HomeMy WebLinkAboutArticle submitted by Sampou - Logan et al. 2021_ESCOVol.:(0123456789)1 3
Estuaries and Coasts
https://doi.org/10.1007/s12237-021-01006-0
A Review of Habitat Impacts from Residential Docks
and Recommended Best Management Practices with an Emphasis
on the Northeastern United States
John M. Logan1 · Alex Boeri1 · Jill Carr2 · Tay Evans3 · Eileen M. Feeney1 · Kate Frew3 · Forest Schenck3 ·
Kathryn H. Ford1
Received: 11 February 2021 / Revised: 3 September 2021 / Accepted: 9 September 2021
This is a U.S. government work and not under copyright protection in the U.S.; foreign copyright protection may apply 2021
Abstract
Small docks and floats are common in estuaries and coastal waters worldwide. These structures serve a role in coastal
recreation by facilitating access to waterways. However, they can impact shoreline ecological function. While individual
environmental impacts are generally minor, increasing dock proliferation and overlap with sensitive coastal resources can
result in cumulative impacts that pose threats at the ecosystem level. Docks promote changes in habitat and aquatic com-
munities through alteration of environmental conditions. Here, we review the potential environmental impacts of docks on
estuarine and coastal flora and fauna and discuss best management practices (BMPs) to avoid or minimize such impacts with
a focus on New England. We consider impacts in relation to the structural components of docks: the piles, decking, and floats.
Impacts to salt marsh and submerged aquatic vegetation are a particular focus given the important ecosystem services these
vegetated habitats provide and their vulnerability to dock-induced habitat alteration. Potential environmental impacts depend
on structure size, design, and location, and can include both short-term (e.g., turbidity from pile installation) and long-term
(e.g., salt marsh loss from chronic shading) effects. Such effects can be minimized through BMPs (e.g., construction outside
sensitive time-of-year periods, designs to reduce shading). As BMPs tend to reduce rather than avoid environmental effects,
cumulative impacts also need to be considered in the permitting process. We recommend that managers develop plans or
bylaws that identify sensitive habitats where dock construction should be avoided as well as BMPs to make remaining dock
proliferation less impactful.
Keywords Boardwalks · Coastal development · Cumulative impacts · Piers · Shading · Walkways
Introduction
A dock is a structure that extends into waterways to enhance
access for a variety of activities. The term dock is similar to
wharf or pier, with distinctions between the terms typically
hinging on size, materials, or whether or not vessels are
accommodated. In this paper, we focus on docks, and we use
the word “dock” to describe small, residential-scale structures,
which typically support recreational vessel berthing for small
motor or sailing craft < 40 ft (12 m) and paddle craft such as
kayaks and canoes. The major structural components of docks
are pile-supported decks connected to shore and floating struc-
tures, or “floats” that are often attached to decks or shore with
gangways (Fig. 1). An alternative to a pile-supported struc-
ture is a solid-fill structure or causeway. Habitat impacts are
higher for a solid-fill structure, which displaces considerably
more area than a pile-supported structure. Due to their habitat
Communicated by Kenneth L. Heck
* John M. Logan
john.logan@mass.gov
1 Massachusetts Division of Marine Fisheries, South Coast
Field Station, 836 South Rodney French Boulevard,
New Bedford, MA 02744, USA
2 Massachusetts Bays National Estuary Partnership, 251
Causeway Street, Suite 800, Boston, MA 02114, USA
3 Massachusetts Division of Marine Fisheries, Annisquam
River Marine Fisheries Field Station, 30 Emerson Avenue,
Gloucester, MA 01930, USA
Estuaries and Coasts
1 3
A
e
f
dc
a b
MHW
MLW
eelgrassramp/
gangway
dock
salt
marsh
N
E
S
W
salt marsh
access ramp
house
deck
gangway/ramp
MLW
MHW
edge of upland vegetation
eelgrass
f
g
e
B
Estuaries and Coasts
1 3
impacts and cost, solid-fill structures are rarely seen as new
construction projects, so this review focuses solely on pile-
supported docks.
Docks are common in estuaries worldwide (Gissy 1985;
Kennish 2002, 2016; Kelty and Bliven 2003). Coastal devel-
opment combined with a desire for water access has led to
dock proliferation along the east coast of the United States
(U.S.) in recent decades resulting in the construction of
thousands of individual structures (Chinnis and Stidham
2001; Kelty and Bliven 2003; Seabrook 2012; Logan et al.
2018a). This proliferation has led to concerns among coastal
resource managers about the cumulative environmental
impacts of these structures (Buchsbaum 2001; Bliven 2003,
2005; Kelty and Bliven 2003). Due to their location along
the shore, docks often conflict with coastal resource areas
including salt marsh, seagrass, oyster reefs, and mudflats.
Many U.S. states have guidance documents that describe
best management practices (BMPs) for docks to avoid or
minimize impacts to coastal resources (Table 1). The BMPs
(Table 1) are either required (e.g., ACOE Programmatic
General Permits) or recommended (e.g., state guidelines)
during dock design and construction. However, most avail-
able guidelines pre-date many recent studies of dock impacts
(e.g., Eriander et al. 2017; Logan et al. 2018a, b). Here,
we provide an updated summary of the existing literature
regarding the environmental impacts of docks with a particu-
lar focus on marine and estuarine resources in the northeast-
ern U.S. region of New England. Our review focuses on the
three main structural components of docks: the piles, deck-
ing, and floats. We also consider impacts of associated boat
use as well as approaches to address cumulative impacts.
We evaluate the existing information to recommend BMPs
that avoid and minimize dock impacts and limit cumulative
impacts at the scale of a waterbody.
Piles
The only structural components of docks that always result
in direct habitat impacts are the support piles. Support piles
are often made of wood due to its low cost, ease of use,
and resistance to corrosion, rust, and spalling (i.e., fragment-
ing) (Lebow et al. 2019). Cement, steel, and composites
are also options as pile materials (Bright and Smith 2002).
