Introduction

Human and natural communities located in the coastal zone are increasingly threatened by climate impacts (e.g., sea level rise (SLR), flooding from coastal storms and coastal erosion) and anthropogenic stressors (e.g., pollution, land use change and development) (e.g., Donnelly and Bertness 2001; Feagin et al. 2005; Erwin et al. 2006; Defeo et al. 2009; Shepard et al. 2012; Fagherazzi et al. 2013; Valle-Levinson et al. 2017; Dahl et al. 2017). In recent years, coastal systems along the United States Atlantic, Gulf of Mexico, and Caribbean coasts have also experienced a number of extreme events (e.g., Hurricanes Katrina, Rita, Ike, Gustav, Sandy, Harvey, Irma, Maria) and accidents (e.g., Deepwater Horizon oil spill). The combined effects of these gradual and acute threats requires innovative, holistic and collaborative approaches to reduce risk (e.g., Adger et al. 2005; NOAA 2015; Wamsler et al. 2016). This includes coastal adaptation strategies that consider short- and long-term climate scenarios, uncertainty, cost-benefit analyses of competing actions, as well as the incorporation of nature and natural elements into shoreline management systems (Stein et al. 2013; RAE 2015).

Inherent to coastal adaptation is the concept of resilience, which is the ability of socio-ecological systems to absorb and recover from disturbances, while retaining or even regaining essential structures, processes or functions (Adger et al. 2005; Folke 2006; Cutter et al. 2008; Fisichelli et al. 2016). Managing for resilience can take several forms, including: 1) resisting change through restoration and maintenance of current conditions, 2) accommodating some level of change after a disturbance, but generally returning to a previous state, or 3) facilitating change either through active management towards a desired new state (e.g., reorganization) or passively allowing for autonomous change (Fisichelli et al. 2016). Natural infrastructure, including natural habitats and features designed to mimic natural processes, can serve as an alternative management approach to traditional grey infrastructure for risk reduction and may provide added benefits to socio-ecological systems (e.g., Arkema et al. 2013; Reguero et al. 2018). These benefits can be characterized as supporting, regulating, culturally sustaining, and provisioning ecosystem services and include enhanced erosion control, recreation and habitat preservation, among others (MEA 2005; Gedan et al. 2010; NOAA 2010; Scyphers et al. 2011; Grabowski et al. 2012; Bridges et al. 2015). While natural infrastructure is becoming more widespread in practice, it often represents a relatively small fraction of a community’s portfolio of coastal risk-reducing strategies when compared to more traditional grey infrastructure (Temmerman et al. 2013; Sutton-Grier et al. 2015; Small-Lorenz et al. 2016; Wamsler et al. 2016; Bilkovic et al. 2016).

The goal of this study is to increase awareness of how natural infrastructure is being used in the coastal zone to enhance socio-ecological resilience to natural and anthropogenic stressors. We assessed the benefits, opportunities and best practices of using four costal habitats, 1) tidal marshes, 2) beaches and barrier islands, 3) biogenic reefs, and 4) mangroves, as natural infrastructure across the U.S. Atlantic, Gulf of Mexico, and Caribbean coasts. In addition, we provide an overview of remaining challenges and information needs that impede systematic consideration of natural infrastructure in coastal planning and management.

The impacts and potential responses of coastal habitats to sea level rise, storms and other stressors

Tidal marshes, beaches and barrier islands, biogenic reefs, and mangroves provide critical nesting, foraging and resting habitat for many fish and wildlife species of high conservation concern, as well as essential nursery and refuge habitat for commercially and recreationally important fishes and invertebrates (NRC 2007; Gedan et al. 2010). These four habitats also benefit coastal communities by providing risk reduction through the attenuation or dissipation of wave energy, breaking of offshore waves, slowing of inland water transfer (NRC 2007; Costanza et al. 2008; Gedan et al. 2010), and sediment stabilization (NRC 2007; Gedan et al. 2010; Scyphers et al. 2011; Gittman et al. 2014; La Peyre et al. 2015). The resilience of these four habitats to rising sea levels, coastal flooding, extreme storm events, and other stressors over the near and long-term depends largely on the exposure to a threat (e.g., rate of local SLR) and sensitivity to a threat based on the surrounding local conditions (e.g., availability of suitable adjacent habitat) to support dynamic response to stressors.

