Towards an urban marine ecology: characterizing the drivers, patterns and processes of marine ecosystems in coastal cities

Abstract

Human population density within 100 km of the sea is approximately three times higher than the global average. People in this zone are concentrated in coastal cities that are hubs for transport and trade – which transform the marine environment. Here, we review the impacts of three interacting drivers of marine urbanization (resource exploitation, pollution pathways and ocean sprawl) and discuss key characteristics that are symptomatic of urban marine ecosystems. Current evidence suggests these systems comprise spatially heterogeneous mosaics with respect to artificial structures, pollutants and community composition, while also undergoing biotic homogenization over time. Urban marine ecosystem dynamics are often influenced by several commonly observed patterns and processes, including the loss of foundation species, changes in biodiversity and productivity, and the establishment of novel assemblages, ruderal species and synanthropes. Further, we discuss potential urban acclimatization and adaptation among marine taxa, interactive effects of climate change and marine urbanization, and ecological engineering strategies for enhancing urban marine ecosystems. By assimilating research findings across disparate disciplines, we aim to build the groundwork for urban marine ecology – a nascent field; we also discuss research challenges and future directions for this new field as it advances and matures. Ultimately, all sides of coastal city design: architecture, urban planning, and civil and municipal engineering, will need to prioritize the marine environment if negative effects of urbanization are to be minimized. In particular, planning strategies that account for the interactive effects of urban drivers and accommodate complex system dynamics could enhance the ecological and human functions of future urban marine ecosystems.
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Accepted 3 May 2019
128: 1215–1242, 2019
doi: 10.1111/oik.0594 6
doi: 10.1111/oik.05946 128 1215–1242
Human population density within 100 km of the sea is approximately three times higher
than the global average. People in this zone are concentrated in coastal cities that are hubs
for transport and trade – which transform the marine environment. Here, we review the
impacts of three interacting drivers of marine urbanization (resource exploitation, pol-
lution pathways and ocean sprawl) and discuss key characteristics that are symptomatic
of urban marine ecosystems. Current evidence suggests these systems comprise spatially
heterogeneous mosaics with respect to articial structures, pollutants and community
composition, while also undergoing biotic homogenization over time. Urban marine
ecosystem dynamics are often inuenced by several commonly observed patterns and
processes, including the loss of foundation species, changes in biodiversity and produc-
tivity, and the establishment of ruderal species, synanthropes and novel assemblages. We
discuss potential urban acclimatization and adaptation among marine taxa, interactive
eects of climate change and marine urbanization, and ecological engineering strategies
for enhancing urban marine ecosystems. By assimilating research ndings across disparate
disciplines, we aim to build the groundwork for urban marine ecology – a nascent eld;
we also discuss research challenges and future directions for this new eld as it advances
and matures. Ultimately, all sides of coastal city design: architecture, urban planning and
civil and municipal engineering, will need to prioritize the marine environment if negative
eects of urbanization are to be minimized. In particular, planning strategies that account
for the interactive eects of urban drivers and accommodate complex system dynamics
could enhance the ecological and human functions of future urban marine ecosystems.
Keywords: climate change, ecological engineering, ocean sprawl, pollution pathways,
resource exploitation
Towards an urban marine ecology: characterizing the drivers,
patterns and processes of marine ecosystems in coastal cities
Peter A.Todd, Eliza C.Heery, Lynette H. L.Loke, Ruth H.Thurstan, D. JohanKotze and ChristopherSwan
P. A. Todd (https://orcid.org/0000-0001-5150-9323) ✉ (dbspat@nus.edu.sg), E. C. Heery and L. H. L. Loke, Experimental Marine Ecology Laboratory,
Dept of Biological Sciences, National Univ. of Singapore, 16 Science Drive 4, Singapore 117558. – R. H. urstan, Centre for Ecology and Conservation,
College of Life and Environmental Sciences, Univ. of Exeter, Penryn, UK. – D. J. Kotze, Faculty of Biological and Environmental Sciences, Ecosystems and
Environment Research Programme, Univ. of Helsinki, Lahti, Finland. – C. Swan, Dept of Geography & Environmental Systems, Univ. of Maryland
Baltimore County, Baltimore, MD, USA.
Forum
Urban ecology has advanced rapidly in recent decades, yet has focused primarily on terrestrial
and freshwater systems. By comparison, urban marine ecology is a eld in its infancy and
lacks the theoretical and empirical foundations underpinning urban ecosystem science on
land. is Forum-article aims to help build such a foundation, by presenting a conceptual
framework of the interacting drivers of marine urbanization, identifying key characteristics
of urban marine ecosystems based on research from disparate disciplines, and highlighting
research priorities that can advance urban marine ecology as a discipline.
Synthesis

1216
Introduction
e world’s population is urbanizing rapidly (Bloom 2011,
Seto et al. 2011, UN 2017) with mass migration towards
coastlines (Creel 2003, McGranahanetal. 2007) and pol-
icy reforms that favour densication (Dallimeretal. 2011,
Kyttäetal. 2013). Population density at the coast (≤100 km
from the sea and ≤100 m above sea level) is approximately three
times higher than the global average and is increasing (Small
and Nicholls 2003). Most people are concentrated in coastal
cities that, as hubs for trade and/or due to a fertile delta, are
frequently situated where river and sea meet (Konishi 2000).
Many of these conurbations have expanded into megacities
of more than ten million people (Nicholls 1995, Li 2003).
For ecologists, coastal cities are of particular interest and con-
cern, not only from a terrestrial perspective, but also in terms
of consequences for, and interactions with, the marine envi-
ronment (Daornetal. 2015, Firthetal. 2016).
Understanding of the eects of urbanization on marine
ecosystems and ecological processes is growing (Burt 2014,
Mayer-Pintoet al. 2015, Firth et al. 2016). Human den-
sity is strongly related to resource exploitation, and one of
the early eects of marine urbanization is the depletion of
nearby shery resources (Li 2003, Kirby 2004). Coastal cit-
ies create marine pollution, including the harmful chemicals,
bacteria and sediments associated sewage and urban runo
(Homanetal. 1983, Nixon 1995, Cornelissenetal. 2008).
ey also lead to nearshore development, usually starting
with a harbour, but also including hard coastal defences to
reduce erosion of valuable land, whether it be pre-existing or
reclaimed (Charlieretal. 2005, Lotzeetal. 2005, Tianetal.
2016). ese articial structures have signicant eects on
the ecology of shorelines, especially when entire habitats are
replaced with novel materials such as concrete and gran-
ite (Firthetal. 2014, Dyson and Yocom 2015, Lokeet al.
2019a).
Several recent reviews have separately highlighted urban-
related pollution and physical modications of urban shore-
lines as critical components of urban marine ecosystem
dynamics (Daornetal. 2015, Firthetal. 2016, Heeryetal.
2018a), but exploitation of marine resources is rarely dis-
cussed in an urban context (though see Li 2003, Baumetal.
2016). e overarching characteristics of urban marine eco-
systems that result from each of these factors and their poten-
tial combined eects have yet to be thoroughly considered.
ere is considerable need to integrate ndings relating to
marine urbanization across subdisciplines of ecology; this
eort would be aided by conceptual frameworks that inte-
grate multiple variables, identify potential interactions and
feedbacks, incorporate historical trajectories, and facilitate the
development of testable hypotheses regarding the response of
urban marine ecosystems to further environmental change.
Frameworks meeting this need would not only broadly
support marine research in the Anthropocene, as nearly all
coastal zones are now strongly impacted by anthropogenic
stressors, but would also help build a foundation for urban
marine ecology – a eld in its nascence. Inevitably, urban
marine ecosystems are coupled social–ecological systems
and are heavily inuenced by what is happening ‘upstream’
in the urban fabric, by physical modications nearshore and
oshore, and by current and future consequences of climate
change, such as sea-level rise and punctuated extreme weather
events. As such, the dynamics and prevailing ecological para-
digms for these systems have yet to be tested experimentally,
and it is only through expanded eld manipulations that it
will be possible to understand the core properties of urban
marine ecosystems: how they are structured, how they func-
tion and the key parameters that drive the ecosystem services
they provide.
In this paper, we outline the primary drivers of marine
urbanization and identify the known patterns exhibited by
marine ecosystems in urban areas. Empirical testing of the
underlying processes that create these patterns and further
research in areas we highlight in this paper can help build a
framework for understanding multifaceted impacts of marine
urbanization, and future trajectories of urban marine ecosys-
tems in the face of climate change.
Three main drivers of marine urbanization
e process of marine urbanization comprises three primary
drivers (Fig. 1). e rst is exploitation of both living and
non-living resources (see ‘Resource exploitation (both liv-
ing and non-living)’) and includes recreational, subsistence
and commercial shing, as well as dredging and mining for
minerals (Table 1). In post-industrialized nations, this may
largely be historical, but with long lasting eects that are still
Figure1. Activities, installations, processes and issues that represent
instances of overlap and interaction among the three major drivers
of marine urbanization: resources exploitation, ocean sprawl and
pollution pathways.