There are several primary impact-producing factors from
piles: habitat displacement, ecosystem interactions with
alternative materials, creation of habitat for invasive spe-
cies, shading and circulation impacts to bordering habitats
and marine resources, contamination of water and fouling
organisms from the leachates from pressure-treated and cre-
osote-treated wood, and construction-related impacts. Each
of these impacts is described below.
Habitat Creation and Alteration
Piles displace existing habitat and create new habitat, so
they represent a direct source of habitat alteration. There
are three variables to consider how piles affect habitats: the
piles themselves, the pile design (standard and bio-friendly
designs), and the pile material. How each of these variables
affects fish and invertebrate diversity and productivity is
described below.
Piles
Habitat displacement can benefit some fish species while
negatively impacting others (Grothues et al. 2016; Brandl
et al. 2017). Piles benefit some fish species by providing
structure and increased foraging habitat (e.g., Gallagher
and Heppell 2010), but the addition of hard structures in a
waterbody may also improve the habitat for invasive species
(Vaselli et al. 2008; Carman et al. 2019). In New England,
wooden piles provide habitat and refuge for a variety of
small fish species, primarily cunner (Tautogolabrus adsper-
sus), rock gunnel (Pholis gunnellus), and grubby (Myoxo-
cephalus aenaeus) (Brandl et al. 2017). Fish abundance was
higher in pile-fields and wrecks than open-water habitat in
the Hudson River estuary due to the increased prevalence
of structure-seeking species like the mummichog (Fundulus
heteroclitus), bay anchovy (Anchoa mitchilli), naked goby
(Gobiosoma bosc), American eel (Anguilla rostrata), and
northern pipefish (Syngnathus fuscus) (Duffy-Anderson et al.
2003). In Australia, large mobile species were found to move
between structures (e.g., marina piles) and open water while
smaller species were only found in close proximity to these
structures (Clynick et al. 2007). Piles can also disrupt school
formation, and the added structure may facilitate ambush
predation relative to open water environments (Grothues
et al. 2016). Juvenile winter flounder (Pseudopleuronectes
americanus) and tautog (Tautoga onitis) held in cages in
open pile fields had similar growth rates to individuals held
in open water (Able et al. 1999), but a sonar survey in the
Hudson River found fewer small fishes in open pile habitats
relative to open water during daytime (Grothues et al. 2016).
Fig. 1 Diagram of the components of a typical dock structure shown as
(A) horizontal and (B) plan view: Pile-supported decking typically con-
nects the bordering upland area to the adjacent shoreline. A ramp or gang-
way extends from the seaward end of the decking to a float located in the
intertidal or subtidal zone. To minimize impacts to aquatic resources, (a)
decking over salt marsh should have a height ≥ 1.5 decking width, (b)
pile spacing should be maximized, (c) pile width minimized, (d) depth
under float at mean low water (MLW) should be 2.5 ft (0.8 m) in shellfish
habitat, (e) float surface area should be minimized, (f) the seaward edge
of the float should be located at least 25 ft (7.6 m) from any bordering
submerged aquatic vegetation (SAV), and (g) decking width minimized.
Docks should be set at a north–south orientation when feasible to
minimize shading to underlying vegetation
◂
Estuaries and Coasts
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Table.1 Summary of best management practice guidelines and general permit conditions for dock design in coastal states in the continental U.S
Maximum dock
width
Minimum dock
height
Decking Minimum float
depth
Materials Orientation
Maine
(Swan and Sowles
2008)
NA 1:1 H:Wa,b 0.75 in (1.9 cm)
spacinga,b
8-12 in (~20-30 cm)
float stopsc
No creosote NA
New Hampshire
(NHDES 2020)
6 ft
(1.8 m)
1:1 H:W 0.75 in (1.9 cm)
spacing
2 ft (0.6 m)Non-toxic N-S
Massachusetts
(Bliven and
Pearlman 2003)
NA 1:1 H:Wk
4 ft (1.2 m)b
0.75 in (1.9 cm)
spacing or
alternative
deckingb,k
1.5 ft (0.5 m)x
2.5 ft (0.8 m)x,y
4 ft (1.2 m)b
No creosote N-Sb,k
Rhode Island
(RICRMP 2007)
4 ft
(1.2 m)
4 ft (1.2 m) or 1:1
H:Wa
NA 1.5 ft (0.5 m)x
No floatb
No creosote NA
Connecticut
(CTDEEP 2015)
4 ft
(1.2 m)a,b
5 ft (1.5 m)n
> 1:1 H:Wa
NA No floatb NA NA
New York
(NYDEC 2013)
4 ft
(1.2 m)
4 ft (1.2 m)a
5 ft (1.5 m)b,o
open grateda 2.5 ft (0.8 m)z No creosote N-S
New Jersey (NJDEP
2020)
4 ft (1.2 m)b
6 ft (1.8 m)a,c
8 ft (2.4 m)d
4 ft (1.2 m)0.38 in (1.0 cm),
0.5 in (1.3 cm),
0.75 in (1.9 cm),
1 in (2.5 cm) or
alternative decking
4 ft (1.2 m)Non-polluting
materialsy
NA
Delaware
(ACOE 2020)
3 ft (0.9 m)a,b,c
4 ft (1.2 m)e,f
5 ft (1.5 m) g
6 ft (1.8 m)h,i
8 ft (2.4 m)j
3 ft (0.9 m)a,b,c
4 ft (1.2 m)e
0.38 in (1.0 cm), 0.5
in (1.3 cm), 0.75 in
(1.9 cm), 1 in (2.5
cm) or alternative
decking
Avoid grounding NA NA
Maryland (ACOE
2011a)
3 ft (0.9 m)a
6 ft (1.8 m)d
3 ft (0.9 m)a,q
4 ft (1.2 m)r
NA NA NA NA
VMRC (1999,
2005)
5 ft (1.5 m)a 3 ft (0.9 m) or
width + 1 ft
(0.3 m)a
NA NA NA NA
North Carolina
(NCDEQ 2021)
6 ft (1.8 m)3 ft (0.9 m)a NA NA NA NA
South Carolina
(SCDHEC 2008)
4 ft
(1.2 m)
3 ft (0.9 m)o NA NA NA NA
Georgia (ACOE
2017)
6 ft (1.8 m)6 ft (1.8 m)s NA 2 ft (0.6 m)NA NA
Florida (Florida
DEP 2020, ACOE
and NMFS 2001)
4 ft
(1.2 m)b,k,l
4 ft (1.2 m)k
5 ft (1.5 m)b,o
0.5 in (1.3 cm)
spacingb
1 ft (0.3 m)x NA N-Sb
Alabama (ACOE
2011b)
5 ft (1.5 m)b,m 1:1 H:Wb,m,o 0.75 in (1.9 cm)
spacing
NA NA NA
Mississippi (ACOE
2018)
6 ft (1.8 m)b,m 1:1 H:Wb,m,o NA NA NA NA
Louisiana (K.