The interactive effects of multiple stressors, such as extreme events and SLR, may push some coastal ecosystems to undergo sudden, rapid and irreversible shifts that result in abrupt or nonlinear changes in an ecosystem quality, property or phenomenon, known as a threshold (CCSP 2009). Quantitative thresholds are important indicators of habitat changes or landscape responses to stressors like SLR and storm surge that could lead to a reduction in a valued resource and related ecosystem services (Powell et al. 2017). A synthesis of existing information on quantitative thresholds and observed or modeled responses to SLR for tidal marshes, beaches and barrier islands, biogenic reefs, and mangroves along the U.S. Atlantic, Gulf, and Caribbean coasts found preliminary information on salt marshes across the geography (Table 1). However, threshold data were scarce for biogenic reefs, mangroves, and beach and barrier island systems, specifically along the northern Gulf of Mexico. Overall, ≥ 50 cm of SLR by 2100 is expected to result in widespread coastal habitat losses along the Atlantic, Gulf, and Caribbean coasts, although losses may vary substantially based on local factors, such as nearshore bathymetry, exposure to severe storms, wave action, and rates of surface elevation change (Fagherazzi et al. 2013; Raposa et al. 2016; Ganju et al. 2017). Some wetlands were predicted to persist in the near term under moderate rates of global SLR (e.g., ~100 cm by 2100) through feedback mechanisms, such as submergence-accretion (wetland inundation with sediment-laden water) and increased plant productivity with submergence (Gedan et al. 2010).

Table 1 Synthesis of responses and quantitative tolerance thresholds related to amounts and rates of sea level rise along the Atlantic, Gulf, and Caribbean coasts for four habitat types: tidal marshes (TM,), beaches/barrier islands (BB,), oyster reefs (OR,), and mangroves (M,). The top figure shows locations where habitat responses and consequences to sea level rise have been observed or modeled across the geography; corresponding details are included in the table below. Icons were obtained from the Integration and Application Network, University of Maryland Center for Environmental Science (ian.umces.edu/imagelibrary/)
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Moderate to high rates of projected SLR (roughly 50–80 cm by 2100) have the potential to substantially impact coastal habitats and degrade, reduce or remove associated ecosystem services (Field 1995; Erwin et al. 2006; Bin et al. 2007; Craft et al. 2009; Melillo et al. 2014). For instance, a 50 cm rise in sea levels by 2100 along the Georgia coast is predicted to convert salt marsh areas to tidal flats and open water, with a concomitant reduction in their productivity and nitrogen sequestration abilities (Craft et al. 2009). Several modeling studies suggest marshes can accrete enough sediment or respond dynamically and keep pace with low to moderate rates of SLR (Lentz et al. 2016; Kirwan et al. 2016). However, empirical studies have shown that, in many places, marsh (Craft et al. 2009; Raposa et al. 2015; Armitage et al. 2015; Watson et al. 2015) and mangrove accretion (Gilman et al. 2008) are not actually keeping pace with current rates of SLR. In South Carolina, a parabolic relationship was demonstrated between inundation and primary production of smooth cordgrass (Spartina alterniflora), suggesting that near-term stability of intertidal salt marsh in response to local SLR depended on marsh elevation (Morris et al. 2002). For oyster reefs, vertical growth on unharvested oyster reefs is generally greater than predicted rates of SLR (Grabowski et al. 2012); however, intertidal oyster reef survival requires inland migration or juvenile recruitment to raise reef elevation and maximize recruitment, growth and survival relative to SLR (Solomon et al. 2014).