1217
relevant today. e second is pollution (see ‘Pollution path-
ways (both industrial and domestic)’), including sediments,
industrial and municipal waste, domestic wastewater a, ani-
mal/slaughterhouse waste, fecal matter, street dust, oil from
automobiles and other contaminant sources, pharmaceuti-
cals, light pollution, and noise pollution (Table 2). e third
is the wholesale conversion of natural habitats into a dierent
state (see ‘Ocean sprawl (both coastal and oshore)’), such
as reclaimed land, seawalls, jetties, piers, marinas, groynes,
breakwaters, port and harbor infrastructure, and bridges
(collectively termed as ‘ocean sprawl’, Table 3). ese three
drivers are presented in the chronological order in which
they often begin to occur, though their timing and relative
scope can vary substantially among cities (Fig. 2). Further,
the three drivers can have interactive eects, with potential
additional consequences for marine ecosystems (see ‘Overlap,
interactions and feedbacks’). Other factors relating to urban-
ization, such as elevated propagule pressure and invasion
risk, can also be particularly intense in coastal cities (Carlton
1996, Ruizetal. 1999, 2000, Mineuret al. 2012, but see
Tanetal. 2018 and Wellsetal. 2019), however, we discuss
these primarily as they relate to one or more of the three driv-
ers presented below.
Resource exploitation (both living and non-living)
It is increasingly well documented that the overexploitation
of living coastal and marine resources is one of the earliest
observable forms of human disturbance within coastal eco-
systems (Jacksonetal. 2001, Pandoletal. 2003, Lotzeetal.
2006). Moreover, coastal systems that have endured the lon-
gest period of intense human impacts and that contain the
highest human populations are among the most degraded
(Lotzeetal. 2006). Yet, awareness of the magnitude of changes
that previously occurred as a result of the exploitation of liv-
ing and non-living marine resources is generally poor. is
is due to exploitation usually commencing prior to regular
monitoring of these systems, coupled with the pervasiveness
of the shifting baseline syndrome, where a lack of knowledge
of past ecological conditions facilitates a gradual ratcheting
down of expectations as to what constitutes a healthy ecosys-
tem (Pauly 1995, Sheppard 1995).
Table 1. Types of marine exploitation and their scope, scale and potential effects.
Exploitation type Scope and scale Potential effects on marine life and habitats
Recreational fishing Estuarine, inshore, offshore;
scale can range from hundreds
to tens of thousands of
participants in a region.
Removal of target fish and shellfish, potentially leading to population-
and/or ecosystem-wide impacts. Delayed mortality from catch and
release practices; mortality of bycatch species; mortality or injury
from boat collisions. Damage or degradation of sensitive habitats
from contact fishing gear or the launching/recovery of boats. Lost
and abandoned fishing gear issues (Table2).
Subsistence fishing and
gleaning
Estuarine, intertidal, inshore;
numbers unknown but likely
to vary greatly by region.
Removal of target fish and shellfish, potentially leading to population-
and/or ecosystem-wide impacts; mortality of bycatch species;
mortality or injury from boat collisions. Damage or degradation of
sensitive habitats from contact fishing gear and the launching/
recovery of boats. Impacts from practices such as cyanide or
dynamite fishing. Lost and abandoned fishing gear issues (Table2).
Commercial fisheries Estuarine, inshore, offshore;
scale variable by fishery and
region but can range from
tens to thousands of
participants.
Removal of large numbers of target species, potentially leading to major
population- and/or ecosystem-wide impacts; mortality of bycatch
species; mortality or injury from boat collisions; damage or degradation
of sensitive habitats from contact fishing gear and the launching/
recovery of boats. Lost and abandoned fishing gear issues (Table2).
Mariculture Estuarine, inshore (offshore
in the future); scale varies
widely and depends upon the
species being farmed.
Transmission of disease and parasites between farmed and native
species; eutrophication due to addition of nutrients (although
shellfish farms may remove nutrients from water column); smothering
of benthic fauna due to build-up of organic material (also leading to
changes to sediment type/chemistry).
Dredging for minerals/
aggregates
Inshore and offshore; scale variable
but can range
from tens to hundreds of km2.
Physical disturbance and removal of the substrate and associated
benthic biota; changes to the composition of the sediment/substrate;
changing bathymetry and sediment transport patterns; smothering of
biota; reduced light and enhanced turbidity due to sediment
suspension, toxicant release (Table2).
Beach mining Inshore; usually conducted at the
local scale but with possible
regional-scale effects.
Direct removal of species and substrate; loss of soft-sediment habitat;
lowering/loss of beach leading to erosion, changing sediment
transport patterns, increased turbidity, changing conditions for fauna/
flora and/or saline water intrusion.
Oil and gas extraction Inshore and offshore (mostly
offshore in recent years);
local to regional-scale effects.
Direct removal of species and substrate; smothering/physical alterations
to habitat/substrate type (i.e. replacement of soft with hard substrate);
chronic and acute toxic pollution events; noise pollution (Table2).
Water extraction for
cooling and
desalination
Inshore, generally
localized effects.
Fish and plankton killed during intake and processing (impingement
and entrainment). Brine and heated water (thermal pollution)
can impact communities near outflows, changing behavior and
physiology. Toxicants can also be released with the effluent.

1218
Coastal population growth and development has
impacted a wide variety of living marine resources (Table 1).
For instance, oyster reefs and maerl beds have dramatically
declined or been extirpated in coastal ecosystems around the
world due to destructive shing methods aimed at provid-
ing food and/or building material for increasingly urbanized
populations (Airoldi and Beck 2007, Claudet and Fraschetti
2010). Human population growth facilitated the establish-
ment and expansion of industrialized commercial harvesting
for marine mammals, turtles and n-sh species, ultimately
resulting in the decline or loss of marine megafauna, and of
diadromous and large demersal sh species (Lotzeetal. 2005,
urstan et al. 2010, Van Houtan and Kittinger 2014).
Targeted n-sh assemblages, although constrained by envi-
ronmental factors (e.g. availability of suitable habitat), have
been shown to decline in abundance and richness along
increasing gradients of human pressure or proximity to urban
centres in a range of habitats (e.g. coral reefs: Williamsetal.
2008, Breweretal. 2009, Aswani and Sabetian 2010; surf
zones of exposed sandy beaches: Vargas-Fonsecaetal. 2016).
Fishing eort also impacts intertidal species abundance, for
example, the majority of known sandy beach invertebrate
shery stocks are fully exploited, overexploited or depleted
due to commercial, subsistence or recreational harvesting
(Defeo and de Alava 1995, Defeo 2003).
Overexploitation often follows a predictable spatio–tem-
poral pattern that is tied to urban growth. is is particularly
evident among exploited sessile species. On the east coast of
the United States, historical oyster shery collapses demon-
strated sequential depletion beginning in urbanized estuar-
ies and spreading along the coast away from urban centres
(Kirby 2004). Many European native oyster reefs adjacent
to urban conurbations became ecologically extinct prior to
the mid-20th century (Korringa 1946, Airoldi and Beck
2007, urstan et al. 2013). Oyster Ostrea angasi reefs in
South Australia disappeared less than 200 years after the rst
records of commercial oyster landings from this region by
early Europeans (Alleway and Connell 2015). A total of ve
species of giant clam were historically recorded in the coastal
seas around Singapore, but now only two remain, and these
only exist in very low abundances (Neo and Todd 2012).
e intensication of giant clam exploitation in the 19th
century, followed by extensive coastal development from the
1960s onwards, are considered to be the main drivers in the
decline and extirpation of these charismatic invertebrates
(Guestetal. 2008, Neo and Todd 2012).
e historical legacy eects of overexploitation, combined
with pollution and coastal development, means that the pres-
ent day commercial exploitation of living marine resources
adjacent to urbanized regions, at least in more economically
Table 2. Pathways and potential effects of pollution on marine life.
Pollutant type Main urban pathways Potential effects on marine life
Sediments Construction sites (on the coast
and within inland urban areas),
dredging, land reclamation.
Turbidity resulting in less light for photosynthesis and visual
predators/prey. Down welling sediments smother benthic
organisms and create a substrate unsuitable for settling larvae.
Nitrogen and phosphorus Industrial discharge, human and
animal waste, detergents,
mariculture.
Eutrophication leading to both micro and macro algal blooms,
reduced water clarity (see ‘sediments’), shifts toward noxious
cyanobacteria and reduced fertilization success in corals.
Plastics (macro and micro),
lost and abandoned
fishing gear
Resin pellets and discarded
end-user products. Fishing
activities.
Ingestion and/or entanglement, leading to internal blockages/
injuries, toxic poisoning, starvation due to false ‘stomach filling’,
suffocation, lacerations, infections, reduced ability to swim.
Compounds from oil Motor vehicles, shipping,
industry.
Impairment of growth and developmental rates, reduced
reproductive output and recruitment rates, increased
susceptibility to disease. Carcinogenic.
Heavy/trace metals Industrial and vehicle emissions,
leaching from landfills, urban
runoff, sewage.
Can inhibit fertilization, recruitment, development, growth in
marine microorganisms, invertebrates and vertebrates.
Carcinogenic. Prone to undergo food chain magnification.
Tributyltin Antifouling paint used in the
maritime industry.
Causes imposex, and reduces growth and larval success, in
various crustaceans and molluscs. Biomagnifies, leading to
endocrine disruption in fishes, marine mammals and humans.
PCBs and PBDEs Discharge from industry, especially
electronics. Used in plastics,
fire retardants and lubricants.
Prone to biomaginification. Interferes with neurological and
hormonal systems of marine organisms and humans. Can lead
to decreases in reproductive capabilities and pose immunotoxic
risk in marine mammals.
Pharmaceuticals Industrial, hospital and domestic
waste.
Interferes with reproduction and development in both animals and
plants. Perturbs fish physiology.
Bacteria and viruses Sewage (from land and boats/ships),
aquaculture.
Diseases, especially acute gastrointestinal illnesses, e.g.
salmonellosis. Viruses can cause hepatitis and respiratory
infections.
Light Streets, private and commercial
buildings, vehicle headlights,
airports.
Encourages unwanted fouling, affects migration and predator–prey
behavior. Disrupts larval settlement. De-synchronization of
broadcast spawning from lunar phase (e.g. corals).
Noise Boat traffic, construction,
machine operation.
Disrupts behavior (e.g. ability to find food, mates or avoid
predators), reduces growth and fecundity.