Morgan, LA CZM
Pers. Comm.)
NA NA 1–2 in (2.5–5.0 cm)
spacingk
NA NA NA
Texas (ACOE 2021)4 ft (1.2 m)5 ft (1.5 m)t 1 in (2.5 cm)u NA NA NA
Washington
(WADOE 2011)
NA NA Grating over at least
50% of surface
areav
4–5 ft (1.2–1.5 m)z Avoid wood
treated with toxic
compoundsA
N-S
Oregon
(ODFW 2016)
6 ft
(1.8 m)
NA NA NA Avoid pressure
treated (PT) wood
NA
Estuaries and Coasts
1 3
The former study suggests that piles can provide suitable
habitat and food source for many fish species, but the latter
study suggests piles may not benefit overall fish production.
Piles displace existing habitat for infaunal invertebrates
(e.g., clams, polychaete worms) but can increase habitat for
a variety of epifaunal invertebrates (e.g., mussels, oysters,
barnacles) by providing surface area for settlement and
attachment. Piles in a Florida lagoon supported a variety
of filter-feeding invertebrates that collectively provided
approximately 30% of the filtration capacity of oyster reefs
in the system (Layman et al. 2014). Fish species diver-
sity and abundance around piles in marinas in Australia
were correlated with the percent cover of pile epibiota in
experimental manipulations (Clynick et al. 2007), and piles
that promoted colonization by mussels, foliose algae, and
other epibiota enhanced fish habitat in some cases. Red king
crabs (Paralithodes camtschaticus) were more abundant on
experimental pile structures in Alaska than adjacent seafloor,
possibly due to habitat provided by colonizing hydroids
(Stevens et al. 2004). Similarly, docks in the southeast U.S.
provided forage from attached fouling organisms growing
on piles for mangrove tree crabs (Aratus pisonii) (Cannizzo
et al. 2018). In Australia, subtidal epibiota diversity and
abundance increased on piles relative to open water, and
unshaded piles had similar communities as natural rocky
habitat (Connell and Glasby 1999).
Table.1 (continued)
a If over wetlands
b If over submerged aquatic vegetation (SAV)
c If over intertidal flats
d If over water
e If municipal or residential community structure over wetlands, SAV, or intertidal flats
f General requirement for piers
g General requirement for docks
h General requirement for docks with boatlifts
i General requirement for community, municipal, or commercial pier structure
j General requirement for community, municipal, or commercial dock structure
k If over salt marsh
l If over mangroves
m If over non-forested wetlands
n Height above substrate at mean high water (MHW) line for public access if stairs or other public access not provided
o Height above MHW
p If walkway over scrub-shrub wetlands
q Height above mean low water (MLW) if decking < 5 ft (1.5 m) width
r Height above MLW if 6 ft (1.8 m) width over open water to protect SAV
s Height above MHW if constructed over a tributary that can be bridged
t Height above MHW or OHW (ordinary high water) measured from top surface of decking if over special aquatic sites
u If over special aquatic sites
v If over nearshore or littoral area
w Measured at low tide in coastal wetland resource areas and ordinary high water or annual high water for inland wetland resource areas
x Depth at MLW
y If over shellfish habitat
z Depth at mean lower low water (MLLW)
A Applies to structures on or over state-owned aquatic lands
B Creosote not recommended in new structures but may be acceptable if adding small number of piles to existing structure
C If PT wood used, ammoniacal copper zinc arsenate (ACZA) recommended if no human or marine mammal contact anticipated, alkaline copper
quaternary (ACQ) if such contact anticipated
Maximum dock
width
Minimum dock
height
Decking Minimum float
depth
Materials Orientation
California
(Metz 2019)
NA NA NA NA No creosoteB; Avoid
PT or use wrap
(piles); if PT used,
ACZA or ACQ
recommendedC
NA
Estuaries and Coasts
1 3
Pile Design and Materials
Incorporation of creative designs and materials in pile instal-
lations can enhance habitat value (Dyson and Yocom 2015;
O’Shaughnessy et al. 2020), but may also have unintended
effects (e.g., spread of invasive species, trophic alterations)
(Dafforn 2017; Malerba et al. 2019). For example, installa-
tion of more textured piles could promote colonization of
epibiota resulting in increased water filtration (Layman et al.
2014) as well as invertebrate and fish diversity and abun-
dance (Clynick et al. 2007; Perkol-Finkel and Sella 2015)
depending on the colonizing species. Given documented use
of piles and other hard structures by invasive species (e.g.,
Cordell et al. 2013), monitoring of invasive species colo-
nization rates and longevity should accompany installation
of these creative designs to better understand the biologi-
cal responses to designs intended to promote colonization
(Dyson and Yocom 2015). Some eco-engineering designs
have demonstrated reductions or prevention of biofouling
and invasive colonization (Paalvast et al. 2012; Perkol-Finkel
and Sella 2015). Certain installation practices (e.g., control-
ling timing of installation, pre-seeding piles with appropriate
native filter-feeders, grazers or predators) show promise in
better controlling the communities that ultimately colonize
piles (Dafforn 2017). Pile material affected habitat value
for epibiota in studies in the U.S. and Europe (Layman et al.