The wide range of studies that have assessed SLR impacts to Atlantic, Gulf, and Caribbean coastal habitats (Table 1) provide a starting point for understanding where and when thresholds may be crossed. These data combined with model outputs that identify where dynamic response (Lentz et al. 2016) or inland migration (Enwright et al. 2015) is most likely can be used to support effective adaptation and resilience actions, such as conserving or restoring viable inland habitats. However, habitat responses are complicated and decisions of which habitat to actively maintain or manage towards transition may not always be straight-forward due to complicated ecosystem interactions. For example, along the Texas coast, marsh areas are decreasing in size in response to local rates of SLR, while mangrove forests are expanding in response to rising winter temperature minima and leading to displacement of salt marshes in some areas (Armitage et al. 2015). While expansion may help to increase the overall extent of mangrove habitat, low island mangroves, which are functionally linked to adjacent coral reefs and experiencing simultaneous decreases in productivity, may suffer from lower sedimentation rates and increased susceptibility to SLR and storms (Gilman et al. 2008). Impediments posed by natural or human features of the surrounding landscape are additional challenges to decision-making. For example, under moderate to high scenarios of SLR, several studies (Table 1) show that habitats can persist by migrating upslope, unless hard coastline features (e.g., bedrock coast), development (e.g., Feagin et al. 2005) or steep slopes block or inhibit habitat movement inland (Lentz et al. 2016). Restricted movement of beaches within narrow zones also has the potential to alter habitat characteristics and interfere with ecological functions that provide protective services to the coast from wave energy, tides and winds (Griggs 2005; NRC 2007; Titus et al. 2009). Consequently, migration corridor planning is especially important in urbanized and high-elevation coastal areas to increase ecosystem connectivity and improve wetland migration (Enwright et al. 2015).

Summary of ecological and human community benefits of management approaches using natural infrastructure

Natural infrastructure is being successfully implemented as part of a suite of coastal adaptation actions along the U.S. Atlantic, Gulf of Mexico, and Caribbean coasts to enhance the resilience of socio-ecological communities to the impacts of SLR and storms (Table 2). The U.S. Army Corps of Engineers (USACE) previously synthesized data on the benefits derived from certain coastal habitats, as well as structural and non-structural coastal risk reduction strategies (USACE 2013). We used the USACE (2013) report as a baseline of information and expanded on its findings through an updated review of the peer-reviewed and grey literature. Our review aimed to provide a more comprehensive treatment of the range of management approaches that incorporate natural infrastructure and derived socio-ecological benefits (i.e., ecosystem services) related to tidal marshes, beaches and barrier islands, biogenic reefs, and mangroves. The socio-ecological benefits offered by these four habitats are organized into six management categories, including 1) restoration, 2) landscape conservation design, 3) living shorelines, 4) facilitated re-location, 5) open space preservation, and 6) land use planning (Table 2).

Table 2 A selection of management approaches and their associated ecosystem services using tidal marshes, beaches and barrier islands, biogenic reefs, and mangroves as natural infrastructure to increase socio-ecological resilience to SLR and storms. Ecosystem services were categorized as: provisioning (♦), regulating (○), cultural (□), and supporting (●), as defined by MEA (2005). Provisioning refers to products obtained from ecosystems, like food and water; regulating describes benefits obtained from ecosystem regulation, such as climate-controlled processes; cultural describes non-material benefits from ecosystems, such as social heritage and a sense of place; and supporting describes the services needed to produce all other ecosystem services, e.g., nutrient and chemical cycling. All ecological benefits are considered supporting (for habitat) services
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The management options described in Table 2 provide a range of ecosystem services that enhance resilience of coastal systems to gradual (e.g., climate change) and episodic (e.g., major storms) threats. For example, landscape conservation design, through assessment, acquisition and management, enhances connectivity while also providing natural corridors for species’ migration, persistence and resilience (Bartuszevige et al. 2016). When used in conjunction with information on climate change and climate refugia (Morelli et al. 2016), as well as with projections of development and population growth, these frameworks can facilitate the identification and prioritization of habitat for conservation and connectivity that best support species under current and future conditions of risk. For instance, establishing a network of protected mangrove areas representing a range of different community types and maturity stages can support mangrove persistence in the face of SLR and other threats (Gilman et al. 2008).

