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Human population growth and accelerating coastal development have been the drivers for unprecedented construction of artificial structures along shorelines globally. Construction has been recently amplified by societal responses to reduce flood and erosion risks from rising sea levels and more extreme storms resulting from climate change. Such structures, leading to highly modified shorelines, deliver societal benefits, but they also create significant socioeconomic and environmental challenges. The planning, design and deployment of these coastal structures should aim to provide multiple goals through the application of ecoengineering to shoreline development. Such developments should be designed and built with the overarching objective of reducing negative impacts on nature, using hard, soft and hybrid ecological engineering approaches. The design of ecologically sensitive shorelines should be context-dependent and combine engineering, environmental and socioeconomic considerations. The costs and benefits of ecoengineered shoreline design options should be considered across all three of these disciplinary domains when setting objectives, informing plans for their subsequent maintenance and management and ultimately monitoring and evaluating their success. To date, successful ecoengineered shoreline projects have engaged with multiple stakeholders (e.g. architects, engineers, ecologists, coastal/port managers and the general public) during their conception and construction, but few have evaluated engineering, ecological and socioeconomic outcomes in a comprehensive manner. Increasing global awareness of climate change impacts (increased frequency or magnitude of extreme weather events and sea level rise), coupled with future predictions for coastal development (due to population growth leading to urban development and renewal, land reclamation and establishment of renewable energy infrastructure in the sea) will increase the demand for adaptive techniques to protect coastlines. In this review, we present an overview of current ecoengineered shoreline design options, the drivers and constraints that influence implementation and factors to consider when evaluating the success of such ecologically engineered shorelines
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Oceanography and Marine Biology: An Annual Review, 2019, 57, 169–228
© S. J. Hawkins, A. L. Allcock, A. E. Bates, L. B. Firth, I. P. Smith, S. E. Swearer, and P. A. Todd, Editors
Taylor & Francis
1National Centre for Coasts and Climate, School of BioSciences,
The University of Melbourne, VIC 3010, Australia
2Experimental Marine Ecology Laboratory, Department of Biological
Sciences, National University of Singapore, 117543, Singapore
3The Swire Institute of Marine Science and School of Biological Sciences,
The University of Hong Kong, Pokfulam, Hong Kong, China
4University of Bologna, Dipartimento di Scienze Biologiche, Geologiche
e Ambientali (BiGeA) and Centro Interdipartimentale di Ricerca per
le Scienze Ambientali (CIRSA), UO CoNISMa, 40126, Italy
5Centre for Marine Socioecology, University of Tasmania, Hobart, TAS 7001, Australia
6Institute for Marine and Antarctic Studies, University of
Tasmania, PO Box 49, Hobart, TAS 7001, Australia
7Department of Biological Sciences, Macquarie University, New South Wales 2109, Australia
8Centre for Research on Ecological Impacts of Coastal Cities, Marine
Ecology Laboratories (A11), School of Life and Environmental Sciences,
The University of Sydney, New South Wales 2006, Australia
9School of Aquatic and Fishery Sciences, University of
Washington, Box 355020, Seattle, WA 98195, USA
10State Key Laboratory of Marine Environmental Science, College of
Ocean and Earth Sciences, Xiamen University, Xiamen, China
11School of Biological and Marine Sciences, University of
Plymouth, Drake Circus, Plymouth, PL4 8AA, UK
12School of Ocean and Earth Science, University of Southampton,
National Oceanography Centre Southampton, SO14 3ZH, UK
13The Marine Biological Association of the United Kingdom, The
Laboratory, Citadel Hill, Plymouth, PL1 2PB, UK
14Georges River Council, Hurstville, New South Wales, 2220, Australia
15Faculty of Architecture, the University of Hong Kong, Pokfulam, Hong Kong, China
16Object Territories, Hong Kong, China
17Simon F S Li Marine Science Laboratory, School of Life Sciences and Earth System
Science Programme, The Chinese University of Hong Kong, Shatin, Hong Kong, China
18Davidson Laboratory, Stevens Institute of Technology, Hoboken, NJ 07030, USA
19ECOncrete Tech LTD, PO Box 51528, Tel Aviv, Israel
20School of Biological, Earth and Environmental Science and Centre for Marine Bio-
Innovation, University of New South Wales, Sydney, New South Wales 2052, Australia
21Sydney Institute of Marine Science, Mosman, New South Wales 2088, Australia
22Singapore Centre for Environmental Life Sciences Engineering,
Nanyang Technological University, 637551, Singapore
23Graduate School of Agriculture, Ehime University, Japan
24State Key Laboratory of Marine Pollution (City University of Hong Kong), Tat Chee Avenue,
Kowloon, Hong Kong, China
*Corresponding author: Kenneth M.Y. Leung
Human population growth and accelerating coastal development have been the drivers for
unprecedented construction of articial structures along shorelines globally. Construction has
been recently amplied by societal responses to reduce ood and erosion risks from rising sea
levels and more extreme storms resulting from climate change. Such structures, leading to highly
modied shorelines, deliver societal benets, but they also create signicant socioeconomic and
environmental challenges. The planning, design and deployment of these coastal structures should
aim to provide multiple goals through the application of ecoengineering to shoreline development.
Such developments should be designed and built with the overarching objective of reducing
negative impacts on nature, using hard, soft and hybrid ecological engineering approaches. The
design of ecologically sensitive shorelines should be context-dependent and combine engineering,
environmental and socioeconomic considerations. The costs and benets of ecoengineered shoreline
design options should be considered across all three of these disciplinary domains when setting
objectives, informing plans for their subsequent maintenance and management and ultimately
monitoring and evaluating their success. To date, successful ecoengineered shoreline projects have
engaged with multiple stakeholders (e.g. architects, engineers, ecologists, coastal/port managers and
the general public) during their conception and construction, but few have evaluated engineering,
ecological and socioeconomic outcomes in a comprehensive manner. Increasing global awareness
of climate change impacts (increased frequency or magnitude of extreme weather events and sea
level rise), coupled with future predictions for coastal development (due to population growth
leading to urban development and renewal, land reclamation and establishment of renewable energy
infrastructure in the sea) will increase the demand for adaptive techniques to protect coastlines. In
this review, we present an overview of current ecoengineered shoreline design options, the drivers
and constraints that inuence implementation and factors to consider when evaluating the success
of such ecologically engineered shorelines.
Introduction and the history of ecological
engineering of shorelines
Humans have been altering coastlines for millennia (Thompson etal. 2002, Dugan etal. 2011, Ellis
2015, Loke etal. 2019), usually with little regard for the environment. Natural habitats have been
replaced with articial structures to create access, reclaim land for agriculture, industry, transport,
residential use and tourism (Gittman et al. 2015, Lai etal. 2015, Chee etal. 2017) and protect
growing coastal populations (Crossett etal. 2004, Nicholls etal. 2007, Firth etal. 2016b) from
erosion and ooding (Burcharth etal. 2007, Kittinger & Ayers 2010, Hinkel etal. 2014). Impacts on
coastal ecosystems from such developments are inevitable and range from chronic to catastrophic
(Airoldi etal. 2005a, Martin etal. 2005, Bulleri & Chapman 2010, Dethier etal. 2016, Bishop
etal. 2017, Heery etal. 2017). Improving environmental conditions along articial shorelines is
critical to compensate for impacts from ocean sprawl—the proliferation of articial structures in the
marine environment (Dafforn etal. 2015a, Dyson & Yocom 2015, Firth etal. 2016b, Bishop etal.
2017, Heery etal. 2017). With climate change and sea level rise, existing shoreline protections are
increasingly proving ineffective and will need substantive reassessment and reconstruction during
the coming decades (Hawkins & Cashmore 1993, Nicholls & Tol 2006, Hallegatte etal. 2013, Hinkel
etal. 2014, Hoggart etal. 2014, Smith etal. 2017a).
The goal of this review is to provide a framework for selecting, applying and tracking the
success of ecologically engineered (hereafter ‘ecoengineered’) shoreline strategies (Mitsch 2012)
for intertidal and shallow subtidal marine environments. The fundamental aim of ecoengineered
shorelines is to build more inclusive, resilient and safe coasts for people and nature, which maximise
benets for ecosystems, society and economies (Airoldi etal. 2005a, Dafforn etal. 2015a). Because
it is a relatively young discipline and innately experimental in its current form, ensuring that it is
developed and applied responsibly requires that parameters and metrics for its success be clearly
dened and monitored (Saleh & Weinstein 2016, Mayer-Pinto etal. 2017). In this review, we take
a multidisciplinary approach in evaluating the benets and challenges of ecoengineered shorelines
and discuss metrics to evaluate success originating in the elds of coastal ecology, engineering,
sociology, economics, urban planning and architecture.
Ecoengineered shorelines are in a sense the most recent stage in the development of marine
infrastructure and coastal armouring, which can be traced back to before the Common Era (i.e. BC)
in Egypt (Loke etal. 2019). Early ports were located in sheltered bays, river mouths and lagoons, with
simple jetties and, eventually, low-crested breakwaters used in combination with natural habitats for
coastal protection (Polanyi 1963, Hoyle 1989, Charlier etal. 2005). In some regions, such as China
and parts of the Mediterranean, technological advances led to more extensive articial shorelines
relatively early (Franco 1996). For instance, with the discovery of pozzolanic ash hydraulic cement,
the Romans started extensive underwater engineering and managed to construct solid breakwaters
to protect fully exposed harbours (Jackson etal. 2017). By the Middle Ages in Europe, planting and
dune belt protection were well established, as were strict environmental regulations. For instance,
legal documents of 1282 and 1339 in Venice prohibited the cutting of coastal trees, picking mussels
and removing sand or vegetation from beaches or dunes (Grillo 1989). In the Renaissance, Leonardo
da Vinci championed the credo of ‘working with Nature’ (‘ne coneris contra ictum uctus: uctus
obsequio blondiuntur’—Nature should not be faced bluntly and challenged, but wisely circumvented)
(Franco 1996). A similar idea was championed by Andries Vierlingh in his manuscript entitled
Tractaet Dyckagie, which remains an important text regarding fundamental errors in land and water
engineering management (Vierlingh etal. 1920).
The formal framing of the concept of ecoengineering is generally attributed to H. T. Odum,
who proposed that ecosystems provide the biological ‘design’ and ‘energy’ required to ‘engineer’
socioeconomic benets of human development (Odum 1975). He stated that ‘the management of nature
is ecological engineering, an endeavour with singular aspects supplementary to those of traditional
engineering’ (Odum 1971). Since then, the concept has gradually expanded, replacing purely ‘ecological
processes’ with ‘natural processes’ to now include physical and other factors such as ‘design’ and
‘energy’ inputs (e.g. the ‘sand engine/sand motor’ approach for beach nourishment and ood protection
that is used along the Dutch coastline; Mulder & Tonnon 2011, Stive etal. 2013, de Schipper etal. 2016).
During the last 10–15 years, the practitioners involved in designing coastal defences have embraced a
‘design with nature’ approach, which has led to various philosophies and programmes for the delivery
of ecologically minded coastal and maritime infrastructure (e.g. the programme of ‘working with
nature’ organised by the World Association for Waterborne Transport Infrastructure (PIANC): https:// US Army Corps of Engineers’ ‘engineering with nature’: https:// EcoShape’s ‘building with nature’:
This transition in perception and practice has been facilitated by the widening recognition of the
pervasive and persistent anthropogenic changes that characterise urban areas (Bulleri & Chapman
2010, Firth etal. 2016b, Mayer-Pinto etal. 2018, Bugnot etal. 2019) and by the emergence of the
‘novel ecosystem’ concept, which emphasises novelty and human agency in the emergent ecological
communities that are without historical precedent (Milton 2003, Hobbs etal. 2006). The novel
ecosystem concept is not without controversy (e.g. Woodworth 2013, Murcia etal. 2014); however,
in novel urban habitats specically, it provides a framework for conceptualising coastal engineering
projects as opportunities for the amelioration of ecosystems that are already heavily shaped by
anthropogenic activity (Perkins etal. 2015, Aguirre etal. 2016). While the distinction is not always
made, urban areas contain both hybrid ecosystems that may have signicant novelty, as well as novel
ecosystems, which, by denition, cannot return to hybrid or historical conditions (Hobbs etal. 2006).
The novel ecosystem concept has arguably changed the perceived value of the urban environment,
particularly in relation to its provisioning of ecosystem services to urban populations, ultimately
facilitating the integration of a broader range of ecological, engineering, social and economic
interests into shoreline development and management (Perring etal. 2013).
There is considerable overlap between the elds of ecoengineering and ecological restoration.
Ecological restoration aims to assist the recovery of degraded ecosystems and put back their
attendant services (Society for Ecological Restoration International Policy Position Statement
(SER 2004). The traditional view of restoration envisaged returning an ecosystem to a predened
historical state (Hobbs & Norton 1996). Given ecosystem variability and complexity, coupled with
the extent of environmental change that has occurred over recent decades, making target baselines
uncertain (Palmer et al. 2016), restoring ecosystems to their former state is often impossible on
human timescales (Hobbs etal. 2006). Thus, many have emphasised the importance of dening
restoration to a dened target state, which is likely to be far from the original pristine condition
of the ecosystem (Hawkins etal. 1999, 2002, SER 2004, Geist & Hawkins 2016, Palmer etal.
2016). Even our best efforts serve primarily to rehabilitate or repair damaged biodiversity and
ecosystem processes and services, rather than comprehensively restore ecosystems (e.g. in a marine
context, ‘nudging nature’—Hawkins etal. 1999, 2002). Conversely reallocation is the process of
assigning an ecosystem to a new use that may not necessarily bear an intrinsic relationship to the
structure or functioning of the predisturbed ecosystem (Aronson etal. 1993). Reconciliation is the
process of modifying and diversifying anthropogenic habitats to harbour a wide variety of native
species (Rosenzweig 2003a; for further discussion, see the section entitled ‘Links to theoretical
and community ecology’, later in this review). While some refer to restoration as the best form of
ecoengineering (Bradshaw 1997), others claim that restoration encompasses ecoengineering (SER
2004). Here, we propose that ecoengineering is a broad approach that can aid restoration (Mitsch
2012), but more typically is used for rehabilitation, reallocation and/or reconciliation.
During the last decade, ecoengineered shorelines have emerged as a promising alternative
to traditional coastal development (Mitsch 2012), although they are by no means a panacea for
countering the negative impacts of shoreline construction, and numerous uncertainties regarding
their efcacy remain (Bouma etal. 2014, Sutton-Grier etal. 2015). We emphasise that these are not
a substitute for natural or even seminatural systems. Neither should such ecoengineering approaches
be used to legitimise developments in sensitive habitats. In this review, our focus is on coastlines that
have already become urbanised or heavily modied by land reclamation for agriculture, industry or
transport infrastructure, including low-density, residential suburban sprawl. The approaches suggested
here are particularly suitable when current coastal developments are being expanded or repaired.
While there are examples of ecoengineered shorelines around the globe and many under development
(Chapman & Underwood 2011, Firth etal. 2014, Elliott etal. 2016, Munsch etal. 2017a), a cohesive
framework for developing ecoengineered shorelines and evaluating their success is presently
lacking (but see relevant work by van Slobbe etal. 2013, van der Nat etal. 2016, Osorio-Cano etal.
2017, Gracia etal. 2018, Whelchel etal. 2018). This review presents the underpinning concepts,
key approaches, recent developments, limitations and future trajectory of ecoengineered shoreline
design. We presentaframework for the development of ecoengineered shorelines by reviewing three
themes: (1) core principles and approaches to ecoengineered shoreline design, (2)factors that inuence
implementation and (3) how to evaluate the success of ecoengineered shorelines.
Design considerations for ecoengineered shorelines
Links to theoretical and community ecology
Ecological concepts and theory provide both a vocabulary and a predictive framework to understand
and manage ecosystems along the spectrum from near-pristine to highly degraded ecosystems. Thus,
theory can inform policy, planning and practice. Here, we briey introduce ecological knowledge
relevant to the design of ecoengineered shorelines. We start with species-area relationships and the
relationship between biodiversity and ecosystem functioning. We then discuss diversity at different
scales, succession and disturbance, environmental gradients and connectivity. The aim of this section
is simply to highlight some of the key theoretical underpinnings of the rapidly emerging eld of
designing ecologically sensitive shorelines.
Species-area relationships
Species-area relationships (SARs)—which involve one of the few widely accepted laws in ecology
(Lawton 1999)—dictate that a larger area will contain a larger number of species. This pattern is
observed at almost every spatial scale (Figure 1; Rosindell & Cornell 2007, O’Dwyer & Green
2010). It is manifest in highly modied shorelines, where the loss of natural habitat leads to direct
loss in species (Rosenzweig 2003a). The same principle may be applied in reverse, however, to
aid modifying, redesigning and diversifying anthropogenic habitats. Reconciliation ecology, which
focusses on how urban and novel habitats may be used to maintain biodiversity and to provide
ecosystem services, provides several examples of this (Hobbs etal. 2013, Perring etal. 2013), in both
terrestrial (Rosenzweig 2003b, Kowarik 2011) and aquatic (Mitsch & Jørgensen 2004, Firth etal.
2016b) environments. While species colonisation in some of these examples occurs by accident (refer
to Rosenzweig 2003b), many other examples are carefully planned and managed (see the section
entitled ‘Case studies and scaling up’, later in this review).
Figure 1 Triphasic SAR. A larger area will include a larger number of species. On a local scale, alpha (α)
diversity describes the number of species within a habitat, while the compositional difference in species
assemblages among habitats (beta, β, diversity) is essential for achieving high gamma (γ) diversity in a region.
Diversity at multiple scales
Diversity is a key theme in ecology, which underpins both the goals and assessment of many
ecoengineered shoreline projects. It is frequently estimated as species richness (i.e. the total number
of species present) and can be quantied at multiple spatial scales. While species richness within
a habitat (alpha diversity) is useful in many instances, the compositional difference in species
assemblages among habitats (beta diversity) is essential for achieving high diversity in a region
(gamma diversity; Figure 1; Whittaker 1960, 1972). Specifying the scale at which ecoengineered
shoreline designs aim to enhance diversity is crucial, as strategies for improving alpha and beta
diversity may differ. Designs focussed on enhancing large-scale (landscape and regional scale
hundreds of metres—tens of kilometres) spatial heterogeneity and creation of various medium-scale
habitat types (1–100 m) will have a positive effect on beta diversity. Those strategies focussed on
increasing local-scale (<1 m) topographic and/or structural complexity (e.g. Coombes etal. 2015,
Loke and Todd 2016) and ameliorating abiotic stressors (e.g. Browne & Chapman 2011, Evans etal.
2016) are more likely to have a positive effect on alpha diversity. Patches of more complex habitat,
especially if they are of different types, among a background landscape of less complex habitats will
also increase beta diversity (Firth etal. 2014, Evans etal. 2016).
Biodiversity and ecosystem functioning
Over the last two decades, there has been a growing realisation that loss of species can compromise
ecosystem functioning (e.g. biomass production, nutrient uptake; Loreau etal. 2002, Tilman etal. 2014,
Oliver etal. 2015) and in turn, ecosystem services (e.g. carbon storage; Díaz etal. 2006, Cardinale
etal. 2012, Isbell etal. 2014). Thus, increasing diversity would enhance the functioning of ecosystems
by elevating production and nutrient cycling, as well as conveying resilience to environmental change
(the insurance hypothesis; Yachi and Loreau 1999) and invasion by nonnative species (Stachowicz
etal. 2002, Arenas etal. 2006, Tan etal. 2018). These processes lead to enhanced goods and services
provided by biodiversity and ecosystems (i.e. provisioning, supporting, and regulating; Raffaelli etal.
2002). For both sedimentary (e.g. Solan etal. 2004) and rocky shore habitats (e.g. O’Connor & Crowe
2005, Grifn etal. 2008, 2010), experimental work has shown that species diversity and identity
inuence ecosystem processes such as productivity and nutrient cycling. The importance of habitat
patch diversity has also been recognised for some time (e.g. Giller etal. 2004, Hawkins 2004) and
more recently has been experimentally demonstrated (e.g. Grifn etal. 2009, Godbold etal. 2011,
Alsterberg etal. 2017). Knowledge of the traits of species (functional diversity) can also be valuable
in translating structural attributes of biodiversity into the maintenance or restoration of ecosystem
processes, and hence services (Rigolet etal. 2014). Ultimately, ecoengineering aims to maximise the
ecosystem services that can be delivered in seminatural and highly modied ecosystems (Dafforn
etal. 2015a). Paying attention to increasing habitat patch diversity, and hence beta species diversity,
is likely to pay greater dividends in terms of ecosystem functioning and provision of services.
Environmental gradients and connectivity
Environmental gradients are a major inuence on community composition and structure in shoreline
environments (see Raffaelli & Hawkins 1996 for a review). The major gradients can be viewed as
being either vertical (e.g. tidal elevation/bathymetry) or horizontal (e.g. exposure to wave action).
These gradients ultimately determine what species can potentially survive in an area, with biological
interactions usually proximately determining the actual assemblage in a location [e.g. experimental
work by Jonsson etal. (2006) on the respective roles of wave action and grazing on breakwaters].
Disturbance regimes also vary along these gradients, but the strength and direction of these effects are
frequently modied by local contingency (Underwood & Chapman 1996, Chapman etal. 2010, Bulleri
etal. 2011), including small-scale topography (Johnson etal. 1998, 2003, Berntsson etal. 2000) and the
vagaries of settlement by algal propagules and the larvae of invertebrates (Butman 1987, Underwood
& Fairweather 1989, Fletcher & Callow 1992, Knights etal. 2012). Designs that minimise the extent
to which environmental gradients are altered or intercepted are less likely to fundamentally change
community structure. For example, installing hard coastal armour in an area naturally typied by
coarse sand or gravel will inevitably lead to scouring of any colonising rocky shore species (Moschella
etal. 2005), preventing communities from developing beyond the early successional stages.
Articial structures often support assemblages analogous to rocky reefs jutting out from high-
energy sandy beaches (Bally etal. 1984) that tend to be dominated by ephemeral early successional
species. Additionally, designs perform best when they account for ecological connectivity at various
scales. Changes to the movement of organisms and resources such as detritus should be avoided
(Huijbers etal. 2013, Bishop etal. 2017). Arrays of structures scaling up to affect whole coastlines
(e.g. Ma etal. 2014, Dong etal. 2016) can act as stepping stones for the spread of native species
responding to climate change (Johannesson & Warmoes 1990, Mieszkowska etal. 2006, Keith etal.
2010, Firth etal. 2015, Huang etal. 2015, Dong etal. 2016), as well as help invasions by nonnative
species (Floerl etal. 2009, Mineur etal. 2012, Airoldi etal. 2015, Bishop etal. 2017).
Succession and disturbance
Novel habitats are colonised by a variety of marine organisms (Wahl 1989). When relatively constant
conditions prevail, communities in novel habitats gradually shift in composition and structure
through the processes of immigration and extinction (MacArthur & Wilson 1967). Connell & Slatyer
(1977) proposed three models of succession: the classical model of positive facilitation of later
stages by earlier colonisers, inhibition of later stages by earlier arrivals and a neutral tolerance
model reecting arrival of propagules and longevity of species. Signicant disturbances, whether
natural or anthropogenic, tend to push a successional sequence back to earlier stages (i.e. system
retrogression; Odum 1985). Biological interactions such as grazing have been shown to break
inhibition by intermediate phases such as green algae (Sousa 1979).
Classical ecological theory predicts the highest levels of diversity under moderate levels of
disturbance, either in terms of intensity or frequency [the intermediate disturbance hypothesis (IDH);
Connell 1978], as a low-disturbance regime leads to dominance by competitive-dominant species and
strong frequent disturbances suppress most species to also result in low overall diversity (e.g. Sousa
1979). Although IDH has been challenged on both a theoretical and empirical basis (Lubchenco 1978,
Fox 2013), it is well established that high levels of disturbance lead to depauperate systems comprising
relatively few early successional or opportunistic taxa (Pulsford etal. 2014). As disturbances are
commonplace in human-dominated environments, ecoengineering strategies that help stabilise the
disturbance intensity or frequency may be benecial (Airoldi etal. 2005b, Bulleri & Airoldi 2005,
Airoldi & Bulleri 2011, Bracewell etal. 2013, Salomidi etal. 2013), increasing the likelihood that
communities progress to a climax (Hawkins etal. 1983) or oscillate to and from intermediate stages.
Additionally, species or groups of species that positively inuence succession can be integrated
into ecoengineered shorelines through seeding (e.g. Perkol-Finkel etal. 2012, Ferrario etal. 2015, Ng
etal. 2015) or targeted designs (Strain etal. 2017a). Studies on the mechanisms of succession have
highlighted the importance of certain species interactions in determining the rate and direction of
community development (Maggi etal. 2011). Positive interactions such as facilitation [e.g. settlement
by mussels or tubicolous polychaetes enhancing the colonisation of meiofauna on hard surfaces
(Dubois etal. 2002, O’Connor & Crowe 2007) or increased accumulation of leaf litter by mangrove
on previously bare substrata, which aids in recruitment of juvenile nekton] may be particularly
important (Bulleri etal. 2018). Organisms that serve as ecosystem engineers and foundation species
also have profound impacts on the way that communities develop and, ultimately, on diversity as well
(e.g. Coleman & Williams 2002, Tolley & Volety 2005, Marzinelli etal. 2014).
Biological invasions
Finally, theoretical and community ecology concepts from invasion biology are also relevant for
ecoengineered shoreline design, as articial structures appear particularly prone to colonisation
by nonnative species (Bulleri & Airoldi 2005, Glasby etal. 2007, Vaselli etal. 2008, Ruiz et al.
2009, Dumont etal. 2011, Dafforn etal. 2012, Mineur etal. 2012, Simkanin etal. 2012, Airoldi
etal. 2015). These concepts are strongly interrelated with many of those already presented in this
review, particularly diversity, connectivity, succession and disturbance. Articial structures provide
novel colonisation opportunities for nonnative species (Dafforn etal. 2012). Those located in urban
areas may be subject to particularly high levels of propagule pressure (‘introduction effort’), given
their close proximity to known invasion hubs such as marinas and shipping facilities (Seebens etal.
2013), which is thought to strongly inuence invasion risk (Lockwood etal. 2005). Urban articial
structures may also be subject to particularly high rates of disturbance (Bulleri & Airoldi 2005, Piola
& Johnston 2008, Kenworthy etal. 2018) or to relatively low levels of predation by mobile consumers
(Simkanin etal. 2013, Astudillo etal. 2016, 2018, Rogers etal. 2016), ultimately favouring nonnative
taxa. Further, additional structural complexity, which is a key component of many ecoengineered
shoreline projects (discussed further in section ‘Environmentally sensitive hard defences, later in
this review), can modify predator-prey interactions and potentially facilitate nonnative prey species
by enhancing refugia from native predators (Barrios-O’Neill etal. 2014).