2014; Paalvast et al. 2012; Perkol-Finkel and Sella 2015). In
a Florida lagoon, filter-feeding communities varied greatly
across pile types. Communities on concrete piles contributed
68% of total filtration capacity even though concrete piles
were found on only 7% of sampled docks. Communities on
wooden piles with pile wrap supported 10% of filtration even
though wooden piles were most common, associated with
69% of sampled docks (Layman et al. 2014). Piles in New
York outfitted with rough textured concrete designed to pro-
mote invertebrate colonization had greater species diversity
than fiberglass piles (Perkol-Finkel and Sella 2015). In both
studies, filter feeders settled more on the rougher texture of
the concrete piles. In the Netherlands, piles outfitted with
synthetic, free-hanging ropes increased colonization by
native species (blue mussels (Mytilus edulis) below mean low
water (MLW), seaweed above MLW line) but not invasives
(Pacific oyster (Crassostrea gigas)) relative to standard piles
(Paalvast et al. 2012). In Florida, dock piles were wrapped
with oyster mats and had oyster bags suspended adjacent to
piles in a pilot study examining potential means of creating a
“living dock” (Weaver et al. 2018). Both dock augmentations
allowed successful colonization of oysters and a variety of
invertebrates (e.g., barnacles, sponges, mussels).
For fishes, it is unclear if pile material influences commu-
nity composition. Pile material (concrete, wood, or polyvinyl
chloride (PVC)) did not affect fish community composition
around docks in Belize and Panama. However, sample sizes
for each material were low (n ≤ 6 for each type) and may
have been inadequate to detect differences (Brandl et al.
2017). Fish communities should be affected indirectly since
pile material affects epibionts (Layman et al. 2014; Perkol-
Finkel and Sella 2015), and the fish community will likely
respond to changes in the epibiota.
Invasive Species
Piles can also serve as habitat for invasive invertebrate spe-
cies (Ruiz et al. 2009). For example, new piles represent
unoccupied hard substrate, and such open spaces promote
the colonization of invasive marine sessile invertebrates
(Stachowicz et al. 1999). When piles are installed in areas
with little or no natural hard substrate, invasive species asso-
ciated with hard substrate may be provided with previously
unavailable habitat. In North America, the greatest numbers
of invasive species in coastal waters were observed in docks
and marinas among all surveyed habitat types followed by
rocky reefs (Ruiz et al. 2009). In Australia, piles supported
approximately twice as many invasive species as rocky
reefs (Glasby et al. 2007) and 10–80% greater percent cover
(Dafforn et al. 2012). While colonization of piles and floats
by invasive species is well documented globally (Lambert
and Lambert 1998, 2003; Lambert 2002; Ruiz et al. 2009;
Simkanin et al. 2012), conditions that promote or limit the
spread of invasives to natural habitats are not well under-
stood (Simkanin et al. 2012). If natural hard bottom habitat
is proximal to piles, these structures may serve as a gateway
for the spread of invasives to neighboring habitats (Simkanin
et al. 2012). However, an experimental predator manipula-
tion study found that the high abundance of predators in
natural hard bottom habitat can inhibit colonization of those
natural habitats and piles by invasive invertebrates (Dumont
et al. 2011; Kimbro et al. 2013). Given that invasive colo-
nization of piles and other artificial hard substrates is well
documented, limiting the proliferation of piles reduces the
potential for the spread of invasive species (Wasson et al.
2005).
Shading and Circulation
Piles can alter the surrounding environment through
shading and alteration of water circulation. Piles cast
shade and reduce the availability of light to the water
column and benthic habitat. Piles can trap floating
debris, which further affects light and habitat (Night-
engale and Simenstad 2001). Structures placed in mov-
ing water can alter circulation (Ramos et al. 2016),
which, depending on the conditions, can either cause
scour and erosion or lead to increased deposition of
sediments (Pentilla and Doty 1990; Kelty and Bliven
2003). Salt marsh vegetation biomass decreases under
Estuaries and Coasts
1 3
dock decking as pile spacing decreases (Logan et al.
2018a), possibly due to increased shading from piles
and associated support structures (e.g., cross bracing,
horizontal stringers) and/or shading by wrack trapped
by piles. Salt marsh losses under a dock system that
used a powered cart atop thin rails rather than tradi-
tional decking may have also been due to indirect pile
impacts as this system required closer pile spacing (10
ft; ~ 3 m) relative to other construction methods with
greater spacing of 12–20 ft (~ 3.7–6.1 m) (Alexander
2012).
Piles can trap materials that smother underlying salt
marsh or submerged aquatic vegetation (SAV). While
wrack naturally accumulates on marsh surfaces, piles
magnify burial effects on marsh vegetation by preventing
tidal action from moving such material and instead caus-
ing it to remain in place for prolonged periods. Pilings
alter marsh wrack distribution in marshes, trapping wrack
and keeping it in place for longer periods of time (Alex-
ander 2008). Wrack burial impacts are greatest for large
mats that remain in place for longer time intervals and
at higher marsh elevations (Valiela and Rietsma 1995).
Spartina patens, the dominant high marsh vegetation in
U.S. east coast salt marshes, is more susceptible to mor-
tality from burial (due to less frequent inundation and
flushing) than the low marsh dominant, S. alterniflora,
but both species die off when subjected to wrack burial
for greater than two months (Bertness and Ellison 1987).
In the absence of vegetation, these wrack-covered areas
can erode and cease to be suitable marsh habitat due to
lowered elevation and increased tidal inundation. Simi-
larly, if scouring occurs in SAV the vegetation community
can be eliminated or altered (Beal and Schmit 2000). Like
areas of marsh loss, denuded subtidal areas around piles
can fill in with shell debris and detritus, making reveg-
etation unlikely due to habitat alteration and smothering
(Beal and Schmit 2000).
Toxic Leachates
Wood piles are commonly treated with anti-microbial
chemicals to lessen decay from insects and rotting and con-
sequently increase pile lifespan (Kelty and Bliven 2003;
Gorenier and Lebow 2006). Anti-microbial treatments
include metal additives absorbed into the wood under pres-
sure (pressure-treated wood) and creosote coatings. While
rarely used in contemporary dock construction, creosote
coatings are still used in commercial applications and a
legacy pollution effect remains (Younie 2015). Both treat-
ment types are known to leach contaminants into the water
and surrounding sediments (Malins et al. 1985; Weis and
Weis 1996). These effects are described in turn.