Increasing coastal connectivity can also enhance storm protection services (Table 2). According to Barbier et al. (2008a), the relationship between wave attenuation and change in habitat area is nonlinear for salt marshes and mangroves, such that increasing habitat areas inland from the shoreline results in quadratic and exponential reductions in wave heights. Simulations using four hypothetical hurricanes at 12 locations along a shoreline transect (approximately 6 km in length) in the Caernarvon Basin in Louisiana, found storm surge levels were reduced by 1 m for every 9.4 to 12.6 km of additional wetlands along the transect (Barbier et al. 2013). Mangroves in Florida were also found to reduce peak surge levels by 0.4–0.5 m per km of mangrove forest width (Zhang et al. 2012). Beach, dune and barrier island restoration (e.g., dune building, beach nourishment), and limiting development to enable dynamic responses (e.g., breaches) to SLR and storms may be other important actions to increase protective services in some places (e.g., Defeo et al. 2009). The level of storm protection provided by beaches largely depends on the slope of the nearshore submerged environment, wave magnitude and sediment supply (NRC 2007; USACE 2013). Dunes also block waves and prevent inland inundation, depending on several similar factors (Barbier et al. 2008a; Temmerman et al. 2013). An exponential relationship was found between the percent cover of dune grasses and size of oceanic waves blocked by sand dunes, such that as the vegetation density and dune height increased, higher waves were needed to overtop the dune (Barbier et al. 2008b). Facilitating the establishment of dune grasses along the backshore of beaches and the use of sand fencing can help trap sands and create and maintain dunes (USEPA 2009); however, these beach stabilization approaches may conflict and thus need to be balanced with the need for sparsely vegetated areas on dynamic beaches for piping plover and other shorebird feeding and nesting areas (Lott et al. 2007).

Restoration of oyster habitat is a primary strategy to restore lost ecological functions and the broader socio-ecological benefits they provide, including storm protection services through wave attenuation (Table 2; Grabowski et al. 2012; Ferrario et al. 2014). However, oyster reef building and restoration is not effective everywhere. Oyster reef sills, which are often built along eroding shorelines, may not support viable oyster populations for protection against SLR if sited in the subtidal zone in high salinity areas (Baggett et al. 2015; Ridge et al. 2015; Walles et al. 2016). Further, the value of shoreline stabilization provided by oyster reef restoration can vary greatly by location, and restoration investments may not be recovered in places where oyster harvesting practices are particularly destructive (Grabowski et al. 2012).

Vegetated coastal habitats provide important carbon sequestration services and have more long-term potential than terrestrial forests due to higher rates of organic carbon burial in sediments (McLeod et al. 2011). Restored tidal marsh and mangroves may offer more carbon benefits relative to newly created wetlands or through passive management approaches (Kroeger et al. 2017). The global value of coastal vegetated sequestration is between $6.1 and $42 billion USD annually, while conversion and degradation of these habitats can release between 0.15 and 1.02 billion tons of carbon dioxide per year (Pendleton et al. 2012). In Massachusetts, a 20-acre restoration project that removed two culverts to restore tidal flows to a salt marsh showed a net increase in carbon sequestration of 76 metric tons of carbon dioxide per year. Another 60-acre restoration project removed over four feet of wetland fill to restore salt marsh and grassland habitat, which led to a net increase in carbon sequestration of 101 metric tons of carbon dioxide per year (MA DER 2012; 2014). The differences in carbon sequestration rates between these restoration sites may be due to the amount of carbon sequestered by various habitat types, such as high versus low saltmarsh and filled uplands versus coastal grasslands (MA DER 2014).

Lastly, adaptive frameworks and decision support tools that allow managers to integrate and continuously update predictions of risk from climate change, land use and human population growth projections can increase the effectiveness of the types of natural infrastructure described in Table 2 and support short- and long-term biological, cultural, social, and economic goals (e.g., Bartuszevige et al. 2016; Anderson and Barnett 2017). The consideration of quantitative thresholds to climate and other stressors can also help establish management targets and timelines (Powell et al. 2017). For example, when sediment augmentation is used as an approach for tidal marsh restoration, SLR and storm projections along with threshold data for marsh habitats can guide the frequency and amount of sediment deposition, monitoring and maintenance needed to keep pace with gradual and episodic changes (Foley et al. 2015). Other management options (e.g., retreat from coasts and open space preservation) focus on risk reduction by moving people and property out of harm’s way, often with economic incentives like flood insurance discounts. When combined with other zoning and land use protections, these actions can create secondary and tertiary benefits of increasing the persistence and resilience of natural habitats and species. For example, managing lands after the re-location of people or infrastructure in the coastal zone can enable the natural migration of coastal systems as needed in response to relative SLR. More information about how to apply these and other adaptation approaches is available through the Massachusetts Wildlife Climate Action Tool (https://climateactiontool.org).

