidal wetland creation has reached the deployment level, the estimates for CO₂ removal are based primarily on assumptions that these approaches attain OC accumulation rates similar to those for natural or restored wetlands.

Management of Wetland Transgression into Uplands

Wetland transgression allows the area of tidal wetlands to expand as sea levels rise and increase annual CO₂ removal capacity, especially if erosion of existing tidal wetlands can also be prevented. Coastal mapping and SLR projections of inundation enables identification of potential opportunities for transgression. Schuerch et al., 2018 modeled that wetland gains of up to 60 percent would be possible if transgression can occur in just 37 percent of wetland areas globally, compared to an expected loss of up to 30 percent due to SLR without transgression. However, knowledge is more limited with regard to how current land cover, ownership, and economic value of upland areas will influence the practical potential for transgression. For example, commercially valuable upland is being armored once the flooding risk becomes apparent. Haer et al., 2013 report that inundated land area will increase between 2.6 and 7.6 million ha by 2100, depending on SLR projection and extrapolation method. Assuming 1ft of vertical accretion by 2100, 2, 4, and 6 ft of SLR above MHHW correspond with 1.5, 0.87, and 0.93 million ha, respectively, of net upland (excluding developed land) inundated (NOAA SLR viewer²). Managed shoreline retreat strategies that discourage development and decrease population in areas with increasing flood risks and that allow them to flood with coastal waters (Kousky, 2014), not only reduce coastal risks but also increase land area for CO₂ removal. Barriers to managed transgression for wetlands into uplands may vary by upland land-use type. Approximately 43 percent of upland land area between MHHWS (Mean Higher High Water Spring) and MHHWS +2ft is cultivated, with another 20 percent as pasture/hay and grassland (NOAA Office for Coastal Management, 2018). Information about change in areas of inundated developed areas was not available. However, because economic costs of inaction increase over time (Hauer et al., 2016; Reed et al., 2016), coastal adaptation policies for coastal risk reduction could foster a predictable trend for not only identifying additional areas for inland migration and capacity for its management, but also additional areas for other NETs. In other words, similar socioeconomic issues may exist when converting other coastal land uses to coastal wetlands (i.e., restoration).

Management strategies would be needed to maximize the potential expansion of tidal wetlands into uplands. Management of wetland transgression could also include policy measures that both reduce flood risk (Brody et al., 2007) and increase CO₂ removal capacity. However, we lack typologies and rigorous methodologies for predicting where transgression will occur, how existing land uses will allow it, and the cost to protect some potentially large expanses of low-lying coast for wetland expansion. The field lacks knowledge of the OC burial rate trajectory associated with

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² See https://coast.noaa.gov/digitalcoast/tools/slr.html.

upland land transgression to wetlands. Upland soils can begin to bury OC at greater rates but productivity of upland species may decline while wetland species invade, with a period of net lower productivity and lower OC burial rates than “natural” coastal wetlands.

Augmentation of Shorelines with Carbon-Rich Materials

Augmenting coastal projects with carbon-rich materials can enhance carbon storage by burying externally-produced carbon-rich materials (e.g., burying wood or biochar) and can increase rates of OC burial by bioengineering wetland species toward lower rates of decomposition (e.g., manipulating lignin content). These approaches could occur in tandem with restoration, coastal adaptation and shoreline protection, and management of wetland transgression.

Direct addition of slow-decomposition forms of OC such as logs and biochar can augment carbon burial. Several studies have evaluated wood burial as a means to increase CO₂ sequestration capacity (Freeman et al., 2012) and biochar addition as a means to reduce nitrogen mineralization of peat and coastal wetlands (Luo et al., 2016; Zheng et al., 2018). The scientific literature also supports the use of carbon-rich materials to increase the CO₂ removal potential of peatlands by “injecting” timber (Freeman et al., 2012). Similarly, decay-resistant conditions of tidal wetlands and seagrass meadows could be harnessed, but further study is needed to evaluate decay rates of different materials and demonstration projects are needed to achieve CO₂ removal at scale. Evidence indicates that the waterlogged wood of ships that has been buried in sediment is often in a very good state of in-situ preservation in marine environments (Gregory et al., 2012). Macchioni et al., 2016 found that all samples from wood foundations in Venice, Italy, showed at least 30 percent residual bulk density, with the earliest known construction in 1854. The state of wood preservation is related to several factors, including thickness of the element, depth of burial, horizontal/vertical position, and wood species (Macchioni et al., 2016). A 3-year wood degradation study showed that wood buried below 43 cm was highly protected from decay, with erosion bacteria promoting the loss of surficial wood (<0.5 mm; Bjordal and Nilsson, 2008). Wood burial is only one nature-based example. Constructing revetments or breakwaters with concrete composed of carbon-rich water, aggregate, or embedded wood is another example of a more engineered approach.

Afforestation management strategies for planting mangroves in marsh areas are also under consideration. SLR and warming are leading to increased expansion of mangroves in some areas, suggesting that this or other tree species could be introduced

in other areas of transgressing wetlands. To avoid double-counting afforestation gains in wetlands with terrestrial areas, accounting methods based on new remote sensing and field validation research must be developed, which has been identified as an area of improvement for future EPA GHG inventory compilation (EPA, 2017).

Genetically engineering wetland macrophytes to increase their lignin content is another option to enhance wetland OC burial. There has been considerable interest in degrading lignin and altering the lignin content of vegetation to improve the efficiency of cellulosic biofuels production (Wei et al., 2001). The interest for biofuels is to decrease lignin content to increase accessibility of plant polysaccharides to microbial and enzymatic digestion (e.g., Ragauskas et al., 2006), but the interest for coastal blue carbon is to increase lignin content of roots and rhizomes and reduce OC decomposition. Biofuels researchers have identified genes that encode the enzymes leading to the building blocks of lignin (Hoffmann et al., 2003), and have achieved downregulation of some of these genes and hence lignin biosynthesis (Chen and Dixon, 2007; O’Connell et al., 2002; Reddy et al., 2005). The predominantly anoxic conditions in coastal peat sediments allow an accumulation of phenolic compounds from lignin, inhibiting decay by suppressing phenol oxidase enzyme activity and microbial enzymatic decomposition of senescent vegetation (Appel, 1993; McLatchey and Reddy, 1998). Promoting production of phenols and decay inhibitors by genetic modification of plants such as Sphagnum that produce phenols has been considered for freshwater environments (Freeman et al., 2012). Physiochemical enhancement to suppress phenol oxidase activity by manipulating oxygen availability, acidification by addition of sulfates, addition of phenolic compounds (peat leachates and polyphenolic waste materials), lowering pH, labile carbon, and inorganic nutrient supply has also been proposed (Freeman et al., 2012).