Pressure‑Treated Wood
The main chemical additives in pressure-treated wood are
chromated copper arsenate (CCA) and various copper-based
alternatives that do not contain arsenic but have higher
concentrations of copper than CCA (e.g., alkaline copper
quaternary (ACQ) and copper azole (CA)) (Lebow et al.
2019). CCA is no longer produced for use in residential set-
tings (i.e., inside a house) due to the toxicity of arsenic to
humans (Gerstein and Zaccaria 2004), but it can be used
for dock construction. Copper is toxic to marine organ-
isms at very low concentrations (Ansari et al. 2004), so it
is effective in anti-fouling compounds but has the potential
to cause adverse environmental impacts (Weis and Weis
1992a, 1996). Pressure-treated wood is known to leach the
metal additives over time in proportion to the original con-
centrations of metals (Hingston et al. 2001); therefore, the
copper-based alternatives to CCA-treated wood leach higher
concentrations of copper (Stook et al. 2005; Temiz et al.
2006; Dubey et al. 2007) than CCA. Metal leachate con-
centrations are also influenced by environmental conditions,
and generally increase with lower pH, higher salinities, and
higher temperatures (Hingston et al. 2001; Moghaddam and
Mulligan 2008). Metal concentrations are higher in recently
(≤ 3 years) constructed structures (Weis and Weis 2002),
and leaching is most rapid in the first month following pile
installation due to an initial release of a pulse of unfixed
or poorly fixed preservatives followed by an exponential
decline towards a lower leaching rate (Breslin and Adler-
Ivanbrook 1998; Lebow et al. 2004). Leached metals are
transferred to sediment, organisms, and waters surround-
ing the wood piles. Metal concentrations are higher in fine-
grained sediments (i.e., silts and clays) and poorly flushed
systems with reduced tidal exchange (Weis and Weis 1992a;
Weis et al. 1993a, 1993b, 1998). For example, copper con-
centrations of oysters (Crassostrea virginica) growing on
pressure-treated bulkheads in a canal with minimal flush-
ing were twelve times higher than background levels while
conspecifics on similar structures in open water were twice
background levels (Weis and Weis 1992b). Typically, trans-
fer of metals to surrounding sediments and sessile organ-
isms is spatially constrained to areas within approximately
33 ft (10 m) of structures containing pressure-treated wood
(Wendt et al. 1996; Weis et al. 1998; Weis and Weis 2002;
Vasilas et al. 2011a), but the spatial range of impact can be
wider under older structures (Weis and Weis 2002; Vasilas
et al. 2011a), possibly due to metal transport after release
through mineralization (Vasilas et al. 2011a).
Negative effects on aquatic biota subjected to metal lea-
chates from pressure-treated wood in laboratory settings
range from reduced fitness to mortality (Weis and Weis
2004). CCA-treated wood leachates slowed limb regen-
eration in fiddler crabs (Uca pugilator) (Weis et al. 1991,
Estuaries and Coasts
1 3
1992) and induced mortality at higher concentrations in
fiddler crabs and mummichog embryos (Weis et al. 1991).
CCA leachates also indirectly induced mortality in mud
snails (Ilyanassa obsoleta) consuming green algae exposed
to leachates (Weis et al. 1991; Weis and Weis 1992b) and
reduced sea urchin (Arbacia punctulata) fertilization (Weis
et al. 1992b). ACQ-treated wood was more toxic than CCA
as well as all other tested materials (e.g., untreated wood,
steel, concrete) for zooplankton (Daphnia magna) and fishes
(rainbow trout (Oncorhynchus mykiss) and threespine stick-
leback (Gasterosteus aculeatus)). Untreated wood was also
toxic to brine shrimp (Artemia franciscana) and oyster (C.
gigas) embryos in laboratory exposure (Libralato et al.
2007). Amphipods (Ampelisca abdita) exposed to wood pile
leachates had reduced survival with untreated but not CCA-
treated wood, possibly due to toxic effects of naturally leach-
ing compounds (e.g., phenols) that might be extracted or
chemically altered during the preservation process (Baldwin
et al. 1996). As noted by Weis et al. (1991), these labora-
tory studies do not mimic the full range of ambient condi-
tions (e.g., dilution from tidal flushing, avoidance for mobile
fauna) and do not necessarily represent impacts to fauna in
field conditions. For example, the proximity of CCA-treated
docks in estuaries with moderate (~ 5–6.5 ft; 1.5–2.0 m) tidal
ranges did not affect oyster growth, survival rates, and metal
concentrations over a 6-week exposure period nor the short-
term (< 1 week) survival rates of mud snails, mummichogs,
juvenile red drum (Sciaenops ocellatus), or juvenile white
shrimp (Penaeus setiferus) (Wendt et al. 1996).
The broader ecological effects of pressure-treated wood
structures are less clear. Epibiota colonizing CCA-treated
wood in New York estuaries had lower species diver-
sity, biomass, and abundance relative to communities on
untreated piles (Weis and Weis 1992a), and epibiota diver-
sity decreased with increasing CCA concentrations in the
Mediterranean Sea (Karayanni et al. 2010). For experimen-
tal wood panels submerged at different European coastal
sites, epibiota diversity (Brown et al. 2003) and abundance
(Brown and Eaton 2001) were greater on CCA-treated wood
compared to untreated wood, possibly due to modifications
to the wood surface and positive responses of dominant taxa,
respectively. Metals could be transported greater distances
away from docks via trophic transfer as organisms acquire
elevated levels of metals indirectly through the consumption
of contaminated prey (Weis and Weis 2004). However, this
uptake by local producers and impacts on associated food
webs are not well understood (Sanders 2008).