A comparison of management approaches utilizing natural infrastructure and traditional grey infrastructure

Grey infrastructure has long been used to protect coastal communities from wave impacts, flooding and erosion. However, the myriad benefits that natural infrastructure can provide to ecological and human communities (Table 2) has made them an increasingly attractive alternative to grey infrastructure. In addition, research shows that restoration and management using natural infrastructure can be equally or more successful than grey infrastructure for flood risk reduction when implemented in appropriate places (e.g., Gedan et al. 2010; Temmerman et al. 2013; Jonkman et al. 2013; Small-Lorenz et al. 2016; Bayraktarov et al. 2016).

Natural infrastructure is naturally dynamic and in many ways resilient to the threats from SLR and storms, because it has some capacity to self-repair with minimal maintenance (Temmerman and Kirwan 2015). Conversely, grey infrastructure requires costly repairs following catastrophic storms, augmentation such as increased seawall heights to keep pace with rising actuarial risks, regular maintenance to delay deterioration and prolong design life, as well as eventual replacement (Temmerman et al. 2013; National Science and Technology Council 2015). In addition, grey infrastructure can adversely impact the surrounding natural environment in many ways, such as through loss of sediment (NRC 2007), decreases in beach volume and dimension (Kraus and Pilkey 1988; Hill 2015), and loss of intertidal habitat (NRC 2007; USEPA 2009; National Science and Technology Council 2015). Grey infrastructure can further lead to habitat fragmentation, declines in biodiversity, increases in invasive species, and reduced habitat migration inland in response to SLR (Bilkovic et al. 2016 and references within). These adverse impacts can degrade or inhibit ecosystem services provided by coastal habitats located adjacent to grey infrastructure.

Hybrid approaches that combine natural and grey infrastructure have been shown to contribute to societal, economic and environmental goals (e.g., National Science and Technology Council 2015). However, more information about the relative benefits and costs is needed to inform decisions on the use of each approach (grey, natural) alone or in concert (Sutton-Grier et al. 2015). As a starting point to address this need, we synthesized examples of coastal management and restoration actions with their corresponding economic and/or ecological derived value estimates (Table 3). We note that while these estimates provide some indication of the relative economic benefits, high uncertainty remains and may have limited transferability from one location to another, particularly if valuations are based on a single site.

Table 3 Selected coastal management actions using grey, natural and hybrid infrastructure and corresponding estimates of their value as ecosystem services for coastal storm protection
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Natural infrastructure

There is an increasing number of studies demonstrating the cost-effectiveness of using natural infrastructure for coastal risk reduction (e.g., Gedan et al. 2010; Grabowski et al. 2012; Temmerman et al. 2013; Barbier 2013; Abt Associates 2014; Temmerman and Kirwan 2015; Martin and Watson 2016; Small-Lorenz et al. 2016; Narayan et al. 2016; Reguero et al. 2018). A synthesis study of restoration projects worldwide found costs and success rates vary by habitat, with mangroves requiring relatively lower investment in comparison to seagrasses, salt marshes and oyster reefs (Bayraktarov et al. 2016). Restoration of salt marshes and coral reefs exhibited the greatest success with annual survival rates of 64.8% and 64.5%, respectively, while seagrass habitats had the lowest post-restoration annual survival rates with a median survival rate of 38% (Bayraktarov et al. 2016). Grabowski et al. (2012) analyzed the cost-benefit ratio of oyster reefs and found that restoration costs are typically recovered in 2–14 years, depending on where restoration occurs and the range of services achieved. Overall, the costs associated with natural infrastructure vary widely and depend on many factors, such as design specifications, size and location of project, materials, maintenance, and disturbances that determine how often and the degree to which maintenance and rebuilding are required (NRC 2014). Restored or created habitats and their ecosystem services generally require several years to decades to become well established (NRC 2007; Temmerman et al. 2013), while hard structures can often be built quickly and offer immediate flood protection to surrounding communities.