This new area of consideration is in the research development and demonstration phase. These types of projects will require increased technological capacity to assess the permanence of materials used and could employ adaptive management-based approaches to “learn while doing.” They could also employ designed experiments, to understand the conditions where they are best applied, and accelerate progress to deployment at scale. With this approach, ecological concerns and off-site negative impacts associated with hard coastal defenses will persist unless the design enhances wetland functions and improves shoreline processes, because shoreline armoring is prohibited or significantly restricted in several U.S. states (O’Connell, 2010). Other opportunities include using species or phenotypes with higher lignin content or afforestation (in mangrove forests). These opportunities have fewer technological barriers but have potential ecological implications.

IMPACT POTENTIAL

Total Carbon Fluxes under Coastal Blue Carbon

Coastal blue carbon is an approach with near-term readiness and low cost when coastal ecosystems are maintained, restored, created, or engineered with minimal hard infrastructure and for other purposes (e.g., coastal risk reduction, fisheries production). Coastal restoration, adaptation, and management offer the potential to maintain and accelerate the rate of negative CO₂ emissions at a scale of 0.02-0.08 Gt/y CO₂ (Table 2.1). The committee expects the timeframe for readiness for each approach to vary. By 2030, the committee estimates that implementation of several of the management approaches described above can result in annual flux of 0.037 Gt/y CO₂. By 2060, annual flux could reach 0.077 Gt/y CO₂ depending on technological developments, improved scientific understanding, and the ability to overcome societal barriers. Technological advancements that enable successful demonstration and deployment of carbon-rich projects supply about 36 percent of annual flux by 2030 and 43 percent by 2060. Wetland transgression becomes a more important flux over time, increasing to 32 percent by 2100. This section describes the source of these estimates.

The total carbon flux per year, and potential carbon impact of coastal blue carbon is most influenced by the total area of coastal carbon ecosystems, the rate at which they bury OC, and the potential to augment projects and strategies to manage wetland transgression. As described above, the committee considered the following key coastal blue carbon approaches: (1) active ecosystem management; (2) restore coastal wetlands where they have been degraded or lost; (3) convert hardened and eroding shorelines to natural shorelines as part of coastal adaptation; (4) manage wetland transgression into uplands with change in SLR and human drivers/impacts; and (5) increase CO₂ removal capacity by augmentation with carbon-rich materials. The total carbon flux potential is based on the maximum area available to implement each approach, and the sequestration rates of 7.98 t/(ha y) CO₂ for tidal wetlands and 1.58 t/(ha y) CO₂ for seagrass meadows (EPA 2017; SOCCR 2).

Active Ecosystem Management

According to NOAA C-CAP, current estuarine wetland area is 0.22 million ha. Estimating areas of seagrass meadow is more challenging because of knowledge gaps in seagrass meadow distribution, extent, and species identity. The current estimate of seagrass meadow in the United States is 0.24 million ha (SOCCR). The total carbon flux

TABLE 2.1 Total U.S. Annual Carbon Flux for Tidal Wetlands and Seagrass Meadows for Key Coastal Blue Carbon Approaches Evaluated in This Report

^a Active ecosystem management is unlike other approaches in that maintaining rather than increasing wetland area constitutes negative emissions. However, proactive management is needed to maintain the rates of flux and area. Because of high uncertainty in estimating the decreasing baseline of the current carbon burial rate, the committee did not deduct the baseline when calculating this CO₂ removal number. ^bTwenty-five percent of potential area restored by 2030, full area restored by 2060; ^c25 percent of potential area adapted by 2030, full potential area adapted by 2060; ^dProjected 0-2 ft SLR by 2060 and 2-4 ft SLR by 2100, land area estimate reflects assumption of 1 ft of accretion through 2100; ^eNo additional areas for any CO₂ removal approach except managed wetland transgression were included in the 2100 scenario. ^fAugmentation of projects with carbon-rich materials implemented at 25 percent potential area for restoration and adaptation projects by 2030, full potential area by 2060 (annual rate based on area of projects implemented by year indicated).

of coastal blue carbon can be estimated from the combined area and rate of OC burial for existing tidal wetlands and seagrass beds. For existing “natural” tidal wetlands and seagrasses, the total carbon flux is currently approximately 0.021 Gt/y CO2 (Table 2.1). To maintain these annual rates, the rate of sequestration per unit area and total area of coastal wetlands would require different levels of management, and in many cases, proactive management. Because of high uncertainty in projecting the future change to the baseline sequestration, the committee did not deduct the baseline when calculating the impact potential for this NET. Instead, the committee used the current coastal carbon burial rate as the CO₂ removal capacity of active ecosystem management to maintain this number.

Restoration of Lost or Degraded Coastal Wetlands and Wetland Creation

The total carbon flux from restoration of coastal wetlands can be estimated using the rate of OC burial applied to the areas within the coastal wetland boundary (under MHHW line) currently in other land uses or where wetland condition has been degraded. Coastal wetlands can be restored through multiple approaches depending on their type, degree of degradation, and geomorphic setting. Each approach has been shown to achieve CO₂ removal at similar or higher rates than existing “natural” tidal wetlands (e.g., Osland et al., 2012). The committee estimated the total potential annual flux from the sum of the annual sequestration for six types of potentially restorable land use types: (1) medium- to high-intensity developed lands (53,938 ha), (2) low-intensity developed lands, open space, and barren land (139,171 ha), (3) cultivated, pasture/hay and grasslands, including tidally restricted lands (317,468 ha), (4) unconsolidated shore (341,721 ha), and (5) recently lost or eroded and converted to open water (125,525 ha; EPA, 2017; NOAA Office for Coastal Management, 2011). Additionally, the area of seagrass meadow estimated to have been lost in the United States provides an opportunity for further restoration (342,943 ha; Waycott et al., 2009). To avoid double-counting of land areas, the committee assumed that developed and cultivated lands approximated the MSL-MHHW extent of the tidal frame, cultivated lands were at approximately MSL, unconsolidated shore occupied the tidal frame from below MSL to 50 ft below mean low water (MLW), and recent open water from 50 ft below MLW to 150 ft below MLW. The committee estimated an annual flux of 0.008 Gt/y CO₂ if all former coastal wetlands were restored and potentially suitable areas used for wetland creation. Because not all former coastal areas could be restored immediately, the committee estimated an annual rate of 0.004 Gt/y CO₂ by 2030 (with 50 percent of the available land area restored) and 0.008 Gt/y CO₂ by 2060 (with all potential area restored). The committee held this rate constant into 2100, even though it depends on maintaining this area of coastal wetlands.