Creosote‑Treated Wood
Creosote is another chemical that is added to wood
under pressure to lessen decay from insects and rotting
(Brooks 1995). Creosote-treated piles release polycyclic
aromatic hydrocarbons (PAHs) into the surrounding envi-
ronment resulting in negative impacts to nearby aquatic
organisms (Malins et al. 1985; Duncan et al. 2017; West
et al. 2019). Though creosote-treated piles have been
banned in many regions, their use still persists in some loca-
tions (e.g., Metro Vancouver) primarily due to their cost-
effectiveness and longevity (Younie 2015). Therefore, BMP
guidelines put forth by U.S. state agencies frequently call for
avoidance of creosote-treated piles or note the unlawfulness
of their use (e.g., Swan and Sowles 2008; Table 1). Creo-
sote-treated piles installed prior to being banned remain
in place in some harbors. For example, 30,000 abandoned
creosote-treated piles were recently mapped across San
Francisco Bay (Werme et al. 2010). In Washington state,
16,000 creosote-treated piles remained after the removal
of more than 25,000 (Robertson 2018).
The release of PAHs from creosote-treated piles var-
ies with environmental conditions (Brooks 1995; Stratus
Consulting Inc. 2006; Perkins 2009). Elevated concentra-
tions of PAHs are mostly localized in close proximity to
the piles (Duncan et al. 2017) and are highest in environ-
ments with poor circulation and low sediment oxygen
concentrations due to limited dilution and reduced rates
of microbial decomposition, respectively (Brooks 1995;
Stratus Consulting Inc. 2006). The release of PAHs from
creosote-treated piles can continue long after installation.
For example, herring (Clupea pallasii) larvae collected
near creosote-treated piles in place for > 100 years had
elevated levels of PAHs (West et al. 2019). Thus, similar
environmental conditions (e.g., low flow, fine sediments)
that elevate heavy metal contamination from pressure-
treated wood piles also increase PAH contamination
from creosote-treated piles, but the release of leachates
of creosote-treated piles continues over much longer time
spans (decades to centuries).
Impacts of creosote on marine organisms have been
documented in both laboratory (e.g., Vines et al. 2000)
and field (e.g., West et al. 2019) settings (Cherr et al.
2017). More vulnerable taxa include benthic fishes and
invertebrates, and egg and larval stages of fishes due
to heightened exposure to sediment-bound PAHs and
greater sensitivity to low PAH concentrations, respec-
tively (Malins et al. 1985; Vines et al. 2000; Duncan et al.
2017). Creosote exposure can result in both diminished
health in adults as well as mortality in larvae. For example,
English sole (Parophrys vetulus), a benthic flatfish, had a
high prevalence of hepatic lesions in habitat with creosote-
contaminated sediment (Malins et al. 1985). In laboratory
exposure, Pacific herring embryos that attached directly to
creosote-treated wood failed to further develop a few days
post-incubation (Vines et al. 2000).
Estuaries and Coasts
1 3
Alternatives to Pressure‑ and Creosote‑Treated Wood:
Challenges and Trade‑offs
Many U.S. states recommend avoiding creosote and
pressure-treated wood (e.g., WADOE 2011, ODFW 2016;
Table 1), but ultimately all pile materials have trade-offs
in terms of cost, efficacy, availability, and environmental
impacts that will need to be weighed for individual projects
(Table 2). These trade-offs among materials often make
identifying the least environmentally damaging construction
materials complex (May et al. 2017). For example, external
wraps that reduce chemical leaching can also reduce pile
biological function by limiting colonization by filter feed-
ers (Layman et al. 2014), and the risk of greater leachate
release from pressure-treated piles without wrap may be out-
weighed by the reward of ecosystem services in an embay-
ment with high flushing (i.e., lower sensitivity to leaching
impacts) and populations of native filter-feeders (e.g., oys-
ters). Using untreated local hardwoods from the U.S. (e.g.,
black locust (Robinia pseudoacacia), cedar (Thuja spp.),
white oak (Quercus alba)) would alleviate metal leaching
effects but such materials may be less readily available and
have a shorter lifespan than pressure treated wood materi-
als (Gorenier and Lebow 2006; Lebow et al. 2019). Even
pressure-treated wood will ultimately still be susceptible to
degradation by fungi and marine borers (Lopez-Anido et al.
2004). Amazonian hardwoods (e.g., Greenheart (Chloro-
cardium rodiei)) have high durability (Crossman and Simm
2004) and will also alleviate metal leaching effects but may
compound broader environmental effects (e.g., deforestation
and unsustainable harvest practices) (Treu et al. 2019).
Life cycle assessments (LCAs) offer a broader, more
holistic evaluation in material selection. An LCA estimated
that wood-plastic composite decking resulted in fourteen
times greater greenhouse gas (GHG) emissions relative to
ACQ-treated lumber (Bolin and Smith 2011) while plas-
tic pile GHG emissions were about 2.5 times greater than
CCA-treated piles (Bolin and Smith 2012) (Table 2; May
et al. 2017). However, these life cycle comparisons were
sensitive to assumptions of relative virgin vs. recycled plas-
tic content and not necessarily representative of all avail-
able composite products (Platt et al. 2005; Bolin and Smith
2011). Estimates of GHG emissions from wood-plastic
composites decreased by 28% when virgin plastic was
substituted with 100% recycled plastic (Fuchigami et al.
2020). Steel and concrete GHG emissions also exceeded
those for the life cycle of CCA-treated piles (Bolin and
Smith 2012). The production of Portland cement, an
essential component of concrete, is highly energy inten-
sive and consequently has a large carbon footprint (Meyer
2009; Cooke et al. 2020). Therefore, each project should
consider the surrounding environment and determine an
appropriate approach on a case-by-case basis within a sys-
tem level analysis. An online model developed by Oregon
State University and the Western Wood Preservers Institute
(WWPI) can be used as a tool to estimate the potential
impact of different construction designs on local water col-
umn and sediment contaminant concentrations (Western
Wood Preservers Institute 2018a). When pressure-treated
wood is used, careful selection of source material that has
been produced using BMPs to reduce leaching will aid in
minimizing environmental impacts (Lebow et al. 2019).