1219
developed countries (MEDCs), is often far lower than its his-
torical peak (Lotzeetal. 2005, 2006). e search for resources
has thus moved further oshore and into less exploited regions
(Swartzetal. 2010, Andersonetal. 2011). Recreational sh-
ing participation rates in MEDCs have also seen a decline in
the last two decades as a result of factors related to urbaniza-
tion, such as increased urban sprawl, demographic change,
and a reduction in shable water resources (Poudyalet al.
2011). In contrast, within less economically developed
countries (LEDCs), small-scale and subsistence shing
often remains a signicant source of livelihood for coastal
communities in or near urban areas (Smitetal. 2017). e
maintenance of these traditional activities is, however, under
pressure from factors such as declining water quality and
coastal development (Smitetal. 2017), as well as enhanced
access to education and alternative employment opportuni-
ties for children of shing families (Teixeiraetal. 2016). In
some cases, urbanization may enhance economic opportuni-
ties for small-scale shing communities. In southern Brazil,
for example, the proximity of small-scale shers to urban
centres has expanded opportunities for subsistence shers to
access additional markets, as the presence of high numbers
Table 3. Types of human-made structures comprising ocean-sprawl, their functions and potential impacts. Note: All of these structures
require some alteration and/or loss of natural habitat.
Structure type Function Potential effects on marine habitats
Reclaimed land and
artificial islands
Alleviation of coastal squeeze and
expansion of land for industry and
development.
Directly results in habitat loss, and fragmentation. Sedimentation
during construction, altered hydrodynamics interferes with
connectivity at landscape and local scales.
Artificial coastal
defenses
Engineered to protect shorelines from
shoreline erosion, flooding and impacts
from waves.
Reduced intertidal extent resulting in steeper slopes. Footprint of the
structure removes existing natural habitat but effects may extend
beyond structure (halo effect). Change in substrate material and
altered hydrodynamics could result in different colonizing
assemblages.
Commercial ports,
docks and marinas
Industry, services and recreation. Elevated risk of species invasions, contaminants (oil, antifouling
coatings, noise, light), disturbances associated with shipping
(sediment resuspension, propeller injuries, etc.).
Oil shipping and
refinery infrastructure
Non-renewable resource mining for
energy.
Footprint of the structure removes existing natural habitat but effects
may extend beyond structure (halo effect). Contaminants, risk of
oil spills, noise and light pollution.
Tidal and wind energy
infrastructure
Energy production. Footprint of the structure removes existing natural habitat but effects
may extend beyond structure (halo effect). Noise and light
pollution, electromagnetic fields.
Submarine cables and
pipelines
Telecommunications, power, water, oil. Concrete mattresses are often used to stabilize and position cables
on seafloor. Fragmentation of soft-sediment habitats due to
introduction of hard substrates. Noise and light pollution during
construction phase. Electromagnetic fields.
Figure2. Trajectories of the three key drivers of marine urbanization over time are dicult to hindcast (or forecast) and are likely to be city-
specic. However, they will almost certainly overlap, potentially creating non-linear interactions that are even more challenging to predict
(and are not represented here). For illustration purposes only: (a) the exploitation of living resources could accelerate rapidly during the early
development of many coastal cities, yet decrease in intensity as the resource is overexploited or inaccessible due to other factors, such as
contaminants. Conversely, ocean sprawl may be more likely to follow an asymptotic trajectory, which reaches saturation as an increasingly
large percentage of natural habitats are converted by the installation of articial structures. (b) A possible alternative conguration of driver
trajectories in a younger city with a shorter but equally intense history of marine urbanization.

1220
of shers enables them to supply enough sh to meet supply
chain demand (Hellebrandt 2008).
Urbanization also coincides with increases in the exploi-
tation of non-living resources, including the extraction of
marine aggregates (sand, gravel, rocks) for use in construc-
tion and beach renourishment, mineral resources for indus-
trial applications, and the extraction of energy resources (oil
and natural gas, and wave and tidal resources). Nearshore
aggregate dredging may occur for mud, rock, shells, corals or
sand for construction purposes, or for the heavy or precious
minerals they contain (Charlier and Charlier 1992). Potential
negative eects arising from the extraction of coastal marine
aggregates include an increased risk of ood events and coastal
erosion. For example, aggregate extraction from the coasts of
Kiribati in the South Pacic resulted in beach structure being
degraded, exposing coastal conurbations to enhanced risk
of ood events (Webb 2005, Holland and Woodru 2006).
Similarly, beach mining, nearshore dredging and quarry-
ing have contributed signicantly to coastal erosion in the
Marshall Islands (Holland and Woodru 2006), France, and
Bali (Charlier and Charlier 1992). e extraction of sand for
the renourishment of urban beaches is commonly undertaken
for aesthetic and erosion control purposes (Fletemeyeretal.
2018). Knowledge of the direct and indirect eects of this
activity on the local biota and ecological processes remains
incomplete (Peterson and Bishop 2005), but beach renour-
ishment has been shown to negatively impact nearshore
coral reefs (Hernández-Delgado and Rosado-Matías 2017),
marine invertebrate prey availability and nesting behavior in
sea turtles (Peterson and Bishop 2005). Coastal urbanization
also facilitates the expansion of maritime port operations,
which often dredge nearshore channels to maintain deep-
water access for commodity and passenger transport (Lemay
1998). Dredging and mining represent a major area of over-
lap between exploitation and pollution (Fig. 1) due to the
release of toxicants and sediments that occurs during these
operations.
e establishment of oil and natural gas rigs can be bro-
ken down into four stages: seismic exploration, exploratory
drilling and installation, operation and decommissioning
(Khan and Islam 2008). Each of these stages involves some
form of extractive activity, although the consequences for
marine life are particularly strong during the installation and
decommissioning stages. e installation and decommission
of rig infrastructure may also degrade or destroy the seabed
(Macreadieetal. 2011). However, their establishment intro-
duces a source of hard substrate, potentially increasing local
biodiversity, as well as non-native species, which can alter
community dynamics at local or regional levels (Burtetal.
2009, Fearyet al. 2011, Macreadieet al. 2011). e estab-
lishment of renewable energy infrastructure presents many of
the same ecological issues and opportunities as oil and gas, yet
the installation of some structures, such as tidal barrages, has
the potential for generating signicant physical and ecologi-
cal impacts at the local scale, including the loss of intertidal
habitats, modication of water ow and sediment resuspen-
sion (Gaoetal. 2013, Hooper and Austen 2013).
Pollution pathways (both industrial and domestic)
Urbanization and pollution are tightly linked; whereas as air
and soil pollution are major concerns for terrestrial conur-
bations, contaminated water and sediments are additional
and often critical pollution issues for coastal cities (Table 2).
Originating from both point (e.g. wastewater discharge) and
non-point (e.g. wind-blown debris and dust) sources, pol-
lution impacts marine life at individual, population and
ecosystem levels, frequently bioaccumulating and then bio-
magnifying up the trophic pyramid (Erftemeijeretal. 2012,
Johnstonetal. 2015, Langston 2017). Chronic marine pol-
lution eects tend to be sub-lethal (Browneetal. 2015), but
they can interact with other stressors in ways that ultimately
cause mortality (Yaakubetal. 2014a, Bårdsenetal. 2018).
Urban sediment pollution, commonly the result of runo
from construction work and disturbance via dredging (Rogers
1990, Eggleton and omas 2004, Erftemeijeretal. 2012),
as well as other sources such as beach nourishment and land-
use changes that alter catchment runo (Colosioetal. 2007,
Zhangetal. 2010), aects marine life in multiple ways. e
resulting increase in turbidity reduces light penetration, pho-
tosynthesis (Falkowskietal. 1990), and the maximum depth
at which photosynthetic organisms can grow (Heeryetal.
2018a). Suspended sediments also reduce sh hatching suc-
cess and larval survival (Auld and Schubel 1978), impede
zooplankton feeding (Sewetal. 2018), aect mobile fauna
that rely on visual cues (Weienetal. 2006), and alter a wide
range of benthic ecosystem processes and patterns (Airoldi
2003), including the settlement and successful recruitment of
organisms, the diversity of species, and competitive interac-
tions – such as those between foundation macrophyte species
and low-lying algal turfs (Gorgula and Connell 2004, Russell
and Connell 2005, Gorman and Connell 2009, Knottetal.
2009, Baumanetal. 2015). Smothering by sediment further
reduces light and physically interferes with the functioning
of benthic organisms like corals (Rogers 1990, Junjieet al.
2014), seagrasses (Erftemeijer and Lewis 2006), and certain
life stages of kelps (Devinny and Volse 1978, Geangeetal.
2014).
High nutrient concentrations are frequently attendant
with sediments but, in urban settings, inputs come also from
wastewater treatment plants, industrial discharges, storm-
water runo, dust from land, domestic detergent use and
human sewage (McClellandetal. 1997, Braga et al. 2000,
Atkinsonetal. 2003, Coleetal. 2004, Gawetal. 2014, Vikas
and Dwarakish 2015) and are particularly hazardous in bays
and harbors with limited circulation (Gomezet al. 1990).
Resultant eutrophication can have positive feedbacks on nutri-
ent loads and localized acidication (Howarthetal. 2011)
and leads to many undesirable ecological eects (Bell 1991,
Orthetal. 2017), for example phytoplankton blooms and/
or shifts toward noxious cyanobacteria, macroalgal blooms
that can outcompete foundation species such as corals, and
increases in the occurrence and severity of marine diseases
(Bowen and Valiela 2001, Balestrietal. 2004, Lapointeetal.
2005, Reopanichkul et al. 2009, Haapkylä et al. 2011,