The ability to realize the CO₂ removal potential of these lands will depend on site conditions, elevation, and degree of disturbance. For example, utilizing lands in low-intensity or nonhuman areas may pose the lowest risks to increasing available land for coastal blue carbon because they are not in high demand. Kroeger et al. (2017a) report that 27 percent of former tidal wetlands on the U.S. Atlantic Coast are currently tidally restricted. Much of this area is likely used for agricultural purposes. Converting to coastal wetland from agricultural use may pose greater societal and economic consequences than converting from lands that are of relatively lower intensity (see Chapter 3 for more detail on implications of significant shifts in land use). Although possibly containing eroded or subsided tidal wetlands, unconsolidated shore area may include valuable near-shore habitat that is not suitable for tidal wetland restoration or creation. Some areas may be considered more appropriate for seagrass restoration. Other considerations include changes in coastal policy; for example, if the National Flood Insurance Program (NFIP) adopts a risk-based model, existing areas of developed land with high frequency of losses due to flooding or storm surge may become more readily available for other uses that can sequester OC. These considerations introduce significant knowledge gaps because of the potential variability in the annual carbon flux of restored and created coastal wetlands.

Conversion of Hardened to Natural Shorelines and Stabilization of Eroding Shorelines

The committee used the length of armored shoreline for the United States (22,000 km) and a width of 61 m (used as an approximate tidal range from MSL to 50 ft below MLW) to compute the potential area that could be converted to living shorelines and bury OC. To avoid double-counting with potential restoration areas currently in cultivation and development, the committee used the approximated area occupied by the tidal range below MSL. Replacing the existing length of hardened shorelines (22,000 km) with living shorelines, the committee estimated an annual rate of 0.001 Gt/y CO₂ by 2030. Employing these natural and nature-based measures to prevent erosion of another 22,000 km of existing shoreline would result in an annual rate of 0.002 Gt/y CO₂ by 2060.

Management of Wetland Transgression into Uplands

The committee approximated the following SLR scenarios: 0.68m by 2060, 1.12m by 2100, and 1.68m by 2130 (NOAA Office for Coastal Management, 2011). If the existing tidal wetland areas are maintained in their current state (i.e., keeping pace with SLR), and areas of tidal wetland increase as uplands become inundated by regular tidal

flooding (either through assisted management or other means), total tidal wetland area will increase. With 0.68m of SLR, approximately 1.5 million ha of new tidal estuarine wetlands are estimated to develop (NOAA Office for Coastal Management, 2018). At an annual rate of 7.98 t/(ha y) CO₂ for tidal marsh, this equates to 0.012 Gt/y CO₂. Under 1.12m of SLR, an additional area of 0.87 million ha is projected (0.007 Gt/y CO₂), for a total annual flux increase to 0.019 Gt/y CO₂. However, significant uncertainties persist with regard to the fate of tidal wetlands under SLR and future coastal management (Kirwan and Megonigal, 2013). As the leading edge of upland migrating inland becomes marsh, the coastal edge may be submerged and/or eroded, depending on coastal management. The carbon lost from submerged and/or eroding may cause a permanent loss of carbon, be emitted to the atmosphere, or be redeposited on the marsh or mangrove shoreline or bank. Significant erosion of coastal wetlands in areas has already been observed, and projected SLR suggests that new areas will be eroded without management interventions (Figure 2.4). Using an accretion rate of 1 ft by 2100, “open water” area is projected to increase on the order of 1.5-2.0 million acres with each foot of SLR (NOAA Office for Coastal Management, 2018).

Augmentation of NET Projects with Carbon-Rich Materials

Approximately 467,246 ha (the sum of unconsolidated shore and recently eroded wetland areas) of land is available for restoration, which could incorporate carbon-rich materials to increase elevation to an appropriate level for persistence of wetland vegetation. If carbon-rich materials were supplied at the same rate that restoration and adaptation projects were implemented, and included as part of managed wetland transgression strategies (50 percent by 2030 and full area by 2060), the annual flux would reach 0.007 Gt/y CO₂ by 2030 and 0.027 Gt/y CO₂ by 2060 (based on augmenting with carbon-rich materials to 1.5 ft for restoration projects in unconsolidated shore areas and 3 ft in recently eroded wetlands, 3 ft for coastal adaptation projects, and equivalent of 1 ft stocks in strategies for managed wetland transgression). The addition of carbon-rich materials is an opportunity for carbon storage and not for carbon removal. Thus, annual rates were obtained by dividing the total storage capacity of this level of project implementation by the number of years the projects will occur (up to 2100 in this scenario).

Summary of Coastal Blue Carbon Estimates

Combining the annual fluxes for each approach based on the potential rate at which they could be implemented, the committee estimated an annual carbon flux rate for

the four time horizons: current (2018), 2030, 2060, and 2100. Total potential annual flux was estimated as 0.021, 0.037, 0.077, and 0.058 Gt/y CO₂, respectively (Table 2.2; Figure 2.6).

Total U.S. potential carbon removal capacity is the magnitude of OC burial rates, should all potential activities identified by the committee be implemented and rates of OC sequestration maintained over the specified time horizons. Although global areas of coastal tidal wetland and seagrass vegetation have been estimated and annual rates applied to derive global carbon capacity of existing areas, little is known about the global capacity if these restoration, creation, and nature-based engineering approaches were applied at a global scale.