There remains high uncertainty in the relative effectiveness of natural infrastructure for services like flood risk reduction compared to traditional grey infrastructure. Existing information on flood risk reduction and erosion control has been largely anecdotal to date; this lack of concrete evidence likely inhibits implementation of natural infrastructure, even in cases where it may be less costly than grey infrastructure over the long term. Consequently, it is important to increase the number of valuation studies that definitively link natural infrastructure to the full suite of potential ecosystem and economic benefits, including those that are not traditionally marketed such as flood protection (Table 3) (Barbier 2013).

The global value of ecosystem services provided by natural infrastructure could decline by as much as $51 trillion USD per year or increase by $30 trillion per year based on four alternative global land use and management scenarios (Kubiszewski et al. 2017). Therefore, better communication and public outreach about the costs and benefits of natural infrastructure is critical to ensure decision makers and planners have all existing options available to them to inform action.

Grey infrastructure

Grey infrastructure, such as seawalls, storm surge barriers, dikes, and levees, have been used for decades for protection from storms and flooding. However, these approaches can have unintended negative impacts to habitats that can ultimately undermine the additional flood protection and other services that coastal habitats provide. Hard structures that parallel shores reflect wave energy and constrain the natural inland migration of the shoreline in response to erosion, ultimately causing beaches to become narrower and the beach seaward of the structure to drown (Defeo et al. 2009). This coastal squeezing (Doody 2004; Torio and Chmura 2013) can disrupt normal sediment dynamics, lower the diversity and abundance of biota, and lead to habitat loss (Galbraith et al. 2002; Defeo et al. 2009). Revetments, which are sloping structures made of riprap, concrete mats, timber, or other materials to stop shoreline erosion, can be effective for erosion control if designed and constructed properly. But if revetments are improperly sited on eroding shores, they can accelerate loss of intertidal habitat behind and adjacent to them, causing the beach to convert to open water (NRC 2007). Groins and breakwaters, which are shore-perpendicular and shore-parallel structures, respectively, can similarly reduce sediment supplies in downdrift beaches, causing or accelerating erosion on the inshore sides of the barrier and narrowing or reducing beach habitat (NRC 2007; USEPA 2009).

Grey infrastructure combined with other coastal development and land use changes can lead to further losses in the ecosystem services that coastal habitats provide to society (Table 2; Bayraktarov et al. 2016). For example, dams restrict natural sediment loads needed for salt marsh accretion and maintenance (Weston 2014), decrease geomorphic stability, and degrade salt marsh habitats (Deegan et al. 2012). Grey infrastructure near mangrove habitat can convert mangrove forests to deep water by causing scouring along the front of structures and to downdrift areas (Gilman et al. 2008). Bulkheads can degrade spawning and nursery habitat, while also increasing shoreline erosion. When bulkheads are used to replace degraded vegetated habitats, the water quality improvement function of native vegetation is lost (NRC 2007; Currin et al. 2010).

The direct costs associated with grey infrastructure are well understood due to their long-term implementation, and from being designed according to well-vetted specifications. However, grey infrastructure can generate hidden costs over their design life due to degradation and gradual failure (RAE 2015). In addition, damages caused by these structures to surrounding ecosystems have not yet been fully quantified and documented (RAE 2015). While habitat restoration can be expensive (e.g., Bayraktarov et al. 2016), when represented as average costs per linear foot, the estimated costs of grey infrastructure are generally greater compared to non-structural and hybrid approaches (CCRM 2014).

Hybrid approaches

Hybrid approaches combine grey and natural infrastructure to varying degrees to maximize flood defenses and additional benefits (e.g., Sutton-Grier et al. 2015; Bridges et al. 2015). In some scenarios and locations, hybrid approaches provide the greatest flood protection benefit to coastal communities (USACE 2013; NOAA 2015; Schuster and Doerr 2015). Breakwaters and sills are common in marshes, mangroves and sandy dunes to help attenuate waves and stabilize sediments (NRC 2007). Sills are typically built of oyster shell or granite and placed on the seaward-side of a marsh or mangrove (Sutton-Grier et al. 2015), while breakwaters are generally made of timber, rock or concrete and placed further offshore than sills (RAE 2015). Mangroves can particularly benefit from hybrid shoreline stabilization approaches that use sills and breakwaters to reduce wave energy and maintain calm, low energy conditions that mangroves need to thrive (NRC 2007 and references within). Living shoreline techniques are often hybrid approaches that pair biogenic species and plantings with hardened infrastructure for shoreline protection. For instance, the creation of fringing marsh through plantings may be augmented by the installation of rock sills or other artificial breakwaters along the seaward edge and parallel to the marsh.