1221
Reddinget al. 2013). Human sewage and wastewater cre-
ates additional problems due to the release of fecal coliforms,
antibiotics and other pharmaceuticals (Jiang et al. 2001a,
Shibataetal. 2004, Watkinsonetal. 2007, Roseetal. 2009,
Jiaetal. 2011, Rizzoetal. 2013, Gawetal. 2014).
Toxic pollutants, including organochlorine compounds
(e.g. PCBs and HCH), heavy metals, tributyltin (TBT),
polybrominated diphenyl ethers (PBDEs) and compounds
from oil (e.g. petrogenic PAHs, plastics and microplastics),
are strongly associated with industrial activities and urban
run-o (Kennish 1997, Shazilietal. 2006, Toddetal. 2010,
Coleetal. 2011, Tayebetal. 2015), as well as from ship-
ping and other sea-based sources (Tornero and Hanke 2016).
Many of these substances bioaccumulate in animals (Tanabe
1988, Woletal. 1993, Bayenetal. 2003) interfering with
cellular and biochemical functions and disrupting hormonal,
reproductive, neurological and nervous systems (Portmann
1975, Woletal. 1993, Frigoetal. 2002, Boschetal. 2016).
Lead, cadmium, copper, tin, nickel and iron are among the
metals commonly found in sediments near industrial areas
(Williamson and Morrisey 2000, Buggy and Tobin 2008,
Amin et al. 2009). Copper is especially toxic to marine
invertebrates, including poriferans, cnidarians, molluscs and
arthropods (Reichelt-Brushett and Harrison 1999, 2000,
Johnston and Keough 2000, Brownetal. 2004, Rainbow
2017). e impacts of lead and cadmium on economically
important invertebrates such as oysters and crabs are also
well established in the literature (Ramachandranetal. 1997),
however, recent studies suggest deleterious eects from a
wide range of metals (Langston 2017), particularly when
combined with other anthropogenic stressors (Burton and
Johnston 2010). Other industrial discharges that are known
to have negative eects, albeit usually localized, include brine
from desalination plants and heat from industrial cooling.
Often the most deleterious impacts from these discharges are
toxicants (especially metals, hydrocarbons and anti-fouling
compounds) that enter the sea with the euent (Lattemann
and Höpner 2008, Robertsetal. 2010).
Urban noise pollution usually originates from boat traf-
c and in-water construction (Middel and Verones 2017)
while urban light pollution comes from street lights, build-
ings, shipping, airports and vehicle headlights (Hölkeretal.
2010). For some sh and marine mammals, noise pollu-
tion inhibits communication, aects predator–prey interac-
tions, and has negative eects on growth and reproduction
(Slabbekoorn et al. 2010, Houghton et al. 2015). It may
also impact various other taxa that are sensitive to sound,
such as oysters (Charietal. 2017), clams (Mosher 1972,
Peng et al. 2016), mussels (Roberts et al. 2015), cepha-
lopods (Andréet al. 2011, Fewtrell and McCauley 2012),
shrimp and other invertebrates (Solan et al. 2016). Night
lighting comprises both direct glare and overall increased
illumination, and can disrupt marine ecosystems in a num-
ber of ways (Hölkeretal. 2010). Organisms that use light
to navigate, such as birds and sea turtles, may become dis-
orientated (Daviesetal. 2014), as may sh and sh larvae.
Articial lighting has also been reproted to aect predator
and prey behavior, disrupt larvae settlement, alter distribu-
tion patterns and de-synchronize broadcast spawning species
from normal lunar phases (Beckeretal. 2013, de Sotoetal.
2013, Waleetal. 2013, Navarro-Barranco and Hughes 2015,
Boltonetal. 2017).
A gradient of decreasing levels of various pollutants
with increasing distance from urban sources has been
described multiple times, particularly for: heavy metals
(Qiao et al. 2013), sediments (Todd et al. 2004), marine
debris (Evansetal. 1995, Andradesetal. 2016), and PAHs
(Assunçãoetal. 2017). Whereas the eects of urban (land-
based) light and noise pollution and some contaminants are
limited to a few decimeters to kilometers from the source
(Zaghdenetal. 2005, Burton and Johnston 2010), other pol-
lutants have impacts that extend much further (Heeryetal.
2017). For example, PCBs have been found in Arctic waters
far from any urban or industrial centres, albeit at very low
levels (Gioiaetal. 2008). An important example of urban
pollution being transported huge distances but still having a
substantial negative impact is marine debris, especially plas-
tics. Like other forms of marine debris, plastics have a very
high dispersal potential (Carltonetal. 2017), mainly because
they can take decades to biodegrade (Moore 2008) and are
often buoyant. ey can maintain their structural integrity
for many years, resulting in negative eects, via ingestion or
entanglement, to animals ranging from seabirds, turtles and
marine mammals to crustaceans and cnidrians (Azzarello and
Van Vleet 1987, Moser and Lee 1992, Bjorndaletal. 1994,
Jones 1995, Laist 1997, Lambetal. 2018, Mecalietal. 2018)
far from their point of origin. Due to ultraviolet rays, mechan-
ical and microbial degradation, plastics eventually fragment
into microplastics (ompsonetal. 2004, Barnesetal. 2009)
that are bioavailable to suspension feeding marine organisms,
including zooplankton (Browneet al. 2008, Wrightet al.
2013, Barbozaetal. 2018, Botterelletal. 2018).
Ocean sprawl (both coastal and offshore)
‘Ocean sprawl’ is a term used to describe the proliferation
of human-made hard structures in the marine environment
(Duarteetal. 2013, Firthetal. 2016, Table 3). is encom-
passes oshore infrastructure (e.g. wind farms, oil and gas
platforms, aquaculture facilities, submarine cables/pipes) and
coastal infrastructure such as articial shore defences (e.g. sea-
walls, breakwaters, groynes), as well as facilities associated with
ports, docks and marinas. Ocean sprawl is a fundamental and
dominant feature of urbanized marine environments (Bulleri
and Chapman 2010, Duarteetal. 2013, Daornetal. 2015,
Firth et al. 2016) with articial structures comprising the
bulk of shorelines in many coastal cities (Bullerietal. 2005,
Todd and Chou 2005, Daornetal. 2015, Laietal. 2015)
and modifying habitats well into the subtidal zone (Airoldi
and Beck 2007, Heeryetal. 2017, Heery and Sebens 2018,
Macuraetal. 2019).
As a habitat, articial shorelines are quite distinct
from natural rocky shores (Connell and Glasby 1999,

1222
Rilov and Benayahu 2000, Perkol-Finkel and Benayahu
2004, Bullerietal. 2005, Moschellaetal. 2005, Clynicketal.
2008, Lametal. 2009, Bulleri and Chapman 2010, Laietal.
2018). One of the most obvious dierences is the slope of
hard substrates; while shoreline armoring structures such as
seawalls are generally very steep, natural rocky shores tend to
be more gently sloping with longer and wider intertidal areas
(Gabriele et al. 1999, Knott et al. 2004, Andersson et al.
2009, Chapman and Underwood 2011, Firthet al. 2015).
e smaller area of intertidal zone typical of seawalls is prob-
ably an important contributor to species loss (Chapman
and Underwood 2011, Perkinsetal. 2015) as it can lead to
greater overlap in the distribution of individuals (Kleinetal.
2011) or to superimposed distributions of species that would
not normally occur (Lam et al. 2009, Loke et al. 2019b).
Wave impact is also more intense on steep shores (Gaylord
1999, Cuomoet al. 2010), potentially dislodging intertidal
organisms and/or impeding their settlement (Blockley and
Chapman 2008, Iveša et al. 2010). Compared to natural
hard-bottom habitats, seawalls are topographically ‘sim-
ple’ (Lokeet al. 2014) – having few microhabitats, such as
pits, rock-pools, overhangs and crevices (Chapman 2003,
Chapman and Bulleri 2003, Moreiraetal. 2007), which are
important for the persistence of many intertidal and benthic
species (Chapman and Underwood 2011, Loke and Todd
2016, Lokeetal. 2017). When considering these multiple
eects in combination, it is unsurprising that many direct
comparisons between rocky shores and seawalls often reveal
the latter host lower species richness, reduced functional and
genetic diversity, and dierent community compositions
(Chapman 2003, Bullerietal. 2005, Moschellaetal. 2005,
Fauvelotetal. 2009, Laietal. 2018).
e consequences of ocean sprawl at large spatial scales
are not yet well understood, but they are likely to be con-
siderable given its prominence and extent (Lotzeetal. 2006,
Airoldi and Beck 2007). In some heavily urbanized regions,
entire habitats have been lost as articial structures prolifer-
ate over vast distances (Dongetal. 2016). Even where coastal
transformation is not ubiquitous, clusters of articial struc-
tures can serve as corridors that facilitate species invasions
(Airoldietal. 2015) and alter ecological connectivity, with
signicant eects on marine assemblages (Bishopetal. 2017).
e spatial scale of impacts from articial structures depends
on the type of structure, local hydrodynamic conditions, and
a variety of other parameters (summarized by Heeryetal.
2017). For instance, uxes of exogenous detritus from arti-
cial structures typically aect marine communities within
meters to tens of meters only (Heery and Sebens 2018), while
infrastructure that creates major impediments to circulation
and sediment transport tends to impact marine assemblages
across a much larger area (Bishopetal. 2017).
Overlap, interactions and feedbacks
e three key drivers described above are not limited to
urban areas, yet their relative magnitude and spatial and tem-
poral overlap is often augmented near high-density coastal
development (Jiangetal. 2001b, Kennish 2002, Finkl and
Charlier 2003, Mayer-Pintoetal. 2015). is overlap can
have important consequences for marine organisms and
communities, as eects from multiple anthropogenic stress-
ors are often cumulative and non-linear in the marine envi-
ronment (Adams 2005, Crain et al. 2008, 2009), leading
to complex changes in ecosystem condition (Conversietal.
2015, Halpernetal. 2015, Möllmannetal. 2015). It can
also feedback to inuence the key drivers themselves, which
are each the result of dynamic, interacting socio-economic
and biophysical forces (sensu Albertietal. 2003), and closely
interrelated in the coupled social-ecological systems that
characterize coastal cities (Liu et al. 2007, Alberti 2008,
Grimmetal. 2008a, Pickettetal. 2011). Such feedbacks and
interactions are widely recognized as shaping urban ecosystem
function (Wu 2014, McPhearsonetal. 2016), and are cen-
tral in nearly all current models of urban ecosystem dynamics
(Pickettetal. 2001, Albertietal. 2003, Grimmetal. 2013).
In this section, we highlight some known and likely inter-
actions among the three drivers (exploitation, pollution and
ocean sprawl) of marine urbanization. Each interaction ts
conceptually within the overlapping regions of the Venn dia-
gram in Fig. 1.
One of the best examples of complex interactions and
feedbacks among the drivers of marine urbanization and
ecosystems is the relationship between habitat conversion,
contaminants and invasion risk. Articial structures associ-
ated with port infrastructure and shoreline protection tend
to both concentrate environmental contaminants by alter-
ing hydrodynamic patterns and reducing water movement
(Walthametal. 2011, Riveroetal. 2013), and by facilitat-
ing increased contaminant inux, for instance from antifoul-
ing paints (Schietal. 2004, 2007, Warnkenetal. 2004,
Simetal. 2015). Copper emissions from antifouling paints
then have both direct and indirect consequences for marine
organisms (Rygg 1985, Perrettetal. 2006). e toxin enters
the food web by accumulating in algal tissues (Johnstonetal.
2011) or being consumed directly by non-selectively feed-
ing animals, which can additionally accelerate the leaching
and burial process in adjacent sediments (Turner 2010).
While toxic eects from copper negatively impact many
marine organisms and reduce diversity (Rygg 1985), dier-
ential responses to copper contamination among inverte-
brates (Piola and Johnston 2006) combined with the novel
colonization habitat that is provided by oating docks and
other marina structures can disproportionately favor non-
indigenous taxa, thus facilitating marine invasions (Piola
and Johnston 2008, Daornet al. 2009, Piolaetal. 2009,
Airoldi and Bulleri 2011, Edwards and Stachowicz 2011,
Cordelletal. 2013, McKenzieetal. 2012).
e trajectory of marine resource exploitation in urban
areas is also closely tied to that of pollution pathways and
ocean sprawl (Inglis and Kross 2000, Jiang et al. 2001b,
Cundyetal. 2003) (Fig. 2). In the early developmental stages
of many cities, shoreline habitats were converted by articial
structures to facilitate resource exploitation industries and