Existing coastal ecosystem management areas yield an estimated potential carbon capacity of 0.233, 0.868, and 1.714 Gt CO₂ by 2030, 2060, and 2100, respectively. Tidal wetland and seagrass areas that could be restored yield another 0.023, 0.265, and 0.591 Gt CO₂ by 2030, 2060, and 2100, respectively. These estimates assume that areas that could be restored within the coastal boundary (below MHHW) are restored at 25 percent total area by 2030 and 100 percent total area by 2060. The estimate does not include any new projects beyond 2060, which would increase the total 2100 capacity. Nature-based adaptation projects yield an additional 0.006, 0.068, and 0.152 Gt CO₂ by 2030, 2060, and 2100, respectively. This estimate also does not include new projects beyond 2060, which would increase the total 2100 capacity. Managed wetland transgression adds another 0.496 Gt CO₂ by 2060 (with 0.68 m SLR) and 0.980 Gt CO₂ by 2100 (with 1.12 m of SLR). Finally, augmenting shorelines with carbon-rich materials, implemented at the same rate as described, yields another 0.148, 1.123, 1.426 Gt CO₂ by 2030, 2060, and 2100, respectively. Thus, the total potential capacity for coastal carbon is 0.410, 2.820, and 5.424 Gt CO₂ by 2030, 2060, and 2100, respectively.

Other Radiative Impacts

Methane

Restoration of tidal wetlands from drained upland and impounded fresh wetlands provides extra benefit to negative carbon emissions, which is the reduction of CH₄ emissions from ditches in these areas. Generally, tidal restrictions such as roads or undersized culverts increase GHG emissions because they either drain the wetland, causing CO₂ emissions, or they cut off the input of the saline water needed to inhibit CH₄ emissions. The supply of tidal saline water is often associated with sulfate that inhibits methanogenesis as sulfate reduction overpasses carbon reduction as electron donors (Poffenbarger et al., 2011). Therefore, salinity has been frequently used as a proxy for estimating CH₄ emissions. Restoration of tidal connections to currently impounded wetlands could decrease CH₄ emissions of 7.9-41.1 (g/(m² y) CH₄) (Kroeger et al., 2017a), or 739-3,843 g/(m² y) CO₂ when converted to CO₂e by a factor of 34 (CH₄ global warming potential, IPCC, 2013). On the U.S. Atlantic coast, if 2,650 km² wetlands could be restored, the CH₄ benefit would be 2.0-10.2 Mt/y CO₂.

Nitrous Oxide

Nitrous oxide (N₂O) is another potent GHG with a global warming potential of 298 (IPCC, 2013). N₂O is produced largely by the denitrification process in wetlands and saturated uplands. The salinity and sulfate in tidal wetlands suppress the production of N₂O. Therefore, the N₂O emissions in tidal wetlands are very small compared with CO₂ and CH₄ fluxes, about an order of magnitude less than CH₄ fluxes (Martin et al., 2018; Murray et al., 2015).

Albedo

Land-use and land-cover changes can change the albedo, or reflectivity, of Earth’s surface, and thereby change the radiation balance. There is little difference in albedo for various wetland plants. The albedo of flooded salt marsh is ~0.089 (Moffett et al., 2010), slightly higher than ocean water. Considering the small total area of coastal wetlands not only for the United States, but also for the world, coastal carbon approaches would have a trivial effect on the overall global radiation balance, considering albedo alone.

SECONDARY IMPACTS

Ecosystem Services

Tidal wetlands and seagrass meadows provide ecosystem services (Barbier et al., 2011), which are broadly defined as “the benefits people obtain from ecosystems” (MEA, 2005a). Examples of services provided by coastal ecosystems include recreation and tourism, key fishery habitats, improved water quality, and flood and erosion mitigation. Each has monetary and nonmonetary value, which can reduce risks to life, property, and economies (Barbier et al., 2011; Duarte, 2000; Lovelock et al., 2017; Mcleod et al., 2011). As described above, these coastal ecosystems are maintained, restored, used for nature-based adaptation, and managed to enable wetland transgression into uplands to provide ecosystem services as well as remove CO₂.

Although ecosystem services in coastal ecosystems are well documented, data on their monetary and nonmonetary values are limited. Where monetary value exists, it has proven difficult to bring the services to the marketplace, with the exception of carbon in mangroves (Jerath et al., 2016). The value of services provided by tidal marsh has been extensively studied, but less is known about seagrass meadows (Table 2.3; Barbier et al., 2011; Craft et al., 2009; Gedan et al., 2009; Orth et al., 2006; Waycott et al., 2009). Even before consideration of its market value for carbon removal, watershed restoration provides a 3:1 (CPRA) to 8:1 (Sklar, presentation to committee) return on investment. The National Science and Technology Council released a report on research needs for coastal green infrastructure to improve assessments of ecosystem services (NSTC, 2015).

Risks of Coastal Blue Carbon

Some of the coastal blue carbon approaches considered here pose risks that will influence where and how they are deployed (e.g., large fill volumes, subtidal areas, coastal landscape processes). These risks include:

• potential for sediment contaminants, toxicity, bioaccumulation and biomagnification in organisms,
• issues related to altering degradability of coastal plants,
• use of subtidal areas for tidal wetland carbon removal,
• effect of shoreline modifications on sediment redeposition and natural marsh accretion, and
• abusive use of coastal blue carbon as means to reclaim land for purposes that degrade capacity for carbon removal.

TABLE 2.3 Ecosystem Service Value Examples for Coastal Blue Carbon

^aGBP: Great Britain pounds; ^bAUD: Australian dollars.

SOURCE: Modified from Barbier et al., 2011.

The U.S. Army Corps of Engineers (USACE) uses Ocean and Inland Testing Manuals (OTM and ITM) and tiered approaches to evaluating toxicity issues associated with dredged materials in accordance with the Food and Drug Administration Action Levels for Poisonous and Deleterious Substances in Fish and Shellfish for Human Food and the Water Resources Development Act of 1999. The USACE follows Section 103(b) of the Marine Protection, Research and Sanctuaries Act in choosing placement sites designated by EPA to reduce human and environmental impacts to the maximum extent possible (EPA, 1991).

The alteration or degradation of the wetland plant community raises significant ecological questions. Most coastal fisheries of the U.S. East and Gulf of Mexico coasts are estuarine-dependent, that is, they rely on tidal wetlands as nursery grounds for juvenile development. At some stage, most of these organisms depend on wetland vegetation-derived detritus. Changing the lignin content of the detrital feedstock could negatively impact secondary production of the entire coastal zone. The social and economic impacts of introducing genetically modified plants into the coastal zone must be evaluated before or alongside any research agenda on lignin modification to investigate at what point, if any, marsh survival based on increased lignin content balances the alteration of coastal fisheries and dependent human livelihood.