More novel hybrid approaches include the use of natural infrastructure to protect permanent and temporary grey infrastructure from storms and waves until the natural features mature and become well established (Sutton-Grier et al. 2015). For instance, oyster reefs located seaward of armored shorelines serve as natural breakwaters that attenuate wave energy and, thus, lessen the impacts of storms, while promoting sediment deposition shoreward of the reef and mitigating habitat loss caused by the existing grey infrastructure (USEPA 2009; Scyphers et al. 2011; Baggett et al. 2015).

When living shorelines are used alone or as hybrid approaches, monitoring results suggest these installments can be effective for enhancing coastal resilience. In Maryland, over 300 marsh fringe sites have been constructed and monitored over a 20-year period, demonstrating they have been effective for erosion control and wetland habitat creation (NRC 2007 and references within). Like many of the approaches discussed in this study, the costs of living shoreline projects can vary greatly with location (RAE 2015) and are not appropriate or effective everywhere. Initial costs can be significantly less than those for grey infrastructure, yet long-term costs will depend on whether and how frequently the living shoreline must be repaired or rebuilt (Titus et al. 2009; Temmerman et al. 2013; Bilkovic et al. 2016).

The future of natural infrastructure: opportunities and limitations

The implementation of natural infrastructure alone or through hybrid approaches to enhance resilience to SLR, storms and other coastal stressors is becoming more widespread in practice (DOI Metrics Expert Group 2015; Abt Associates 2015, 2016; MARCO and NWF 2017). Currently, however, managers have limited opportunities to directly compare risk reduction benefits and costs with traditional grey infrastructure. Regulatory barriers, coupled with lack of public awareness and contractor knowledge of the long-term services provided by natural infrastructure, have impeded permitting processes, which remain cumbersome compared to grey infrastructure. Streamlined guidance for the implementation of natural infrastructure, especially following storms and other extreme events, could give communities greater confidence and advance their application. In particular, best practices for site selection and the conditions where natural infrastructure can be most effective for maximizing socio-ecological benefits are still needed (Bayraktarov et al. 2016; Jahn 2016).

Ecosystem service valuation represents a growing opportunity for enhanced socio-ecological resilience planning and more informed decision-making. Additional studies with greater geographical coverage, consistent terminology, and methodologies for quantifying and valuing services, particularly non-marketed and indirect ecosystem services, would help increase awareness of the total benefits related to natural infrastructure (Barbier 2013; Olander et al. 2015). Lastly, uncertainty about how climate change will impact ecosystem services, including linking changes in ecosystem structure and function to the production of goods and services, limits management and decision making in this arena (Barbier 2013).

Preliminary, yet rapidly, maturing information on natural infrastructure and hybrid approaches can be used to take action and integrate their benefits into resilience planning guidelines. Here, we present potential actions that can increase information and reduce uncertainty around the use of natural infrastructure in coastal planning processes.

  1. 1)

    Substantially increase performance evaluation and widespread monitoring of natural infrastructure.

  • Identify and develop best practices, clear monitoring goals and standardized post-implementation performance evaluations.

  • Communicate and disseminate results more widely, particularly when surprises and complications occur (Jahn 2016; MARCO and NWF 2017).

  • Explore innovative approaches and funding mechanisms for increasing data and long-term monitoring of restoration before and after project implementation, such as using citizen science networks (MARCO and NWF 2017).

  • Where possible, use quantitative thresholds to stressors to inform decisions regarding site selection, design, and implementation of natural infrastructure and hybrid approaches (Adger et al. 2009; Stein et al. 2013; Stein et al. 2014; Powell et al. 2017).