1223
the economic growth they fueled (Squires 1992). Overwater
structures that housed cannery facilities and seafood markets
were prominent drivers of early waterfronts in San Francisco
(Walker 2001), Singapore (Chang and Huang 2011), and
many other coastal cities globally (West 1989, Portmanetal.
2011). Various shoreline armoring structures were also part
of facilities for resource exploitation industries, such as oil
and gas (Minca 1995), and remain important drivers in adap-
tation plans for protecting these industries from future sea
level rise (Frenchetal. 1995, Ng and Mendelsohn 2005).
Pollution associated with resource exploitation and habi-
tat conversion continues to be problematic in many urban
and suburban areas, for instance surrounding shellsh
aquaculture farms, oil reneries, port infrastructure and
dredged waterways that harbor contaminants (Board 1997,
Pereiraet al. 1999, Jones et al. 2001, Strand and Asmund
2003, Tolosa et al. 2004, Medeiros et al. 2005, Casado-
Martínezetal. 2006, Paisséetal. 2008, Knottetal. 2009),
and alters system dynamics via multiple biogeophysical path-
ways, trophic levels and functional groups (Paisséetal. 2008,
Weisetal. 2017).
As coastal cities grow, and eects from various aspects of
marine urbanization increasingly overlap (Fig. 2), the system’s
potential for feedbacks appears to intensify (Fernando 2008,
Grimmetal. 2008b). For instance, as impervious surfaces
proliferate on land, increased delivery of stormwater can
accelerate the accumulation of contaminants in receiving
waterbodies (Leeetal. 2006, Jartunetal. 2008, 2009, Jartun
and Pettersen 2010, Walshetal. 2012). Similarly, as resource
exploitation and shoreline alteration expand, so too does the
spatial extent and magnitude of marine debris and contami-
nants (Garcia-Sandaetal. 2003, Wake 2005, Ng and Song
2010, Märkletal. 2017), which can in turn impact exploit-
able marine resources (Islam and Tanaka 2004). Additional
biogeochemical and ecological feedbacks have also been
important historically, in some cases leading to losses in a
system’s capacity to absorb urban impacts over time (Cloern
2001, Nyströmetal. 2012). For instance, the loss of oyster
reefs due to overharvesting and eutrophication is thought to
have reduced the ltration capacity of urban estuaries in the
United States (Zimmerman and Canuel 2000, Kempetal.
2005, Wilbergetal. 2011, zu Ermgassenetal. 2013), poten-
tially inhibiting their ability to accommodate further pollu-
tion inux. Similar feedbacks surrounding challenges such as
harmful algal blooms and marine diseases may be increas-
ingly likely as ecosystems are further altered by marine urban-
ization (Prinsetal. 1997, Sunda et al. 2006, Heisleretal.
2008, Crainetal. 2009). However, such feedbacks can be
dicult to predict and may obfuscate eorts to eectively
anticipate ecosystem response to further environmental
change (Elmqvistetal. 2003).
Key ecological patterns
e convergence of exploitation, pollution and ocean sprawl
that typies urban marine environments may lead to shifts in
ecosystem characteristics and several key ecological patterns,
which are just beginning to emerge in the literature.
Homogenized systems, comprising heterogeneous
mosaics
A common theme in the terrestrial urban ecology literature
is the spatial heterogeneity that occurs across landscapes as
a result of urbanization (Pickett et al. 1997, Dow 2000,
Cadenasso et al. 2007, Pickett and Cadenasso 2008). e
resulting ‘mosaics’ of habitat types, biophysical charac-
teristics, and land use are temporally dynamic and inu-
enced by multiple interacting social and ecological drivers
(Pickettetal. 2017). At the same time, there are consider-
able similarities across cities in the underlying processes and
trajectory of urbanization, leading to an overall homogeniza-
tion among urban ecosystems regionally and globally (Alberti
2005, McKinney 2006). Even though research supporting
these concepts is far more comprehensive in terrestrial envi-
ronments, there are several indications of comparable pat-
terns among urban marine ecosystems based on the current
literature (Daornetal. 2015).
Most coastal cities are positioned in estuaries and bays that
were historically dominated by soft sediments. As articial
structures are added to these sedimentary environments, a
checkerboard of hard and soft habitats is created, with each
supporting distinct biotic assemblages (Connell and Glasby
1999, Glasby 2000, Connell 2001, Barros et al. 2001).
is can alter ecosystem dynamics in several ways. In some
regions, articial structures support a larger standing stock
of benthic macroalgae and other hard-bottom organisms,
which then enter adjacent sediments as detritus and may
alter sedimentary community dynamics (Boehlert and Gill
2010, Heery 2018, Heery and Sebens 2018). Articial struc-
tures has been shown to o act as ‘stepping stones’ for disper-
sal, particularly of non-indigenous taxa (Bulleri and Airoldi
2005, Glasbyetal. 2007, Vasellietal. 2008, Sheehy and Vik
2010, Airoldietal. 2015, Fosteretal. 2016) and alter genetic
population structure of marine fauna (Fauvelotetal. 2012).
Marine species vary in dispersal potential, and many taxa
encounter barriers to dispersal at relatively small spatial scales
(Darlingetal. 2009, Costantinietal. 2013, Maasetal. 2018,
Sefbomet al. 2018). Dispersal limitation can therefore also
interact with local stressors and abiotic conditions to result in
compositionally very dierent assemblages across patches of
hard substrata (Bulleri and Chapman 2004, Munari 2013).
is may be accentuated where urban habitat conversion has
signicantly altered hydrodynamic patterns, created other
additional barriers to dispersal and subsequent settlement
(Bishopetal. 2017) or changed the conguration of habitats
at the landscape scale (Lokeetal. 2019c).
Spatially heterogeneous mosaics also form in urbanized
seascapes as a result of ne-scale gradients in nutrient enrich-
ment and sediment pollution (Airoldi 2003, Baum et al.
2015, Ling et al. 2018), particularly in low ow environ-
ments and enclosed estuaries and embayments (Balls 1994,
Daueret al. 2000). For instance, physical disturbance from

1224
swing moorings, which are ubiquitous in shallow sedimen-
tary environments in Sydney Harbor, leads to depressed con-
centrations of metal contaminants within a highly localized
area (Hedgeet al. 2017). is may result in complex, ne-
scale spatial patterns in microbial, meiofaunal and macrofau-
nal taxa that are sensitive to metal contamination (Coull and
Chandler 1992, Stark 1998, Lindegarth and Hoskin 2001,
Muchaet al. 2003, Gillanetal. 2005, Sunet al. 2012). It
is likely this is complicated further by localized gradients in
other abiotic conditions, such as granularity, that commonly
occur in the vicinity of articial structures (Martinetal. 2005,
Seitzetal. 2006). While swing moorings and other struc-
tures that increase physical disturbance and scour increase
sediment grain size (Hedgeetal. 2017), structures such as
pilings that reduce ow speeds and increase deposition tend
to reduce the grain size of nearby sediments (Heeryet al.
2018b). Grain size, contaminant concentrations, and a vari-
ety of other ow-related metrics are known to have strong
eects on sedimentary composition and diversity (Mannino
and Montagna 1997, Hewittetal. 2005), which likely varies
considerably in urban seascapes over small spatial scales.
Studies of marine diversity and connectivity relative to
urbanization remain relatively limited, and there is need for
expanded work in this area. In particular, study designs that
allow for the assessment of alpha, beta and gamma diver-
sity could be helpful for beginning to distinguish between
the ecological processes that shape marine assemblages in
spatially heterogeneous urban seascapes. In their eDNA
study on seagrass beds, Kellyetal. (2016) found decreases
in beta diversity even while species richness increased with
the intensity of urbanization. Landscape-scale homogeniza-
tion in urban assemblages has some precedents in freshwa-
ter and terrestrial systems (McKinney and Lockwood 1999,
Holway and Suarez 2006, Urbanetal. 2006, Gromanetal.
2017), but less so in the marine literature (Balataetal. 2007).
For instance, by creating urban freshwater reservoirs/dams
many cities have inadvertently fragmented their catchments
and resulted in biotic homogenization (Olden and Rooney
2006, Oldenetal. 2008). e straightening or ‘linearization’
of shorelines through armoring (Dyl 2009) could homog-
enize intertidal communities at certain scales, though this
has not been demonstrated empirically. Sedimentation may
also cause marine communities to become more homogenous
under certain conditions (Balataetal. 2007). However, more
thorough characterization of diversity measures relative to
resource exploitation, pollution and ocean sprawl should
advance understanding of ecological processes in urban
marine environments.
Loss of foundation species
Urban stressors can be particularly detrimental for sensitive
foundation species such as oysters, reef-building corals, sea-
grasses, mangroves and canopy-forming kelps, which structure
marine ecosystems via the provisioning of biogenic habitat
(Dayton 1972, Bertness and Callaway 1994). Even though
the dynamics of decline vary among taxa and across locations
(Terradosetal. 1998, Waycott etal. 2009, Polidoroetal.
2010, Heeryetal. 2018a), loss in foundation species is gen-
erally tied to one or more of the three major drivers of marine
urbanization (Rogers 1990, Hastings et al. 1995, Airoldi
2003, Balestrietal. 2004, Kirby 2004, Connelletal. 2008,
Strainetal. 2014, Yaakubetal. 2014a, Alleway and Connell
2015). In temperate areas, nutrient-rich, high sedimentation
conditions can limit the recruitment and survival of canopy-
forming kelps while supporting opportunistic, turf-forming
algal species that can act as kelp competitors (Airoldi 1998,
Benedetti-Cecchi et al. 2001, Gorgula and Connell 2004,
Russell and Connell 2005, Colemanet al. 2008, Gorman
and Connell 2009). Similarly, in the tropics, sediment pol-
lution has multiple negative eects on corals. ese decrease
coral cover and disproportionately impact competitive,
branching coral genera such as Acropora, which ultimately
lowers reef complexity in urban areas (Heeryet al. 2018a).
Ocean sprawl can also be an important driver of foundation
species loss. For instance, despite the numerous ecosystem
services they provide to urban communities (Benzeevetal.
2017), mangrove forests are cleared in many coastal areas
to make way for urban development (Harper et al. 2007,
Martinuzziet al. 2009, Laietal. 2015, Richards and Friess
2016). Where urban mangroves are left intact, they are
vulnerable to deleterious eects from articial structures con-
structed nearby; mangrove forests adjacent to seawalls tend to
be narrower, with less leaf litter and fewer saplings than those
without seawalls (Heatherington and Bishop 2012). Coral
reefs and seagrass beds are also frequently built over (Chou
2006, Burtetal. 2013, Yaakubetal. 2014b). Furthermore,
urban losses in foundation species often involve feedbacks
that prevent subsequent population recovery (Altieri and
Witman 2006, de Boer 2007, Moore et al. 2014). For
instance, seagrass loss can be tied to sediment pollution and
eutrophication (Waycottetal. 2009, Orthetal. 2017) and
deforestation and altered hydrodynamic regimes from coastal
construction (da Silvaetal. 2004), as well as possible indirect
eects from top to down reductions in grazers that control
seagrass epiphyte loads (Duyetal. 2005, Myersetal. 2007).
e reduction of seagrass bed cover can lead to destabilization
of sedimentary substrata, which then further increases tur-
bidity (de Boer 2007) and potentially inhibits recolonization
(Mooreetal. 2014).
ere is increasing evidence that multiple, often interact-
ing, urban-related drivers aect both foundation species and
ecological response to foundation species loss (Lenihan and
Peterson 1998, Jackson 2008, Claudet and Fraschetti 2010,
Nyströmetal. 2012, Strainetal. 2014, Ferrarioetal. 2016,
Orthetal. 2017), although studies evaluating multiple urban
stressors simultaneously are rare (O’Brienet al. 2019). e
abundance of kelps and other important habitat-forming mac-
roalgae is negatively correlated with human population den-
sity in several regions, including temperate coasts in Australia
and North America (Connell et al. 2008, Scherner et al.
2013, Feist and Levin 2016), and this is likely linked to gra-
dients in sedimentation and nutrients (Fowleset al. 2018).