Societal Impacts

There may also be significant societal barriers to conversion of flood-prone lands to wetlands despite the increasing coastal risks of flooding. To overcome these barriers, managers must plan and design the coast in a way that allows for continued development by humans while enhancing carbon removal capacity of larger areas of wetlands (Stark et al., 2016). However, the rising risk of flooding and storm damage in coastal zones may lead to regulatory changes that reduce the financial burden on the federal government and disincentivize coastal development. Coastal hazards significantly increase risks to people and infrastructure. For example, the cumulative cost of the $16 billion weather events in the United States in 2017 was $306.2 billion (NOAA National Centers for Environmental Information, 2018). As a result of Superstorm Sandy, Hurricane Katrina, and the events of 2017, the Federal Emergency Management Agency (FEMA) has incurred more than $24 billion in debt (CBO, 2017) with additional borrowing expected.

As regulatory agencies better enable restoration and nature-based adaptation projects, efforts to safeguard against land reclamation may create a socioeconomic barrier (Chee et al., 2017). The difference between wetland creation for coastal adaptation

(wetland reclamation) and land reclamation is distinct but may be blurred by uninformed policy or unenforced policies. As a best practice, assurances that coastal restoration and nature-based adaptation projects are intended to build wetlands for effective carbon removal rather than to develop land in the coastal zone should be built into the regulatory framework.

Other Barriers

This report addresses the permitting mechanisms that exist for deployment of NETs. However, the permitting process should be improved. For example, the USACE process takes on average more than 300 days to complete (USACE, 2017). In 2016, a NW54 was approved to expedite living shoreline or coastal bioengineering projects. Projects eligible for this expedited permit review are limited to 30 ft depth from shoreline and 500 ft length (USACE, 2017). Identification of ways to accelerate regulatory approval of projects that also safeguard robust and effective carbon removal is a critical need.

ESTIMATED COSTS OF IMPLEMENTING COASTAL BLUE CARBON

The costs to implement the different coastal blue carbon approaches vary widely and largely depend on project size, intervention type, design and construction costs, materials costs, costs to transport materials and equipment, and monitoring carbon removal. However, if such projects occur regardless of carbon removal potential, because of the multiple ecosystem services and coastal adaptation functions they confer, then only the incremental costs for monitoring carbon removal needs to considered.

Incremental Costs for Monitoring Coastal Blue Carbon

If projects are implemented for purposes other than or in addition to carbon removal, then costs are reduced to the incremental cost of monitoring coastal carbon removal. Such costs approximate $0.75/t CO₂ for tidal wetlands and $4/t CO₂ for seagrass meadows) for all coastal blue carbon approaches, except those augmented with carbon-rich materials (estimated at $1-30/t CO₂) depending on the material and construction method used. To estimate those monitoring costs, the committee considered the costs associated with existing coastal and terrestrial monitoring networks that include coupled remote sensing and plot-based measurements. For example, the California Forest Change Detection Program has achieved a cost of $0.004/ha through agency efficiencies and leveraging staff across programs (Fisher et al., 2007a). The monitoring

needs for coastal blue carbon would not be limited to land change detection, which represents the minimum cost. A relevant example of a comprehensive monitoring program is the Coastwide Reference Monitoring System (CRMS), which is a mechanism to monitor and evaluate the effectiveness of projects conducted under the Coastal Wetlands Planning, Protection, and Restoration Act (CWPPRA) in Louisiana at the project, region, and coastwide levels (Steyer et al., 2003). The network provides multiple forms of data and research for a variety of user groups, including resource managers, academics, landowners, and researchers. The estimated cost of this research and monitoring program is $80/ha or $6/(ha y) based on projects funded thus far.³ Monitoring system costs reported for national-level programs were reported as $0.50 to $5.50/ha (Böttcher et al., 2009). Recognizing that the CRMS includes research, the incremental cost of monitoring may be lower.

RESEARCH AGENDA

The committee developed a research agenda with the overarching goal to preserve and enhance the high rates of OC sequestration in existing tidal wetlands and seagrasses in the coastal zone and to expand the area covered by these ecosystems. As discussed earlier, carbon sequestration rates can be enhanced through a combination of management activities that depend on

• increasing the OC density in soils of coastal systems,
• retarding edge erosion of existing wetlands,
• increasing aerial expanse of wetlands through transgression into upland areas as these areas become flooded by the sea,
• augmenting mineral sediment availability to ensure wetland elevation remains in balance with increasing rates of SLR,
• hybrid “engineering” (restoration, creation, coastal adaptation) approaches that enhance carbon removal and maintain or improve coastal ecosystem services, and
• augmenting soils with high concentrations of slowly degrading OC such as biochar or logs.

The research agenda also examines technological needs and the feasibility of carbon removal as part of coastal protection projects designed to minimize exposure of human systems to the risk of storms and floods. The research is recommended to reduce the barriers to reaching 1 GT CO₂ scale deployment in the United States as well as the

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³ Based on projects listed on the CWPPRA website for 2005-2019. See https://www.lacoast.gov/new/Projects/List.aspx.

largest uncertainties associated with the processes that most affect annual carbon removal and capacity.

The research agenda spans basic science research on carbon removal potential, carbon transformation, and permanence under different approaches as well as socioeconomics and policies associated with converting land for carbon removal. A proposed framework to accelerate deployment of approaches that increase wetland area through restoration, creation, nature-based adaptation, and managed wetland transgression at scale combines designed experimental research, demonstration sites, and adaptive management of engineered projects. Designed research enables rigorous tests of approaches, demonstration sites enable tests of augmenting projects with carbon-rich materials, and adaptive management enables modification of technology and engineering or management parameters when projects do not meet expected performance criteria. Some proposed approaches will have implications for measurement of species diversity and productivity in the coastal zone. Finally, social science research will investigate both the implications of these projects and the social barriers to achieving them at scale. The research agenda components are described in detail below, and their costs are summarized in Table 2.4.

Basic Research

Research is needed that will address some of the key uncertainties in understanding and using coastal ecosystems as a NET, including but not limited to (1) the fate of OC fixed in coastal ecosystems; (2) the changes in areal extent of coastal ecosystems in response to climate change, SLR, and human disturbance, and (3) genetic engineering or selection of high OC density materials and coastal plants that decay slowly in coastal sediments. The committee envisions a comprehensive research program at a similar scale to NOAA’s national sea grant program budget.