Table.2 Summary of benefits and negative aspects of different potential construction materials for dock decking and pile supports
Material Pros Cons
CCA •Commonly used
•Can use local wood supply
•Long lifespan
•Low lifecycle greenhouse gas (GHG) emissions
•Release of arsenic
•Release of chromium
•Release of copper
ACQ •Commonly used
•Can use local wood supply
•Long lifespan
•Low lifecycle GHG emissions
•Release of higher copper concentrations than CCA
Creosote •Long lifespan
•Can use local wood supply
•Release of PAHs
Local (U.S.) hardwood
(e.g., black locust, cedar, white oak)
•No leaching of metals or PAHs •Less commonly available
•Shorter lived than pressure-treated wood
Amazonian hardwood •No leaching of metals or PAHs •GHG emissions
•Sustainability concerns for many sources
Chemically-modified wood •No leaching of metals or PAHs
•High durability
•Less commonly available
Concrete •No leaching of metals or PAHs
•Can promote colonization by filter-feeders
•May promote spread of invasive species
•High lifecycle GHG emissions
Composite •No leaching of metals or PAHs
•Widely available as decking material
•High lifecycle GHG emissions
Estuaries and Coasts
1 3
Several piling designs and materials have future potential
as replacements of conventional pressure-treated wood, but
currently require further development. Chemical modifica-
tion of wood involves covalently bonding a chemical group
to a wood component, resulting in a stable bond between the
reagent and wood cell wall components (Mantanis 2017).
Chemically modified wood, which has been shown to have
greater durability than CCA-treated wood in marine trials
(Westin et al. 2016), is a potential future alternative that
avoids the release of toxic leachates (Mantanis 2017). Fiber
reinforced composites (FRPs), which often use high percent-
ages of recycled plastic, are also highly durable and corrosion
resistant, but further assessment of long-term resilience is
needed (Guades et al. 2010; Zyka and Mohajerani 2016) and
the potential for plastic contamination needs to be studied.
Finally, eco-engineering approaches show promise in enhanc-
ing the ecological function of support piles (e.g., Paalvast
et al. 2012; Perkol-Finkel and Sella 2015; O’Shaughnessy
et al. 2020), but still need to be further evaluated in terms of
the potential spread of invasive species (Vaselli et al. 2008;
Carman et al. 2019).
Construction Impacts
Pile installation via pile driving, jetting, or blasting can
temporarily impact fishes and invertebrates through direct
mortality as well as the introduction of turbidity and noise
(Mulvihill et al. 1980; Iafrate et al. 2016). These latter
impacts may adversely affect fish and shellfish eggs, lar-
vae, and adults. Eggs and larvae of estuarine fishes exhibit
some of the most sensitive responses to elevated turbid-
ity. Impacts from suspended sediments include egg burial,
hatching delay, reduced feeding success, behavioral changes,
and mortality (Wilber and Clarke 2001; Berry et al. 2011).
For shellfish, pumping rates and growth can be impacted
(Loosanoff 1962; Wilber and Clarke 2001). In New Eng-
land, winter flounder spawn in coastal embayments during
winter months (Pereira et al. 1999), and their demersal eggs
are also vulnerable to physical damage from pile installa-
tion and associated turbidity (Berry et al. 2011). Installation
during the spawning season could negatively impact species
like alewife (Alosa pseudoharengus) that spawn along pond
shorelines (as reviewed in Mather et al. 2012) due to direct
impacts to spawning substrates. Pile-driving in areas of con-
taminated sediment may negatively impact nearby fishes that
are attracted to the increased suspension of benthic organ-
isms as the water column and substrate surfaces are dis-
turbed (Nightengale and Simenstad 2001). Noise from pile
driving can trigger avoidance behavior in some fish species
(Hawkins et al. 2014; Iafrate et al. 2016) and limit their use
of habitat near the dock site during construction, potentially
interfering with migratory pathways.
Noise and turbidity impacts can be reduced with the use
of modified timing and installation procedures. Construc-
tion-related impacts can be minimized through time-of-year
restrictions that avoid sensitive life history stages of species
present at a dock installation site that are more vulnerable to
turbidity or noise-related impacts of pile installation (Evans
et al. 2015). Noise levels vary among piling installation
methods with vibratory hammers producing lower noise
vibrations than impact hammers (Denes et al. 2016). For
any pile driving activity, potential impacts to mobile fauna
can be reduced with a “soft-start” approach, which initi-
ates the construction process with lower noise level activity
(Robinson et al. 2007). This approach can potentially allow
mobile organisms to leave the area before greater noise pro-
duction occurs. Turbidity associated with pile-driving can be
reduced by avoiding jetting (Gabr et al. 2004) and contained
through the use of silt curtains.
In addition to direct mortality, turbidity, and noise, jet-
ting or blasting methods for pile installation can cause
scouring in the benthic habitat surrounding installed piles
that can persist for years post-construction (Shafer and
Robinson 2001) and can impact SAV (Beal and Schmit
2000). In Florida, pile installation using jetting produced
a halo with a diameter up to 1.6 ft (0.5 m) around the piles,
and the resulting bare area was found to deepen over time
due to scouring (Beal and Schmit 2000). Bare areas around
piles ranged from approximately 3.0–6.5 ft (0.9–2.0 m) in
diameter in St. Andrew Bay, Florida, even though the age
of the piles varied widely and thus some areas should have
presumably revegetated. In many cases, bare areas from
adjacent piles were observed to overlap and coalesce into
continuous expanses of bare sediments causing fragmenta-
tion of the SAV meadow (Shafer and Lundin 1999; Shafer
and Robinson 2001).
Installation of new as well as removal of existing piles
can also have environmental impacts depending on the pile
material. Complete removal of piles is generally preferable
in restoring aquatic habitat relative to pile cutting. For creo-
sote piles, cutting can release additional PAHs into the envi-
ronment possibly by exposing previously unweathered wood
material (West et al. 2019). To minimize additional PAH
release, the Washington Department of Natural Resources
(WADNR) recommends full removal as the preferred option
and collection of any splintered wood, debris, and fragments
(WADNR 2017). Similar debris collection and proper dis-
posal practices are recommended in the removal of pres-
sure-treated wood piles as well as during installation if any
such debris is generated (Western Wood Preservers Institute
2018b). When cutting is the only option during pile removal,
the WADNR recommends cutting occur at least 2 ft (0.6 m)
below the mudline in intertidal and shallow (< − 10 ft (3 m)
mean lower low water (MLLW) subtidal waters and at least
Estuaries and Coasts
1 3
1 ft (0.3 m) below the mudline in deeper water environments
for creosote piles (WADNR 2017; Abercrombie 2018). The
Oregon State Marine Board (OSMB) recommends a mini-
mum 1 ft (0.3 m) below mudline cutoff depth for all pile
materials and marine habitats. For creosote piles, the OSMB
recommends the addition of clean fill to cover the remaining
stump (OSMB 2012).