Fundamental design options
The construction of articial structures such as seawalls, groynes and breakwaters has historically
been the default approach to coastal protection (Cooper etal. 2016). Articial structures are
economically and environmentally costly (Jones 1994, Airoldi etal. 2005a, Kittinger & Ayers 2010,
Hinkel etal. 2014), however, contributing to the growing interest in alternatives that compensate for
the negative impacts of traditional coastal infrastructure and provide multiple functions (Chapman &
Underwood 2011, Francis & Lorimer 2011, Mitsch 2012, Evans etal. 2017). These alternatives can be
categorised into three fundamental approaches: (1) building in design features in new or modifying
existing articial structures (termed hard ecological engineering; Chapman & Underwood 2011,
Firth etal. 2014); (2) replacing articial structures with sediments, vegetation and/or other habitat-
forming organisms (called soft ecological engineering; Temmerman etal. 2013, Morris etal. 2018b)
and (3) applying both soft and hard engineering approaches in combination (known as hybrid
ecological engineering; Chapman & Underwood 2011, Bilkovic & Mitchell 2013) (Figure 2).
Many terms are used to refer to both traditionally and ecologically engineered shores. Common
synonyms for articial structures and coastal defences include coastal infrastructure, shoreline
armouring, hard structures and urban structures. The term Ecoengineered shorelines is used to
describe shorelines that have been developed or retrotted using ecoengineering principles, via
either hard, soft or hybrid approaches. Created or restored habitats are those that have undergone
soft ecoengineering, and are commonly referred to as nature-based coastal defence (van der Nat
Figure 2 Approaches to ecoengineered shorelines include (a) hard (tile-enhancement units installed along the
seawalls at Changi, Singapore), (b) hybrid (rock llet with saltmarsh in Chesapeake Bay, United States) and (c)
soft ecological engineering (bagged oyster shell reef with saltmarsh in Chesapeake Bay).
etal. 2016, Osorio-Cano etal. 2017, Gracia etal. 2018, Sutton-Grier etal. 2018). In the United States,
soft and some hybrid ecological engineering projects also fall under the commonly used term living
shorelines (Bilkovic & Mitchell 2017).
Although the overall objective of ecoengineering is similar regardless of approach (i.e. to build
shorelines for the benet of both humans and nature; Mitsch 2012), there is a fundamental difference
between hard and soft ecoengineering from a coastal defence perspective (Morris etal. 2018b).
The goal of hard ecoengineering is usually to compensate for the negative impacts of articial
structures (e.g. seawall or breakwater) through enhancing biodiversity and ecological functioning
while maintaining their physical integrity (e.g. Moschella etal. 2005, Burcharth etal. 2007, Chapman
& Underwood 2011, Firth etal. 2014, Pioch etal. 2018). Conversely, while soft ecoengineering is
usually proposed from a conservation and/or sustainability perspective (Dafforn etal. 2015a, Mayer-
Pinto etal. 2017), the created or restored habitat needs to provide adequate coastal defence in addition
to ecological and landscape values (e.g. biodiversity enhancement) if this technique is to replace or
complement articial structures. Although hard ecoengineering is less challenging to implement
from a coastal defence perspective, there is growing evidence that created or restored habitats using
soft engineering can also provide protection that is equivalent to or better than traditional engineered
hard structures (Gittman etal. 2014, Smith etal. 2017a, Smith etal. 2018). As mentioned previously,
soft engineering approaches and hard structures may even operate synergistically to enhance coastal
protection (Smallegan etal. 2016, Vuik etal. 2016)—as successful hybrid schemes in some cases.
Goals and objectives
Dening the objectives or goals is considered the most fundamental step in any restoration (e.g.
Hobbs & Norton 1996, SER 2004, McDonald etal. 2016, Palmer etal. 2016) or ecoengineering
project (Mayer-Pinto etal. 2017). Ecoengineered shorelines are multifunctional and are valued for
their potential to accrue benets across multiple stakeholders (Evans etal. 2017), each of whom has
their own set of objectives (e.g. economic, engineering, social, ecological) that typically drive or
inuence the objectives of a project. Sometimes these objectives overlap—for instance, creating a
wide beach could also generate space for recreation, protect inland infrastructure and form habitat
for nesting shore birds (Temmerman etal. 2013). Due to the way in which projects are conceived and
funded (see the section entitled ‘Implementation of ecoengineered shorelines’, later in this review),
one or more of these objectives may take precedence over the others, which can lead to trade-offs.
Nevertheless, multifunctionality is a key goal of ecoengineered shorelines, and thus engineering,
ecological and socioeconomic outcomes should be optimised. Evaluation of the success or failure
of projects needs to focus on whether these multiple objectives are achieved (see the section entitled
‘Evaluating the ecoengineered shorelines approach in practice’, later in this review).
Identifying the approach
Ecoengineered shorelines range considerably in form and function and may be conceptualised
along a spectrum (Moosavi 2017), from created (i.e. a habitat not historically present) or restored
natural habitats at one end to ecologically engineered hard structures at the other. The most suitable
approach for a site will depend on socioeconomic variables and the physical, environmental and
the ecological contexts (Figure 3; Spalding etal. 2014). Shorelines should be defended only where
there is something to protect (Airoldi etal. 2005a). This is usually property or infrastructure, or
in some instances, ecological systems (Moody etal. 2016) or archaeological or cultural heritage
(Reimann etal. 2018) threatened by inundation or erosion (Figure 3). Land reclaimed for agriculture
via empoldering almost always needs some form of protection (Woltjer & Al 2007). Any form
of shoreline construction or modication will pose inherent risks to marine communities and the
coupled natural-human systems of which they are a part. The need for shoreline protection, therefore,
must be sufciently great to warrant assuming such risks (Airoldi etal. 2005a).
In general, selecting from soft ecoengineering, a hybrid approach and environmentally sensitive hard
defences requires consideration of the physical environment (particularly wave conditions; Leonardi etal.
2016) and the respective cost of each approach (Arkema etal. 2017) in the context of the main purpose
of the project (e.g. creating a bathing beach or a surng bay, as opposed to protecting an agricultural
land or an ecosystem of conservation interest). At low-wave-energy sites, soft and hybrid ecoengineering
approaches are often feasible (Leonardi etal. 2016, 2018, Sanford & Gao 2018). At higher-energy sites,
with signicant wave action from either wind or shipping, certain types of soft engineering may not persist
over time or provide adequate shoreline protection (Bouma etal. 2014). Such cases necessitate hybrid
approaches, or at the highest energy end of the spectrum, hard defences (Moosavi 2017).
Despite growing evidence suggesting that soft and hybrid approaches are often cost effective
and sufcient for protection from ooding and erosion (Spalding etal. 2014, Gittman & Scyphers
2017, Sutton-Grier etal. 2018), considerable uncertainties (both perceived and realised) remain
surrounding their efcacy as coastal defences (Gedan etal. 2011, Perkins etal. 2015, Sutton-Grier
etal. 2015, Arkema etal. 2017). Thus, at least some degree of hard ecoengineering is often required
(Cooper etal. 2016). The value and/or importance of the property, infrastructure or ecosystem to
be protected will have some bearing on the choice of approach (soft, hard or hybrid). For example,
if the asset to be protected is essential to national or regional infrastructure, such as an airport,
hospital or railway, there may be less willingness to accept (WTA) approaches perceived to carry
greater uncertainty with respect to inundation and ood risk. Often, there is stronger support for
hard engineering approaches (Gray etal. 2017) that have been more thoroughly tested over time
(Sutton-Grier etal. 2015), even when they are unnecessary or less cost effective than soft or hybrid
alternatives (Scyphers etal. 2014, Gittman & Scyphers 2017). In contrast, where the asset is less
crucial for social and economic health or primarily of local signicance, there may be a higher
chance for adopting more experimental (and potentially riskier) ecoengineered shoreline approaches
that provide protection under most (but not the most severe) storm events, or for which the evidence
of efcacy in terms of shoreline protection is limited (but see Gittman etal. 2014, Smith etal. 2018).
Figure 3 A owchart for identifying the suitable approach to ecoengineered shoreline protection. Blue
diamonds indicate a decision (dark blue is the starting point), and yellow boxes indicate the ecoengineered
shoreline approach to take. Biogenic habitat refers to natural or restored (i.e. soft ecoengineering) habitats.
Environmentally sensitive hard defence refers to hard ecoengineering approaches.
Which ecoengineered shoreline approaches can and should be adopted will largely be inuenced
by the degree to which the environment has already been modied by anthropogenic activities
(Bouma etal. 2014). In heavily modied environments, environmental degradation (e.g. by pollution
and land reclamation) may result in conditions that are no longer suitable for the establishment of
soft or hybrid ecoengineered shorelines (Chee etal. 2017), necessitating use of hard defences, at least
in the interim, if cleanup efforts or long-term mitigation of environmental conditions are part of the
development plan. The implementation of soft or hybrid approaches may also be restricted by the lack
of available space, particularly in heavily urbanised settings (Saleh & Weinstein 2016). In contrast, in
environments where intact and functional habitat-forming species are present, the priority should be
to protect these and, where necessary, enhance their efcacy in protecting shorelines by increasing
their habitat area or encouraging growth forms that are more effective in stabilising sediments,
dissipating wave energy or both. Particularly at sites where the natural habitat type is sedimentary
(e.g. seagrass beds, mangrove forests, saltmarshes, and dune systems), use of hard defences should
be applied only where absolutely necessary, as they result in a qualitative state change from a soft- to
hard-bottomed systems (Airoldi etal. 2005a).
Furthermore, the selected approach will also depend on site-specic factors such as tidal range and
whether the shoreline is natural or reclaimed land. These factors will inuence the space available for
interventions, as well as the size and conguration required for an intervention to protect from erosion
and ooding effectively (Stark etal. 2016). Importantly, as sea level rise continues to shift intertidal
zones landward through coastal squeeze (e.g. Jackson & McIlvenny 2011, Torio & Chmura 2015, Luo
etal. 2018), ecoengineering approaches utilising intertidal species will also need the necessary space
(and capacity) for the species to retreat (Bilkovic etal. 2016). Where the tidal range is small, or the
site is the result of land reclamation, there may be little intertidal area for ecoengineering approaches
utilising littoral vegetation (Strain etal. 2017b), necessitating the use of subtidal species (such as
shellsh, macroalgae or coral reefs) for shoreline stabilisation (Temmerman etal. 2013).
Environmentally sensitive hard defences
Where hard ecoengineering approaches are deemed most appropriate, several factors typically
inuence the specics of their design. The options for environmentally sensitive hard defences will
depend on whether the structure is being designed de novo or whether an enhancement of an existing
structure is planned.
New structures should be designed to maximise the intertidal surface area available for
colonisation and to include habitat structural complexity (e.g. introducing microhabitats) that increase
diversity (Loke etal. 2014, 2017) and provide refuges for target species (or their prey resources) from
predators and environmental stressors (Moschella etal. 2005, Chapman & Underwood 2011, Loke
etal. 2015). A greater and more heterogeneous intertidal area may be achieved by stepping structures
(e.g. Barangaroo, in Sydney, Australia), sloping structures (i.e. a 45° angle as opposed to 90°), adding
macroscale habitat features such as horizontal ns (e.g. Seattle seawall; for more information, see the
section entitled ‘Case studies and scaling up’, later in this review) or by integrating holes of varying
sizes as shelter for sh (Sella & Perkol-Finkel 2015) or invertebrates (Martins etal. 2010, Witt etal.
2012). Porous structures not only are advantageous from an engineering perspective in dampening
waves (Burcharth etal. 2007), but also provide much habitat for colonisation, even if much of the
additional biodiversity is hidden (Sherrard etal. 2016).
Microhabitats can be incorporated in structures by utilising heterogeneous revetment boulders
rather than smooth panelled seawalls or by casting pits, grooves, water-retaining structures and other
features into the concrete panel or block (Chapman & Blockley 2009, Chapman & Underwood 2011,
Firth etal. 2014, 2016a, Loke etal. 2016, Strain etal. 2018b). Where possible, new structures should
utilise materials that encourage the recruitment of target organisms. Recruitment will be dependent
on the surface chemistry (Pomerat & Weiss 1947, Harlin & Lindbergh 1977, Anderson 1996,
Neoetal. 2009) and roughness (microtopography, Köhler et al. 1999, Nandakumar etal. 2003,
Coombes etal. 2015) of the structure. The colour of structures may also inuence colonisation (James
& Underwood 1994, Satheesh & Wesley 2010) by affecting brightness (many larvae are negatively
phototactic; Bayne 1964, Raimondi & Keough 1990) and temperature (dark colours absorb while light
colours reect heat, Kordas etal. 2015, McAfee etal. 2017), although this has received little research
attention. Use of local quarried rock may provide a substratum that is physically and chemically most
similar to natural rocky shores (Green etal. 2012). Alternatively, the chemical composition of concrete
may be altered to encourage recruitment (Sella & Perkol-Finkel 2015, McManus etal. 2018).
Options for existing structures will be more limited. They may include subtractive approaches
where microhabitats and area are added by drilling holes, pits and grooves in structures (Chapman &
Underwood 2011) or coring larger rock pools (Evans etal. 2016). Alternatively, additive approaches
include retrotting articial structures with units such as complex tiles (Goff 2010, Loke & Todd 2016,
Strain etal. 2017a) and water-retaining owerpots and articial rock pools (Chapman & Blockley
2009, Browne & Chapman 2011, Chapman & Underwood 2011, Firth etal. 2014, 2016a, Evans etal.
2016, Morris etal. 2017), which may be constructed from bioenhancing concrete materials (Sella
& Perkol-Finkel 2015) or the transplantation of living, habitat-forming species (Perkol-Finkel etal.
2012, Ng etal. 2015), see review by Strain etal. (2018b).
Additionally, it may be possible to add new habitats to existing seawalls by placing boulders
at the toe of each seawall (Green etal. 2012, Liversage & Chapman 2018). This also minimises
scouring of the main structure by wave driven sand, gravel and cobbles (Moschella etal. 2005).
Although some ecoengineered shorelines have been shown to reduce the proportion of nonnative
species compared with traditional articial structures (Sella & Perkol-Finkel 2015), an important
aspect of monitoring is to ensure that the ecoengineered habitats are not favouring the establishment
or spread of nonnative species (e.g. Morris etal. 2018a). Strategic choices for ecoengineered shoreline
design and management surrounding invasion prevention are detailed by Dafforn (2017) and
include techniques such as preseeding, managing for strong native grazer and predator populations
and modifying the chemical and physical properties of ecoengineering features to facilitate the
establishment of native assemblages.
The specic design of environmentally sensitive hard defences should be informed by
socioeconomic and ecological goals (see the section ‘Evaluating the ecoengineered shorelines
approach in practice’, later in this review). For example, if there is an ecological goal of biodiversity
enhancement, it might be important to incorporate a diversity of microhabitats into structures, so
as to maximise the number of niches (Strain etal. 2018b) and enhance beta diversity (cf. Kelly
etal. 2016). In contrast, if the goal is enhancing water quality through ltration, microhabitats and
chemistry that favour bivalves (or other lter feeders) should be targeted (see Strain etal. 2017a for
an example of protective microhabitats that enhance bivalve survival). Morris etal. (2018c) provide
a framework for deciding what interventions to apply in order to target particular functional groups
of sh. Design interventions can be targeted to enhance exploited species, such as limpets in the
Azores (Martins etal. 2010). Enhancing grazers will lead to less slippery and dangerous seaweeds
(Jonsson etal. 2006) in areas where structures abut beaches used for recreation.
Soft approaches and the selection of habitat-forming species
Soft ecologically engineered shorelines are typically founded on habitat-forming species. A variety
of these species may be suitable for shoreline stabilisation, including (but not limited to) reef-
forming invertebrates such as corals (Ferrario etal. 2014), oysters (Scyphers etal. 2011), mussels and
worms (Moody 2012), intertidal vegetation such as mangroves and saltmarsh (Kumara etal. 2010,
Gittman etal. 2014) or subtidal vegetation, such as seagrass or kelp (Dubi & Tørum 1994, Ondiviela
etal. 2014). The choice of habitat-forming species for shoreline protection projects will depend on
physicochemical, ecological and socioeconomic factors. First, the selected species must be able to
persist, grow and reproduce under both present and projected future environmental conditions, given
ongoing climate change and coastal development (Walles etal. 2016). Second, the species must
be capable of forming habitat that is able to adequately protect shorelines through stabilisation of
sediments or dampening of wave energy under present and projected future environmental conditions
(Narayan etal. 2016, Morris etal. 2018b).
Where possible, native species should be selected over nonnative species. However, in rare
instances, nonnative species may be more suitable choices due to extirpation of their native
functional analogues or their greater capacity than natives to tolerate future environmental change.
In the Netherlands, where native at oysters (Ostrea edulis) are now functionally extinct (Beck etal.
2011) and native mussel beds (Mytilus edulis) have become largely overgrown by nonnative Pacic
oysters, Crassostrea gigas (Markert etal. 2010), the nonnative oyster is being utilised as a device
for shoreline stabilisation. This is because it is more suited to the present environment and is, to a
large degree, functionally equivalent to the native species (da Vriend etal. 2014). Interestingly, an
experimental study by Borsje etal. (2011) found that nonnative oyster beds were more effective in
wave attenuation than were native mussel beds. Nonnative species, however, should not be introduced
for shoreline protection purposes if they are not already naturalised in the targeted environment, or if
their presence will lead to negative ecological or socioeconomic impacts that outweigh their benets
(Bunting & Coleman 2014). Ultimately, the habitat-forming species need to be self-sustaining. The
probability of this may be maximised where there are nearby patches of extant natural habitat to
which the population being used for ecoengineering is connected.
Socioeconomic considerations should also be taken into account when selecting species. For
example, shellsh reefs may be a less desirable alternative for ecoengineering in areas subject to
high human recreational activity due to risk of cuts and injury (and potential infection) from razor-
sharp shells (O’Donnell 2016). Similarly, mangroves may be a socially unacceptable alternative if they
facilitate the proliferation of mosquito populations, especially in areas where they are carriers of disease
(Temmerman etal. 2013, Friess 2016). Conversely, certain species can be particularly desirable because
they provide valued ecosystem goods or services. For instance, in areas where improvement of water
quality is a key goal, living shorelines based on shellsh may not only serve to stabilise shorelines,
but also improve water clarity (Allen etal. 1992, Wilkinson etal. 1996, Coen etal. 2007, Grabowski
& Peterson 2007). Where the habitat-forming species provides food or raw materials, it may need
to be protected from harvest so as not to compromise its role in protecting the shoreline (O’Donnell
2016). For example, the success of oyster reef restoration projects in the United States, including
those for shoreline protection, is enhanced by their protection within no-take reserves (Powers etal.
2009). In Southeast Asia, rewood collection may compromise the long-term persistence of mangrove
forests unless appropriate protective measures are put in place (Malik etal. 2015). Another notable
socioeconomic goal can be achieved through habitat-forming species that secrete calcareous skeletons,
such as oysters, corals and tube worms, through bioprotection of underlying rock or concrete, which
potentially increases strength and longevity of structures (Coombes etal. 2013, 2017, Coombes, 2014).
Where habitat-forming species are being established from a baseline of zero at the shoreline site
and the shoreline is already under threat from erosion or inundation, temporary protective measures
may need to be in place while the habitat establishes. An example of this is the use of cultivator pots,
protective matting and structures for mangrove seedlings (Krumholz & Jadot 2009, Tamin etal.
2011). These can help shield the habitat-forming species from erosion and stimulate establishment
through enhanced sediment accretion, as well as protecting the shoreline from erosion until the
habitat-forming species is able to attain sufcient size and/or density to full the shoreline protection
function. Knowledge of the congurations and morphologies of the habitat-forming species that best
serve the functional role of shoreline protection will assist in determining which populations of the
habitat-forming species to transplant to the site and the optimal way in which transplantation should
proceed in order to achieve the desired goal. In some instances, transplantation of multiple habitat-
forming species that exist in adjacent or nested congurations may best stabilise the shoreline and
offer enhanced ecological cobenets where they support unique species assemblages or promote
distinct ecosystem functions (Gribben etal. this issue).
Hybrid approaches
Hybrid approaches combine hard defences with natural elements, such as restoring key species or
habitats by plantings and seeding. Such approaches can provide considerable benets by helping to
support broader management and conservation goals, enhancing ecosystem services and/or providing
added protection beyond that of traditional hard structures (O’Donnell 2016). They enable ecological
communities to be incorporated into defences at sites where they would not survive on their own or
would, on their own, provide inadequate coastal protection (Sutton-Grier etal. 2015).
The conguration of natural and hard elements in a hybrid approach depends on the wave
exposure of the environment, the amount of space available for transplants and the species involved.
At sites where natural habitats persist and provide shoreline protection from moderate (but not
large) storms, they may be fringed on their landward side by hard defences, such as seawalls or
revetments, which provide the required protection from major storm events. At sites where wave
action inhibits establishment of unprotected living habitat, the hard defence may be placed on the
seaward side to protect the habitat from erosion. Low-crested rock sills are used to stabilise landward
sediments in many instances, so that saltmarsh (Benoit etal. 2007, Bilkovic & Mitchell 2013) or
mangroves (Hashim etal. 2010) can establish and grow. Hard defences and living habitat can also
be interspersed. In Florida, for example, mangroves were planted in concrete cultivars that attenuate
waves, accrete sediment and are designed to provide more favourable environments for survival and
growth (Krumholz & Jadot 2009). This technique is also currently being implemented in Victoria,
Australia (le/0018/123507/Climate-
Generally, as wave energy increases, a greater ratio of hard defence to soft defence via a living
habitat will be required. In environments with limited intertidal space, the opportunity for living
habitat may be limited to small habitat pockets among hard defences (see the pocket beach example in
Seattle, described in Toft etal. 2013). Nevertheless, the inclusion of these can still lead to signicant
biodiversity enhancement and ecosystem service provision over those provided by the hard defences
alone (Spalding etal. 2013).
Implementation of ecoengineered shorelines
The implementation of the ecoengineered shoreline approach is a complex process that is inuenced
by politics, public safety, community values and cost. Thus, it is not without potential risks and
challenges. A major concern is dealing with communicating uncertainties and the need for adequate
data (Chapman & Underwood 2011). Poor understanding of the underpinning mechanisms of common
ecoengineering techniques (Gedan etal. 2011, Bouma etal. 2014) can limit the effectiveness of
ecoengineered shoreline designs and lead to unanticipated changes in important ecosystem processes
(Bilkovic & Mitchell 2013). There may be unintended consequences of standard ecoengineering
strategies, such as enhanced structural complexity that may lead to increased accumulation of litter
by shoreline structures (Aguilera etal. 2016). There is a lack of accepted and standardised criteria
for evaluating their success (Perkins etal. 2015). A reputational risk is the misuse of ecoengineered
shoreline concepts and terminology to justify and market expanded loss and conversion of natural
habitats as ecologically desirable (Pilkey etal. 2012). However, the ecoengineered shoreline approach
also has numerous potential benets, which have already been discussed in this review, and which
have been strongly emphasised in the recent research literature (Pioch etal. 2018).
The concept of working with nature has existed for centuries, but it was little less than half a
century ago when the term ecoengineering was framed and the concept recognised (see the section
entitled ‘Introduction and the history of ecological engineering of shorelines’, earlier in this review).
Although older work on the ecology of articial habitats exists (e.g. Plymouth and Port Erin breakwaters,
United Kingdom; Southward & Orton 1954, Kain & Jones 1975) and a broader interest in the ecology
(Hawkins & Cashmore 1993) and restoration of articial marine habitats emerged in the 1980s and
1990s (e.g. Russell etal. 1983, Allen etal. 1992, 1995, Hawkins etal. 1992a,b, 1993, 1999, 2002, Allen
& Hawkins 1993), research on ecoengineering of coastal structures only began to accelerate in the early
2000s (Chapman & Underwood 2011, Perkins etal. 2015, Strain etal. 2018b). In part, this increase has
been prompted by growing recognition of the benets of the ecoengineered shoreline approach (Pioch
etal. 2018) the importance and relevance of which are particularly evident when considered within
the context of major ecological, social and economic challenges currently faced by coastal ecosystems
and human communities. It is within this context that we identify several drivers of the recent growing
interest in ecoengineered shorelines, which we have broadly classied into ve categories (discussed in
the next sections): (1) climate change and associated sea level rise and increased ood risk; (2) coastal
urbanisation, economic development and urban renewal; (3) integration in governance and policy; (4)
growing public outreach and awareness and (5) advances in science and technology.
Climate change, the associated sea level rise and increased ood risk
Anthropogenic climate change is placing coastal communities at risk; consequently, adaptation
strategies are likely to be urgently required, especially in low-elevation coastal areas (Neumann
etal. 2015, Dangendorf etal. 2017). Rising sea levels and increasingly frequent or more intense
storms are leading to coastal erosion and ooding, necessitating either relocation or defence of
low-lying settlements and infrastructure. By the end of this century, a rise of at least 0.25–0.55 m
is projected, and about 95% of the ocean area will be affected (IPCC 2014). It is estimated that in
the United States, approximately 1500 homes will be lost to coastal erosion each year for several
decades, at a cost to property owners of $530 million per year (Heinz Centre 2000). These impacts
of climate change are exacerbated by continued population growth in the coastal zone (Nicholls etal.
1999). Worldwide, it is predicted that by 2070, 37 million people and assets worth $13 trillion will
be exposed to coastal hazards such as storms, ooding and climate variability (Nicholls & Hanson
2013). With carbon emissions still not stabilised globally, development of climate change adaptation
strategies is now a priority for many jurisdictions (IPCC 2014).
Disasters such as the Indian Ocean tsunami in December 2004 and the hurricanes Katrina
(2005), Sandy (2012) and Michael (2018) in the United States have drawn attention to the important
role that coastal vegetation and reef-forming invertebrates play in shielding populations and property
from sea level rise and storms (e.g. Danielsen etal. 2005, Gedan etal. 2011, Arkema etal. 2013). In
some instances, the role that natural ecosystems play in reducing ooding is more effective than that
of hard defences (Gittman etal. 2014). For example, along the east coast of the United States, damage
to shorelines protected by marsh was less than along shoreline protected by bulkheads following
category 1 hurricane events (Gittman etal. 2014). In several countries, including the United States,
these observations, coupled with the growing need for coastal settlements and infrastructure to adapt
to the threat of climate change, has led to policy that focusses attention away from hard to soft (i.e.