The fate of OC produced and buried in soils/sediment of coastal ecosystems

Basic research is required to reduce uncertainties in how changes in sea level, climate, and human activities will impact the primary production, ecosystem respiration, and long-term burial of OC in coastal wetland ecosystems. Our current understanding and ability to predict carbon burial rates in existing coastal wetland ecosystems is limited, and predicting how rates will change is an important challenge. Research should be designed to take advantage of existing strong gradients in OC burial rates, tidal amplitude, biogeographic province/wetland species composition, climate, sediment

TABLE 2.4 Costs and Components of a Coastal Blue Carbon Research Agenda

NOTE: CMS = Carbon Monitoring System; LTER = Long-Term Ecological Research; NASA = National Aeronautics and Space Administration; NET = Negative Emissions Technology; OC = organic carbon; SLR = sea level rise.

availability, direct local human activities (e.g., nitrogen enrichment) to develop models with universal application. A combination of field, lab experiments, and modeling activities would be appropriate in conducting this research. Potential funders: National Science Foundation Division of Environmental Biology (NSF DEB), Chemistry & Materials Science), USACE, Department of Energy (DOE), industry research and development (R&D), architecture and engineering (A&E) firms, foundations. Research budget: $2M/y over 10 years.

The change in areal extent of coastal ecosystems through the remainder of the 21st century in response to changes in major controlling drivers, such as climate change, sediment availability, SLR, and human disturbances

Coastal wetland ecosystems are being heavily impacted, and their areal extent is changing rapidly as a result of rapidly changing rates of SLR, sediment availability, and other factors. The fate of existing systems and whether they will decrease in areal expanse because of edge erosion or drowning or increase in areal expanse in conjunction with transgression into upland areas as they flood with rising seas is poorly understood. The ability to predict areal expanse is of paramount importance because of its effect on the future carbon removal trend. Research should be conducted to develop mechanistic and predictive understanding of these dynamics in the future. A research program should be developed to improve understanding and prediction of the extent of the coastal wetlands under multiple stresses. Potential funders: NSF DEB, NSF Division of Ocean Sciences (OCE), DOE, NASA. Research budget: $2M/y over 10 years.

Selection of plants/phenotypes able to produce high OC–density tissues as well as other OC materials that resist decay in coastal ecosystem sediment carbon and slow-decay species/phenotypes used in enhancing coastal carbon

There is little knowledge about the preservation of OC-rich materials in coastal sediments and the capacity to increase their production by new strains/phenotypes of coastal plants. Research is needed to better understand the decomposition and preservation capacity of OC-rich materials such as wood logs and biochar. Research should be conducted to investigate the feasibility and ecological costs and benefits of introducing new plants/phenotypes/genotypes able to produce greater amounts of less degradable tissues. Genetic research could increase lignin content of existing wetland plants. Also needed is a program to improve the technological readiness of

using carbon-rich materials, in some cases with new lab-scale experiments, including materials science. Multi-omics research should be supported to better understand how lignin decomposition is related and controlled by microorganisms (Billings et al., 2015) and environmental variation. Potential funders: NSF (engineering and infrastructure programs, DEB, Chemistry & Materials Science), USACE, DOE, industry R&D, A&E firms, foundations. Research budget: $2M/y over 5 years.

Development

Mapping transitions in wetland and seagrass land cover and land use due to SLR and other drivers

More work is needed to develop and refine remote-sensing approaches to estimate potential areas for restoration, nature-based adaptation, and wetland elevation gain, wetland productivity, wetland OC burial, edge erosion, and transgression into uplands in the future. Current approaches are labor intensive in the field and lab. Remote sensing offers the opportunity to scale efficiently, to improve accuracy, and be able to better predict and map existing and potentially available coastal wetland areas for carbon removal. For existing, transgressing, and restored tidal wetlands, knowledge gaps persist in mapping salinity boundaries, which reduces the ability to predict CH₄ and CO₂ emissions and uptake. Similarly, for seagrass meadows, OC accumulation rates are roughly understood, but the ability to map and monitor their areal extent is limited. Other key research needs include development of robust typologies that coincide with scenario modeling to identify and project vulnerable land areas where management and restoration should be focused. Finally, developing new methods to account for forested/afforested wetlands along tidal boundaries (i.e., differentiating them from terrestrial forest) in areas where wetlands are transgressing is another key research need.

Further technological development of mapping and remote-sensing tools, as well as field-validated modeling to derive more robust relationships and constrain variability in detecting coastal classifications fundamental for carbon removal (e.g., project types, salinity, soil accretion), are recommended. Research efforts should prioritize long-term research sites. Research sites should be augmented with long-term data to support remote-sensing applications. Potential funders: NASA, DOE, NOAA, U.S. Forest Service (USFS), EPA. Research budget: $2M/y (tidal wetland: $1.5M; seagrass: $500K) for 5 years.

Development of a Comprehensive Coastal Blue Carbon Projects Database

A strategy is needed to develop a robust means to verify and catalog number, types, size, cost, engineering specifications, and performance of coastal blue carbon approaches. Best management practices may include design criteria and performance functions associated with technological and ecological specifications of projects. The database will be useful for developing and testing predictive models of coastal blue carbon ecosystem extent and CO₂ sequestration rate. It will also be critical for evaluating the conditions under which carbon removal is optimized, so that adaptive management can be facilitated.

The committee recommends:

• development of an integrated project data repository
• design guidelines for coastal blue carbon with emphasis on coastal restoration, nature-based adaptation, and management of wetland transgression approaches that are most cost-effective, including levels of technical readiness and best management practices
• policy instruments for implementation, monitoring, and assessment, including a catalog of the number, types, size, ecological and geophysical site conditions, and performance of coastal blue carbon approaches and assessment of associated resource requirements

The “Blue Carbon Research Coordination Network” is developing a comprehensive database of carbon data and some online resources catalog living shoreline projects, but with only limited quantitative data and no comprehensive project and data repositories of this type. Potential funders: NSF, EPA, U.S. Fish and Wildlife Service (USFWS), USACE, NOAA, state agencies. Research budget: $2M/y for 20 years, with an interagency–academia–nongovernmental organization (NGO)–industry program work group.