1225
Yet, ocean sprawl may also be an important factor in mac-
roalgal community dynamics. Reduced topographic com-
plexity, changes in substrate type, and altered substrate
proles are all factors that can limit kelp abundance (Toohey
2007, Schroeteretal. 2015) and correlate with urban habi-
tat conversion. Articial structures not only support distinct
macroalgal assemblages compared with natural rocky shores
(Glasby 1999) – the kelps that inhabit them also support dis-
tinct epifaunal and microbial communities and erode at dif-
ferent rates (Marzinellietal. 2009, 2018, Mayer-Pintoetal.
2018). Habitat conversion thus likely inuences ecological
processes in urban areas where canopy-forming kelps persist.
e interaction of resource extraction, pollution and ocean
sprawl as drivers of foundation species loss, and the ecological
responses to this loss, are important future areas of research.
Importantly, these processes are highly dynamic, with ecolog-
ical legacies from past impacts, and future scenarios linked to
rising temperatures and pCO2, that are challenging to ascer-
tain (Ramalho and Hobbs 2012, Davisetal. 2017, Gaoetal.
2017, Heldtetal. 2018, Fig. 2).
Changes in biodiversity and productivity
Patterns of biodiversity in urban marine environments are
complex. Resource extraction, sediment pollution and habi-
tat modication are important drivers of marine biodiversity
declines globally (Sala and Knowlton 2006), and there are
many examples from the literature of reduced species rich-
ness and altered community composition at heavily urban-
ized sites (Pearson and Rosenberg 1978, Longetal. 1995,
Lindegarth and Hoskin 2001, Lotzeetal. 2006, Airoldi and
Beck 2007, Poquita-Du 2019). Even though the diversity
of marine assemblages in some regions is negatively corre-
lated with human population density (Scherneretal. 2013,
Neoetal. 2017), this pattern is not universal, and varies con-
siderably between regions, cities, the taxa and type of diver-
sity considered, and the methods used. For instance, using
eDNA from water samples, Kellyetal. (2016) found that
species richness was positively correlated with land-based
urbanization in intertidal seagrass beds. Similarly, while some
studies have reported higher species diversity on articial
shorelines than on their natural counterparts (Chou and Lim
1986, Connell and Glasby 1999, Munschetal. 2015), others
have found articial shorelines to be relatively depauperate
(Firthetal. 2013, Aguileraetal. 2014, Laietal. 2018).
ere are similar complexities surrounding productivity in
urban marine environments. In nutrient-rich marine estuar-
ies, like those in most coastal cities, climate variables, such
as major precipitation events and interannual uctuations in
weather patterns, tend to be particularly important drivers of
temporal patterns in primary production (Mallinetal. 1993,
Rodrigues and Pardal 2015), as these events deliver land-
based sources of nitrogen to coastal waters. However, the
relationship between nutrient load and primary production is
highly variable (Borum and Sand-Jensen 1996), and urban-
related increases in nutrient loads can have dierent eects
depending on tidal regimes, the system’s trophic structure,
as well as other factors (Alpine and Cloern 1992, Monbet
1992). Nutrient loading therefore does not manifest com-
parable, elevated marine production across cities. Moreover,
broader ecosystem responses to primary production also vary
across urban marine ecosystems. In some locations, nutri-
ent enrichment can trigger micro- and macroalgal blooms
that are highly detrimental to important foundation species
(McGlathery 2001) while, in other places, the same process
may increase secondary production (Leslieetal. 2005) and
species richness (Whittaker and Heegaard 2003).
Novel assemblages
Novel assemblage structure tends to emerge as species move
and change in abundance and dynamics in response to envi-
ronmental change (Hobbset al. 2018). e most obvious
manifestation of this phenomenon in urban marine environ-
ments is among sessile assemblages on articial shorelines.
Conversion from natural shores to hard articial struc-
tures creates new habitats for colonization and supports
novel assemblages of hard-bottom organisms (Chou and
Lim 1986, Connell and Glasby 1999, Bulleriet al. 2005,
Moschellaetal. 2005, Clynicketal. 2008, Lametal. 2009,
Airoldietal. 2015, Munschetal. 2015). ese assemblages
dier from nearby rocky shores with respect to composition
(Chapman 2003, Bulleri and Chapman 2010, Airoldietal.
2015, Laietal. 2018) and genetic diversity (Fauvelotetal.
2009). Dierences in species abundance between articial
and natural rocky shores may be biased towards some func-
tional groups, such as motile primary consumers (Chapman
2003, Pister 2009). However, human-made habitats in urban
areas also provide a foothold for a variety of non-indigenous
species, many of which are non-motile (Glasbyetal. 2007,
Vasellietal. 2008, Ruizetal. 2009, Sheehy and Vik 2010,
Simkaninetal. 2012, Airoldietal. 2015, Fosteretal. 2016).
Ruderal species and potential synanthropes
On land, urbanization is strongly associated with the pro-
liferation of ruderal and synanthropic species (McKinney
2006). Ruderal species, those that grow in contaminated
soils or human wastes, typically include a variety of weedy
plant species (Haigh 1980), while ‘synanthropes’ is a term
typically applied to mid-level consumers, such as raccoons
and coyotes, that have higher densities and abundances in cit-
ies than in adjacent rural areas (McKinney 2002). Although
not well studied, there is evidence of analogue taxa exploit-
ing urban marine environments. Polluted sediments in urban
areas appear to generate opportunities for certain marine
microbes (Córdova-Kreylos et al. 2006, Cetecioğlu et al.
2009, Nogueiraet al. 2015). For instance, Alteromonadales,
Burkholderiales, Pseudomonadales, Rhodobacterales and
Rhodocyclales bacteria that are involved in the degradation
of hydrocarbons, were found to be more abundant in pol-
luted urban mangrove forests in Brazil (Marcial Gomesetal.
2008). Some macroalgae also respond opportunistically to
polluted urban waters (Valielaetal. 1990, Raven and Taylor