‘living shoreline’) defences as a climate change adaptational strategy (Finucane etal. 2014).
Despite the catastrophic effects that natural disasters may have on urban settlements, they can
also bring about changes in public sentiment and government action. For example, Hurricane Sandy
caused extensive damage to more than 650,000 homes along the east coast of the United States and
claimed over 150 lives (Stewart, 2014). In response, the U.S. government provided more than $50
billion towards disaster relief and established the Hurricane Sandy Rebuilding Task Force to examine
how funds might be best used to improve the resilience of communities and infrastructure to future
and existing threats, including those exacerbated by climate change. The resulting Infrastructure
Resilience Guidelines (Finucane etal. 2014) called for ‘environmentally sustainable and innovative
solutions that consider natural infrastructure options in all Federal Sandy infrastructure investments’.
As a result, large numbers of living shoreline projects have now been implemented across the
coastline (e.g. Tom’s Cove and Assateague Beach), which are anticipated to provide not only coastal
defence, but also other societal benets.
Coastal urbanisation, economic development and urban renewal
As the human population living in the coastal zone continues to increase, rapid urbanisation is
transforming natural shorelines, expanding the built environment and increasing land reclamation.
Developed urban shorelines are characteristically dominated by articial structures, which have
nite service life spans and need to be maintained regularly and replaced eventually (Mahmoodian
etal. 2015). Urban renewal provides opportunities for converting existing coastal defences into more
ecologically sustainable designs. For example, the Seattle seawall (see the section entitled ‘Case
studies and scaling up’, later in this review) was failing and at risk of collapse should an earthquake
hit, posing a public safety hazard. This, combined with the importance of the shoreline as a migratory
corridor for endangered juvenile Chinook salmon, provided the opportunity to incorporate habitat
improvements to the seawall reconstruction. Ecoengineered shoreline approaches will be adopted
most readily where they provide clear benets, which are in the public interest.
It is not unusual for areas of a city to be intentionally or unintentionally deindustrialised or
deurbanised back to a more natural situation, and in such cases, ecoengineered shorelines can
potentially play an important role. The Barangaroo Reserve in Sydney, Australia, is an example of a
project driven by deindustrialisation and redevelopment. This reserve was once a concrete container
terminal, but it underwent a major transformation to become a harbour foreshore park. Projects of
this sort have great potential and open up huge opportunities for ecoengineered shorelines. Obsolete,
locked dock basins in macrotidal estuaries in Europe have been the focus of ambitious urban renewal
schemes, usually retaining the dock basins as water features or marinas and turning warehouses
into housing, ofces, hotels, shops, museums, art galleries and entertainment venues. One of the
best studied has been the Liverpool dock system, where water quality was improved by articial
mixing and bioltration by articially (Russell etal. 1983) and naturally settled mussels (Allen etal.
1992, Hawkins etal. 1992a, Allen & Hawkins 1993, Wilkinson etal. 1996), leading to increases
in biodiversity and healthy inner-city ecosystems of considerable conservation interest (Allen etal.
1995). The docks were much used in outreach and public engagement (Hawkins etal. 1993). Had
the ecoengineered shoreline concept existed in the 1980s, considerable impetus and a body of good
practice could have been deployed at the planning stage of urban renewal in Liverpool.
Integration into governance and policy
Political will and shifts in governance and policy represent a strong and effective driver of
ecoengineered shoreline development (Arkema etal. 2017, Sutton-Grier etal. 2018). For example,
the central government of China issued a document on the use and protection of the country’s coast in
2017, which specied that at least 35% of its shoreline should be maintained in a natural state by 2020
and that all proposed reclamation projects should be halted. This new national policy has sparked
interest among provincial governments to invest in ecoengineered shoreline development and to
convert concrete shorelines into more environmentally friendly ones by means of ecoengineering
(Duan etal. 2016). At a state level, Maryland in the United States introduced the Living Shorelines
Protection Act in 2008, which requires that property owners use living shorelines as the default for
protection unless the owner can demonstrate the need for an articial structure (e.g. a bulkhead)
instead (Kochnower etal. 2015). Following up on European Union Directives and policy guidelines,
the UK government and its devolved administrations (e.g. Wales) have given considerable emphasis
on forming policy to incorporate habitat restoration and using ecoengineering in new coastal
developments, including coastal defences in emerging planning guidance (see Evans etal. 2017,
2019 for detailed reviews of these efforts).
Global concerns for climate change and maintenance of biodiversity as seen in the participation of
different governments in international agreements and conventions may also lead to ecoengineering
options. The Paris Climate Agreement states that signatory parties ‘should take action to conserve
and enhance, as appropriate, sinks and reservoirs of greenhouse gases’ (United Nations 2015),
and the Convention on Biodiversity Conservation aims to achieve ‘the conservation of biological
diversity, the sustainable use of its components’ (United Nations 1992). Ecoengineering is potentially
one of the means that can help achieve the goals of these agreements and conventions.
Growing public outreach and awareness
Public awareness of and concern about marine environmental issues are positively correlated with
being well informed (Gelcich etal. 2014). Recent expanded interest in ecoengineered shorelines thus
may be tied to greater education and understanding of the topic, as well as greater experience with
coastal hazards that necessitate coastal defence infrastructure (Kochnower etal. 2015, Gray etal.
2017). Interviews with coastal residents in the United States suggest that awareness of and preference
for ecoengineered shoreline approaches over traditional engineering are common, although so too
are various misconceptions about the topic (Scyphers etal. 2014). In a recent survey of residents
of four harbours in Australia and New Zealand, Kienker etal. (2018) found that level of education
was positively correlated with support for marine ecoengineering. Additionally, prior knowledge
about the dominant marine articial structures in their harbour inuenced how much participants
in the study were willing to pay for ecoengineering (Kienker etal. 2018). A study of a variety of
different stakeholders in Wales, ranging from engineers through coastal ecologists to the general
public highlighted a willingness to embrace a multifunctional approach to the design of sea defences,
especially if ecological objectives could also be delivered (Evans etal. 2017).
Education may result from formal programmes, such as in schools, or from community
engagement in outreach activities and media articles (Figure 4). Emphasis on raising public
awareness has been central to several recent ecoengineered shoreline projects. For example, the
Billion Oyster Project in New York integrated curriculum development and restoration-based
education programmes about the importance of oyster reefs and their ecosystem services for primary
and high school students (Janis etal. 2016), which likely contributed to public awareness of the
project. Local media attention has also recently increased the visibility of various ecoengineered
shoreline projects, such as owerpots on Sydney seawalls (e.g. Chapman & Blockley 2009, Browne
& Chapman 2011, Morris etal. 2017).
Advances in science and technology
Over the past two decades, there has been rapid growth in the understanding of how shoreline defences
affect coastal ecosystems (e.g. loss of biodiversity, increase in invasive species) (Airoldi etal. 2005a,
Martin etal. 2005, Bulleri & Chapman 2010), and how these negative effects may be compensated
for through ecoengineered shoreline designs (Moschella etal. 2005, Perkins etal. 2015, Firth etal.
2016a, Munsch etal. 2017b, Strain etal. 2018b). It is now acknowledged that some articial defences
can cause new erosion problems or loss of public amenities (e.g. beach loss in front of seawalls;
Fletcher etal. 1997). Recent research has increasingly emphasised ecoengineered shorelines as a
potentially more cost-effective strategy than traditional approaches to coastal protection (Sutton-Grier
etal. 2018). As studies on ecoengineered shorelines have multiplied, the accumulation of data and
research ndings have enabled meta-analyses pinpointing the precise ecoengineering strategies and
specic circumstances under which ecoengineered shorelines are likely to be successful at meeting
their stated aims (Gedan etal. 2011, Bugnot etal. 2018, Strain etal. 2018b).
In addition, many ecoengineered shoreline strategies have themselves advanced and become more
cost effective because of recent innovations in fabrication (3-dimensional printing) and materials science
(e.g. ecofriendly cement mixes; Perkol-Finkel & Sella 2014). Further, advances in ecology and ecosystem
science have brought insights that clarify ecological processes and the functional traits of taxa that are
central to ecosystem services provided by shoreline habitats. For example, greater understanding of
the role of oysters in bioltration, habitat provisioning, sediment stabilisation and wave attenuation in
New York Harbour (e.g. Coen etal. 2007, Grabowski etal. 2012) ultimately facilitates ecoengineered
shoreline projects that are of ecological and social value (e.g. the Billion Oysters Project). As knowledge
of coastal ecological processes and the science of ecoengineering continue to grow and support advances
in technology, it is likely to drive the demand for ecoengineered shorelines.
Barriers to implementation
In general, barriers may include (1) a lack of awareness by decision-makers of the available options
and a failure to engage with the multiple stakeholders involved in communicating the benets
of ecoengineered shoreline approaches, (2) government policies that are insufciently exible to
accommodate new approaches and ways of thinking and (3) an absence or paucity of science-based
evidence for the success, benets and long-term durability of innovations (see the section entitled
Figure 4 Various signs at ecoengineered shorelines to raise awareness and public education.
‘Evaluating the ecoengineered shoreline approach in practice’, later in this review). Additionally,
there may be cultural, nancial or logistic (e.g. extreme environmental degradation) barriers to
adopting ecoengineered shoreline approaches. To incentivise the innovation and development of
ecoengineered shoreline designs and technologies, a consistent and supportive framework is required
to improve risk-to-reward ratios, particularly during the demonstration stage of technology and
design development. This framework may enhance positive expectations, stimulate learning, improve
the design and increase the likelihood of successful project implementation (Foxon etal. 2005).
Awareness and communication among key stakeholders
For hard ecoengineered shorelines, small-scale trials have been carried out in different parts of
the world over the last three decades, but medium- to large-scale implementation of ecoengineered
shorelines is still in its infancy (but see the Seattle case study as one exception). For soft engineering,
larger-scale implementation is more common; for instance, there has been a great effort to restore
mangroves in Asia, with coastal defence as a driver in many projects (Saenger & Siddiqi 1993,
Benthem etal. 1999, IFRC 2011). However, in contrast to hard ecoengineering, which has often
been research driven, these larger-scale soft ecoengineering projects frequently lack scientic rigour
and robust ecological-monitoring regimes (Narayan etal. 2016). Despite hot spots of research and/
or implementation of ecoengineered shorelines globally, one barrier to their implementation more
generally may be the wider transfer of information to relevant local, state or national governments.
Even in countries that are driving ecoengineered shoreline research (like Australia, European
countries, and the United States), regional (state or provincial) governments may not know about
the rationale, concept and potential value of ecoengineered shoreline implementation. A better
understanding of these innovations by government leaders, key stakeholders and the general public
may lead to support for small-scale project trials. In turn, the success of these trials can lead to the
implementation of large-scale ecoengineered shoreline projects.
Effective communication and partnerships among architects, engineers, ecologists,
socioeconomic scientists and other relevant stakeholders are also important. Cross-disciplinary
teams generate holistic projects through collaboration to achieve a common goal and build synergy.
In particular, there is a need to build trust and better understanding among stakeholders, who often
do not share the same perspective and interests (Prati etal. 2016), so as to develop ecoengineered
shorelines that can full both engineering requirements (i.e. shore protection and structure integrity)
and ecological goals, while also providing a social benet (Evans etal. 2017).
A better international strategy for enhancing the awareness of ecoengineered shorelines and
their benets to people and the environment is needed. The closest international tool to address
this need currently is the development of an online coastal resilience mapping interface, driven
by The Nature Conservancy in the United States, which has attempted to synthesise data on the
coastal protection provided by intact natural habitats and to identify soft engineering projects (http:// This tool does not include hard ecoengineered shorelines.
The National Oceanic and Atmospheric Administration (NOAA) has developed a useful webpage
to introduce the concept of living shorelines, with some examples (https://www.
insight/living-shorelines) in the United States. In Australia, the state government of New South Wales
has developed an Environmentally Friendly Seawalls guide (
resources/estuaries/pubs/090328-Seawall-Guide-2012-Reprint.pdf), which summarises local projects.
Ideally, an internationally oriented ecoengineered shoreline organisation and educational website
should be established, accompanied by a campaign aimed at increasing awareness at a global level.
This effort could capitalise on existing networks, such as Restore America’s Estuaries and the World
Harbour Project, and include international organisations such as The Nature Conservancy and World
Wide Fund for Nature to enhance content delivery. The strategies could include using multimedia,
both conventional and Internet-based, to introduce ecoengineered shorelines and highlight illustrative
projects from around the world.
Policies, incentives, regulations and nancial instruments
At present, specic policies that call for integrating ecological considerations into the design and
construction of coastal defence and coastal development schemes through ecological engineering
are generally lacking in most countries or jurisdictions. Nevertheless, ecoengineered shorelines can
potentially make a signicant contribution towards a wide variety of current policy objectives, such
as those that support sustainable development and the maintenance of biodiversity (Naylor etal.
2012, Dafforn etal. 2015b). For example, under the United Nations Convention on the Law of the
Sea (UNCLOS), states are required to protect and preserve the marine environment (Ban etal. 2014).
In this context, ecoengineered shorelines could be proposed as a tool for generating more adaptive,
resilient coastlines for the mutual benet of the environment and society. Having ecoengineered
shorelines on the international policy agenda may incentivise the implementation of local legislation
to further support their application (e.g. the Living Shorelines Act, Maryland).
The few large-scale ecoengineered shoreline projects emerging recently (e.g. Tong King Delta and
Mekong Delta, Vietnam; Grenville Bay, Grenada; Barangaroo Reserve, Carss Park and Olympic Park
in Sydney; the Seattle Seawall Project, Harlem River Designing the Edge Project, Brooklyn Bridge
Park, and Vancouver Conference Centre Project in the USA), if monitored efciently for successes and
failures, could provide valuable evidence to inform the adoption of ecoengineered shoreline designs,
but also offer some essential information for doing a benet-cost analysis (BCA) of any proposed
ecoengineered shoreline project. Insufcient resources are usually the main barrier for any innovative
development and its commercialisation (Hadjimanolis 1999). This situation is exacerbated by the lack
of full accounting of the benets associated with ecoengineered shorelines, which has rarely been
performed. This is due primarily to the difculty in assigning monetary values to the ecosystems
and social services that the projects are designed to perform (see the section entitled ‘Measuring
socioeconomic outcomes’, later in this review; but see Narayan etal. 2016, Reguero etal. 2018).
To date, many of the documented ecoengineered shoreline projects have been on a small scale,
partly due to insufcient nancial support. In most cases, local governments play a key role by
providing some seed funding for conducting feasibility studies on small-scale trials of various
ecoengineered shoreline designs. If a trial project proves successful, the government might proceed to
conduct a large-scale implementation, depending on the policy agenda, site availability and nancial
resources. Unfortunately, even when money is available, in some cases the type of monitoring that
is performed focusses only on proving whether a particular trial concept can or should be scaled up
at a particular site (J. Miller, pers. obs.). In these cases, the opportunity to identify benets that may
apply more broadly to ecoengineered shorelines constructed in other locations is lost.
Certain countries have structured regulations that may encourage implementation of
ecoengineered shoreline designs. For example, in the United States, there are regulations that
demand quantitative mitigation measures for development projects, including in coastal and marine
environments. These typically call for the restoration of habitat loss and/or ecosystem functions
affected by the project and are often signicantly higher in cases where on-site mitigation is not
possible in terms of the investment expected (e.g. Ambrose 1994). Currently, mitigation measures
include actual restoration work (e.g. marsh restoration, construction of articial reefs) or purchasing
mitigation credits from available mitigation banks. Ecoengineered shoreline implementation can be
viewed as on-site mitigation measures, potentially offsetting some of the overall compensation and
mitigation costs. This can effectively incentivise the implementation of ecoengineered shorelines;
however, documentation of such cases is lacking within the United States, and currently this approach
is not widely adopted in other countries.
For green buildings, there are certication and grading systems, such as those by the Leadership
in Energy and Environmental Design (LEED) initiated in the United States (Humbert etal. 2006) and
nowadays applied in many countries worldwide, the Building Research Establishment Environmental
Assessment Method (BREEAM) in the United Kingdom (Crawley & Aho 1999) and Building
Environmental Assessment Methods (BEAM) in Hong Kong (Lee etal. 2007) in order to recognise
the environmental friendliness and energy-saving efforts being adopted in the building as an incentive
for the project proponent or developer to achieve the highest rating (e.g. platinum award in LEED).
Recently, the Waterfront Edge Design Guidelines (WEDG) in New York and New Jersey Harbour
have been developed by the Waterfront Alliance ( The WEDG is a
point-based rating system with a set of guidelines for waterfront projects in New York and New Jersey
Harbour, with a view to creating resilient, ecological and accessible waterfronts (Box 1). During the
evaluation process, actions taken in the planning, design, construction and even postconstruction
(monitoring) process are tallied to determine how successful a project is in achieving benets in the
three core areas: resilience, ecology and access. Any project proponents can voluntarily go through
the WEDG assessment to gain certication and receive advice from professional assessors to further
improve the project in the plan’s three aspects: shoreline protection, ecological enhancement and
diverse uses of the waterfront by various stakeholders.
Thus far, the WEDG has been used to certify over a half-dozen projects spanning a range of
shoreline uses, from parks to heavy industrial areas in New York City (http://wedg.waterfrontalliance.
org). Recently, the WEDG has undergone a revision intended to streamline the ratings system
and make it more readily applicable to shorelines outside the New York/New Jersey metropolitan
The WEDG employs an evidence-based approach, focussing on three key pillars of excellent
waterfront design in New York and New Jersey Harbour. They include the following:
Resilience: The waterfront project should reduce risks or be adaptable to the effects of
sea level rise and increased coastal ooding, through setbacks, structural protection
and other integrative landscaping measures.
Ecology: The waterfront project should protect existing aquatic habitats and use
designs, materials and shoreline congurations to improve the ecological function of
the coastal zone and strive to be consistent with regional ecological goals.
Access: The waterfront project should be equitable and informed by the community,
enhancing public access, supporting a diversity of uses, from maritime, recreation
and commerce where appropriate, thereby maximising the diversity of the harbour
and waterfront.
The credit points made by the professional assessors would help guide the design process,
from conceptual design through operations, and provide design performance goals for
resilience, ecology and access in the following six categories:
Category 0: Site Assessment and Planning
Category 1: Responsible Siting and Coastal Risk Reduction
Category 2: Community Access and Connections
Category 3: Edge Resilience
Category 4: Natural Resources
Category 5: Innovation
area. The WEDG is the rst example of this kind to pave the way for developing an international
certication system that will play a key role in incentivising the development of ecoengineered
shorelines around the globe. Furthermore, the question is how to use the certication system to
further incentivise ecoengineered shorelines. Potentially, if waterfront projects have successfully
addressed the Sustainable Goals of the United Nations (
sustainable-development-goals) and received certication by a system like the WEDG, then the
project proponent could receive a tax reduction or rebate and banks could provide better loans with
a lower interest rate for the project proponent. However, such arrangements are likely to depend on
individual governments and banks. There is certainly a need to build more successful precedent
cases to convince governments and nancial institutions.
Engineering concerns and paucity of science-based evidence for success
One of the biggest obstacles from an engineering standpoint is liability (Slate etal. 2007). Although
regulations differ from country to country, engineers are generally responsible for ensuring public
safety and can be held responsible if their projects fail. Findings of fault can result in severe economic,
professional or even criminal sanctions (Baura 2006). As a result, engineers tend to be cautious
and view unproven technologies/approaches with an appropriate degree of trepidation. When it
comes to the design of traditional infrastructure such as buildings and bridges, engineers can rely
on established codes such as the International Building Code for guidance.
For less traditional infrastructures such as coastal defence systems (i.e. seawalls, breakwaters
and revetments), codes are often lacking. As a result, engineering design is often guided by a
set of standard practices/approaches advocated in respected design manuals such as the Coastal
Engineering Manual (U.S. Army Corps of Engineers 2002) and The Rock Manual (CIRIA etal.
2007). Absent codes, following well-vetted engineering practices provides a level of protection
from professional liability should a project fail. For the engineering community to more openly
embrace ecoengineered shorelines, design manuals need to be developed (Burcharth etal. 2007), or
perhaps more likely, sections need to be added to existing design manuals to deal specically with
ecoengineered shoreline approaches. Such an effort needs to be led by an internationally recognised
body such as the U.S. Army Corps of Engineers or Construction Industry Research and Information
Association (CIRIA).
A second, related concern from an engineering standpoint is the real or perceived lack
of predictability of the behaviour of certain types of ecoengineered shoreline projects with
time (Bouma etal. 2014). Ecoengineered shorelines have a temporal component that is more
difcult to deal with than traditional engineering materials. For example, fatigue is a well-
known engineering phenomenon in which materials lose strength over time due to repetitive
loading. Most traditional engineering materials have gone through enough testing that this
fatigue behaviour can be predicted and planned for during design. The living element of
ecoengineered shoreline projects is much less predictable (and certainly much less well studied)
than the temporal behaviour of traditional materials (Bouma etal. 2014). Furthermore, some
ecoengineered shoreline approaches may actually depend on the living portion of the project for
properties critical to the performance of the structure. For example, if oyster growth is critical
to the wave attenuation or chloride penetration prevention characteristics of a project, the threat
of environmental phenomena affecting growth has a potentially devastating, cascading impact
on the engineering characteristics. This can be related back to liability and the reluctance of
engineers to design something whose behaviour they cannot predict. A robust evidence base on
the efcacy of ecoengineered shoreline approaches was identied by some stakeholders, however,
as a barrier to fuller implementation (Evans etal. 2017). Part of the problem is that the evidence
is accumulating, but it is not necessarily being delivered in a readily accessible and digestible
form to key practitioners, statutory bodies and legislators.
Evaluating the ecoengineered shoreline approach in practice
Traditional articial defence structures are constructed to protect land from erosion and ooding.
Their success is typically measured from an engineering standpoint of whether protection was
achieved or not. In contrast, ecoengineered shorelines are valued for their potential to provide
multifunctional benets, and thus measuring their success calls for a wider analysis. We suggest three
main domains of objectives for which the success of ecoengineered shorelines may be measured:
Ecological objectives, often stated in terms of enhancement of biodiversity or a particular
species (e.g. salmon in Seattle; Munsch etal. 2017b) or the reestablishment of a habitat that
provides ecosystem services such as water ltration or biodiversity provision (e.g. oyster
re efs)
Engineering objectives, which relate to the function of an engineered insertion over its
projected life span
Socioeconomic objectives, such as public acceptance, politics, recreational value, property
prices and development goals (Table 1)
Table 1 Examples of criteria for measuring the success of ecoengineered shorelines
Category Goals: Functions/services Measure Controls/reference
Ecology Native species biodiversity At species level: species richness,
biomass, abundance, percentage of cover,
percentage, community assemblage
At habitat level: Habitat diversity
At genetic level: Genetic diversity
Articial structure
Natural shoreline
Invasive species Number and abundance of species
Ratio of native to invasive species
Articial structure
Natural shoreline
Target species Enhancement/recovery of abundance or
survival of target species
Articial structure
Natural shoreline
Ecological functioning and
Integrity of biological assemblage (e.g.
functional groups)
Water quality
Primary productivity
Ecosystem engineers/habitat-forming
Carbon sequestration
Articial structure
Natural shoreline
Fisheries production Enhancement of sheries supply
Usage of habitat/refuge by larvae
Articial structure
Natural shoreline
Connectivity Enhancement and/or reduction of
Articial structure
Natural shoreline
(risk reduction)
Energy attenuation Wave height reduction
Current reduction
Natural shoreline
Articial structure
Shoreline stabilisation Horizontal shoreline location Natural shoreline
Articial structure
Metrics relating to each of these domains are essential for assessing multifunctionality and
overall benets of ecoengineered shorelines. Although a combination of these domains is becoming
increasingly common in monitoring plans for ecoengineered shorelines (Yepsen etal. 2016), we show
that in practice, few projects have incorporated all three (marine scientists/managers, engineers and
the public; see Figure 5).
Measuring ecological outcomes
Ecoengineered shorelines can be designed to meet a variety of ecological objectives, including
biodiversity enhancement, invasion resistance, facilitation of specic target species, specic ecosystem
Table 1 (Continued) Examples of criteria for measuring the success of ecoengineered shorelines
Category Goals: Functions/services Measure Controls/reference
Achieving adequate/desirable
sedimentation dynamics
Regulating sediment accumulation Natural shoreline
Articial structure
Reduce ooding Surge extent/height Natural shoreline
Articial structure
Structural integrity Structural integrity
Resistance to extreme weather events
Through time
Natural shoreline
Articial structure
Social Aesthetics Appeal to people Before-after
Areas with and
without interventions
Tourism and recreation Waterfront accessibility
People’s awareness and use of the
Areas with and
without intervention
Education People’s knowledge and awareness of
coastal biodiversity and ecoengineered
Areas with and
without intervention
and policy
Facilitate use of
ecoengineered shorelines
Funding incentives, permits,
recommendations, and regulations
Hazard mitigation Protection of property and life Before-after
Natural shoreline
Articial structure
Economic Shore protection
Reduction of economic loss due to
Reduction of insurance premium
Increase in property value
Natural shoreline
Articial structure
Creation of jobs Increase in job opportunities in landscape
architecture, marine ecology,
construction, tourism and shery sectors
Business opportunity Nurturing new businesses associated with
successful waterfront
Project performance Construction, maintenance and services
Articial structure
Note: The term articial structure refers to a traditional engineering solution (e.g. seawalls). Natural shoreline refers to the
intact equivalent of the created habitat (e.g. mangroves, rocky shores). A bare reference without a structure (e.g.
mudats) may also be an appropriate control for some measurements.
processes and general functioning, with the ultimate aim of provision of services (Table 1). The objectives
selected depend on the type of ecoengineered shoreline project, the ecological and environmental
context in which the project takes place and local societal needs and interests. Ecological objectives
for hard ecoengineered shorelines tend to focus on ameliorating abiotic and/or biotic stressors that
limit diversity in general, thereby enhancing focal species of conservation or commercial interest
or increasing native populations or functional groups that enhance ecosystem services (Chapman &
Blockley 2009, Dugan etal. 2011, Perkol-Finkel etal. 2012, Mayer-Pinto etal. 2018, Strain etal. 2018b).