  • Develop scenarios to increase understanding of how resilient natural infrastructure may be at different locations, under future conditions, and timelines to thresholds that may affect habitats and ecosystem services.

  1. 2)

    Increase information on the potential benefits and costs of natural, grey and hybrid approaches for decision-making in the coastal zone.

  • Develop and expand standardized long-term monitoring protocols and common metrics to clarify how natural infrastructure performs relative to traditional armoring practices.

  • Expand cost-benefit analyses to account for the cumulative services accrued by natural infrastructure in comparison to grey infrastructure, as well as long-term maintenance and augmentation costs of both approaches.

  • Increase research to clarify the link between natural infrastructure and potential ecosystem services, as well as the extent to which grey infrastructure affects ecosystem services.

  • Increase synthesis of studies on ecosystem service valuation related to natural infrastructure to stimulate new research and reduce data gaps.

  • Increase assessment of the potential benefits of natural infrastructure for carbon sequestration (Bianchi et al. 2013; Jerath et al. 2016; Yando et al. 2016).

  1. 3)

    Increase coordination and planning around socio-ecological resilience goals.

  • Better communicate the potential benefits of coastal habitats and biodiversity as part of a broader community adaptation strategy (Mawdsley et al. 2009; NASEM 2016).

  • Increase outreach to landowners as part of planning processes to facilitate prioritization of areas where land acquisition may be the best option for autonomous change.

  • Seek opportunities to communicate and integrate the full range of ecosystem services derived from natural infrastructure into community resilience planning and decision making (e.g., Grimm et al. 2013; Nelson et al. 2013; Grimm et al. 2016).

  • Engage state agencies as part of project planning and implementation processes (USEPA 2009). For instance, a best practice for development of State Wildlife Action Plans (SWAPs) is to invite representatives of municipal, county and/or regional planning entities to serve on conservation plan committees (Association of Fish and Wildlife Agencies 2012).

  • Take advantage of regular planning cycles (e.g., SWAPs, hazard mitigation, comprehensive/land use) to coordinate socio-ecological resilience goals. While timeframes differ, these activities bring multiple partners and stakeholders together to identify shared priorities, develop strategies, and inform each other’s benchmarks, successes and challenges.

Conclusions

Coastal management strategies that incorporate natural infrastructure and hybrid approaches provide opportunities for risk reduction and coordination around shared socioeconomic and ecological goals. Studies on climate resilience have grown rapidly in recent years (Fisichelli et al. 2016) and are increasingly being considered in practices related to coastal protection, restoration and management (Mawdsley et al. 2009; NOAA 2010; Stein et al. 2014; Schuster and Doerr 2015; Staudinger et al. 2015; NOAA 2016; USFWS 2016), as well as national assessments and agency operations (e.g., USACE 2014; iCASS 2016). Recent federal efforts, such as the Gulf Coast Ecosystem Restoration Council (2013) and the U.S. Department of the Interior’s Hurricane Sandy Coastal Resiliency Competitive Grant Program that were established in response to major disasters, have helped to advance implementation of coastal adaptation strategies for enhanced socio-ecological resilience at federal, state and local levels. Nonetheless, challenges to systematic implementation of natural infrastructure for enhanced coastal resilience remain, and many practitioners have limited resources to keep up with this rapidly advancing field.

To meet this need, this study provides a comprehensive overview of the current knowledge of how tidal marshes, beaches and barrier islands, biogenic reefs, and mangroves have been used as natural infrastructure to enhance coastal resilience in response to SLR and coastal storms along the United States Atlantic, Gulf, and Caribbean coasts. Our summary demonstrates that investments in natural infrastructure in the coastal zone can have measured value for coastal communities while increasing ecological persistence and resilience. However, information is highly nuanced and spatially variable. More research is needed to develop best practices for where a particular natural infrastructure may be most effectively applied and what can realistically be expected in terms of performance and derived ecosystem services. Natural infrastructure may not be the best option in some locations, and grey infrastructure or hybrid approaches may perform better depending on the local landscape and socio-ecological goals. Regardless, ensuring coastal managers and planners are aware of all potential options and of the short- and long-term costs and benefits is key for advancing this field.