Demonstration

Coastal blue carbon demonstration projects and field experiment network

Demonstration of carbon-rich projects will serve to justify and prioritize approaches that are the most cost-effective for increasing CO₂ removal capacity. Field-scale experiments should take advantage of the extreme range in environmental conditions and settings in which U.S. coastal carbon ecosystems exist to facilitate understanding of basic processes, such as OC burial rates and long-term decay rates of carbon-rich materials under waterlogged, saline conditions. Multiyear studies of low-risk,

high-carbon species or phenotypes should also be conducted. Field experiments should involve augmenting a subset of existing natural, restoration, eroding, and adaptation demonstration sites to test carbon-rich approaches in mesoscale field projects. Potential funders: NSF (engineering and infrastructure programs, DEB), NOAA, USACE, DOE, industry R&D, A&E firms, foundations. Research budget: new field experimental network of demonstration projects that accelerate carbon removal: $10M/y for 20 years.

Carbon removal systems field experiments and adaptively managed site network (managed and engineered sites)

For “engineered” projects (i.e., wetland restoration and creation and nature-based coastal adaptation), knowledge gaps exist about the various management approaches to sustain the rates of OC accumulation reported in case studies. Although the number and coverage of wetland restoration projects have increased, few studies have considered the long-term OC rates under different settings and project types, typologies guiding implementation (Dürr et al., 2011), and best management practices for monitoring and evaluating the carbon removal performance or response to SLR. Such projects have been implemented since the 1950s, but few studies have answered the suite of research questions that have been so rigorously applied for “natural” coastal systems. For example, intervention projects that integrate artificial and natural features for coastal defense often focus on biological habitat or diversity rather than carbon sequestration (Firth et al., 2014). Given the likelihood of increasing numbers of projects to protect coastlines, this research infrastructure serves to address both technological issues with engineering coastlines for increasing carbon removal capacity and testing and demonstrating performance of existing and new projects. Further, an understanding of how adaptation projects that reduce risks to people and infrastructure can increase and accelerate carbon removal over time, while also preserving and enhancing other ecosystem services, is needed. What remains uncertain is the ability of large-scale shoreline modifications, in areas reliant on sediment resuspension and redeposition for marsh accretion, to maintain the marsh platform and wetland function in the long-term. The committee is unaware of any surface elevation tables (SETs) or long-term estimates of OC accumulation from projects that have applied nature-based approaches to coastal risk reduction, and few estimates from created wetlands.

Research is also needed to understand better how to manage wetland transgression and erosion. Such research should investigate the underlying processes for enhancing OC burial with different management activities or that enable wetland transgression to occur, as well as the costs, benefits, and socioecological system scenarios and

responses to inform prediction of trajectories and variability in OC burial over time. Although management of wetland transgression could provide immense new areas of wetlands to bury OC and remove carbon, less is known about OC accumulation rates achieved under these conditions. Using adaptive management, corrective action can be applied if management activities do not maintain OC accumulation rates.

Uncertainties about predictions of engineered wetland performance and wetland survival may create a barrier to any coastal blue carbon approach without sustained and new research, utilizing an adaptive management approach that enables planning, trial and error, and corrective actions if performance criteria are not met (Zedler, 2017). Although soil OC burial for coastal blue carbon is considered an incremental cost of coastal resilience projects, a monitoring network would need to be implemented. This network would consist of landscape-level research on management interventions, coupled soil-decay models to forecast outcomes, and empirical field measurements for validation. All such research should be designed to maximize transfer of technologies and approaches to multiple coastal settings. The research objectives include (1) optimizing of OC burial in transgressing wetlands with a suite of structural and nonstructural management approaches, (2) forecasting effects of SLR and sediment availability on trajectories and thresholds of wetland elevation gain and OC burial, and (3) understanding radiative gas emissions in new marshes created by SLR-induced transgression. The field would also benefit from better understanding of the OC balance of uplands as they transition to tidal wetlands. Little is known of the trajectory of upland plant productivity, OC accumulation, and OC fate during the onset of tidal flooding, exposure to saline seawater, and conversion to tidal wetland. Demonstrations/deployments should be conducted across the major upland land-use and land-cover categories likely to be inundated by the sea during the 21st century.

Development of research infrastructure across new engineered and managed sites, and augmentation at existing sites, is needed for a combination of field experiments, networked across adaptively managed projects for accelerated deployment, and coupled with modeling and synthesis efforts. Designed research and field experiments would include observations across networked sites with statistical sampling applied across gradients of project type, coastal biogeomorphic conditions, watershed management, SLR, and climate change exposure. At a minimum, the specific research infrastructure needs includes plot-to-atmosphere measurements of CO₂ and CH₄ fluxes coupled with remote sensing, eddy flux measurement sites for CO₂ and CH₄, a micrometeorological station, water level and salinity instrumentation, an array of SETs, and vegetation plots. An integrated research network across existing NSF/LTER coastal sites, NOAA/National Estuarine Research Reserve Association (NERRA) sites, and NSF/National Ecological Observatory Network (NEON) coastal sites is recommended,

with consistent and comparable monitoring of carbon removal rates across sites, by augmenting existing sites with some research infrastructure already in place. Long-term research sites in managed areas anticipated for wetland transgression with up to 2 ft SLR, near-shore submerged aquatic areas (i.e., seagrass), and engineered projects should be prioritized. Efforts should leverage known adaptive management research infrastructure such as CRMS in coastal Louisiana. This research should identify and prioritize approaches that are most cost-effective (minimal infrastructure and negative impacts) for improving CO₂ removal capacity and other ecosystem services. Data synthesis activities are needed to evaluate knowledge gaps, update the state of knowledge as new research results become available, and scale knowledge up to the regional levels. Synthesis products would inform the efforts of a multiregion, inter-organization (agency–academic–NGO–industry) work group to develop a framework for an adaptive management plan, identify projects and common research protocols for carbon removal and other ecosystem services, and identify a common set of SLR and climate change scenarios and models. Potential funders: NOAA Sea Grant, NSF multiple directorates, DOE, NASA, USFWS, USACE, state agencies, foundations. Research budget: $30M/y for 20 years.