In contrast, where soft or hybrid ecoengineered shorelines can be constructed by establishing coastal
plants, shellsh or coral reefs (Arkema etal. 2013), the main ecological focus is typically on selecting
and monitoring the efcacy of specic habitat-forming organisms for coastal protection via the
dampening of waves or stabilisation of sediment (Borsje etal. 2011). However, measures of ecological
cobenets, such as enhancement of biodiversity or sheries productivity, carbon sequestration and
improvement of water quality (Barbier etal. 2011, Davis etal. 2015, Sutton-Grier etal. 2015) may also
be important for quantifying the success of soft and hybrid ecoengineered shorelines.
Measuring the e cological benets of ecoengineered shorelines requi res prespecifying a monitoring
framework (prior to initiation) that effectively assesses changes in a specic intervention or habitat
of interest (Michener 1997, Block etal. 2001). Ideally, monitoring plans compare ecoengineered
shorelines with controls (i.e. unaltered articial shorelines), as well as with ‘reference areas’ (i.e.
natural habitats that the ecoengineered shorelines are seeking to mimic). In this way, planned
ecoengineered shoreline interventions function as experiments (Chapman etal. 2018). Appropriate
controls and reference areas depend on the type of ecoengineered shorelines being constructed. For
hard ecoengineered shorelines, the relevant control is typically a new or cleared section (i.e. with any
existing ora and fauna removed) of an articial structure at the same site or at a nearby site with
comparable environmental conditions, depending on the scale of the manipulation (Chapman etal.
2018), and reference sites are frequently rocky shores, the closest natural hard-substrate analogues
(Chapman & Underwood 2011, Evans etal. 2016).
Figure 5 The proportion of key stakeholders (marine scientists/managers, engineers, public and all three)
involved in ecoengineered shoreline projects in select locations in the United States, Australia, Europe and
the United Kingdom. The data were provided by researchers in these places, and the locations presented were
based on where the information about stakeholder participation was available.
For soft-engineering interventions, controls should be established patches of the habitat type
being constructed. There have also been some attempts to compare soft and hybrid interventions
with ‘hard’ reference sites (e.g. coral reefs with breakwaters and saltmarsh with and without sills
with bulkheads; Ferrario etal. 2014, Gittman etal. 2014). However, it is important to note that
reference sites may have differences in physiochemical parameters, community succession and
other environmental drivers, which can make comparisons between ecoengineered shorelines and
reference sites difcult to interpret (Chapman etal. 2018), especially on highly modied coastlines.
Data regarding the ecological performance of reference sites can provide helpful insight into the
expected natural variability in recruitment (often due to chance events in plankton) or survival of
organisms (e.g. due to heat stress and predation), as well as the possible impacts from other environmental
stressors over time (e.g. climate change or shing; see Figure 6). At least two natural reference sites
should be surveyed multiple times both before and after ecoengineered shoreline installation (Beyond
BACI; Underwood 1991). For example, if the ecoengineered shoreline is located between two rocky
shores, then both rocky shores should be used as reference sites and surveyed regularly alongside the
monitoring of the ecoengineered shoreline. Equally, the effects of ecoengineered seawalls should be
compared with multiple unmanipulated seawall sites (Morris etal. 2017).
Monitoring and evaluation programmes can be qualitative, semiquantitative or quantitative and may
employ various sampling techniques, including visual surveys, eld sensors and destructive sampling
(Table 2; Burcharth etal. 2007). Generally, the time required for sampling increases as sampling
methods become more quantitative. Which destructive or remote-sensing methods are appropriate
will depend on the scale of assessment (e.g. whether changes in alpha, beta or gamma diversity are
being evaluated) and the variables being assessed. Core variables for monitoring commonly include the
number, diversity and abundance of species (Mayer-Pinto etal. 2018, Strain etal. 2018b), the growth
and structure of vegetation or habitat-forming organisms and functional measurements such as ltration
rates or primary production (Mayer-Pinto etal. 2018). These variables are typically compared between
ecoengineered shorelines, controls and reference areas based on differences in means (Osenberg etal.
1999, Munkittrick etal. 2009, Smith etal. 2017b) and assemblage structures (Anderson 2001). Measures
of effect size, such as Cohens D or log response ratio, can also be helpful (Cohen 1988, Hedges etal.
1999, Coleman etal. 2006, Nakagawa & Cuthill Innes 2007) for quantifying the magnitude of change
resulting from the intervention relative to controls and reference sites.
Figure 6 A hypothetical illustration of plausible temporal changes in ecological performance of an ecoengineered
shoreline and multiple control and reference sites (solid and dashed lines indicate the various sites). The shaded area
around each line represents var iation in a single hypothetical site or plot. It is expected that ecoengineered shorelines
will enhance the targeted ecological response to be more similar to the reference habitat of interest (e.g. natural
habitat), in comparison to the control (i.e. the unaltered structure), which will remain the same.
To date, much of the experimental research on ecoengineered shorelines has been undertaken
at small spatial (i.e. ranging from centimetres to metres) and temporal scales (<12 months) (Bishop
etal. 2017, Chapman etal. 2018, Morris etal. 2018b, Strain etal. 2018b). Because of the effects
that the shoreline form can have on ecological connectivity (Bishop etal. 2017), and with growing
interest by local government agencies, nongovernmental organisations (NGOs) and developers in
implementing ecoengineered shorelines at the larger scales required for coastal protection, there is a
need to develop appropriate monitoring protocols for addressing ecological changes at these scales.
The 12-month monitoring period undertaken by many organisations is arbitrary and probably much
too short for demonstrating impacts on community structure or ecosystem functioning. Neither
is there sufcient time for natural succession to take its course. For example, in an early study
of colonisation of large concrete blocks placed on the outside of a Plymouth Breakwater in the
United Kingdom, communities and assemblages took more than 5 years to stabilise to typical small-
scale patchiness and low-amplitude uctuations, with populations of the key grazing limpets taking
considerable time (>7 years) to stabilise (Hawkins etal. 1983).
In many cases, the outcomes of ecoengineered shoreline projects at 12 months depend heavily
on initial conditions (e.g. created wetland with greater cover will report success in shorter time
frames than created wetlands with less cover; Mitsch & Wilson 1996). Research on wetlands suggest
that 1520 years are required to judge the success of ecoengineered shoreline projects (Mitsch &
Wilson 1996). The spatial and temporal scales of ecoengineered shoreline monitoring should be
linked to the generation time and geographic distribution of the communities being investigated.
Developing standardised monitoring protocols and promoting their use across multiple users (e.g.
the Shoreline Monitoring Toolbox in Puget Sound, Washington State;
toolbox and the Hudson River in New York: Findlay etal. 2018) can help to facilitate collaborations
between ecologists and the broader public (Toft etal. 2017) and may help to extend the monitoring
time frames of ecoengineered shoreline projects beyond 12 months. But one size is unlikely to
t all cases, and monitoring programmes need to be tailored to the objectives of a project and its
environmental and socioeconomic context.
A lack of clear project goals and appropriate monitoring and evaluation of these has hampered
assessments of ecological successes and failures. For example, in an analysis of oyster reef restoration
projects undertaken in the Chesapeake Bay for shoreline stabilisation or other purposes between
1990 and 2007, only 43% of projects included both the restoration and monitoring required to assess
their success or failure (Kennedy etal. 2011). Similarly, of all documented marine coastal restoration
projects over a 40-year period, only 61% provided information on survival of the restored organisms
(Bayraktarov etal. 2016). Where monitoring has been carried out for ecological enhancements of hard
defences, it has typically focussed on the success of interventions in enhancing species richness and/or
the abundance of key functional groups relative to controls (see Strain etal. 2018b). Although overall
Table 2 Evaluation techniques to measure the ecological and effectiveness of ecoengineered
Technique Method Time scale Effort and precision Expectations
Visual Photographs
Online outreach
Hours Low Initial documentation of project
success or problems to address
Survey data Structural amount of components
(e.g. vegetation, algae, sessile
invertebrates, sediments)
Year Medium Inform management on initial
meeting of project goals
Statistical analysis Evaluation based on a structured
experimental design of target
Years High Validation of project goals and
recommendations on further
species richness tends to increase with the addition of microhabitats, responses among functional
groups can vary. The extent to which increased species richness is due to native versus nonnative
species (Morris etal. 2018a, Strain etal. 2018b), as well as effects on ecosystem services (Mayer-
Pinto etal. 2018), are rarely addressed, and yet they are important considerations for monitoring,
particularly where ecoengineered shorelines are designed to meet multiple ecological objectives.
An additional limitation in measuring the success of ecoengineered shorelines in meeting
ecological objectives is that unsuccessful (or only partially successful) projects are rarely documented,
making it difcult to assess their frequency (see comments and an example in Firth etal. 2016a). It
is likely that the incidence of failure is high, particularly for soft or hybrid ecoengineered shoreline
projects in which plantings are conducted without rst assessing adequate environmental suitability
of the site. For example, between 1989 and 1995, only 1.52% of the 9050 ha of mangroves planted
in West Bengal, India, survived (Sanyal 1998). An evaluation of the success of a the $35 million
World Bank funded Central Visayas Regional Project in the Philippines indicated that only 18.4%
of the 2,927,400 mangroves planted over 492 ha survived (Silliman University 1996). Plantings are
especially likely to fail at sites that have not previously supported the planted species, underscoring
the importance of collecting data and identifying appropriate ecological objectives prior to the
implementation of ecoengineered shorelines.
Measuring engineering outcomes
Ecoengineered shorelines are generally conceived to address one or more engineering objectives,
which may or may not be directly related. These include shoreline stabilisation, ood mitigation,
energy attenuation and sediment control. Depending on a number of factors, including the physical
setting, funding source and socioeconomic considerations, one or more of these engineering
objectives may take precedence and must be identied clearly so that appropriate success metrics
can be dened. As with ecological objectives, the success of engineering objectives then requires the
development and implementation of an appropriate framework for monitoring (Yepsen etal. 2016,
Findlay etal. 2018).
Ecoengineered shorelines are generally considered successful from an engineering perspective
if they achieve predetermined engineering objectives and they remain intact, such that they continue
to achieve that objective over time. The former is typically referred to as the engineering function
of the project, and the latter the engineering form. Engineering function and form are often (but not
always) related to each other.
The specic metrics used to evaluate function depend on the engineering objectives of the
project (Thayer etal. 2003). Shoreline position is an effective measure of function where shoreline
stabilisation is a core engineering objective (Yepsen etal. 2016). Because shoreline position tends
to be highly 3-dimensional and extremely dynamic, comparison with control sites is essential
(Underwood 1994). It may be possible, however, to use an alternative, less variable measure of
shoreline stabilisation, such as the marsh edge or vegetation line (Kreeger etal. 2015). Projects in
which ood mitigation is an objective typically utilise ood height and/or extent as a metric for
success (Yepsen etal. 2016). Although reduction in ood height compared with surrounding areas is
typically easy to measure, reduction in ood extent is often more difcult and may require numerical
simulations with the ecoengineered shoreline project compared to without based on past ood events
(Yepsen etal. 2016). This can be particularly difcult in built environments, as even small changes
in development patterns can drastically alter ood pathways and the extent of ooding.
Projects with energy attenuation as an objective typically use wave height or current velocity
attenuation as a measure of success. These parameters are measured in front of (offshore) and behind
(inshore) the designed wave attenuation feature and used to calculate transmission coefcients (i.e. the
ratio of their measured value inshore versus offshore), which is inversely related to the dissipation of
energy (e.g. Garvis 2009, Manis etal. 2015). Even though it is common for the transmission through
a traditional engineering structure to increase over time as the structure breaks down, certain types
of ecoengineered shorelines, such as those incorporating ecosystem engineers or biogenic taxa, are
expected to become more effective over time as the biological growth helps to attenuate the incident
energy (Manis etal. 2015). Projects aimed at controlling sediment deposition/erosion typically use
areal extent or sediment volume to judge success. Areal extent (the net increase or decrease in land
area), which is related to shoreline position, is an essential metric in projects intended to accrete or
trap sediment (i.e. with features such as sills, stream barbs, and groynes). While volume change is
perhaps a more complete measure of sediment control, it typically requires more effort and expense
to collect and therefore it is sometimes forgone.
Metrics related to engineering form provide information about the physical integrity of specic
features of ecoengineered shoreline projects. The most appropriate metrics of form depend on the
type of ecoengineered shorelines being constructed. Crest elevation refers to the elevation of the top
of an engineered or natural shoreline feature. Several of the engineering objectives dened here (e.g.
wave height attenuation and ood mitigation) are strongly dependent on elevation (Armono & Hall
2003, Allen & Webb 2011, Webb & Allen 2015). Thus, stable crest elevation over time is critical
to the success of an ecoengineered shoreline project. In addition to elevation, changes in slope or
orientation can be indicative of potential problems. For shore-detached rock structures, changes
in slope have a direct impact on the stability of the structure and its ability to dissipate waves. For
shore-attached vertical structures, deections can be indicative of problems related to scour and
overtopping. For shore-attached sloping structures, changes in slope directly inuence the stability
of stone armouring, as well as run-up. Scour is a generic term referring to the erosion that occurs
adjacent to a hard feature. Scour occurs naturally in front of cliff faces and around rocks, but it also
can be caused by built features such as bulkheads and seawalls. It can occur in front of (toe scour),
adjacent to (anking or end effects) or behind (leeside scour) structures.
Another important component of engineering monitoring is inspection of structural
performance in terms of integrity of materials or construction units (Burcharth et al. 2006).
Structural integrity can be broadly used to describe the physical integrity of an ecoengineered
shoreline project, considering the wide range of types of structure this includes. Despite being
designed in a manner that facilitates ecological enhancement, often the main objective of an
ecoengineered shoreline project is to provide a structural solution. When considering a composite
structure, such as a riprap or revetment, structural integrity can be evaluated according to the
displacement of individual units from the structure (Burcharth etal. 2006). For solid structures
such as a seawall, structural integrity can be about the physical degradation of the overall structure
and the constituent components. A component is dened as an individual structural member, and
the collection of such members such as piles, pile caps, decks and ttings makes up the structure.
When dealing with soft and hybrid solutions, such as marshes or marsh sills, some of these
components are living, including marsh plantings and items derived from natural materials such as
coir logs. In addition to assessing the overall condition of the structure, each individual component
should be rated (NYCEDC 2016). Monitoring techniques include visual site inspections (described
in further detail later in this review), taking core samples of concrete/rock armour and more
standard concrete testing such as chloride or water penetration and compressive strength (Sagoe-
Crentsil etal. 2001). In certain cases, more sophisticated measures can be made using probes for
corrosion progress and density (Castaneda etal. 2017).
The techniques used to evaluate the metrics discussed here vary depending on the project type,
the monitoring budget and the technical abilities of the monitor. Some of the more common tools
and techniques are described here; however, the list is far from exhaustive (Yepsen etal. 2016). The
simplest approach for measuring distances, areas and volumes is to establish a localised survey
grid and to measure distances and elevations with respect to known, xed points. Although more
expensive and requiring more expertise, utilising a survey-grade global positioning system (GPS)
can improve both the accuracy (on the order of centimetres, both horizontally and vertically) and
the efciency of data collection compared with simpler approaches. Light detection and ranging
(LiDAR), a remote-sensing method, can increase the accuracy and efciency of surveys even
further, but this technology is often out of reach for most ecoengineered shoreline projects due to
the expense and expertise required. An emerging approach, which offers many of the advantages of
LiDAR at a fraction of the cost, is the utilisation of recreational grade drones and ‘structure from
motion’ techniques. Structure from motion utilises photogrammetric techniques to piece together
3-dimensional information from a series of 2-dimensional photographs (Turner etal. 2016). Some
techniques are capable of providing information about slope/inclination, and simple inclinometers
can be used to complement low-cost/low-tech techniques.
Simple water-level/wave staffs can be used to estimate water levels and wave heights (Lapann-
Johannessen etal. 2015). More advanced techniques include low- and high-frequency pressure
gauges, ultrasonic wave/water level gauges and electronic wave/water-level staffs (high frequency is
required for wave measurements in shallow water). Current attenuation is probably the most difcult
parameter to measure, due to the expense and difculty involved in obtaining current measurements.
On large-scale schemes, satellite imaging can also be used for monitoring the structural integrity of
the project (Yepsen etal. 2016).
Ecoengineered shorelines are often designed for a life span of 25, 50 or even 100 years, and thus
they require longer periods for success evaluation. Nonetheless, short-term monitoring (1–2 years
postdeployment) of key structural performance criteria (such as cracking or corrosion) can provide
more immediate measures of success (Thayer etal. 2003). Levels I, II and III describe standard levels
of structural inspection and are based on the American Society of Civil Engineers (ASCE) Waterfront
Facilities Inspection and Assessment Manual (Waterfront Facility Inspection Committee 2015). Level
I examination is a visual inspection that does not require any structural components to be cleaned,
and so it can be completed most rapidly. Level II examination is focussed on identifying damaged or
deteriorated areas that may be hidden by marine organisms or surface materials. Level III examination
is more involved and often entails the use of both nondestructive testing (NDT) techniques and partially
destructive techniques. This can include extracting material samples (cores) of the concrete or wooden
structure for off-site testing or in situ surface hardness testing (NYCEDC 2016). For example, the
structural performance of bioenhanced concrete pile encasements constructed in Brooklyn Bridge Park
to restore the load-bearing capacity of the pier piles, while generating valuable habitat for marine ora
and fauna, was monitored 1 and 2 years postdeployment through Level II inspections (Perkol-Finkel
& Sella 2015). These ndings, as well as the results of the ecological monitoring for the project, were
submitted to the New York State Department of Environmental Conservation to validate the structural
and ecological performance of the structure. This was required for acceptance of the bioenhanced
structure to provide mitigation credits towards the overall project.
As the eld of ecoengineering is relatively new and project budgets are often limited after
construction, longer-term monitoring data (>5 years postdeployment) are rare. Nonetheless, as
risk reduction and engineering performance are key components of many ecoengineered shoreline
projects, it is important to conduct such tests and revisit installations several years after deployment.
A recent example is the biological and structural monitoring of a bioenhanced breakwater section
in Haifa Port, Israel, where Sella and Perkol-Finkel (unpubl. data) revisited the installation 6 years
postdeployment. Biological monitoring included on-site visual and photographic surveys, as well as
structural monitoring via a combination of visual inspection and core samples for structural integrity
of the bioenhanced concrete.
Long-term monitoring plans should also aim to identify and quantify changes in the structural
performance due to the increased presence of marine life. As these natural systems develop and
their presence increases over time, they will affect the structure’s performance and interact with the
surrounding environment. For example, marine growth could increase the crest height of a breakwater
and further dissipate local wave energy, or alternatively marine growth on a pile encasement could
alter ow patterns and change drag forces around the structure (Yacouby-Al etal. 2014).
Measuring socioeconomic outcomes
Few ecoengineered shoreline projects have achieved socioeconomic success, despite the capacity
of healthy coastal habitats to provide a plethora of ecosystem services that are valuable to humans
(Barbier etal. 2011). For instance, the complex habitat structure provided by vegetated coastal
habitats or biogenic reefs creates a nursery habitat for commercially important invertebrate and
sh species (Beck etal. 2001, Heck Jr etal. 2003). Shellsh reefs lter nutrients and metals from
the water, which can increase water quality (Gifford etal. 2004, 2005, Kellogg etal. 2013, Smyth
etal. 2013, Onorevole etal. 2018); similarly, seaweeds can sequester nutrients and heavy metals, as
well as carbon (Chung etal. 2011, Henriques etal. 2015, Yang etal. 2015, Krause-Jensen & Duarte
2016). Further, the basis of soft ecoengineering or living shorelines comes from the observation
that natural, intact habitats can protect against erosion and ooding (Shepard etal. 2011, Arkema
etal. 2013, Temmerman etal. 2013).
There is growing evidence of the efcacy of ecoengineered shorelines in providing and enhancing
many of the ecosystem services associated with natural ecosystems (Gittman etal. 2014, 2016, Davis
etal. 2015, Onorevole etal. 2018). Nevertheless, their value in providing these services will depend
on the identity of the habitat types included in ecoengineered shoreline designs, their density and
conguration and the characteristics of the environment in which they are placed (Piehler & Smyth
2011, Smyth etal. 2015). For example, Smyth etal. (2015) found that the relationship between
oyster density and denitrication was weakly positive at ambient nitrogen concentrations, but at
elevated nitrogen concentrations, it was nonlinear, increasing from low to medium oyster densities
but declining at high densities. Overall, it is expected that the ecosystem services provided by an
ecoengineered shoreline will increase over time as the habitat becomes established and grows.
Even well-established ecoengineered shorelines are unlikely to attain the same level of ecosystem
functioning as natural, intact habitats. This is due to trade-offs between engineering, ecological and
socioeconomic objectives that result in a design that is suboptimal for ecosystem service provision.
Simultaneous monitoring of ecological, engineering and socioeconomic objectives is needed to
investigate trade-offs, but it is seldom done.
While there have been attempts to value economically the ecosystem services provided by
estuarine and coastal systems (e.g. Costanza etal. 1997, 2014, Barbier etal. 2011), attempts to value
the services provided by ecoengineered shorelines holistically have been lacking (Beck etal. 2018,
Reguero etal. 2018). Economic valuation involves assigning quantitative values to the goods and
services provided by various habitats and is a useful way to compare benet-cost ratios of management
scenarios (Spurgeon 1999). For ecoengineered shorelines, an ultimate economic goal may be that its
benet-cost ratio exceeds that of alternative articial structures. However, economic valuation can be
challenging for ecosystem services because not all of them have a market value. Goods and services
with a market value can be evaluated based on revenue increases—for instance, when ecoengineered
shorelines increase the production of commercially important seafood species (e.g. Martins etal.
2010), marine-derived pharmaceuticals and raw products, such as rewood. Where there are no
markets for services, indirect pricing methods can be used to establish the (revealed) willingness to
pay (WTP) or willingness to accept (WTA) compensation for the availability or lack of these services
(de Groot etal. 2002).
A variety of indirect methods may be applied to the valuation of ecoengineered shorelines. First,
the economic value of shoreline protection may be determined using avoided damages or replacement
costs (de Groot etal. 2002, Rao etal. 2015), such as the avoided damage to coastal properties after
a storm in areas with an ecoengineered shoreline protecting the coast. This evaluation could be in
comparison to a nearby, unprotected shoreline or, alternatively, a nearby articial structure with a
similar exposure (Gittman etal. 2014). The latter contrast enables an analysis of how maintenance
and replacement costs following a storm compare between ecoengineered shorelines and articial
Second, the hedonic pricing model, which works on the premise that the price of a marketed
good (e.g. houses) is related to structural, neighbourhood and environmental characteristics, may
be applied to ecoengineered shorelines (Landry & Hindsley 2006). Coastal property protected with
ecoengineered shorelines may differ in value compared to property with no coastal protection or
property protected with hard structures (Landry & Hindsley 2006). House prices are directly related
to beach width (Pompe & Rinehart 1994) and other environmental attributes, such as the open/green
space provided by wetlands, which can positively inuence prices in urban areas (Mahan etal. 2000)
and negatively inuence them in rural areas (Bin & Polasky 2005). Additionally, house prices are
inuenced by perceived risk to ooding and erosion (Jin etal. 2015).
A third method for quantifying the value of ecoengineered shorelines is the willingness of a
representative sample of individuals to pay for services they provide (de Groot etal. 2002). This
can be evaluated in terms of the time and cost that recreational users spend travelling to a site to
take advantage of its amenities (Spurgeon 1999). As with the other indirect methods, WTP can be
compared between ecoengineered shorelines, shorelines with traditional structures and shorelines
without structures. Ideally, WTP may also be compared for a given site before and after ecoengineered
shoreline construction.
In addition to such economic valuations, ecoengineered shorelines may be valued in terms of
their social benets and ability to effect policy change. Throughout history, cultures, knowledge
systems, religions, heritage values and social interactions have been inuenced by the nature
of ecosystems (Hassan etal. 2005). Moreover, humans have always shaped the environment to
enhance the availability of such services. Ecoengineered shorelines may enable the public to
reclaim the water’s edge, bringing people back to urban waterfronts (Sairinen & Kumpulainen
2006) for leisure activities and experiences. They can improve shoreline aesthetics, promote
public visitation, increase cultural output (Hortig etal. 2001, Reise 2003, Claesson 2011) and
be valuable for scientic investigation and environmental education (Krauss etal. 2008, Mitsch
etal. 2008). Social impacts can be assessed using a variety of methods, ranging from simple
counts of visitation numbers and frequencies and research and education programmes, through
to more complex assessments of changing perceptions (Maas & Liket 2011). Questionnaires
and participatory mapping of such elements as use values (Brown & Hausner 2017, Strain etal.
2019) are common tools for assessing use and perceptions. Increasingly, technologies such as
immersive landscape theatres and virtual reality are also being used to investigate perceptions
of aesthetic change (Wang etal. 2013, Miller etal. 2016) and may be suitable for before-and-
after assessments.
Policy impacts of ecoengineered shoreline projects take a number of forms (Keck & Sikkink
1998), which makes their measurement challenging. New, highly visible and successful ecoengineered
shorelines can provide the rst step in inuencing policy where ecoengineered shorelines are
not currently part of the political agenda. In drawing attention to the utility of ecoengineered
shoreline approaches, such projects may persuade governmental or other inuential stakeholders to
endorse international declarations or conventions, securing procedural change at the domestic and
international levels and changes in policy (e.g. changes in legislation or budget allocations). This
route could start with the avocation of national or international guidelines for best practices in the
construction of shoreline protection schemes.
Procedural changes may include the development of permitting processes or funding schemes
that enable the use of and incentivise investment in ecoengineered shorelines for coastal defence.
This has the potential to lead to policy changes to support ecoengineered shoreline investments
as part of the increased investment in coastal infrastructure. These political effects need to lead
to on-the-ground changes in key stakeholders, such as private landholders that are implementing
coastal protection works or developers that can declare ecoengineered shorelines as part of the
tender (Keck & Sikkink 1998). In the United States, there has been some progress towards investing
in living shoreline approaches through policy drivers securing procedural (e.g. green bonds and
infrastructure banks used to fund nature-based coastal defence in Massachusetts) and legislative
changes (e.g. the Living Shoreline Act in Maryland; Sutton-Grier etal. 2018). There are guides to
monitoring and evaluating policy inuence (e.g. Reisman etal. 2007, Jones 2011). The method starts
with the development of a theory of change (Reisman etal. 2007) that describes a set of activities
(e.g. ecoengineered shoreline pilot project), outputs (e.g. design guidelines for ecoengineered
shorelines), outcomes (e.g. change to ecological or social system as a result of activity), impacts
(i.e. the overall contribution of the outcomes to the goal) and goals (i.e. overall project objective).