1226
2003). For instance, transplant experiments have demon-
strated that the photosynthetic capacity of sea lettuce Ulva
lactuca increases while that of canopy-forming brown sea-
weed Sargassum stenophyllum decreases in response to urban
waters (Scherner et al. 2012). Dierential photosynthetic
responses to copper contaminants among dierent species of
Ulva may connote a competitive advantage in contaminated
urban areas (Hanetal. 2008). Similarly, the combination of
elevated sediment and nutrient loads increases the cover of
lamentous turf-forming macroalgae in eld manipulations
(Gorgula and Connell 2004) and is thought to be central
to turf proliferation in metropolitan areas (Airoldi 1998,
Connelletal. 2008, Strainetal. 2014).
Evidence for synanthropic marine consumer species is
more limited. Most of the studies on sh distribution patterns
in urban areas and relative to coastal population density sug-
gest primarily negative impacts of urbanization on major sh
groups (Toftetal. 2007, Williamsetal. 2011, Kornisetal.
2017, Munschetal. 2017, Cinner et al. 2018). Although
several well-recognized terrestrial synanthropes, including
raccoons and rats, are known to forage in intertidal habitats
(Carlton and Hodder 2003), degraded intertidal resources
in urban areas are unlikely to be a major driver of synan-
thropic distribution patterns for these species. ere is at least
one record, however, of rats occurring in higher densities on
articial breakwaters than on natural shorelines (Aguilera
2018). Heeryetal. (2018c) found that deep-dwelling giant
Pacic octopus were more common in urban than in rural
areas of Puget Sound (northeast Pacic), and suggested this
may be a function of the amount of marine debris in the
urban benthos, which octopus utilize as shelter (Katsanevakis
and Verriopoulos 2004, Katsanevakisetal. 2007). Articial
structures, such as docks and buoys, are widely used as haul
out sites by pinnipeds (Heath and Perrin 2009) and could
similarly inuence localized pinniped distribution patterns
in urban areas (DeAngelisetal. 2008). Duarteetal. (2013)
noted that oating structures associated with coastal develop-
ment could play a key role in facilitating jellysh blooms, by
expanding the available habitat for polyp recruitment. ese
lines of evidence suggest that, where synanthropic distribu-
tion patterns do exist among marine consumers, ocean sprawl
may be an important underlying mechanism (Heery et al.
2018c).
In addition to ruderal macrophytes and synathropic con-
sumers, the interacting drivers of marine urbanization appear
to facilitate the establishment of opportunistic sessile inver-
tebrates, many of which are non-indigenous. Opportunistic
responses to multiple urban drivers may provide a particular
advantage. For instance, the bryozoans, Bugula neritina and
Watersipora subtorquata, and the ascidian, Botrylloides vio-
laceus, have particularly high tolerances for copper toxicity
(Piola and Johnston 2006), which could partially explain their
successful invasion of urban marine environments beyond
their endemic range (Piolaetal. 2009, McKenzieetal. 2011,
Osborneetal. 2018). In addition, larval dispersal for these
taxa is aided by shipping activities between coastal cities, and
they readily utilize articial structures, such as oating docks,
as habitat for settlement (Lambert and Lambert 1998, 2003,
Piola and Johnston 2008, Daornetal. 2009, Piolaetal.
2009, Airoldi and Bulleri 2011, Edwards and Stachowicz
2011, Gittenberger and van der Stelt 2011, McKenzieetal.
2012, Simkaninetal. 2012, Cordelletal. 2013, Zhanetal.
2015). In this way, simultaneous positive responses to mul-
tiple urban drivers may help to facilitate invasion success
in urban areas, although the strength of these responses
likely vary between cities, taxonomic groups, and latitudes
(Canning-Clodeetal. 2011).
Acclimatization and adaptation
Urbanization is considered a major selective pressure
(Alberti 2015, Donihue and Lambert 2015) leading to phe-
notypic changes at both the organismal and species levels
(Albertietal. 2017a). ese changes are either phenotypically
plastic (i.e. within-lifetime) responses such as acclimatization,
or (population-level) adaptation via genetic change over mul-
tiple generations (Albertietal. 2017b, Johnson and Munshi-
South 2017). Recent advances in understanding evolutionary
responses to urbanization have been driven largely by work
in terrestrial systems (Parteckeet al. 2006, Miranda et al.
2013, Johnson and Munshi-South 2017). However, there is
ample precedent for rapid evolutionary change and pheno-
typic plasticity in response to anthropogenic stressors in the
marine environment (Todd 2008, Sanford and Kelly 2011).
All three of the key drivers of marine urbanization are
known to structure population genetics among a variety of
marine taxa (examples – resource exploitation: Smithet al.
1991, Hauser et al. 2002; pollution: Suchanek 1993,
López-Barea and Pueyo 1998, Naccietal. 1999, Maet al.
2000, Virgilio et al. 2003, Virgilio and Abbiati 2004,
McMillan et al. 2006, Galletly et al. 2007, Moraga and
Tanguy 2009; ocean sprawl: Street and Montagna 1996,
Fauvelotet al. 2012). In many cases, resource exploitation,
pollution and ocean sprawl lead to population bottlenecks
and reduced genetic diversity (Nevoetal. 1986, Maltagliati
2002, Fauvelotetal. 2009, Unghereseetal. 2010, Neo and
Todd 2012, Pinsky and Palumbi 2014). Yet evidence of
micro-evolution in urban marine environments has been lim-
ited. Some of the best examples come from the ecotoxicology
literature (Medinaetal. 2007). For instance, McKenzieetal.
(2011) showed heritable copper tolerance in the bryozoan
Watersipora subtorquata. Similarly, Galletly et al. (2007)
found a signicant geneotype × environment interaction in
hatching success of the ascidian, Styela plicata, under dif-
ferent copper concentrations, yet hatching success at high
concentrations had a dierent genetic basis than that at low
concentrations, suggesting dierent genetic mechanisms for
adaptation depending on pollution levels.
Trait plasticity in response to marine urbanization has
been much more widely documented. Many marine organ-
isms exhibit substantial capacity for acclimatization that
may provide a tness advantage; this could include changes
in morphology, physiology, behavior and/or life history

1227
(West-Eberhard 1989, Foo and Byrne 2016). Goiranetal.
(2017) observed melanism in sea snakes inhabiting urban
sites that they proposed facilitates the excretion of trace pol-
lutants. Phenotypically plastic responses to light in corals are
well documented and can benet colonies where sediment
pollution and associated turbidity is prevalent (Todd et al.
2003, Hoogenboometal. 2008, Ow and Todd 2010). Some
marine invertebrates also exhibit transgenerational plasticity,
wherein parents alter the phenotypes of gametes in response
to factors such as copper and salinity to maximize gamete per-
formance (Marshall 2008, Jensenetal. 2014). Several other
examples of trait plasticity from natural rocky shores may be
additionally relevant in the abiotically stressful environments
created by seawalls and other articial structures (Strainetal.
2018). For example, dog whelks Nucella lapillus and other
gastropods have larger feet in high wave energy environments
so they can adhere better to the substrate (Etter 1988, Trussell
1997), potentially an advantage on steep seawalls that inten-
sify wave shock. Similarly, local adaptation for thermal tol-
erance in acorn barnacles Semibalanus balanoides (Bertness
and Gaines 1993) and acclimatization to high temperatures
in various intertidal gastropods (Williams and Morritt 1995,
Marshalletal. 2010) may facilitate survival in novel thermal
environments associated with ocean sprawl.
Urbanization-driven trait changes can have important
eects on community interactions (Palkovacset al. 2012,
Albertietal. 2017a), yet much work remains to understand
the nature of these eects in the marine environment, as well as
their ultimate consequences for functioning in urban marine
ecosystems. is work needs to be conducted across multi-
ple organismal scales to account for potential urban-related
acclimatization at the level of holobionts – host–microbial
assemblages that function as an ecological unit (Ziegleretal.
2016, Evansetal. 2017). Further, the heritability of urban-
driven adaptation should be considered through both genetic
and epigenetic approaches, as acclimatization responses can
be inherited via transgenerational maternal eects and meth-
ylation patterns (Sunetal. 2014, Suarez-Ulloaetal. 2015).
Climate change and marine urbanization
e eects of climate change interacting with marine urban-
ization range from reasonably established to complex and
speculative possibilities. Atmospheric warming from green-
house gases leads to the thermal expansion of the oceans and
melting of glacial and polar ice, and is well-documented as the
cause of current and predicted sea-level rise (Neumannetal.
2015). Increases in the severity, and possibly occurrence,
of major storms have also been attributed to global warm-
ing (Walshetal. 2016). is combination of rising seas and
extreme weather pose direct ooding and erosion threats to
coastlines and, together with coastal development, represent
the main drivers of the current proliferation of sea defenses
(Daornetal. 2015). Elevated temperatures, altered rainfall
patterns, and other changes associated with climate change
(Duyetal. 2015, Donatetal. 2016) pose challenges for
marine organisms that inhabit coastal defense structures
(Ngetal. 2017), as well as for marine communities that pro-
vide sources of food and natural defenses for coastal cities,
such as coral reefs and mangrove forests (Wardetal. 2016,
Hoegh-Guldbergetal. 2017). Of course, coastal cities are
also part of the problem as they contribute to climate change
via high levels of greenhouse gas emissions, energy consump-
tion and changes in land use, hydrology and biodiversity
(Grimmetal. 2008a), but these additional impacts of marine
urbanization are beyond the scope of the current review.
One of the better studied interactions between urbaniza-
tion and climate change is ‘coastal squeeze’, rst reported
by Doody (2004), but later rened and dened by Pontee
(2013, p. 206) as: ‘one form of coastal habitat loss, where
intertidal habitat is lost due to the high water mark being
fixed by a defence or structure (i.e. the high water mark resid-
ing against a hard structure such as a sea wall) and the low
water mark migrating landwards in response to SLR’ (sea
level rise). Loss and/or fragmentation of tidal wetlands means
a concomitant reduction in ecosystem services, including
ood and erosion abatement, biodiversity support, water
quality, carbon sequestration and benets to coastal sher-
ies (Torio and Chmura 2013). Managed retreat (or realign-
ment), where infrastructure is relocated inland to escape the
eects of erosion and ooding (Alexandreaetal. 2012), can
alleviate coastal squeeze by moving back or removing hard
articial defences, thereby elimitaing the xed high water
mark back-stop. However, the distances required for coasal
habitats to successfully move inland can be considerable –
potentially being meters per year depending on rate of sea
level rise (Pethick 2001).
Temperature is a critical stressor on rocky shores (Helmuth
and Hofmann 2001) but little is known regarding the ther-
mal landscape of articial coast defenses (Zhaoetal. 2019).
e homogeneity of articial structures may create thermal
barrens that challenge intertidal organisms (Perkins et al.
2015) or, alternatively, provide refugia from thermally-lim-
ited predators. Helmuthetal. (2006), based on a compren-
sive study of the spatial and temporal patterns in the body
temperature of the mussel Mytilus californianus on natural
rocky shores, concluded that interacting factors such as tidal
regime and wave splash can create complex thermal mosaics
of temperature that are potentially more important locally
than those of large-scale (e.g. latitudinal) climate eects.
Hence, it will be dicult to predict or measure the broader
impacts of global warming on the intertidal area of seawalls
and similar structures. Climate associated shifts in patterns
of rainfall and runo, e.g. heavier rainfall and/or more pro-
longed rainfall (Wallaceetal. 2014), could overwhelm drain-
age systems leading to peaks in the inux of pollutants. ese
unusual pollution spikes would likely be concurrent with
increased sedimentation, eutrophication and low salinity, all
of which could moderate species and community response
and the toxicity of pollutants (Pearson and Rosenberg 1978,
Šolić and Krstulović 1992, Verslyckeetal. 2003).
Climate change is also likely to impact natural coastal
defenses. Healthy coral reefs and mangrove forests are eective