Deployment

Socioecological and economic research to quantify the costs and benefits of coastal blue carbon

Numerous social and policy aspects emerge from coastal blue carbon approaches, particularly related to land-cover and land-use change. Social barriers are not well understood, particularly in the context of new policy initiatives to reduce costs incurred with increasing coastal risk. Social acceptance of allowing wetlands to transgress into upland areas is highly uncertain (Brody et al., 2012). Other socioecological knowledge gaps include how anticipatory decision-making and managed shoreline retreat strategies could be incentivized (e.g., “true” value of risk reduction rather than insurance based) for coastal risk reduction. As regulatory agencies better enable restoration and nature-based adaptation projects, efforts to safeguard against land reclamation may create a socioeconomic barrier to implementation (Chee et al., 2017). The difference between wetland creation for coastal adaptation (wetland reclamation) and land reclamation is distinct but may be blurred by uninformed policy or unenforced policies. As a best practice, assurances that coastal restoration and nature-based adaptation projects are intended to build wetlands for effective carbon removal rather than to develop land in the coastal zone should be built into the regulatory framework.

Research is needed on the barriers to, and incentives for, increasing the area available for coastal blue carbon and potential impacts of coastal wetland reclamation as a type of carbon removal (e.g., numerous studies illustrate biodiversity risks associated with sea walls that reduce tidal mudflat area). Such social science and policy research could address questions, such as:

• If the ecosystem services value derived from coastal wetland reclamation is added, what are the trade-offs and what governance/policies are necessary to manage against unintended outcomes?
• What would be added costs for permitting (and to USACE/state/local agencies)?

If coastal restoration projects are implemented to enhance ecosystem services other than carbon removal, then the $/t CO₂ cost for carbon removal is merely the incremental cost of monitoring. However, costs may differ between coastal projects that do and do not consider carbon removal. For example, it may cost more to implement a coastal adaptation structural strategy that sequesters CO₂ (natural or nature-based infrastructure) than one that is more typical (seawalls). Continued research at the local and regional levels is needed to identify approaches to project coastal vulnerability that integrates robust information based on modeling and field measurements of OC accumulation on lands made available through managed shoreline retreat.

Building and promoting coastal carbon as a NET does not depend on a carbon price, because most wetland restoration and coastal adaptation projects are conducted without concern for CO₂ mitigation. Research that quantifies the cost-benefits of coastal blue carbon may incentivize governments to convert lands exposed to coastal flooding to wetlands or private property owners to abandon vulnerable properties (e.g., NFIP repetitive loss properties) and land uses (e.g., agriculture, low-density lands) in lieu of other ecosystem services, and ultimately release those lands for carbon removal. Further, scientists should evaluate the economic co-benefits of wetland restoration/creation/protection. They also need to enhance their understanding of ways to align the timing of expenditures on wetland actions to decrease vulnerability with the benefits accrued to local governments and citizen taxpayers.

Given the paucity of social science, economic, and policy research to date, numerous socioeconomic issues pose barriers to deploying coastal blue carbon at scale. Researchers should explore socioecological and economic linkages among actions to protect, restore, and expand tidal wetlands and human well-being:

• What combinations of federal, state, and local decision-making are most effective, cost-efficient, and protective of local community property rights and values while decreasing community vulnerability to SLR and storm damages?
• How will market forces influence land use in coastal areas? What are the costs and benefits of regulatory vs market-based approaches to wetland management?

The field would benefit from a typology to better predict the likelihood of individual parcels of property by county becoming wetland (actively and passively) as sea level rises. The typology will likely consider factors such as

• trends in local land-use change, property values, major sources of local government revenue, and extent to which marsh management can reduce property damage from storms and SLR,
• current extent of development, existing open space, and future cost of maintaining infrastructure under a scenario of climate change and SLR, and
• politics or people’s attitudes about the conversion of property to wetlands.

Finally, research is needed on how:

• decisions on risk avoidance are made and whether that varies in relation to the risk timeline,
• availability of resources influences community/local adaptive opportunity and capacity, and
• adaptive options involving wetlands impact local government revenue and finances.

Potential funders: NSF multiple directorates, NOAA, EPA, state agencies, foundations. Research budget: $5M per year for 10 years.

Monitoring and Verification and Research Management

The cost of coastal blue carbon ($/tCO₂) includes the incremental cost of monitoring (described above).

Time delays in fundraising, plan review, and construction should be considered when estimating planning timelines. Research oversight and coordination is embedded within the research agenda cost. Recommendations include developing (1) a multiagency (federal, state, local agencies and academic institutions) working group charged and funded to oversee and integrate research efforts and (2) wetland

mitigation-type policies that assess performance based on criteria for best implementation/management practices.

SUMMARY

Coastal wetlands are extremely productive ecosystems; they act as long-term carbon sinks by removing carbon from the atmosphere through photosynthesis and storing it in their soils for long time periods. Compared to other ecosystems, coastal wetlands sequester a very high amount of carbon per unit area: 7.98 t/(ha y) CO₂ for tidal wetlands and 1.58 t/(ha y) CO₂ for seagrass meadows.

This chapter identifies approaches for tidal wetland and seagrass ecosystem management that could contribute to carbon removal and reliable sequestration gains against the expected future baseline of natural sequestration. These include restoration of former wetlands, use of nature-based features in coastal resilience projects, managed migration as sea levels rise, augmentation of engineered projects with carbon rich materials, and management to prevent expected future losses in carbon capacity. The committee believes that implementation of these approaches could result in the storage of an additional 5.424 Gt CO₂ in coastal wetlands by 2100. Carbon sequestration is one among many ecosystem services provided by coastal wetlands that drive interest in their conservation and restoration. Ongoing restoration activities and the expected growth of coastal resilience efforts provide an opportunity to leverage carbon benefits at no or low marginal cost.

Although wetland restoration is under way is some coastal zones, uncertainty regarding its projected capacity for coastal carbon removal renders it immature as a long-term NET. Biological and geomorphic controls on the rate and permanence of carbon accumulation and sequestration are not well enough understood to predict its future impact alongside high rates of sea level rise and future coastal management practices. Significant unknowns remain about the influence of future changes in coastal watersheds on the fundamental controls on OC burial and of human interventions.

In this chapter, the committee presents a research agenda of basic research, pilot deployment, and monitoring with the aim to better understand the controls on the baseline level of carbon sequestration for the remainder of the century as well as to enhance sequestration rates above the shifting baseline. In addition, societal decisions about climate adaptation and coastal development will influence the ability to maintain or increase coastal wetland area (Schuerch et al., 2018). Therefore, the research agenda also outlines information needs regarding societal responses to land-use change in the coastal zone.