The outcomes are the target for measuring success, which may be through the impact of research,
media uptake, or stakeholder surveys (see Jones 2011).
Adaptive management in response to monitoring outcomes
Assessment of the efcacy of ecoengineered shorelines in meeting objectives, using monitoring
and evaluation programmes, is essential to the success of present and future projects. Due to the
relative immaturity of the discipline, there is a paucity of data demonstrating how project design
and environmental and socioeconomic context inuence project success. Such data are critical
to informing the design of future ecoengineered shorelines and ne-tuning existing projects to
increase their efcacy. Ecoengineering of community structure and ecosystem functioning along
urban shorelines is complicated by uncertainties over costs and potential benets and long-
term sustainability (Temmerman etal. 2013), making it difcult for governments to legislate for
ecoengineered shorelines. Uncertainty increases over time (Bouma etal. 2014), as natural variability,
management objectives and economic constraints can all inuence the success of ecoengineering.
Adaptive management is a structured decision-making strategy to govern social-ecological
systems that embraces their complexity and uncertainty, providing opportunities for learning
and adapting to change (Folke etal. 2005). It rests on the recognition that urban landscapes and
seascapes need to be understood and managed as complex, adaptive social-ecological systems
and points to the importance of actively managing resilience, here dened as the capacity to
persist with functional integrity under changing social and environmental conditions. The
feedback between learning and decision-making is a dening feature of adaptive management,
with learning contributing to management by helping to inform decision-making and management
contributing to learning through interventions that are useful for investigating resource processes
and impacts (Williams 2011). Fundamental steps in the process include the articulation of clear
objectives (ecological, engineering and socioeconomic), identication of management alternatives,
predictions of management consequences, recognition of key uncertainties and quantitative
assessment of outcomes.
Interest in and application of adaptive management have grown steadily over the last few
decades (Chafn etal. 2014), and this tool has been recommended as particularly useful in urban
ecosystem restoration (Hychka & Druschke 2017) and ecoengineering (Mayer-Pinto etal. 2017).
Indeed, the urban setting offers unique opportunities to address environmental issues, while at the
same time delivering wide-reaching benets to the multiple stakeholders and users of marine urban
waterfronts. There is also an urgent imperative to create urban infrastructure and environments that
are more resilient to climate-related risks. Even though examples of real-scale implementations are
limited, scientic knowledge and technologies are being developed to provide adaptive solutions
for ecoengineering. Examples include ecosurfaces that can be ecologically retrotted with further
enhancement if monitoring data suggest that the original objectives are not being met (Morris
etal. 2018c), optimising the timing and frequency of maintenance interventions to infrastructure
by incorporating knowledge of the life histories of species (Airoldi & Bulleri 2011), limiting the
social barriers to ecoengineering via better understanding of the social perceptions and expectations
through social surveys and interviews (Kienker etal. 2018) and/or providing quantitative data on the
cost-effectiveness of various management options (Reguero etal. 2018).
Case studies and scaling up
As is the case for any signicant coastal or urban development, scaling up ecoengineered shorelines
at any site should make economic sense—so long as all costs and benets are factored into that
calculation, including valuing environmental and societal benets and the costs of traditional
development. As with valuation of environmental assets generally, this is still a new eld, and
the ecoengineered shoreline concept includes a diversity of approaches and technologies. Thus, it
is difcult to make any global statements about the cost-benet ratio for living shorelines across
all possible sites and approaches. However, it is instructive to examine projects where the balance
between ecology, engineering and socioeconomic factors has played out with varying degrees of
success. Three case studies (in Seattle, the East and Gulf coasts of the United States, and Sydney) are
described next, and they represent hard, soft and hybrid designs, respectively. Their implementation
and evaluation are also outlined. They cover a range of project types, physical settings, engineering
objectives and habitats. Although all the projects are multiobjective in nature, the drivers of each
project and the metrics used to evaluate them differ.
Case Study 1: Hard ecoengineering: Seattle, Washington
After a large earthquake in 2001, the Elliott Bay Seawall in Seattle, built between 1916 and 1934,
was failing and at risk of collapse in the event of another moderate-sized to large earthquake, posing
a public safety hazard. Because of the scale of the city’s infrastructure, replacement of the wall had
an intense, hard engineering focus. However, the need to replace the failing wall with a structure
designed to modern earthquake standards also provided the opportunity to ecologically enhance the
heavily modied foreshore areas (Cordell etal. 2017). The adjacent estuarine waters are an important
migratory corridor for juvenile Chinook salmon (Oncorhynchus tshawytscha), which is listed as
threatened under the U.S. Endangered Species Act, and the new structure was designed to provide
ecological complexity and a well-lit environment conducive to migration, foraging and refuge for
salmon and other species, thus incorporating ecological components into traditional poured concrete
walls (Munsch etal. 2017a).
Key to the success of this project was social buy-in, as habitat improvement for juvenile salmonids
has great cultural, recreational and economic importance in the region. The Seattle waterfront is
also a vibrant hub of the city, including many business, tourism, recreation, transportation, shipping
and port activities. This social aspect is considerable, especially given the public funding involved
on the project and the sense of place that a broad array of user groups depend upon for a connection
to the water.
The design and implementation of the Seattle seawall habitat enhancements developed during
a decade of studying habitat enhancements for juvenile salmon in Elliott Bay. Information gathered
from testing seawall design options with experimental panels that had enhanced texture and slope
and overhead light penetrating surfaces (LPSs) allowed for proof of concept at the metre scale that
was incorporated into the nal seawall design at the kilometre scale (Cordell etal. 2017, Figure7).
Additional components incorporated into the nal seawall design came from the success of a
project at the nearby Olympic Sculpture Park, which at the 100-m scale created an intertidal bench
and a pocket beach (Toft etal. 2013). Data gathered from these smaller-scale projects allowed for
scaling-up of ecoengineered shoreline implementation (see the section entitled ‘Process of scaling-up
successful projects’, later in this review, for more discussion).
The nal design for Seattle’s seawall included: (1) placement of intertidal benches (stone-lled
marine mattresses) and creation of an articial beach, designed to mimic a shallow-water, low-
gradient habitat; (2) incorporation of crevices and ledges into the seawall face, with the goal of
increasing complexity and enhancing production of invertebrates and algae and (3) addition of LPSs
(glass blocks) in the cantilevered sidewalk above the seawall, intended to provide a light corridor to
enhance juvenile salmon outmigration and feeding and to potentially improve productivity under
piers (Figure 7).
There is an associated time scale with such a large, hard ecoengineering scheme, highlighting
that implementation of ecoengineering projects is not just a matter of spatial scale. Experimental
seawall panels were deployed in 2008 and monitored for 4 years, experimental LPSs were deployed
in 2013 and monitored for 1 year, the Olympic Sculpture Park was created in 2007 and monitored
both before and after enhancement from 2005 to 2011 and the rst phase of the Seattle seawall
rebuild was completed in 2017, with the second phase yet to be initiated. The experimental
stages were needed to ensure the ecoengineered habitats met ecological (i.e. enhanced salmon
habitat), engineering (i.e. no compromise to structural stability of seawall) and socioeconomic (i.e.
community acceptance) outcomes.
Case Study 2: Soft ecoengineering: living
shorelines on the East and Gulf coasts
The ‘living shorelines’ approach emerged in the 1970s through many examples of bank erosion
control and shoreline stabilisation using saltmarsh plantings in Chesapeake Bay (Garbisch &
Garbisch 1994). This technique is regarded as successful at providing shoreline protection that is
adaptive to sea level rise, in addition to cobenets such as habitat and food resources for sh and
other wildlife or improved water quality (Craft etal. 1999). However, it was recognised that the
Figure 7 (A) Monitoring algae and invertebrates on experimental habitat test panels on the Seattle seawall
before reconstruction; (B) habitat enhancements incorporated into the new seawall, including projections and
texturing on the wall and marine mattress benches; (C) snorkelling to survey sh at a mean tide when the bench
is inundated; (D) view from an adjacent pier showing a cantilevered sidewalk with light penetrating surfaces
consisting of glass blocks.
creation or restoration of saltmarsh alone only provided sufcient coastal protection in low-energy
environments. This resulted in the development of the ‘hybrid living shoreline’ approach, which
incorporated varying degrees of hard structure to reduce energy and allow living shorelines to be
applied in a wider range of environmental conditions (Garbisch & Garbisch 1994; Figure 8).
Key to the living shorelines technique is that construction must conserve, create or restore
natural shoreline functions, in addition to providing coastal defence. In order to retain ecological
functioning, strong emphasis is placed on maintaining the connection between terrestrial and
marine systems (Bilkovic etal. 2017), a link that is severed by traditional articial defences, such as
bulkheads. In Chesapeake Bay, rock sills are widely used in combination with saltmarsh plantings for
additional coastal protection (Figure 8). However, although such a design is regarded as successful
for coastal defence, the addition of these hard structures in a soft-bottomed environment results in
ecological trade-offs through enhancing lter feeders that colonise the sill at the expense of deposit-
feeding infauna, with potential consequences for nutrient cycling (Bilkovic & Mitchell 2013).
There is increasing interest in using oyster-reef living shorelines for coastal protection. In
response to the loss of native oyster populations (Crassostrea virginica; Beck etal. 2011), restoration
techniques, originally applied for shery enhancement and increasing oyster populations, have been
adopted with the primary objective of other services, such as coastal defence and enhancement of
water quality. Oysters need a hard substratum for settlement, which led to a diversity of oyster-reef
structures being developed. These include loose shell, bagged shell and oyster mats and a number
of precast concrete or steel structures (e.g. Oyster Castles®, Ready ReefTM, ReefballTM, ReefBLKSM,
ShoreJAXTM, OysterbreakTM and Wave Attenuating Units®) (Hernandez etal. 2018). Unfortunately
there is a lack of scientic or engineering data to support the use of one approach over another. This
Figure 8 Examples of living shorelines projects along the East and Gulf coasts, United States. (A) Restored
oyster reefs using oyster mats; (B) bagged shell oyster reef living shorelines with saltmarsh; (C) rock sill with
saltmarsh and (D) wave-attenuating units with saltmarsh.
has prevented the development of universal design guidelines appropriate for living shorelines, which
has limited their wider application (Narayan etal. 2016, Morris etal. 2018b). Recent studies have
tried to gauge public perceptions of living shorelines to help inform management of coastal hazards
(Gray etal. 2017). Future research to optimise living shoreline design for coastal protection using
the minimum amount of hard material or optimising the placement of hard material for species
colonisation (e.g. oysters) will further enhance their ability to provide multiple functions and services.
Case Study 3: Hybrid ecoengineering: Sydney
The emphasis of many foreshore projects is to construct efciently engineered armouring structures
that protect infrastructure, but these frequently have negative impacts on the intertidal environment
(Bulleri & Chapman 2010, Ma etal. 2014, Firth etal. 2016b). Projects within the Georges River
estuary, including Carss Bush Park Seawall in New South Wales, Australia, have examined foreshore
infrastructure design from both scientic and engineering approaches, to not only develop an
ecologically responsive foreshore design, but also ensure the successful construction and long-term
structural integrity of projects (Figure 9).
The Carss Bush Park project aimed to design and construct infrastructure that allows natural
coastal processes, such as tidal ingress, to inuence the foreshore. Numerous habitat structures were
introduced into what was a highly degraded environment. These included rock pools at varying
intertidal levels, longer foreshore slopes and crevices, mudats for mangrove and benthic organisms,
a ‘naturalised’ creek line and endangered saltmarsh benches (Heath 2017, Strain etal. 2018a). The
objective of creating such a diverse intertidal zone is to connect the foreshore with previously
Figure 9 Carss Bush Park Foreshore before construction (A and C) and after construction, including
saltmarsh (B) and intertidal habitats (D).
developed ecoengineering projects and natural foreshore areas in order to expand the connectivity
of estuarine/marine ecosystems within the Georges River (Strain etal. 2018a).
The design and construction of this project have led to improvements in the natural sh and
seaweed biodiversity in the articial rock pools (Heath and Moody 2013, Bugnot etal. 2018, Strain
etal. 2018b), while also increasing the social values of the Georges River foreshore (Heath, pers.
obs.). Combined with other successful ecoengineering projects along the Georges River, a cumulative
improvement in habitat availability has been achieved. The project also incorporates adaptability to
future climate change, ensuring its aesthetics and ecological benets are not affected by sea level rise.
The project design reected its heavy recreational use and the highly urbanised nature of the
location by encouraging the reconnection of the community to the foreshore and Kogarah Bay.
This was critical to ensuring the local community’s support (Heath, pers. obs.). Landscaping was
important in making the ecoengineered shorelines scheme aesthetically pleasing to a broad sector
of the community and to promote usability.
The designs also provided environmental benets, not only with the introduction of intertidal
habitats, but with the utilisation of existing seawall sections to create large saltmarsh benches (Heath
2017, Strain etal. 2018a). The use of sections of existing infrastructure in stage 1 reduced the volume
of material leaving the site, while also creating the tidal barrier necessary for the saltmarsh to thrive.
Critical to the design was ensuring that the saltmarsh was sporadically inundated with saline water
during king (spring)tide events but not affected by all tidal events. By keeping sections of the
existing seawall in place, it was not necessary to build new structures or reclaim land from Kogarah
Bay as alternative designs required.
While restoration of the Georges River intertidal foreshore to natural conditions is not achievable
due to its highly urbanised state (Alyazichi etal. 2015), the foreshore ecoengineering projects outlined
previously show how ecological enhancement along a highly urbanised shoreline can be achieved.
These ecoengineering projects protect public land from ooding and erosion, improve native
biodiversity (Heath & Moody 2013, Bugnot etal. 2018, Strain etal. 2018b) and habitat connectivity
(Strain etal. 2018a), with the potential to restore other ecosystem services (e.g. nutrient cycling and
carbon sequestration), while developing relationships with the community through educational and
recreational engagement (Heath 2017).
Process of scaling up successful projects
To date, most hard and soft ecoengineered shorelines have focussed on rehabilitating or adding a
single type of habitat at the site scale (Scyphers etal. 2011, Davis etal. 2015, Strain etal. 2018b).
Enhancements to seawalls are typically at the scale of metres or less (Strain et al. 2018b), while
living shorelines are typically constructed at scales of less than a hectare (e.g. Scyphers etal. 2011,
Davis etal. 2015). In contrast, the scale at which marine ecosystems are degraded by shoreline
development is typically in the order of 10–1 million ha (Edwards & Gomez 2007). To benet
large-scale processes and a broad range of species (Bishop etal. 2017), the scale and complexity of
ecoengineered shorelines, as for restoration projects more generally, need to be much larger than is
commonly considered today (Naveh 1994, Hobbs & Norton 1996, Soulé & Terborgh 1999).
The scaling-up of ecoengineered shorelines may, conceivably, be achieved in several ways. First,
benecial approaches piloted at small scales may simply be applied to larger areas and more sites. For
example, the 1.2-km Elliott Bay seawall (in Seattle; see the section entitled ‘Case studies and scaling
up’, earlier in this review) includes textured and sloped surfaces that were initially experimented on
at the metre scale (Cordell etal. 2017). Similarly, the approximately 1-km-long New York Living
breakwaters project, due to be built following the ‘Rebuild by Design’ competition (http://www., incorporates
features such as bioenhanced concrete blocks and tide-pool armouring units that have previously
been deployed at demonstration scales (Perkol-Finkel & Sella 2015, Sella & Perkol-Finkel 2015).
It should be noted, however, that because many ecological processes are highly scale dependent
(Turner 1989, Wiens 1989, Levin 1992), failure of a particular approach to enhance a desired function
at a small scale does not necessarily preclude its efcacy at a larger scale. Hence, interventions need
to be designed with the scale of the targeted ecological processes in mind.
There are several reasons why ecoengineered shorelines of larger areas may be expected to yield
greater ecological benets than their smaller counterparts. Ecological theory suggests that the number
of species supported by an ecosystem increases with habitat area (see the section entitled ‘Links to
theoretical and community ecology’, earlier in this review; Arrhenius 1921, Gleason 1922). Projects of
larger scale are also more likely to incorporate the heterogeneity required to facilitate ecological processes
critical to ecosystem function (MacArthur & Wilson 1967, Rosenzweig 1995, Tews etal. 2004). Small
areas of habitat are generally considered to be more susceptible to disturbances, as the probability that an
entire habitat patch is affected decreases with increasing size (MacArthur & Wilson 1967, Meurant 2012).
Additionally, related to living shoreline projects, wave attenuation increases with distance inside
habitat patches, although the magnitude of such protection is the greatest in the rst few metres, with
the benets of additional width much diminished (Peterson etal. 2004, Bradley & Houser 2009,
Manca etal. 2012). To date, these predictions remain largely untested and the question of how benets
of ecoengineering scale with extent remains a key unknown. Some functions may be expected to scale
linearly with areal extent; for example, water ltering capacity may increase linearly with the density
or biomass of lter-feeders supported by the ecoengineered shoreline. In contrast, biodiversity, and
the ecosystem functions that depend on it, may display a nonlinear pattern of increase to a maximum
value, beyond which further increases in area have no effect (Yachi & Loreau 1999, Thébault &
Loreau 2006). Monitoring common sets of variables across projects of smaller and larger area would
enable future meta-analyses that address relationships between ecoengineered shoreline extent and
community structure and function. Furthermore, studies that monitor the benets of ecoengineered
shorelines also require the costs (e.g. facilitation of nonnative species) to determine the optimal area of
interventions. Modelling studies that include data on the dispersal capabilities of target and nontarget
pest species might assist in identifying the locations at which ecoengineering initiatives may be
applied to the greatest benet (Bishop etal. 2017).
Second, scaling-up may involve increasing the complexity of projects to include multiple habitats.
Ecoengineered shoreline projects that include multiple components will increase not only alpha-, but
also beta-diversity, and hence ecosystem function (see the section entitled ‘Links to theoretical and
community ecology’, earlier in this review). Many marine species require multiple habitat types in
which to complete their life history (Thorpe 1988, Krumme 2009, Sheaves 2009). Hence, building
ecoengineered shorelines of a single habitat type may fail to provide the resources needed for species
across their life history (Bishop etal. 2017, Morris etal. 2018c). However, whether the inclusion of
multiple habitats in ecoengineered shorelines is desirable or not will depend on the identity (i.e. native
versus nonnative) and function of the species they support, goals of the intervention (e.g. bolstering the
population of a species that depends on multiple habitat types increasing ecosystem function in a severely
degraded urban environment, restoring natural habitat congurations) and, potentially, also whether the
ecoengineered shoreline is present in an area that historically supported multiple or a single habitat type.
Both the Seattle waterfront (see the section entitled ‘Case studies and scaling up’) and the New York
Living Breakwaters projects exemplify how multiple habitat types may be included in hard-engineering
projects. The breakwaters include ‘Reef Streets’, which add habitat for marine life, as well as intertidal
crests, which further increase the complexity and diversity of available habitats. Bioenhancing concrete
units are embedded into the breakwaters’ fabric, providing targeted enhancement measures. These
include multifunction armour units and tide-pools, which by their surface complexity and macro
design increase the available surface area to be utilised as habitat and enhance the ecosystem services
provided by structure (
ny-living-breakwaters). The GreenShores project in Pensacola, Florida, which created more than 12 ha
of oyster reefs, saltmarsh and seagrass habitat along 3.2 km of urban waterfront, is an example of a
living shorelines project targeting multiple habitat types (DEP 2012). The sizes of, distances between
and identity of habitat patches should be informed by knowledge of the seascape ecology of target
species and communities. In addition to the ecological benets of large-scale projects, there may be
signicant socioeconomic benets. As the scale of ecoengineered shoreline projects increases and
they are progressively more grounded in science, an economy of scale is expected to emerge. At
present, however, the evidence for economies of scale in restoration projects is weak due to their almost
universally small scale (Bayraktarov etal. 2016). In some instances, there may even be an inverse
economy of scale caused by large-scale projects proceeding without adequate systems knowledge
(Turner & Boyer 1997). Additionally, large-scale projects may generate more media and public interest,
which may in turn lead to greater opportunities for cofunding and investment.
Several core elements typify projects that have successfully scaled up. These include vision, prior
knowledge, technological capacity, nancial viability and social licence (Menz etal. 2013). Manning
etal. (2006) identify the preemptive constraint of vision as a major impediment to scaling up restoration
more generally and suggest stretch goals and backcasting as two potential approaches to overcome
this. Stretch goals are ambitious, long-term goals that can be used to inspire creativity and innovation
(Manning etal. 2006). Backcasting involves visualising a desired end point and then retrospectively
developing a pathway to that point (Manning etal. 2006). Prior knowledge and technological capacity
are commonly built through small-scale pilot studies. For example, the Seattle Waterfront project (see
the section entitled ‘Case studies and scaling up’, earlier in this review), which, among other goals,
sought to provide a suitable habitat for juvenile salmon foraging and migration, was designed following
years of studying the mechanisms by which urban structures affect juvenile salmon (e.g. Munsch etal.
2014, 2015, 2017a). It built on small-scale demonstration projects, experimenting on the efcacy of
proposed interventions at mitigating these impacts (Goff 2010, Toft etal. 2013).
The division of large-scale projects into multiple stages can also assist in ensuring their success.
Lessons from earlier stages can be used to inform later stages using an adaptive management
approach (see the section entitled ‘Adaptive management in response to monitoring outcomes’, earlier
in this review). Such an approach was used in the Pensacola Green Shores project (DEP 2012).
Worthwhile questions include: What happens if there is failure at the experimental stage? Does it
feed back in a different way in the process of scaling-up? Arguably, success and failure are opposite
ends of the spectrum of learning that an experiment can provide (Firth etal. 2016a), and both can be
used to inform implementation at a larger scale. Instrumental to this line of thought is representing
failure in a way that acknowledges the causal mechanisms and generates reasons with an eye to
future applications. Social licence is built through engaging stakeholders in all steps of the project,
from the planning to the implementation and ongoing monitoring and evaluation phases.
The availability of funding can be a major impediment to large-scale restoration projects
(Manning etal. 2006). As previously noted, most research on the ecoengineered shoreline approach
has occurred in developed countries (in particular Australia, European countries and the United
States). Nevertheless, investment in restoration projects in developing countries can achieve up to
30 times more unit area of habitat than developed countries on a dollar-for-dollar basis, or up to 200
times more unit area when accounting for the local value of the U.S. dollar in developing nations
(Bayraktarov etal. 2016). Data on the success of ecoengineered shorelines are urgently needed from
developing countries to inform the development of large-scale ecoengineering in these areas.
Concluding comments
In summary, while the protection of all-natural ecosystems remains the ideal conservation strategy,
the presence of urban and novel articial habitats should not be ignored. Given their ubiquity in
many countries and increasing prevalence in general, there is a real imperative to compensate for
the negative impacts that seawalls and other articial coastal defences have on shorelines. Examples
of ecological engineering efforts around the world have shown that this is possible, but in general,
few projects have taken an interdisciplinary approach to setting goals, monitoring and evaluation.
This means that there is a paucity of information on the multifunctionality of various ecoengineered
shoreline approaches, inhibiting the development of benets-cost analyses and/or guidelines that
could be adopted at national or international scales.
Most of the projects to date have been done on a small scale and in isolation. The adoption of
more standardised monitoring techniques that included ecological, engineering and socioeconomic
evaluation would enable a more holistic approach to the evaluation and wider implementation of
the ecoengineered shoreline approach globally. Proven techniques are urgently needed so that
governments, policymakers and international organisations seeking sustainable development goals
related to coastlines and oceans have the capacity to develop guidelines, policies and regulations
that promote ecoengineered shoreline strategies, where all-natural solutions are not feasible. Greater
investment in design, experimentation and development of ecoengineered shorelines could have
multiple benets in improving habitat, biodiversity, water quality and reductions in seaborne waste,
resulting in improved access to and appreciation of waterfronts. Doing so would move current
options beyond a prime focus on structures with industrial or defensive functions and to a holistic
approach with environmental goals with multiple applications at a range of scales.
More examples of successful ecoengineered shorelines are required to demonstrate their potential
to provide coastal protection, enhance biodiversity and ecosystem services and increase people’s
enjoyment. In order to change the way that future coastal defences look and function, ecoengineered
shoreline designs generated through interdisciplinary collaborations (e.g. between architects,
engineers, ecologists, social scientists and coastal managers) need to be encouraged. Nonetheless,
it is important to note that, as in most large-scale engineered intervention, the success of a project
at one locality (e.g. Seattle Seawall Project) does not necessarily mean that the same approach will
be successful in others (i.e. one size is unlikely to t all), due to differences in government policies,
ecological goals, site-specic environmental conditions, nancial support and other socioeconomic
factors. As the success of projects from engineering, ecological and socioeconomic standpoints is
demonstrated by a sound evidence base, stakeholders will become more invested in ecoengineered
shoreline development, initiating the support of government and the development of relevant policy.
This article represents a joint effort and collective views from an architect, ecologists, engineers,
a social scientist and a governmental ofcer who participated in the 2nd International Workshop
on Eco-shoreline Designs for Sustainable Coastal Development, held at the University of Hong
Kong, Hong Kong in May 2018. This work was a joint effort from the invited participants plus other
colleagues. The authors would like to thank the Civil Engineering and Development Department
of the Government of the Hong Kong Special Administrative Region of China, the HKU School of
Biological Sciences and the World Harbour Project for supporting the organisation of the workshop.
This research is partially supported by the National Research Foundation, Prime Minister’s
Ofce, Singapore, under its Marine Science Research and Development Programme (Award No.
Aguilera, M.A., Broitman, B.R. & Thiel, M. 2016. Articial breakwaters as garbage bins: structural complexity
enhances anthropogenic litter accumulation in marine intertidal habitats. Environmental Pollution 214,
Aguirre, J.D., Bollard-Breen, B., Cameron, M., Constantine, R., Duffy, C.A.J., Dunphy, B., Hart, K., Hewitt,
J.E., Jarvis, R.M., Jeffs, A., Kahui-McConnell, R., Kawharu, M., Liggins, L., Lohrer, A.M., Middleton,
I., Oldman, J., Sewell, M.A., Smith, A.N.H., Thomas, D.B., Tuckey, B., Vaughan, M. & Wilson, R. 2016.
Loved to pieces: toward the sustainable management of the Waitematā Harbour and Hauraki Gulf.
Regional Studies in Marine Science 8, 220–233.