1228
at protecting coastlines from wave impact and associated ero-
sion in tropical and subtropical regions, but both are vul-
nerable to climate change. Extended periods of warmer than
average sea temperatures causes coral bleaching that, when
severe, kills colonies (Hoegh-Guldberg 1999) resulting in the
loss of wave-absorbing reef complexity (Alvarez-Filipetal.
2009, Graham and Nash 2013). As mangroves live within
a narrow band of suitable habitat determined by local tidal
regimes, they are susceptible to sea level rise if it exceeds
the rate of soil accumulation, leading to shoreline retreat
(Lovelocketal. 2015). Many tropical and subtropical towns
and cities benet from the protection that coral reefs and
mangroves provide (Ferrarioetal. 2014), and their loss can
lead directly to the installation of alternative coastal defense
measures, of which hard amour such as seawall, rip-rap and
gabion are frequently chosen. ere is also strong potential
for additive or synergistic eects as coral reefs and mangroves
near urban areas are likely to be heavily exploited as well as
impacted by pollution (Wells and Ravilious 2006). In addi-
tion to these rather more predictable consequences of climate
change, urban marine environments – as part of urban eco-
systems – are shaped by a multitude of interacting social and
ecological drivers (Albertietal. 2003) and are likely to exhibit
non-linear dynamics characteristic of complex adaptive sys-
tems (Scheeretal. 2001, Alberti 2008). e three major
drivers of marine urbanization have gradually altered urban
marine ecosystems in ways that may have reduced their capac-
ity to absorb disturbance; for instance to a 100-year storm
event, a sudden change in socio-economic variables such as a
rapid loss in food security, a major marine disease epidemic,
or various other pulse perturbations. Without considerably
more research, it is unclear how urban marine ecosystems
will respond to such disturbances, whether they are suscep-
tible to future phase shifts, and what such shifts might mean
for ecosystem functions and ecosystem services. While these
should be focal points of future research (discussed below),
approaches such as scenario planning (Petersonetal. 2003)
that integrate and accommodate uncertainties directly into
management of urban marine environments would be highly
benecial (Albertietal. 2003).
Ecological engineering
It is predicted that by the next decade approximately three
quarters of the world’s population will reside in coastal
zones (Small and Nicholls 2003, Bulleri and Chapman
2015). Coastal land is therefore in high demand and devel-
opment and reclamation are occurring at unprecedented
scales (Yeung 2001, Duarteetal. 2008, Duanetal. 2016,
Chee et al. 2017, Sengupta et al. 2018). In addition, the
risks of climate change, as outlined in the previous section,
have resulted in an urgent need for greater shoreline protec-
tion, especially in low-elevation coastal zones (LECZ) (sensu
Neumannetal. 2015). For instance, in China, Japan and
Korea alone, 28% of the global population are currently liv-
ing in LECZ and it is predicted that by 2070, 37 million
people and assets worth $13 trillion are going to be exposed
to coastal hazards such as storms, ooding and climate vari-
ability (Nichollsetal. 2013). Strategies that mitigate risk and
help coastal cities adapt to sea level rise and climate change
are already being implemented in many parts of the world
(Zimmerman and Faris 2010, Hayes et al. 2018) and are
predicted to increase in the coming decades (Neumannetal.
2015, Dangendorfetal. 2017). Such strategies, though mul-
tifaceted, include expanded coastal armoring (French and
Spencer 2001, Hinkeletal. 2014), the integration of new
stormwater capture and treatment systems, and a wide vari-
ety of other modications to increase the resilience of urban
infrastructure (Zimmerman and Faris 2010).
If the past is any indication, future proliferation of marine
urbanization will further facilitate the formation of novel
assemblages of marine organisms on an unprecedented scale.
Currently, there is considerable debate in ecology regard-
ing the concept of ‘novel ecosystems’ (Hobbsetal. 2014,
Murciaetal. 2014), i.e. ecosystems shaped by human inter-
vention that are distinct from their historical state, and that
cannot be returned to their historical trajectory (Hallettetal.
2013). It is presently unclear whether urban marine ecosys-
tems meet all criteria of ‘novel ecosystems’ (Morseetal. 2014),
but their trajectory is undeniably shaped by the way in which
coastal cities develop and modify the marine environment
(Daornetal. 2015). Given the potential of marine assem-
blages to provide ecosystem services to urban populations,
as well as recent success in the realm of eco-shoreline design
(Toftetal. 2013, Morrisetal. 2019), it may be more helpful
to consider urban marine ecosystems and their future trajec-
tory within the framework of ‘designed ecosystems’ (Higgs
2017) or ‘reconciliation ecology’ (Rosenzweig 2003a). While
both of these frameworks arose with the realization that some
systems have been so severely altered and/or degraded it is
practically impossible to apply conventional restoration prac-
tices (or expect the system to shift back towards a ‘historic’
or ‘pre-disturbed’ state), conceptually they are fundamen-
tally dierent in their intent, starting point and develop-
mental trajectory (Hunter and Gibbs 2007, Higgs 2017).
For instance, ‘designed ecosystems’ often involve large-scale
intervention eorts to create and sustain the system whereas
‘reconciliation ecology’ is less reliant on long-term interven-
tion and more based on the idea that ‘if you build it, they will
come’ (Rosenzweig 2003b, p. 6). ‘Ecological engineering’, i.e.
the design and engineering of urban infrastructure congru-
ent with ecological principles, can be viewed as straddling
between these frameworks, as it often requires huge initial
intervention but with less emphasis on subsequent man-
agement and maintenance (see recent review by Lokeet al.
2019a).
Ecological engineering is currently being trialed, or
attempted in earnest, in many locations around the world
(Chapman and Blockley 2009, Mitsch 2012, Strainet al.
2018). Nature-based or soft-engineering approaches using
‘green infrastructure’ for coastal defense are preferred over
hard engineering approaches in many coastal cities as they
have been shown to be more cost-eective in the longer term

1229
and can serve multiple functions in addition to ood risk
reduction (Temmerman et al. 2013, Spalding et al. 2014,
Regueroetal. 2018). However, these solutions are often not
adopted due to feasibility (e.g. mangrove planting at sites
with high wave energy or ow) or socio-economic reasons
(e.g. lack of political will, support or resources). In addition,
hard articial coastal defenses have frequently already been
constructed and cannot realistically be removed. Given that
more human-made shorelines are expected to be built in
the foreseeable future, it is critical to nd ways to increase
their ecological and social value while maintaining their
engineering function (Borsjeetal. 2011, Lokeetal. 2019a).
e ecological engineering of human-made shoreline struc-
tures is a new but dynamic eld, and there is often a tradeo
between taking time to understand these habitats as a system,
and the urgency or desire to implement practical solutions
(Morriset al. 2019). Knowledge of urban shoreline ecosys-
tems and of strategies that eectively enhance ecosystem
functioning and services should improve over time, as eco-
logical enhancement and blue/green infrastructure projects
become more common and are applied in a broader variety of
urban marine environments (Ponteeetal. 2016). Developing
and maintaining research collaborations with industry will be
essential to ensure that lessons from each of these projects are
shared and translated into subsequent designs and engineer-
ing solutions (Mayer-Pintoetal. 2017). Further, partnerships
with city governments and planners will be needed if eco-
logical enhancement projects are to be applied concurrently
with broader improvements in water quality and at a su-
cient scale to have long-standing benets, and then carefully
monitored over time.
Critical challenges and research directions
Awareness of the impacts of exploitation, marine pollu-
tion and ocean sprawl, as well as of eco-engineering coun-
termeasures, is growing (Chapman and Underwood 2011,
Morriset al. 2016, Lotze et al. 2018, Strainet al. 2019).
However, there remain many emerging issues, knowledge
gaps and research needs at numerous scales for understand-
ing the dynamics of urban marine ecosystems (Airoldietal.
2005, Kueer and Kaiser-Bunbury 2014) and building urban
marine ecology as a discipline. Here, we oer some critical
research questions and areas for investigation that have yet to
be fully addressed.
1. What are the interactive eects of multiple stressors in
urbanized coastal areas, including feedbacks and changes
over time?
2. Stronger characterization of spatial and temporal patterns
of biodiversity in urbanized marine environments.
3. What are the mechanisms driving marine synanthropy?
4. Key marine urban ecosystem functions, their most essen-
tial drivers, and likely future trajectories – including
implications for current and future provisioning of eco-
system services.
5. Assessment of the evidence for urban-driven trait selection
in the marine environment.
6. What ecological enhancement approaches (ecological
engineering, green- and blue-infrastructure, etc.) are most
eective in urban settings?
ere are also numerous questions related to the key ecolog-
ical processes discussed in the section ‘Key ecological patterns’
that need to be elucidated, especially disentangling co-varying
stressors and determining the long-term responses of organ-
isms and populations to marine urbanization. Ultimately, all
aspects of coastal city design: architecture, urban planning
and civil and municipal engineering, will need to prioritize
the marine environment if the negative eects of urbanization
are to be minimized. In particular, planning strategies that
account for the interactive eects of drivers and accommo-
date complex system dynamics should enhance the ecological
and human functions of future urban marine ecosystems.
Acknowledgements – We would like to thank Oikos for the
opportunity to write this Forum article. Any opinions, ndings
and conclusions or recommendations expressed in this material are
those of the authors and do not necessarily reect the views of the
U.S. National Science Foundation.
Funding – is research is supported by the National Research
Foundation, Prime Minister’s Oce, Singapore under its Marine
Science Research and Development Programme (award no.
MSRDP-05). is research was also supported with resources
from the U.S. National Science Foundation Long-term Ecological
Research (LTER) Program (grant no. DEB-1027188 to CMS).
Author contributions – PAT and ECH contributed equally to this
paper and are joint rst authors.
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