Airoldi, L., Abbiati, M., Beck, M.W., Hawkins, S.J., Jonsson, P.R., Martin, D., Moschella, P.S., Sundelöf, A.,
Thompson, R.C. & Åberg, P. 2005a. An ecological perspective on the deployment and design of low-
crested and other hard coastal defence structures. Coastal Engineering 52, 1073 –1087.
Airoldi, L., Bacchiocchi, F., Cagliola, C., Bulleri, F. & Abbiati, M. 2005b. Impact of recreational harvesting
on assemblages in articial rocky habitats. Marine Ecology Progress Series 299, 55–66.
Airoldi, L. & Bulleri, F. 2011. Anthropogenic disturbance can determine the magnitude of opportunistic
species responses on marine urban infrastructures. PLOS ONE 6, e22985.
Airoldi, L., Turon, X., Perkol-Finkel, S. & Rius, M. 2015. Corridors for aliens but not for natives: effects of
marine urban sprawl at a regional scale. Diversity and Distributions 21, 755–768.
Allen, J.R. & Hawkins, S.J. 1993. Pelagic-benthic interactions in an enclosed coastal ecosystem: the impact
of Mytilus edulis populations on water quality in the South Docks, Liverpool. In Bivalve Filter Feeders
in Estuarine and Coastal Ecosystem Processes NATO ISI Series G., R.F. Dame (ed.). Berlin: Springer-
Verlag, 515 516.
Allen, J.R., Hawkins, S.J., Russell, G.R. & White, K.N. 1992. Eutrophication and urban renewal: problems and
perspectives for the management of disused docks. In Marine Coastal Eutrophication, R.A. Vollenweider
etal. (eds). Amsterdam: Elsevier, 1283–1295.
Allen, J.R., Wilkinson, S.B. & Hawkins, S.J. 1995. Redeveloped docks as articial lagoons: the development of
brackish-water communities and potential for conservation of lagoonal species. Aquatic Conservation:
Marine and Freshwater Ecosystems 5, 299 –309.
Allen, R.J. & Webb, B.M. 2011. Determination of wave transmission coefcients for oyster shell bag
breakwaters. Coastal Engineering Practice 2011, 684–697.
Alsterberg, C., Roger, F., Sundbäck, K., Juhanson, J., Hulth, S., Hallin, S. & Gamfeldt, L. 2017. Habitat diversity
and ecosystem multifunctionality—the importance of direct and indirect effects. Science Advances 3,
Alyazichi, Y.M., Jones, B.G. & McLean, E. 2015. Source identication and assessment of sediment
contamination of trace metals in Kogarah Bay, NSW, Australia. Environmental Monitoring and
Assessment 187, 20.
Ambrose, R.F. 1994. Mitigating the effects of a coastal power plant on a kelp forest community: rationale and
requirements for an articial reef. Bulletin of Marine Science 55, 694–708.
Anderson, M.J. 1996. A chemical cue induces settlement of Sydney rock oysters, Saccostrea commercialis, in
the laboratory and in the eld. The Biological Bulletin 190, 350–358.
Anderson, M.J. 2001. A new method for non-parametric multivariate analysis of variance. Austral Ecology
26, 32–46.
Arenas, F., Sánchez, I., Hawkins, S.J. & Jenkins, S.R. 2006. The invasibility of marine algal assemblages: role
of functional diversity and identity. Ecology 87, 2851–2861.
Arkema, K.K., Grifn, R., Maldonado, S., Silver, J., Suckale, J. & Guerry, A.D. 2017. Linking social, ecological,
and physical science to advance natural and nature-based protection for coastal communities. Annals of
the New York Academy of Sciences 1399, 5–26.
Arkema, K.K., Guannel, G., Verutes, G., Wood, S.A., Guerry, A., Ruckelshaus, M., Kareiva, P., Lacayo, M.
& Silver, J.M. 2013. Coastal habitats shield people and property from sea-level rise and storms. Nature
Climate Change 3, 913.
Armono, H.D. & Hall, K.R. 2003. Wave Transmission on Submerged Breakwaters Made of Hollow
Hemispherical Shape Articial Reefs. A paper presented at the 1st Coastal, Estuart and Offshore
Engineerign Speciality Conference of the Canadian Society for Civil Engineering, at Moncton, Nouveau-
Brunswick, Canada (4–7 June 2003), CSC-185-1–CSC-185-10.
Aronson, J., Floret, C., Le Floc’h, E., Ovalle, C. & Pontanier, R. 1993. Restoration and rehabilitation of
degraded ecosystems in arid and semi-arid lands. II. Case studies in Southern Tunisia, Central Chile and
Northern Cameroon. Restoration Ecology 1, 168–187.
Arrhenius, O. 1921. Species and area. Journal of Ecology 9, 95–99.
Astudillo, J.C., Bonebrake, T.C. & Leung, K.M.Y. 2018. Deterred but not preferred: predation by native whelk
Reishia clavigera on invasive bivalves. PLOS ONE 13, e0196578.
Astudillo, J.C., Leung, K.M.Y. & Bonebrake, T.C. 2016. Seasonal heterogeneity provides a niche opportunity
for ascidian invasion in subtropical marine communities. Marine Environmental Research 122, 1–10.
Bally, R., McQuaid, C.D. & Brown, A.C. 1984. Shores of mixed sand and rock—an unexplored marine
ecosystem. South African Journal of Science 80, 500–503.
Ban, N.C., Maxwell, S.M., Dunn, D.C., Hobday, A.J., Bax, N.J., Ardron, J., Gjerde, K.M., Game, E.T., Devillers,
R., Kaplan, D.M., Dunstan, P.K., Halpin, P.N. & Pressey, R.L. 2014. Better integration of sectoral planning
and management approaches for the interlinked ecology of the open oceans. Marine Policy 49, 127–136.
Barbier, E.B., Hacker, S.D., Kennedy, C., Koch, E.W., Stier, A.C. & Silliman, B.R. 2011. The value of estuarine
and coastal ecosystem services. Ecological Monographs 81, 169–193.
Barrios-O’Neill, D., Dick, J.T.A., Emmerson, M.C., Ricciardi, A. & Macisaac, H.J. 2014. Predator-free space,
functional responses and biological invasions. Functional Ecology 29, 377–384.
Baura, G.D. 2006. Engineering Ethics: An Industrial Perspective. Amsterdam: Academic Press.
Bayne, B.L. 1964. The responses of the larvae of Mytilus edulis L. to light and to gravity. Oikos 15, 162–174.
Bayraktarov, E., Saunders, M.I., Abdullah, S., Mills, M., Beher, J., Possingham, H.P., Mumby, P.J. & Lovelock,
C.E. 2016. The cost and feasibility of marine coastal restoration. Ecological Applications 26, 1055–1074.
Beck, M.W., Brumbaugh, R.D., Airoldi, L., Carranza, A., Coen, L.D., Crawford, C., Defeo, O., Edgar, G.J., Hancock,
B., Kay, M.C., Lenihan, H.S., Luckenbach, M.W., Toropova, C.L., Zhang, G. & Guo, X. 2011. Oyster reefs at
risk and recommendations for conservation, restoration, and management. BioScience 61, 107–116.
Beck, M.W., Heck, K.L., Able, K.W., Childers, D.L., Eggleston, D.B., Gillanders, B.M., Halpern, B., Hays,
C.G., Hoshino, K., Minello, T.J., Orth, R.J., Sheridan, P.F. & Weinstein, M.P. 2001. The identication,
conservation, and management of estuarine and marine nurseries for sh and invertebrates. BioScience
51, 633 – 641.
Beck, M.W., Losada, I.J., Menéndez, P., Reguero, B.G., Díaz-Simal, P. & Fernández, F. 2018. The global ood
protection savings provided by coral reefs. Nature Communications 9, 2186.
Benoit, J., Hardaway, J.C.S., Hernandez, D., Holman, R., Koch, E., McLellan, N., Peterson, S., Reed, D. & Suman,
D. 2007. Mitigating Shore Erosion Along Sheltered Coasts, Washington, DC: National Research Council.
Benthem, W., van Lavieren, L.P. & Verheugt, W.J.M. 1999. Mangrove rehabilitation in the coastal Mekong
Delta, Vietnam. In An International Perspective on Wetland Rehabilitation, W. Streever (ed.). Dordrecht,
the Netherlands: Springer Netherlands, 29–36.
Berntsson, K.M., Jonsson, P.R., Lejhall, M. & Gatenholm, P. 2000. Analysis of behavioural rejection of
micro-textured surfaces and implications for recruitment by the barnacle Balanus improvisus. Journal
of Experimental Marine Biology and Ecology 251, 59–83.
Bilkovic, D.M. & Mitchell, M.M. 2013. Ecological tradeoffs of stabilized salt marshes as a shoreline protection
strategy: effects of articial structures on macrobenthic assemblages. Ecological Engineering 61, 46 9– 481.
Bilkovic, D.M. & Mitchell, M. 2017. Designing living shoreline salt marsh ecosystems to promote coastal
resilience. In Living Shorelines: The Science and Management of Nature-Based Coastal Protection,
D.M. Bilkovic etal. (eds). Boca Raton, Florida: CRC Press, 293–316.
Bilkovic, D.M., Mitchell, M., Mason, P. & Duhring, K. 2016. The role of living shorelines as estuarine habitat
conservation strategies. Coastal Management 44, 161–174.
Bilkovic, D.M., Mitchell, M.M., Toft, J.D. & La Peyre, M.K. 2017. A primer to living shorelines. In Living
Shorelines: The Science and Management of Nature-based Coastal Protection, D.M. Bilkovic etal.
(eds). Boca Raton, Florida: CRC Press, 3–10.
Bin, O. & Polasky, S. 2005. Evidence on the amenity value of wetlands in a rural setting. Journal of Agricultural
and Applied Economics 37, 589–602.
Bishop, M.J., Mayer-Pinto, M., Airoldi, L., Firth, L.B., Morris, R.L., Loke, L.H., Hawkins, S.J., Naylor, L.A.,
Coleman, R.A. & Chee, S.Y. 2017. Effects of ocean sprawl on ecological connectivity: impacts and
solutions. Journal of Experimental Marine Biology and Ecology 492, 7–30.
Block, W.M., Franklin, A.B., Ward, J.P., Jr, Ganey, J.L. & White, G.C. 2001. Design and implementation of
monitoring studies to evaluate the success of ecological restoration on wildlife. Restoration Ecology 9,
Borsje, B.W., van Wesenbeeck, B.K., Dekker, F., Paalvast, P., Bouma, T.J., van Katwijk, M.M. & De Vries, M.B.
2011. How ecological engineering can serve in coastal protection. Ecological Engineering 37, 113 –122 .
Bouma, T.J., van Belzen, J., Balke, T., Zhu, Z.C., Airoldi, L., Blight, A.J., Davies, A.J., Galvan, C., Hawkins,
S.J., Hoggart, S.P.G., Lara, J.L., Losada, I.J., Maza, M., Ondiviela, B., Skov, M.W., Strain, E.M.,
Thompson, R.C., Yang, S.L., Zanuttigh, B., Zhang, L.Q. & Herman, P.M.J. 2014. Identifying knowledge
gaps hampering application of intertidal habitats in coastal protection: opportunities and steps to take.
Coastal Engineering 87, 147–157.
Bracewell, S.A., Robinson, L.A., Firth, L.B. & Knights, A.M. 2013. Predicting free-space occupancy on novel
articial structures by an invasive intertidal barnacle using a removal experiment. PLOS ONE 8, e74457.
Bradley, K. & Houser, C. 2009. Relative velocity of seagrass blades: implications for wave attenuation in low-
energy environments. Journal of Geophysical Research: Earth Sur face 114 , F01004.
Bradshaw, A.D. 1997. What do we mean by restoration. In Restoration Ecology and Sustainable Development,
K.M. Urbanska etal. (eds). Cambridge, UK: Cambridge University Press, 8–14.
Brown, G. & Hausner, V.H. 2017. An empirical analysis of cultural ecosystem values in coastal landscapes.
Ocean & Coastal Management 142, 49– 60.
Browne, M.A. & Chapman, M.G. 2011. E cologically informed engineeri ng reduces loss of intertidal biodiversity
on articial shorelines. Environmental Science & Technology 45, 8204–8207.
Bugnot, A.B., Hose, G.C., Walsh, C.J., Floerl, O., French, K., Dafforn, K.A., Hanford, J., Lowe, E.C. & Hahs,
A.K. 2019. Urban impacts across realms: making the case for inter-realm monitoring and management.
Science of the Total Environment 648, 711–719.
Bugnot, A.B., Mayer-Pinto, M., Johnston, E.L., Schaefer, N. & Dafforn, K.A. 2018. Learning from nature to
enhance Blue engineering of marine infrastructure. Ecological Engineering 120, 611– 621.
Bulleri, F. & Airoldi, L. 2005. Articial marine structures facilitate the spread of a non-indigenous green
alga, Codium fragile ssp. tomentosoides, in the north Adriatic Sea. Journal of Applied Ecology 42,
Bulleri, F. & Chapman, M.G. 2010. The introduction of coastal infrastructure as a driver of change in marine
environments. Journal of Applied Ecology 47, 26–35.
Bulleri, F., Cristaudo, C., Alestra, T. & Benedetti-Cecchi, L. 2011. Crossing gradients of consumer pressure
and physical stress on shallow rocky reefs: a test of the stress-gradient hypothesis. Journal of Ecology
99, 335–344.
Bulleri, F., Eriksson, B., Queirós, A., Airoldi, L., Arenas, F., Arvanitidis, C., Bouma, T., Crowe, T., Davoult,
D., Guizien, K., Iveša, L., Jenkins, S., Michalet, R., Olabarria, C., Procaccini, G., Serrão, E., Wahl, M.
& Benedetti-Cecchi, L. 2018. Harnessing positive species interactions as a tool against climate-driven
loss of coastal biodiversity. PLOS Biology 16, e2006852.
Bunting, D. & Coleman, R.A. 2014. Ethical consideration in invasion ecology: a marine perspective. Ecological
Management & Restoration 15, 64–70.
Burcharth, H.F., Hawkins, S.J., Zanuttigh, B. & Lamberti, A. 2007. Environmental Design Guidelines of Low
Crested Coastal Defence Structures. Amsterdam: Elsevier.
Burcharth, H.F., Kramer, M., Lamberti, A. & Zanuttigh, B. 2006. Structural stability of detached low crested
breakwaters. Coastal Engineering 53, 381–394.
Butman, C.A. 1987. Larval settlement of soft-sediment invertebrates - the spatial scales of pattern explained by
active habitat selection and the emerging role of hydrodynamical processes. Oceanography and Marine
Biology 25, 113 –165.
Cardinale, B.J., Duffy, J.E., Gonzalez, A., Hooper, D.U., Perrings, C., Venail, P., Narwani, A., Mace, G.M.,
Tilman, D., Wardle, D.A., Kinzig, A.P., Daily, G.C., Loreau, M., Grace, J.B., Larigauderie, A., Srivastava,
D.S. & Naeem, S. 2012. Biodiversity loss and its impact on humanity. Nature 486, 59 67.
Castaneda, A., Albear, J.J.H., Corvo, F. & Marrero, R. 2017. Concrete quality assessment before building
structures submitting to environmental exposure conditions. Revista de la Construccion 16, 3 74 –387.
Chafn, B.C., Gosnell, H. & Cosens, B.A. 2014. A decade of adaptive governance scholarship: synthesis and
future directions. Ecology and Society 19, 56.
Chapman, M.G. & Blockley, D.J. 2009. Engineering novel habitats on urban infrastructure to increase intertidal
biodiversity. Oecologia 161, 625– 635.
Chapman, M.G., Tolhurst, T., Murphy, R. & Underwood, A. 2010. Complex and inconsistent patterns of
variation in benthos, micro-algae and sediment over multiple spatial scales. Marine Ecology Progress
Series 398, 33– 47.
Chapman, M.G. & Underwood, A.J. 2011. Evaluation of ecological engineering of ‘armoured’ shorelines to
improve their value as habitat. Journal of Experimental Marine Biology and Ecology 400, 302–313.
Chapman, M.G., Underwood, A. & Browne, M.A. 2018. An assessment of the current usage of ecological
engineering and reconciliation ecology in managing alterations to habitats in urban estuaries. Ecological
Engineering 120, 560–573.
Charlier, R.H., Chaineux, M.C.P. & Morcos, S. 2005. Panorama of the history of coastal protection. Journal
of Coastal Research 21, 79 111.
Chee, S.Y., Othman, A.G., Sim, Y.K., Mat Adam, A.N. & Firth, L.B. 2017. Land reclamation and articial
islands: walking the tightrope between development and conservation. Global Ecology and Conservation
12, 80 95.
Chung, I.K., Beardall, J., Mehta, S., Sahoo, D. & Stojkovic, S. 2011. Using marine macroalgae for carbon
sequestration: a critical appraisal. Journal of Applied Phycology 23, 877–886.
CIRIA, CUR & CETMEF. 2007. The Rock Manual. The Use of Rock in Hydraulic Engineering, London:
CIRIA, 2nd edition.
Claesson, S. 2011. The Value and valuation of maritime cultural heritage. International Journal of Cultural
Propert y 18, 61– 8 0.
Coen, L.D., Brumbaugh, R.D., Bushek, D., Grizzle, R., Luckenbach, M.W., Posey, M.H., Powers, S.P. &
Tolley, S.G. 2007. Ecosystem services related to oyster restoration. Marine Ecology Progress Series
341, 303–307.
Cohen, J. 1988. Statistical Power Analysis for the Behavioral Sciences, Hillsdale, NJ: Erlbaum Associates,
2nd edition.
Coleman, F.C. & Williams, S.L. 2002. Overexploiting marine ecosystem engineers: potential consequences
for biodiversity. Trends in Ecology & Evolution 17, 40–44.
Coleman, R.A., Underwood, A.J., Benedetti-Cecchi, L., Åberg, P., Arenas, F., Arrontes, J., Castro, J., Hartnoll,
R.G., Jenkins, S.R. & Paula, J. 2006. A continental scale evaluation of the role of limpet grazing on rocky
shores. Oecologia 147, 556–564.
Connell, J.H. 1978. Diversity in tropical rain forests and coral reefs. Science 199, 1302–1310.
Connell, J.H. & Slatyer, R.O. 1977. Mechanisms of succession in natural communities and their role in
community stability and organization. The American Naturalist 111, 1119–1144.
Coombes, M.A. 2014. The rock coast of the British Isles: weathering and biogenic processes. In Rock Coast
Geomorphology: A Global Synthesis, D.M. Kenney etal. (eds). London: The Geological Society.
Coombes, M.A., La Marca, E.C., Naylor, L.A. & Thompson, R.C. 2015. Getting into the groove: opportunities
to enhance the ecological value of hard coastal infrastructure using ne-scale surface textures. Ecological
Engineering 77, 314–323.
Coombes, M.A., Naylor, L.A., Viles, H.A. & Thompson, R.C. 2013. Bioprotection and disturbance: seaweed,
microclimatic stability and conditions for mechanical weathering in the intertidal zone. Geomorphology
202, 4 –14.
Coombes, M.A., Viles, H.A., Naylor, L.A. & La Marca, E.C. 2017. Cool barnacles: do common biogenic
structures enhance or retard rates of deterioration of intertidal rocks and concrete? Science of the Total
Environment 580, 1034 –1045.
Cooper, J.A.G., O’Connor, M.C. & McIvor, S. 2016. Coastal defences versus coastal ecosystems: a regional
appraisal. Marine Policy.
Cordell, J.R., Toft, J.D., Munsch, S.H. & Goff, M. 2017. Benches, beaches, and bumps: how habitat monitoring
and experimental science can inform urban seawall design. In Living Shorelines: The Science and
Management of Nature-based Coastal Protection, D.M. Bilkovic etal. (eds). Boca Raton, Florida: CRC
Press, 421438.
Costanza, R., d’Arge, R., de Groot, R., Farber, S., Grasso, M., Hannon, B., Limburg, K., Naeem, S., O’Neill,
R.V., Paruelo, J., Raskin, R.G., Sutton, P. & van den Belt, M. 1997. The value of the world’s ecosystem
services and natural capital. Nature 387, 253–260.
Costanza, R., de Groot, R., Sutton, P., van der Ploeg, S., Anderson, S.J., Kubiszewski, I., Farber, S. & Turner,
R.K. 2014. Changes in the global value of ecosystem services. Global Environmental Change 26,
Craft, C., Reader, J., Sacco, J.N. & Broome, S.W. 1999. Twenty-ve years of ecosystem development of
constructed Spartina alterniora (Loisel) marshes. Ecological Applications 9, 1405–1419.
Crawley, D. & Aho, I. 1999. Building environmental assessment methods: applications and development trends.
Building Research & Information 27, 300–308.
Crossett, K.M., Culliton, T.J., Wiley, P.C. & Goodspeed, T.R. 2004. Population Trends along the Coastal
United States: 1980 2008. Government Printing Ofce.
da Vriend, H., Koningsveld, M.V. & Aarninkhof, S. 2014. ‘Building with nature’: the new Dutch approach to
coastal and river works. Proceedings of the Institution of Civil Engineers 167, 18–24.
Dafforn, K.A. 2017. Eco-engineering and management strategies for marine infrastructure to reduce
establishment and dispersal of non-indigenous species. Management of Biological Invasions 8, 153 –161.
Dafforn, K.A., Glasby, T.M., Airoldi, L., Rivero, N.K., Mayer-Pinto, M. & Johnston, E.L. 2015a.
Marine urbanization: an ecological framework for designing multifunctional articial structures.
Frontiers in Ecology and the Environment 13, 82–90.
Dafforn, K.A., Glasby, T.M. & Johnston, E.L. 2012. Comparing the invasibility of experimental ‘reefs’ with
eld observations of natural reefs and articial structures. PLOS ONE 7, e38124.
Dafforn, K.A., Mayer-Pinto, M., Morris, R.L. & Waltham, N.J. 2015b. Application of management tools to
integrate ecological principles with the design of marine infrastructure. Journal of Environmental
Management 158, 61–73.
Dangendorf, S., Marcos, M., Wöppelmann, G., Conrad, C.P., Frederikse, T. & Riva, R. 2017. Reassessment of
20th century global mean sea level rise. Proceedings of the National Academy of Sciences 114, 594 6–5951.
Danielsen, F., Sørensen, M.K., Olwig, M.F., Selvam, V., Parish, F., Burgess, N.D., Hiraishi, T., Karunagaran,
V.M., Rasmussen, M.S. & Hansen, L.B. 2005. The Asian tsunami: a protective role for coastal vegetation.
Science 310, 643643.
Davis, J.L., Currin, C.A., O’Brien, C., Raffenburg, C. & Davis, A. 2015. Living shorelines: coastal resilience
with a blue carbon benet. PLOS ONE 10, e 0142595.
de Groot, R.S., Wilson, M.A. & Boumans, R.M. 2002. A typology for the classication, description and
valuation of ecosystem functions, goods and services. Ecological Economics 41, 393–408.
de Schipper, M.A., de Vries, S., Ruessink, G., de Zeeuw, R.C., Rutten, J., van Gelder-Maas, C. & Stive, M.J.
2016. Initial spreading of a mega feeder nourishment: observations of the Sand Engine pilot project.
Coastal Engineering 111, 23–38.
DEP. 2012. Project GreenShores. Florida: U.S. Department of Environmental Protection.
Dethier, M.N., Raymond, W.W., McBride, A.N., Toft, J.D., Cordell, J.R., Ogston, A.S., Heerhartz, S.M. &
Berry, H.D. 2016. Multiscale impacts of armoring on Salish Sea shorelines: evidence for cumulative and
threshold effects. Estuarine, Coastal and Shelf Science 175, 106 117.
Díaz, S., Fargione, J., Chapin, F.S., III & Tilman, D. 2006. Biodiversity loss threatens human well-being.
PLOSBiology 4, e277.
Dong, Y.-W., Huang, X.-W., Wang, W., Li, Y. & Wang, J. 2016. The marine ‘great wall’ of China: local-
and broad-scale ecological impacts of coastal infrastructure on intertidal macrobenthic communities.
Diversity and Distributions 22, 731–74 4.
Duan, H., Zhang, H., Huang, Q., Zhang, Y., Hu, M., Niu, Y. & Zhu, J. 2016. Characterization and environmental
impact analysis of sea land reclamation activities in China. Ocean & Coastal Management 130, 12 8–13 7.
Dubi, A.M. & Tørum, A. 1994. Wave damping by kelp vegetation. In Proceedings of 24th International
Conference on Coastal Engineering. Kobe, Japan: ASCE.
Dubois, S., Retière, C. & Olivier, F. 2002. Biodiversity associated with Sabellaria alveolata (Polychaeta:
Sabellariidae) reefs: effects of human disturbances. Journal of the Marine Biological Association of the
United Kingdom 82, 817–826.
Dugan, J., Airoldi, L., Chapman, M., Walker, S., Schlacher, T. 2011. Estuarine and coastal structures:
environmental effects, a focus on shore and nearshore structures. In Treatise on Estuarine and Coastal
Science, Vol. 8, E. Wolanski & D. McLusky (eds). Waltham, Massachusetts: Academic Press, 17–41.
Dumont, C.P., Harris, L.G. & Gaymer, C.F. 2011. Anthropogenic structures as a spatial refuge from predation
for the invasive bryozoan Bugula neritina. Marine Ecology Progress Series 427, 95–103.
Dyson, K. & Yocom, K. 2015. Ecological design for urban waterfronts. Urban Ecosystems 18, 189–2 08.
Edwards, A.J. & Gomez, E.D. 2007. Reef Restoration Concepts and Guidelines: Making Sensible Management
Choices in the Face of Uncertaint y. St Lucia, Australia: Coral Reef Targeted Research & Capacity
Building for Management Programme.
Elliott, M., Mander, L., Mazik, K., Simenstad, C., Valesini, F., Whiteld, A. & Wolanski, E. 2016.
Ecoengineering with ecohydrology: successes and failures in estuarine restoration. Estuarine, Coastal
and Shelf Science 176, 12–35.
Ellis, E.C. 2015. Ecology in an anthropogenic biosphere. Ecological Monographs 85, 287–331.
Evans, A.J., Firth, L.B., Hawkins, S.J., Hall, A.E., Ironside, J.E., Thompson, R.C. & Moore, P.J. 2019. From
ocean sprawl to blue-green infrastructure – A UK perspective on an issue of global signicance.
Environmental Science & Policy 91, 60 69.
Evans, A.J., Firth, L.B., Hawkins, S.J., Morris, E.S., Goudge, H. & Moore, P.J. 2016. Drill-cored rock pools:an
effective method of ecological enhancement on articial structures. Marine and Freshwater Research
67, 123–130.
Evans, A.J., Garrod, B., Firth, L.B., Hawkins, S.J., Morris-Webb, E.S., Goudge, H. & Moore, P.J. 2017.
Stakeholder priorities for multi-functional coastal defence developments and steps to effective
implementation. Marine Policy 75, 143–155.
Ferrario, F., Beck, M.W., Storlazzi, C.D., Micheli, F., Shepard, C.C. & Airoldi, L. 2014. The effectiveness of
coral reefs for coastal hazard risk reduction and adaptation. Nature Communications 5, 3794.
Ferrario, F., Iveša, L., Jaklin, A., Perkol-Finkel, S. & Airoldi, L. 2015. The overlooked role of biotic factors
in controlling the ecological performance of articial marine habitats. Journal of Applied Ecology 53,
Findlay, S.E.G., Miller, J.K., William, A. & Hauser, E.E. 2018. Hudson River Sustainable Shorelines
Rapid Assessment Protocol – Manual. Staatsburg, NY: Hudson River Sustainable Shorelines Project.
Finucane, M.L., Clancy, N., Willis, H.H. & Knopman, D. 2014. The Hurricane Sandy Rebuilding Task Force’s
Infrastructure Resilience Guidelines: An Initial Assessment of Implementation by Federal Agencies,
Santa Monica, California: RAND Corporation.
Firth, L.B., Browne, K.A., Knights, A.M., Hawkins, S.J. & Nash, R. 2016a. Eco-engineered rock pools: a
concrete solution to biodiversity loss and urban sprawl in the marine environment. Environmental
Research Letters 11, 094015.
Firth, L.B., Knights, A.M., Bridger, D., Evans, A.J., Mieszkowska, N., Moore, P.J., O’Connor, N.E., Sheehan,
E.V., Thompson, R.C. & Hawkins, S.J. 2016b. Ocean sprawl: challenges and opportunities for biodiversity
management in a changing world. Oceanography and Marine Biology: An Annual Review 54, 189–262.
Firth, L.B., Mieszkowska, N., Grant, L.M., Bush, L.E., Davies, A.J., Frost, M.T., Moschella, P.S., Burrows,
M.T., Cunningham, P.N., Dye, S.R. & Hawkins, S.J. 2015. Historical comparisons reveal multiple drivers
of decadal change of an ecosystem engineer at the range edge. Ecology and Evolution 5, 3210–3222.
Firth, L.B., Thompson, R.C., Bohn, K., Abbiati, M., Airoldi, L., Bouma, T.J., Bozzeda, F., Ceccherelli, V.U.,
Colangelo, M.A., Evans, A., Ferrario, F., Hanley, M.E., Hinz, H., Hoggart, S.P.G., Jackson, J.E., Moore,
P., Morgan, E.H., Perkol-Finkel, S., Skov, M.W., Strain, E.M., Van Belzen, J. & Hawkins, S.J. 2014.
Between a rock and a hard place: environmental and engineering considerations when designing coastal
defence structures. Coastal Engineering 87, 122–135.
Fletcher, C.H., Mullane, R.A. & Richmond, B.M. 1997. Beach loss along armored shorelines on Oahu,
Hawaiian Islands. Journal of Coastal Research 13, 209–215.
Fletcher, R.L. & Callow, M.E. 1992. The settlement, attachment and establishment of marine algal spores.
British Phycological Journal 27, 303–329.
Floerl, O., Inglis, G.J., Dey, K. & Smith, A. 2009. The importance of transport hubs in stepping-stone invasions.
Journal of Applied Ecology 46, 37–45.
Folke, C., Hahn, T., Olsson, P. & Norberg, J. 2005. Adaptive governance of social-ecological systems. Annual
Review of Environment and Resources 30, 441– 473.
Fox, J.W. 2013. The intermediate disturbance hypothesis should be abandoned. Trends in Ecology & Evolution
28, 86–92.
Foxon, T.J., Gross, R., Chase, A., Howes, J., Arnall, A. & Anderson, D. 2005. UK innovation systems for new
and renewable energy technologies: drivers, barriers and systems failures. Energy Policy 33, 2123 –2137.
Francis, R.A. & Lorimer, J. 2011. Urban reconciliation ecology: the potential of living roofs and walls. Journal
of Environmental Management 92, 1429 1437.
Franco, L. 1996. History of coastal engineering in Italy. In History and Heritage of Coastal Engineering, N.
Kraus (ed.). New York: ASCE, 275–335.
Friess, D.A. 2016. Ecosystem services and disservices of mangrove forests: insights from historical colonial
observations. Forests 7, 183.
Garbisch, E.W. & Garbisch, J.L. 1994. Control of upland bank erosion through tidal marsh construction on
restored shores: application in the Maryland portion of Chesapeake Bay. Environmental Management
18, 677– 691.
Garvis, S.K. 2009. Quantifying the impacts of oyster reef restoration on oyster coverage, wave dissipation and
seagrass recruitment in Mosquito Lagoon, Florida. MSc, University of Central Florida, Orlando.
Gedan, K.B., Kirwan, M.L., Wolanski, E., Barbier, E.B. & Silliman, B.R. 2011. The present and future role
of coastal wetland vegetation in protecting shorelines: answering recent challenges to the paradigm.
Climatic Change 106, 7–29.
Geist, J. & Hawkins, S.J. 2016. Habitat recovery and restoration in aquatic ecosystems: current progress and
future challenges. Aquatic Conservation: Marine and Freshwater Ecosystems 26, 942–962.
Gelcich, S., Buckley, P., Pinnegar, J.K., Chilvers, J., Lorenzoni, I., Terry, G., Guerrero, M., Castilla, J.C.,
Valdebenito, A. & Duarte, C.M. 2014. Public awareness, concerns, and priorities about anthropogenic
impacts on marine environments. Proceedings of the National Academy of Sciences 111, 15042–15047.
Gifford, S., Dunstan, H., O’Connor, W. & MacFarlane, G. 2005. Quantication of in situ nutrient and heavy
metal remediation by a small pearl oyster (Pinctada imbricata) farm at Port Stephens, Australia. Marine
Pollution Bulletin 50, 41742 2.
Gifford, S., Dunstan, R., O’Connor, W., Roberts, T. & Toia, R. 2004. Pearl aquaculture—protable
environmental remediation? Science of the Total Environment 319, 27–37.
Giller, P., Hillebrand, H., Berninger, U.-G., Gessner, M.O., Hawkins, S., Inchausti, P., Inglis, C., Leslie,
H., Malmqvist, B., Monaghan, M.T., Morin, P.J. & O’Mullan, G. 2004. Biodiversity effects on
ecosystem functioning: emerging issues and their experimental test in aquatic environments. Oikos
104, 423–436.
Gittman, R.K., Fodrie, F.J., Popowich, A.M., Keller, D.A., Bruno, J.F., Currin, C.A., Peterson, C.H. & Piehler,
M.F. 2015. Engineering away our natural defenses: an analysis of shoreline hardening in the US. Frontiers
in Ecology and the Environment 13, 301–307.
Gittman, R.K., Peterson, C.H., Currin, C.A., Joel Fodrie, F., Piehler, M.F. & Bruno, J.F. 2016. Living shorelines
can enhance the nursery role of threatened estuarine habitats. Ecological Applications 26, 249–263.
Gittman, R.K., Popowich, A.M., Bruno, J.F. & Peterson, C.H. 2014. Marshes with and without sills protect
estuarine shorelines from erosion better than bulkheads during a Category 1 hurricane. Ocean & Coastal
Management 102, 94 –102.
Gittman, R.K. & Scyphers, S.B. 2017. The cost of coastal protection: a comparison of shore stabilization
approaches. Shore and Beach 85, 19 –2 4.
Glasby, T.M., Connell, S.D., Holloway, M.G. & Hewitt, C.L. 2007. Nonindigenous biota on articial structures:
could habitat creation facilitate biological invasions? Marine Biology 151, 887–895.
Gleason, H.A. 1922. On the relation between species and area. Ecology 3, 158–162 .
Godbold, J.A., Bulling, M.T. & Solan, M. 2011. Habitat structure mediates biodiversity effects on ecosystem
properties. Proceedings: Biological Sciences 278, 2 510 –2518.
Goff, M. 2010. Evaluating habitat enhancements of an urban intertidal seawall: ecological responses and
management implications. MSc, University of Washington, Seattle.
Grabowski, J.H., Brumbaugh, R.D., Conrad, R.F., Keeler, A.G., Opaluch, J.J., Peterson, C.H., Piehler, M.F.,
Powers, S.P. & Smyth, A.R. 2012. Economic valuation of ecosystem services provided by oyster reefs.
BioScience 62, 900–909.
Grabowski, J.H. & Peterson, C.H. 2007. Restoring oyster reefs to recover ecosystem services. In Theoretical
Ecology Series, K. Cuddington etal. (eds). Cambridge, Massachusetts: Academic Press, 281–298.
Gracia, A., Rangel-Buitrago, N., Oakley, J.A. & Williams, A.T. 2018. Use of ecosystems in coastal erosion
management. Ocean & Coastal Management 156, 277–289.
Gray, J.D.E., O’Neill, K. & Qiu, Z. 2017. Coastal residents’ perceptions of the function of and relationship
between engineered and natural infrastructure for coastal hazard mitigation. Ocean & Coastal
Management 146, 144–156.
Green, D.S., Chapman, M.G. & Blockley, D.J. 2012. Ecological consequences of the type of rock used in the
construction of articial boulder-elds. Ecological Engineering 46, 1–10.
Grifn, J.N., de la Haye, K.L., Hawkins, S.J., Thompson, R.C. & Jenkins, S.R. 2008. Predator diversity and
ecosystem functioning: density modies the effect of resource partitioning. Ecology 89, 298–305.
Grif n, J.N., Jenkins, S.R., Gamfeldt, L., Jones, D., Hawkins, S.J. & Thompson, R.C. 2009. Spatial heterogeneity
increases the importance of species richness for an ecosystem process. Oikos 118, 1335–1342.
Grifn, J.N., Noel, L., Crowe, T.P., Burrows, M.T., Hawkins, S.J., Thompson, R.C. & Jenkins, S.R. 2010.
Consumer effects on ecosystem functioning in rock pools: roles of species richness and composition.
Marine Ecology Progress Series 420, 45–56.
Grillo, S. 1989. Venice Sea Defences. Venice: Arsenale Publishing. (In Italian).
Hadjimanolis, A. 1999. Barriers to innovation for SMEs in a small less developed country (Cyprus).
Technovation 19, 561–570.
Hallegatte, S., Green, C., Nicholls, R.J. & Corfee-Morlot, J. 2013. Future ood losses in major coastal cities.
Nature Climate Change 3, 802–806.
Hassan, R., Scholes, R. & Ash, N. 2005. Ecosystems and human well-being: Current state and trends, Volume
1. Millennium Ecosystem Assessment, Global Assessment Reports. Washington: Island Press. Harlin,
M.M. & Lindbergh, J.M. 1977. Selection of substrata by seaweeds: Optimal surface relief. Marine
Biology 40, 33– 40.
Hashim, R., Kamali, B., Tamin, N.M. & Zakaria, R. 2010. An integrated approach to coastal rehabilitation:
mangrove restoration in Sungai Haji Dorani, Malaysia. Estuarine, Coastal and Shelf Science 86, 118 –124.
Hawkins, S., Allen, J., Russell, G., White, K., Conlan, K., Hendry, K. & Jones, H. 1992a. Restoring and
managing disused docks in inner city areas. In Restoring the Nation’s Marine Environment, S.J. Hawkins
etal. (eds). College Park, Maryland: Schiffer, 473–542.
Hawkins, S., Cunningham, P., Dolan, B., Evans, L., Holmes, G., O’Hara, K., Russell, G., Walmsley, A. &
White, K. 1992b. Culture of mussels in Sandon dock, a disused dock basin in Liverpool. Journal of
Medical and Applied Malacology 4, 165–178.
Hawkins, S.J. 2004. Scaling up: the role of species and habitat patches in functioning of coastal ecosystems.
Aquatic Conservation: Marine and Freshwater Ecosystems 14 , 217–219.
Hawkins, S.J., Allen, J.R. & Bray, S. 1999. Restoration of temperate marine and coastal ecosystems: nudging
nature. Aquatic Conservation: Marine and Freshwater Ecosystems 9, 23–46.
Hawkins, S.J., Allen, J.R., Ross, P.M. & Genner, M.J. 2002. Marine and coastal ecosystems. In Handbook
of Ecological Restoration, M.R. Perrow & A.J. Davy (eds). Cambridge: Cambridge University Press.
Hawkins, S.J., Allen, J.R., Russell, G., Eaton, J.W., Wallace, I., Jones, H.D., White, K.N. & Hendry, K. 1993.
Disused docks as a resource in urban conservation and education. In Urban Waterside Regeneration:
Problems and Prospects, K.N. White etal. (eds). London: Ellis Horwood.
Hawkins, S.J. & Cashmore, D. 1993. An ecological view of coastal defences. In Irish Sea Forum Seminar, Rising
Sea Level and Coastal Protection, Belfast, H. Davies (ed.). Liverpool, UK: University of Liverpool Press.
Hawkins, S.J., Southward, A.J. & Barrett, R.L. 1983. Population structure of Patella vulgata L. during
succession on rocky shores in Southwest England. Oceanologica Acta Special volume, 103 –107.
Heath, T. 2017. Carss Bush Park environmentally friendly seawall – Stage 1. Fish Friendly Marine
Infrastructure. New South Wales, Australia: NSW Department of Primary Industries.
Heath, T. & Moody, G. 2013. Habitat development along a highly urbanised foreshore. In Proceedings of
New South Wales Coastal Conference 2003. Available at:
papers2013/Tom%20Heath.pdf (Accessed: 01/04/2019).
Heck K., Jr, Hays, G. & Orth, R.J. 2003. Critical evaluation of the nursery role hypothesis for seagrass meadows.
Marine Ecology Progress Series 253, 123–136.
Hedges, L.V., Gurevitch, J. & Curtis, P.S. 1999. The meta-analysis of response ratios in experimental ecology.
Ecology 80, 1150 1156.
Heery, E.C., Bishop, M.J., Critchley, L.P., Bugnot, A.B., Airoldi, L., Mayer-Pinto, M., Sheehan, E.V., Coleman,
R.A., Loke, L.H. & Johnston, E.L. 2017. Identifying the consequences of ocean sprawl for sedimentary
habitats. Journal of Experimental Marine Biology and Ecology 492, 31– 48.
Heinz Centre. 2000. The Hidden Costs of Coastal Hazards: implications for Risk Assessment and Mitigation.
Washington, DC: Island Press.
Henriques, B., Rocha, L.S., Lopes, C.B., Figueira, P., Monteiro, R.J., Duarte, A., Pardal, M. & Pereira, E. 2015.
Study on bioaccumulation and biosorption of mercury by living marine macroalgae: prospecting for a
new remediation biotechnology applied to saline waters. Chemical Engineering Journal 281, 759–770.
Hernandez, A.B., Brumbaugh, R.D., Frederick, P., Grizzle, R., Luckenbach, M.W., Peterson, C.H. & Angelini,
C. 2018. Restoring the eastern oyster: how much progress has been made in 53 years? Frontiers in
Ecology and the Environment 16, 1–9.
Hinkel, J., Lincke, D., Vafeidis, A.T., Perrette, M., Nicholls, R.J., Tol, R.S., Marzeion, B., Fettweis, X., Ionescu,
C. & Levermann, A. 2014. Coastal ood damage and adaptation costs under 21st century sea-level rise.
Proceedings of the National Academy of Sciences 111, 3292–3297.
Hobbs, J.-P.A., van Herwerden, L., Jerry, D.R., Jones, G.P. & Munday, P.L. 2013. High genetic diversity in
geographically remote populations of endemic and widespread coral reef angelshes (genus: Centropyge).
Diversity 5, 39–50.
Hobbs, R.J., Arico, S., Aronson, J., Baron, J.S., Bridgewater, P., Cramer, V.A., Epstein, P.R., Ewel, J.J., Klink,
C.A., Lugo, A.E., Norton, D., Ojima, D., Richardson, D.M., Sanderson, E.W., Valladares, F., Vilà, M.,
Zamora, R. & Zobel, M. 2006. Novel ecosystems: theoretical and management aspects of the new
ecological world order. Global Ecology and Biogeography 15, 1–7.
Hobbs, R.J. & Norton, D.A. 1996. Towards a conceptual framework for restoration ecology. Restoration
Ecology 4, 93–110.
Hoggart, S.P.G., Hanley, M.E., Parker, D.J., Simmonds, D.J., Bilton, D.T., Filipova-Marinova, M., Franklin,
E.L., Kotsev, I., Penning-Rowsell, E.C., Rundle, S.D., Trifonova, E., Vergiev, S., White, A.C. &
Thompson, R.C. 2014. The consequences of doing nothing: the effects of seawater ooding on coastal
zones. Coastal Engineering 87, 169–182.
Hortig, J.H., Kerr, J.K. & Breederland, M. 2001. Promoting soft engineering along Detroit River shorelines.
Land and Water-Fort Dudge Iowa 45, 24 27.
Hoyle, B.S. 1989. The port—City interface: trends, problems and examples. Geoforum 20, 429– 435.
Huang, X.-W., Wang, W. & Dong, Y.-W. 2015. Complex ecology of China’s seawall. Science 347, 1079.
Huijbers, C.M., Schlacher, T.A., Schoeman, D.S., Weston, M.A. & Connolly, R.M. 2013. Urbanisation alters
processing of marine carrion on sandy beaches. Landscape and Urban Planning 119, 1–8.
Humbert, S., Abeck, H., Bali, N. & Horvath, A. 2006. Leadership in energy and environmental design (LEED),
a critical evaluation by LCA and recommendations for improvement. International Journal of Life Cycle
Assessment 12, 46 57.
Hychka, K. & Druschke, C.G. 2017. Adaptive management of urban ecosystem restoration: learning from
restoration managers in Rhode Island, USA. Society & Natural Resources 30, 1358–1373.
IFRC. 2011. Breaking the Waves. Impact Analysis of Coastal Afforestation for Disaster Risk Reduction in Viet
Nam. Geneva, Switzerland: International Federation of Red Cross and Red Crescent Societies.
IPCC. 2014. Climate change 2014: Synthesis report. Contribution of Working Groups I, II and III to the Fifth
Assessment Report of the Intergovernmental Panel on Climate Change. Geneva, Switzerland: IPCC.
Isbell, F., Tilman, D., Polasky, S. & Loreau, M. 2014. The biodiversity-dependent ecosystem service debt.
Ecology Letters 18, 119–13 4.
Jackson, A. & McIlvenny, J. 2011. Coastal squeeze on rocky shores in northern Scotland and some possible
ecological impacts. Journal of Experimental Marine Biology and Ecology 400, 314 321.
Jackson, M.D., Mulcahy, S.R., Chen, H., Li, Y., Li, Q., Cappelletti, P. & Wenk, H.-R. 2017. Phillipsite and
Al-tobermorite mineral cements produced through low-temperature water-rock reactions in Roman
marine concrete. American Mineralogist: Journal of Earth and Planetary Materials 102, 1435–1450.
James, R.J. & Underwood, A.J. 1994. Inuence of colour of substratum on recr uitment of spirorbid tubeworms to
different types of intertidal boulders. Journal of Experimental Marine Biology and Ecology 181, 105–115.
Janis, S.P., Birney, L.B. & Newton, R. 2016. Billion Oyster Project: linking public school teaching and learning
to the ecological restoration of New York Harbor using innovative applications of environmental and
digital technologies. International Journal of Digital Content Technology and Its Applications 10, 1–14.
Jin, D., Hoagland, P., Au, D.K. & Qiu, J. 2015. Shoreline change, seawalls, and coastal property values. Ocean
& Coastal Management 114 , 185 193.
Johannesson, K. & Warmoes, T. 1990. Rapid colonization of Belgian breakwaters by the direct developer,
Littorina saxatilis (Olivi) (Prosobranchia, Mollusca). Hydrobiologia 193, 99–108.
Johnson, M.P., Frost, N.J., Mosley, M.W.J., Roberts, M.F. & Hawkins, S.J. 2003. The area-independent effects
of habitat complexity on biodiversity vary between regions. Ecology Letters 6, 126–132.
Johnson, M.P., Hughes, R.N., Burrows, M.T. & Hawkins, S.J. 1998. Beyond the predation halo: small
scale gradients in barnacle populations affected by the relative refuge value of crevices. Journal of
Experimental Marine Biology and Ecology 231, 163 170.
Jones, G. 1994. Global warming, sea level change and the impact on estuaries. Marine Pollution Bulletin 28,
Jones, H. 2011. A Guide to Monitoring and Evaluating Policy Inuence. London: Overseas Development Institute.
Jonsson, P.R., Granhag, L., Moschella, P.S., Åberg, P., Hawkins, S.J. & Thompson, R.C. 2006. Interactions
between wave action and grazing control the distribution of intertial macroalgae. Ecology 87, 1169–1178.
Kain, J.M. & Jones, N.S. 1975. Algal recolonization of some cleared subtidal areas. Journal of Ecology 63,
Keck, M.E. & Sikkink, K. 1998. Transnational advocacy networks in the movement society. In: The Social
Movement Society: Contentious Politics for a New Century. Meyer, D.S. & Tarrow, S. (eds), Lanham:
Rowman & Littleeld Publishers, Inc. 217–238.
Keith, S.A., Herbert, R.J.H., Norton, P.A., Hawkins, S.J. & Newton, A.C. 2010. Individualistic species
limitations of climate-induced range expansions generated by meso-scale dispersal barriers. Diversit y
and Distributions 17, 275–286.
Kellogg, M.L., Cornwell, J.C., Owens, M.S. & Paynter, K.T. 2013. Denitrication and nutrient assimilation on
a restored oyster reef. Marine Ecology Progress Series 480, 1–19.
Kelly, R.P., O’Donnell, J.L., Lowell, N.C., Shelton, A.O., Samhouri, J.F., Hennessey, S.M., Feist, B.E. &
Williams, G.D. 2016. Genetic signatures of ecological diversity along an urbanization gradient. PeerJ
4, e2444.
Kennedy, V.S., Breitburg, D.L., Christman, M.C., Luckenbach, M.W., Paynter, K., Kramer, J., Sellner, K.G.,
Dew-Baxter, J., Keller, C. & Mann, R. 2011. Lessons learned from efforts to restore oyster populations
in Maryland and Virginia, 1990 to 2007. Journal of Shellsh Research 30, 719–731.
Kenworthy, J.M., Rolland, G., Samadi, S. & Lejeusne, C. 2018. Local variation within marinas: effects of
pollutants and implications for invasive species. Marine Pollution Bulletin 133, 96–106.
Kienker, S., Coleman, R., Morris, R., Steinberg, P., Bollard, B., Jarvis, R., Alexander, K. & Strain, E. 2018.
Bringing harbours alive: assessing the importance of eco-engineered coastal infrastructure for different
stakeholders and cities. Marine Policy 94, 238–246.
Kittinger, J. & Ayers, A. 2010. Shoreline armoring, risk management, and coastal resilience under rising seas.
Coastal Management 38, 634 653.
Knights, A.M., Firth, L.B. & Walters, K. 2012. Interactions between multiple recruitment drivers: post-
settlement predation mortality and ow-mediated recruitment. PLOS ONE 7, e35096.
Kochnower, D., Reddy, S.M.W. & Flick, R.E. 2015. Factors inuencing local decisions to use habitats to protect
coastal communities from hazards. Ocean & Coastal Management 116 , 277–290.
Köhler, J., Hansen, P.D. & Wahl, M. 1999. Colonization patterns at the substratum-water interface: how does
surface microtopography inuence recruitment patterns of sessile organisms? Biofouling 14, 237–248.
Kordas, R.L., Dudgeon, S., Storey, S. & Harley, C.D. 2015. Intertidal community responses to eld-based
experimental warming. Oikos 124, 888–898.
Kowarik, I. 2011. Novel urban ecosystems, biodiversity, and conservation. Environmental Pollution 159, 19 74–1983.
Krause-Jensen, D. & Duarte, C.M. 2016. Substantial role of macroalgae in marine carbon sequestration. Nature
Geoscience 9, 737.
Krauss, K.W., Lovelock, C.E., McKee, K.L., López-Hoffman, L., Ewe, S.M. & Sousa, W.P. 2008. Environmental
drivers in mangrove establishment and early development: a review. Aquatic Botany 89, 105–127.
Kreeger, D., Moody, J., Katkowski, M., Boatright, M. & Rosencrance, D. 2015. Marsh Futures: Use of Scientic
Survey Tools to Assess L ocal Salt Marsh Vulnerability and Chart Best Management Practices and
Interventions. Wilmington, Delaware: Partnership for the Delaware Estuary.
Kru mholz, J. & Jadot, C. 2009. Demonstration of a new technology for restoration of Red Mangrove (Rhizophora
mangle) in high-energy environments. Marine Technology Society Journal 43, 64–72.
Krumme, U. 2009. Diel and tidal movements by sh and decapods linking tropical coastal ecosystems. In
Ecological Connectivity Among Tropical Coastal Ecosystems, I. Nagelkerken (ed.). Dordrecht, the
Netherlands: Springer, 271–324.
Kumara, M.P., Jayatissa, L.P., Krauss, K.W., Phillips, D.H. & Huxham, M. 2010. High mangrove density
enhances surface accretion, surface elevation change, and tree survival in coastal areas susceptible to
sea-level rise. Oecologia 164, 545–553.
Lai, S., Loke, L.H.L., Hilton, M.J., Bouma, T.J. & Todd, P.A. 2015. The effects of urbanisation on coastal
habitats and the potential for ecological engineering: a Singapore case study. Ocean & Coastal
Management 103, 78 85.
Landry, C. & Hindsley, P. 2006. Willingness to pay for risk reduction and amenities: applications of the hedonic
price method in the coastal zone. In: Management, Policy, Science, and Engineering of Nonstructural
Erosion Control in the Chesapeake Bay. Proceedings of the 200 6 Living Shoreline Summit. Erdle, S.Y.,
Davis, J.L.D. & Sellner, K.G. (eds). Williamsburg. 105–110.
Lapann-Johannessen, C., Miller, J., Rella, A. & Rodriquez, E. (eds). 2015. Hudson River Wake Study.
Staatsburg, NY: Stevens Institute of Technology, TR- 2947; in association with and published by the
Hudson River Sustainable Shorelines Project. Available at :
sites/9/2012/07/Hudson-River-Wake-Report.pdf (Accessed: 01/04/2019).
Lawton, J.H. 1999. Are there general laws in ecology? Oikos 84, 177–192.
Lee, W.L., Yik, F.W.H. & Burnett, J. 2007. Assessing energy performance in the latest versions of Hong Kong
Building Environmental Assessment Method (HK-BEAM). Energy and Buildings 39, 343–354.
Leonardi, N., Carnacina, I., Donatelli, C., Ganju, N.K., Plater, A.J., Schuerch, M. & Temmerman, S. 2018.
Dynamic interactions between coastal storms and salt marshes: a review. Geomorphology 301, 92–107.
Leonardi, N., Ganju, N.K. & Fagherazzi, S. 2016. A linear relationship between wave power and erosion
determines salt-marsh resilience to violent storms and hurricanes. Proceedings of the National Academy
of Sciences 113, 6468.
Levin, S.A. 1992. The problem of pattern and scale in ecology: the Robert H. MacArthur Award lecture.
Ecology 73, 1943 –1967.
Liversage, K. & Chapman, M.G. 2018. Coastal ecological engineering and habitat restoration: incorporating
biologically diverse boulder habitat. Marine Ecology Progress Series 593, 173–185.
Lockwood, J.L., Cassey, P. & Blackburn, T. 2005. The role of propagule pressure in explaining species
invasions. Trends in Ecology & Evolution 20, 223–228.
Loke, L.H.L., Bouma, T.J. & Todd, P.A. 2017. The effects of manipulating microhabitat size and variability
on tropical seawall biodiversity: eld and ume experiments. Journal of Experimental Marine Biology
and Ecology 492, 113–120.
Loke, L.H.L., Heery, E.C. & Todd, P.A. 2019. Shoreline defenses. In World Seas: An Environmental Evaluation,
C. Sheppard (ed.). London: Academic Press, 2nd edition, 491–504.
Loke, L.H.L., Jachowski, N.R., Bouma, T.J., Ladle, R.J. & Todd, P.A. 2014. Complexity for articial substrates
(CASU): Software for creating and visualising habitat complexity. PLOS ONE 9, e87990.
Loke, L.H.L., Ladle, R.J., Bouma, T.J. & Todd, P.A. 2015. Creating complex habitats for restoration and
reconciliation. Ecological Engineering 77, 307–313.
Loke, L.H.L., Liao, L.M., Bouma, T.J. & Todd, P.A. 2016. Succession of seawall algal communities on articial
substrates. Rafes Bulletin of Zoology 32, 1–10.
Loke, L.H.L. & Todd, P.A. 2016. Structural complexity and component type increase intertidal biodiversity
independently of area. Ecology 97, 383–393.
Loreau, M., Naeem, S. & Inchausti, P. (eds). 2002. Biodiversity and Ecosystem Functioning: Synthesis and
Perspectives, New York: Oxford University Press.
Lubchenco, J. 1978. Plant species diversity in a marine intertidal community: importance of herbivore food
preference and algal competitive abilities. The American Naturalist 112, 23–39.
Luo, S.X., Shao, D.D., Long, W., Liu, Y.J., Sun, T. & Cui, B.S. 2018. Assessing ‘coastal squeeze’ of wetlands
at the Yellow River Delta in China: a case study. Ocean & Coastal Management 153, 193–202.
Ma, Z., Melville, D.S., Liu, J., Chen, Y., Yang, H., Ren, W., Zhang, Z., Piersma, T. & Li, B. 2014. Rethinking
China’s new great wall. Science 346, 912.
Maas, K. & Liket, K. 2011. Talk the walk: measuring the impact of strategic philanthropy. Journal of Business
Ethics 100, 445464.
MacArthur, R.H. & Wilson, E.O. 1967. The Theory of Island Biogeography. Princeton, New Jersey: Princeton
Maggi, E., Bertocci, I., Vaselli, S. & Benedetti-Cecchi, L. 2011. Connell and Slatyer’s models of succession in
the biodiversity era. Ecology 92, 1399–1406.
Mahan, B.L., Polasky, S. & Adams, R.M. 2000. Valuing urban wetlands: a property price approach.
LandEconomics 76, 100–113.
Mahmoodian, M., Li, C.Q. & Setunge, S. 2015. Service life prediction of coastal infrastructures in Australia.
In E-Proceedings of the 36th IAHR World Congress, Held During, 28 June–3 July, 2015, in the Hague,
the Netherlands.
Malik, A., Fensholt, R. & Mertz, O. 2015. Mangrove exploitation effects on biodiversity and ecosystem services.
Biodiversity and Conservation 24, 3543–3557.
Manca, E., Cáceres, I., Alsina, J.M., Stratigaki, V., Townend, I. & Amos, C.L. 2012. Wave energy and wave-
induced ow reduction by full-scale model Posidonia oceanica seagrass. Continental Shelf Research
50–51, 100–116.
Manis, J.E., Garvis, S.K., Jachec, S.M. & Walters, L.J. 2015. Wave attenuation experiments over living
shorelines over time: a wave tank study to assess recreational boating pressures. Journal of Coastal
Conservation 19, 1–11.
Manning, A.D., Lindenmayer, D.B. & Fischer, J. 2006. Stretch goals and backcasting: approaches for
overcoming barriers to large-scale ecological restoration. Restoration Ecology 14, 487–492.
Markert, A., Wehrmann, A. & Kröncke, I. 2010. Recently established Crassostrea-reefs versus native Mytilus-
beds: differences in ecosystem engineering affects the macrofaunal communities (Wadden Sea of Lower
Saxony, southern German Bight). Biological Invasions 12, 15.
Martin, D., Bertasi, F., Colangelo, M.A., de Vries, M., Frost, M., Hawkins, S.J., MacPherson, E., Moschella, P.S.,
Satta, M.P., Thompson, R.C. & Ceccherelli, V.U. 2005. Ecological impact of coastal defence structures
on sediment and mobile fauna: evaluating and forecasting consequences of unavoidable modications
of native habitats. Coastal Engineering 52, 1027–1051.
Martins, G.M., Thompson, R.C., Neto, A.I., Hawkins, S.J. & Jenkins, S.R. 2010. Enhancing stocks of the
exploited limpet Patella candei d’Orbigny via modications in coastal engineering. Biological
Conservation 143, 203 211.
Marzinelli, E., Campbell, A., Vergés, A., Coleman, M., Kelaher, B.P. & Steinberg, P. 2014. Restoring seaweeds:
does the declining fucoid Phyllospora comosa support different biodiversity than other habitats? Journal
of Applied Phycology 26, 1089–1096.
Mayer-Pinto, M., Cole, V., Johnston, E., Bugnot, A., Hurst, H., Airoldi, L., Glasby, T. & Dafforn, K. 2018.
Functional and structural responses to marine urbanisation. Environmental Research Letters 13, 014009.
Mayer-Pinto, M., Johnston, E., Bugnot, A., Glasby, T., Airoldi, L., Mitchell, A. & Dafforn, K. 2017.
Building ‘blue’: an eco-engineering framework for foreshore developments. Journal of Environmental
Management 189, 109–114.
McAfee, D., O’Connor, W.A. & Bishop, M.J. 2017. Fast-growing oysters show reduced capacity to provide a
thermal refuge to intertidal biodiversity at high temperatures. Journal of Animal Ecology 86, 1352–1362.
McDonald, T., Jonson, J. & Dixon, K.W. 2016. National standards for the practice of ecological restoration in
Australia. Restoration Ecology 24, S4S32.
McManus, R.S., Archibald, N., Comber, S., Knights, A.M., Thompson, R.C. & Firth, L.B. 2018. Partial
replacement of cement for waste aggregates in concrete coastal and marine infrastructure: a foundation
for ecological enhancement? Ecological Engineering 120, 665–667.
Menz, M.H.M., Dixon, K.W. & Hobbs, R.J. 2013. Hurdles and opportunities for landscape-scale restoration.
Science 339, 526 –527.
Meurant, G. 2012. The Ecology of Natural Disturbance and Patch Dynamics. London: Academic Press.
Michener, W.K. 1997. Quantitatively evaluating restoration experiments: research design, statistical analysis,
and data management considerations. Restoration Ecology 5, 324–337.
Mieszkowska, N., Leaper, R., Moore, P., Kendall, M., Burrows, M., Lear, D., Poloczanska, E., Hiscock, K.,
Moschella, P., Thompson, R., Herbert, R., Laffoley, D., Baxter, J., Southward, A. & Hawkins, S. 2006.
Marine Biodiversity and Climate Change: Assessing and Predicting the Inuence of Climatic Change
Using Intertidal Rocky Shore Biota. Plymouth, UK: Marine Biological Association of the United Kingdom.
Miller, D., Ode Sang, Å., Brown, I., Munoz-Rojas, J., Wang, C. & Donaldson-Selby, G. 2016. Landscape
modelling and stakeholder engagement: participatory approaches and landscape visualisation for conict
resolution. In A Review on the State of the Art in Scenario Modelling for Environmental Management,
N. Sang & A. Ode Sang (eds). Alnarp, Sweden: Swedish Environmental Protection Agency.
Milton, S.J. 2003. ‘Emerging ecosystems’ - a washing-stone for ecologists, economists and sociologists?
SouthAfrican Journal of Science 99, 404–406.
Mineur, F., Cook, E.J., Minchin, D., Bohn, K., MacLeod, A. & Maggs, C.A. 2012. Changing coasts: marine
aliens and articial structures. Oceanography and Marine Biology: An Annual Review 50, 189 23 4.
Mitsch, W.J. 2012. What is ecological engineering? Ecological Engineering 45, 5–12.
Mitsch, W.J. & Jørgensen, S.E. 2004. Ecological Engineering and Ecosystem Restoration. Hoboken,
NewJersey: John Wiley & Sons.
Mitsch, W.J., Tejada, J., Nahlik, A., Kohlmann, B., Bernal, B. & Hernández, C.E. 2008. Tropical wetlands for
climate change research, water quality management and conservation education on a university campus
in Costa Rica. Ecological Engineering 34, 276–288.
Mitsch, W.J. & Wilson, R.F. 1996. Improving the success of wetland creation and restoration with know-how,
time, and self-design. Ecological Applications 6, 77–83.
Moody, J., Kreeger, D., Bouboulis, S., Roberts, S. & Padeletti, A. 2016. Design, Implementation, and Evaluation
of Three Living Shoreline Treatments at the DuPont Nature Center, Mispillion River, Milford, DE.
Wilmington, DE: Partnership for the Delaware Estuary.
Moody, J.A. 2012. The relationship between the ribbed mussel (Geukensia demissa) and salt marsh shoreline
erosion. MSc thesis, Rutgers University, New Brunswick, New Jersey.
Moosavi, S. 2017. Ecological coastal protection: pathways to living shorelines. Procedia Engineering 196,
Morris, R.L., Chapman, M.G., Firth, L.B. & Coleman, R.A. 2017. Increasing habitat complexity on seawalls:
investigating large- and small-scale effects on sh assemblages. Ecology and Evolution 7, 9567–9579.
Morris, R.L., Golding, S., Dafforn, K.A. & Coleman, R.A. 2018a. Can coir increase native biodiversity and reduce
colonisation of non-indigenous species in eco-engineered rock pools? Ecological Engineering 120, 622– 630.
Morris, R.L., Konlechner, T.M., Ghisalberti, M. & Swearer, S.E. 2018b. From grey to green: efcacy of eco-
engineering solutions for nature-based coastal defence. Global Change Biology 24, 1827–1842.
Morris, R.L., Porter, A.G., Figueira, W.F., Coleman, R.A., Fobert, E.K. & Ferrari, R. 2018c. Fish-smart
seawalls: a decision tool for adaptive management of marine infrastructure. Frontiers in Ecology and
the Environment 16, 278 –28 7.
Moschella, P.S., Abbiati, M., Åberg, P., Airoldi, L., Anderson, J.M., Bacchiocchi, F., Bulleri, F., Dinesen, G.E.,
Frost, M., Gacia, E., Granhag, L., Jonsson, P.R., Satta, M.P., Sundelöf, A., Thompson, R.C. & Hawkins,
S.J. 2005. Low-crested coastal defence structures as articial habitats for marine life: using ecological
criteria in design. Coastal Engineering 52, 1053–1071.
Mulder, J.P. & Tonnon, P.K. 2011. ‘Sand engine’: background and design of a mega-nourishment pilot in the
Netherlands. Coastal Engineering Proceedings 32, 35.
Munkittrick, K.R., Arens, C.J., Lowell, R.B. & Kaminski, G.P. 2009. A review of potential methods of
determining critical effect size for designing environmental monitoring programs. Environmental
Toxicology and Chemistry 28, 1361–1371.
Munsch, S.H., Cordell, J.R. & Toft, J.D. 2015. Effects of seawall armouring on juvenile Pacic salmon diets
in an urban estuarine embayment. Marine Ecology Progress Series 535, 213–229.
Munsch, S.H., Cordell, J.R. & Toft, J.D. 2017a. Effects of shoreline a rmouring and overwater str uctures on coastal
and estuarine sh: opportunities for habitat improvement. Journal of Applied Ecology 54, 1373–1384.
Munsch, S.H., Cordell, J.R. & Toft, J.D. 2017b. A quantitative chronology of diurnal feeding in juvenile Pacic
salmon. Transactions of the American Fisheries Society 14 6, 222–229.
Munsch, S.H., Cordell, J.R., Toft, J.D. & Morgan, E.E. 2014. Effects of seawalls and piers on sh assemblages
and juvenile salmon feeding behaviour. North American Journal of Fisheries Management 34, 814 –827.
Murcia, C., Aronson, J., Kattan, G.H., Moreno-Mateos, D., Dixon, K. & Simberloff, D. 2014. A critique of the
‘novel ecosystem’ concept. Trends in Ecology & Evolution 29, 548–553.
Nakagawa, S. & Cuthill Innes, C. 2007. Effect size, condence interval and statistical signicance: a practical
guide for biologists. Biological Reviews 82, 591–605.
Nandakumar, K., Matsunaga, H. & Takagi, M. 2003. Microfouling studies on experimental test blocks of steel-
making slag and concrete exposed to seawater off Chiba, Japan. Biofouling 19, 257–267.
Narayan, S., Beck, M.W., Reguero, B.G., Losada, I.J., Van Wesenbeeck, B., Pontee, N., Sanchirico, J.N.,
Ingram, J.C., Lange, G.M. & Burks-Copes, K.A. 2016. The effectiveness, costs and coastal protection
benets of natural and nature-based defences. PLOS ONE 11, e0154735.
Naveh, Z. 1994. From biodiversity to ecodiversity: a landscape-ecology approach to conservation and
restoration. Restoration Ecology 2, 180–189.
Naylor, L.A., Coombes, M.A., Venn, O., Roast, S.D. & Thompson, R.C. 2012. Facilitating ecological enhancement
of coastal infrastructure: the role of policy, people and planning. Environmental Science & Policy 22, 36– 46.
Neo, M.L., Todd, P.A., Teo, S.L.-M. & Chou, L.M. 2009. Can articial substrates enriched with crustose
coralline algae enhance larval settlement and recruitment in the uted giant clam (Tridacna squamosa)?
Hydrobiologia 625, 83–90.
Neumann, B., Vafeidis, A.T., Zimmermann, J. & Nicholls, R.J. 2015. Future coastal population growth and
exposure to sea-level rise and coastal ooding - a global assessment. PLOS ONE 10, e0118571.
NYCEDC. 2016. Waterfront Facilities Maintenance Management System. Inspection Guidelines Manual.
NewYork: New York City Economic Development Corporation.
Ng, C.S.L., Lim, S.C., Ong, J.Y., Teo, L.M.S., Chou, L.M., Chua, K.E. & Tan, K.S. 2015. Enhancing the
biodiversity of coastal defence structures: transplantation of nursery-reared reef biota onto intertidal
seawalls. Ecological Engineering 82, 480–486.
Nicholls, R.J. & Hanson, S. 2013. Planning for changes in extreme events in port cities throughout the 21st
century. In Climate Adaptation and Flood Risk in Coastal Cities, J. Aerts etal. (eds). Routledge, 9–26.
Nicholls, R.J., Hanson, S., Herweijer, C., Patmore, N., Hallegatte, S., Corfee-Morlot, J., Chateau, J. & Muir-
Wood, R. 2007. Ranking of the world’s cities most exposed to coastal ooding today and in the future.
OECD Environment Working Paper. Organisation for Economic Co-operation and Development
Nicholls, R.J., Hoozemans, F.M. & Marchand, M. 1999. Increasing ood risk and wetland losses due to global
sea-level rise: regional and global analyses. Global Environmental Change 9, S69–S87.
Nicholls, R.J. & Tol, R.S.J. 2006. Impacts and responses to sea-level rise: a global analysis of the SRES scenarios
over the twenty-rst century. Philosophical Transactions of the Royal Societ y A: Mathematical, Physical
and Engineering Sciences 364, 1073.
O’Connor, N.E. & Crowe, T.P. 2005. Biodiversity loss and ecosystem functioning: distinguishing between
number and identity of species. Ecology 86, 1783–1796.
O’Connor, N.E. & Crowe, T.P. 2007. Biodiversity among mussels: separating the inuence of sizes of mussels
from the ages of patches. Journal of the Marine Biological Association of the United Kingdom 87,
O’Donnell, J.E. 2016. Living shorelines: a review of literature relevant to New England coasts. Journal of
Coastal Research 33, 435 –451.
O’Dwyer, J.P. & Green, J.L. 2010. Field theory for biogeography: a spatially explicit model for predicting
patterns of biodiversity. Ecology Letters 13, 87–95.
Odum, E.P. 1985. Trends expected in stressed ecosystems. Bioscience 35, 419422.
Odum, H.T. 1971. Environment, Power, and Society. New York: John Wiley & Sons, Inc.
Odum, H.T. 1975. Combining energy laws and corollaries of the maximum power principle with visual systems
mathematics. In Ecosystem Analysis and Prediction, S.A. Levin (ed.). Philadelphia: Society for Industrial
and Applied Mathematics, pp. 239–263.
Oliver, T.H., Heard, M.S., Isaac, N.J.B., Roy, D.B., Procter, D., Eigenbrod, F., Freckleton, R., Hector, A., Orme,
D.L., Petchey, O.L., Proenca, V., Raffaelli, D., Suttle, K.B., Mace, G.M., Martin-Lopez, B., Woodcock,
B.A. & Bullock, J.M. 2015. Biodiversity and resilience of ecosystem functions. Trends in Ecology &
Evolution 30, 673–684.
Ondiviela, B., Losada, I.J., Lara, J.L., Maza, M., Galvan, C., Bouma, T.J. & van Belzen, J. 2014. The role of
seagrasses in coastal protection in a changing climate. Coastal Engineering 87, 158–168.
Onorevole, K.M., Thompson, S.P. & Piehler, M.F. 2018. Living shorelines enhance nitrogen removal capacity
over time. Ecological Engineering 120, 238–248.
Osenberg, C.W., Sarnelle, O., Cooper, S.D. & Holt, R.D. 1999. Resolving ecological questions through meta-
analysis: goals, metrics, and models. Ecology 80, 1105–1117.
Osorio-Cano, J.D., Osorio, A.F. & Peláez-Zapata, D.S. 2017. Ecosystem management tools to study natural
habitats as wave damping structures and coastal protection mechanisms. Ecological Engineering, http://
Palmer, M.A., Zedler, J.B. & Falk, D.A. (eds). 2016. Foundations of Restoration Ecology. Washington, DC:
Island Press.
Perkins, M.J., Ng, T.P.T., Dudgeon, D., Bonebrake, T.C. & Leung, K.M.Y. 2015. Conserving intertidal habitats:
what is the potential of ecological engineering to mitigate impacts of coastal structures? Estuarine,
Coastal and Shelf Science 167, 504–515.
Perkol-Finkel, S. & Sella, I. 2014. Ecologically active concrete for coastal and marine infrastructure: innovative
matrices and designs. From Sea to Shore–Meeting the Challenges of the Sea. London: Institute of Civil
Enginee rs Publish ing, 1139 –1149.
Perkol-Finkel, S. & Sella, I. 2015. Harnessing urban coastal infrastructure for ecological enhancement.
Proceedings of the Institution of Civil Engineers-Maritime Engineering 168. 102–110. https://doi.
Perkol-Finkel, S., Ferrario, F., Nicotera, V. & Airoldi, L. 2012. Conservation challenges in urban seascapes:
promoting the growth of threatened species on coastal infrastructures. Journal of Applied Ecology 49,
Perring, M.P., Manning, P., Hobbs, R.J., Lugo, A.E., Ramalho, C.E. & Standish, R.J. 2013. Novel urban
ecosystems and ecosystem services. In Novel Ecosystems: Intervening in the New Ecological World
Order, R.J. Hobbs etal. (eds). Oxford: Wiley-Blackwell, 310 325.
Peterson, C.H., Luettich, R.A., Jr, Micheli, F. & Skilleter, G.A. 2004. Attenuation of water ow inside seagrass
canopies of differing structure. Marine Ecology Progress Series 268, 81–92.
Piehler, M.F. & Smyth, A.R. 2011. Habitat-specic distinctions in estuarine denitrication affect both
ecosystem function and services. Ecosphere 2, 117.
Pilkey, O.H., Young, R., Longo, N. & Coburn, A. 2012. Living Shorelines: A Good Idea Gone Bad.
DukeUniversity, Durham, North Carolina.
Pioch, S., Relini, G., Souche, J.C., Stive, M.J.F., De Monbrison, D., Nassif, S., Simard, F., Allemand, D.,
Saussol, P., Spieler, R. & Kilfoyle, K. 2018. Enhancing eco-engineering of coastal infrastructure with
eco-design: moving from mitigation to integration. Ecological Engineering 120, 574–584.
Piola, R.F. & Johnston, E.L. 2008. Pollution reduces native diversity and increases invader dominance in
marine hard-substrate communities. Diversity and Distributions 14, 329–342.
Polanyi, K. 1963. Ports of trade in early societies. Journal of Economic History 23, 30 45.
Pomerat, C.M. & Weiss, C.M. 1947. The inuence of texture and composition of surface on the attachment of
sedentary marine organisms. The Biological Bulletin 91, 57–65.
Pompe, J.J. & Rinehart, J.R. 1994. Estimating the effect of wider beaches on coastal housing prices. Ocean &
Coastal Management 22, 141–152 .
Powers, S.P.H. Peterson, C., Grabowski, J. & Lenihan, H.S. 2009. Success of constructed oyster reefs in
no-harvest sanctuaries: implications for restoration. Marine Ecology Progress Series 389, 159–170.
Prati, G., Albanesi, C., Pietrantoni, L. & Airoldi, L. 2016. Public perceptions of beach nourishment and conict
management strategies: a case study of Portonovo Bay in the Adriatic Italian Coast. Land Use Policy
50, 422– 428.
Pulsford, S.A., Lindenmayer, D.B. & Driscoll, D.A. 2014. A succession of theories: purging redundancy from
disturbance theory. Biological Reviews 91, 148167.
Raffaelli, D., van der Putten, W., Persson, L., Wardle, D., Petchey, O., Koricheva, J., van der Heijden, M.,
Mikola, J. & Kennedy, T. 2002. Multi-trophic dynamics and ecosystem processes. In Biodiversity and
Ecosystem Functioning: Synthesis and Perspectives, M. Loreau etal. (eds). New York: Oxford University
Press, 147–154.
Raffaelli, D.G. & Hawkins, S.J. 1996. Intertidal Ecology. London: Chapman and Hall.
Raimondi, P.T. & Keough, M.J. 1990. Behavioural variability in marine larvae. Australian Journal of Ecology
15, 427–437.
Rao, N.S., Ghermandi, A., Portela, R. & Wang, X. 2015. Global values of coastal ecosystem services: a spatial
economic analysis of shoreline protection values. Ecosystem Services 11, 95–105.
Reguero, B.G., Beck, M.W., Bresch, D.N., Calil, J. & Meliane, I. 2018. Comparing the cost effectiveness of
nature-based and coastal adaptation: a case study from the Gulf Coast of the United States. PLOS ONE
13, e0192132.
Reimann, L., Vafeidis, A.T., Brown, S., Hinkel, J. & Tol, R.S.J. 2018. Mediterranean UNESCO World Heritage
at risk from coastal ooding and erosion due to sea-level rise. Nature Communications 9, 4161.
Reise, K. 2003. More sand to the shorelines of the Wadden Sea harmonizing coastal defense with habitat
dynamics. In Marine Science Frontiers for Europe, G. Wefer etal. (eds). Berlin and Heidelberg: Springer
Berlin Heidelberg, 203–216.
Reisman, J., Gienapp, A. & Stachowiak, S. 2007. A Guide to Measuring Advocacy and Policy. Baltimore:
Organisational Research Services.
Rigolet, C., Dubois, S.F. & Thiébaut, E. 2014. Benthic control freaks: effects of the tubiculous amphipod
Haploops nirae on the specic diversity and functional structure of benthic communities. Journal of
Sea Research 85, 413 –427.
Rogers, T.L., Byrnes, J.E. & Stachowicz, J.J. 2016. Native predators limit invasion of benthic invertebrate
communities in Bodega Harbour, California, USA. Marine Ecology Progress Series 545, 161–173.
Rosenzweig, M.L. 1995. Species Diversity in Space and Time. London: Cambridge University Press.
Rosenzweig, M.L. 2003a. Reconciliation ecology and the future of species diversity. Oryx 37, 194–205.
Rosenzweig, M.L. 2003b. Win-Win Ecology: How the Earth’s Species Can Survive in the Midst of Human
Enterprise. New York: Oxford University Press.
Rosindell, J. & Cornell, S.J. 2007. Species–area relationships from a spatially explicit neutral model in an
innite landscape. Ecology Letters 10, 586–595.
Ruiz, G.M., Torchin, M.E. & Grant, K. 2009. Using the Panama Canal to test predictions about tropical marine
invasions. In Proceedings of the Smithsonian Marine Science Symposium, Washington, DC: Smithsonian
Institution Scholarly Press, 291–299.
Russell, G., Hawkins, S.J., Evans, L.C., Jones, H.D. & Holmes, G.D. 1983. Restoration of a disused dock basin
as a habitat for marine benthos and sh. Journal of Applied Ecology 20, 43–58.
Saenger, P. & Siddiqi, N.A. 1993. Land from the sea: the mangrove afforestation program of Bangladesh. Ocean
& Coastal Management 20, 2