Regardless of timing, all pile construction activities
can impact habitat beyond the footprint of the structures.
Operation of the work vessel in shallow water can result in
grounding or propeller scour, which could damage benthic
resources including shellfish and SAV, and also cause turbid-
ity (Sagerman et al. 2020). Anchored barges also limit light
availability. These vessel impacts can occur in areas adjacent
to the dock footprint as well as nearby areas encountered in
transit to and from the work site. This is a particular concern
for SAV and other shallow water resources that could be
present in the general project vicinity.
Recommendations
• Avoid sensitive habitat when siting piles: When feasible,
pile locations should be adjusted to avoid or minimize
direct impacts to the most sensitive and ecologically
valuable areas.
• Minimize area of direct impact: The total pile footprint
should be reduced to the maximum extent practicable
within the constraints of the structure’s engineering
requirements through the use of monopiles or by maxi-
mizing paired pile spacing and/or reducing pile diameter.
• Choose appropriate pile materials: Most appropriate pile
materials will vary as a function of project site habitat
characteristics and biological communities. Among
pressure-treated wood alternatives, materials with lower
concentrations of copper, such as CCA, are recommended
over those with higher concentrations of copper, such as
ACQ. Creosote treatment of untreated wood is not recom-
mended under any conditions.
• When using pressure-treated piles, the amount of leach-
ing can be reduced with project-specific BMPs relating to
both product selection and installation including purchas-
ing BMP-certified wood, minimizing on-site cutting or
including proper containment, and adding external pile
wrap.
• Guidance on when to use pressure-treated wood alter-
natives: Traditional, pressure-treated wood appears to
provide a relatively low impact use for applications in
locations with greater flushing and coarser, more sandy
sediment, but consider wraps or non-leaching materials
in poorly flushed systems.
• Consider cumulative effects: Highly developed embay-
ments with low flushing represent areas of particular con-
cern as they pose the potential to compound localized
impacts.
• Adhere to time-of-year restrictions in the project
sequencing: Construction-related impacts, such as tur-
bidity or noise, can be minimized through careful coor-
dination of timing to avoid sensitive life history stages of
species present at a dock installation site.
• Use construction methods that minimize habitat distur-
bance: Vibratory hammer use is recommended with a
slow-start approach for pile driving while jetting should
be avoided. Silt curtains or other containment methods
should be used in fine-grained sediment to contain tur-
bidity.
• Construction vessel BMPs: For docks installed in shal-
low water environments, construction activities should
be staged to only occur near high tide to reduce the risk
of vessel grounding or propeller scouring.
Decking
The decking for a dock is elevated with pile supports and
so does not cause any direct impacts. However, decking
can cause a variety of indirect impacts to marine resources
through leaching (Weis and Weis 1996) and shading
(Burdick and Short 1999; Able and Duffy-Anderson 2005;
Logan et al. 2018a). In particular, salt marsh and seagrasses
and the variety of associated ecosystem services provided
by these habitats (Barbier et al. 2011) can be negatively
impacted by shading from dock decking (Burdick and Short
1999; Logan et al. 2018a, b).
Leaching
While some guidelines only recommend avoidance of pres-
sure-treated wood for in-water structures (e.g., WADOE
2011), decking composed of pressure-treated wood also
results in the leaching of heavy metals into the aquatic
environment through weathering and rainfall events (Khan
et al. 2006; Shibata et al. 2007; Lebow 2014). In general,
leaching and rainfall rates are inversely related since water
has less contact time with wood during heavier rain events
(Lebow 2014). Leaching rates from decks are lower than
from piles since piles have routine exposure to water (Lebow
et al. 2004). Like pile impacts, leaching impacts from deck-
ing are higher in the first year after construction, in poorly
flushed environments, and for animals and habitats in close
proximity to the pressure-treated wood structures (Weis and
Weis 1996, 2002, 2004).
Estuaries and Coasts
1 3
Shading
The primary indirect impact imparted by decking is shading.
Reduced light levels under decking adversely affect produc-
tivity of salt marsh and seagrasses and can alter fish behavior
(Burdick and Short 1999; Munsch et al. 2017; Logan et al.
2018a, b). Decking that shades salt marsh vegetation can
reduce underlying production by decreasing stem density and
biomass (Kearney et al. 1983; Sanger et al. 2004; Alexander
and Robinson 2006; Vasilas et al. 2011b; Logan et al. 2018a,
b). Relative impacts vary among decking designs (Kearney
et al. 1983; Colligan and Collins 1995; Sanger and Holland
2002; Alexander and Robinson 2004; Sanger et al. 2004;
Vasilas et al. 2011b; Alexander 2012; Logan et al. 2018a,
b). Spartina stem density under docks along the U.S. east
coast was reduced by approximately 30 to 70% (Alexan-
der and Robinson 2004, 2006; Sanger et al. 2004; Alexan-
der 2012; Logan et al. 2018a) relative to unshaded habitat.
Biomass effects were more variable across dock studies and
ranged from a 63% reduction to a 23% increase, with the lat-
ter observed increase occurring for tall-form S. alterniflora
(Alexander and Robinson 2006; Alexander 2012; Logan et al.
2018a).
Light penetration increased with decking height in field
studies conducted in both the northeast and southeast U.S.
(Alexander 2012; Logan et al. 2018a, b). Salt marsh vegeta-
tion under northeast U.S. docks also responded positively
to increased decking height with clear increases in above-
ground production as decking height increased towards 4 ft
(1.2 m) and higher (Kearney et al. 1983; Logan et al. 2018a,
b). Docks positioned closer to the ground frequently have
large bare patches along the underlying marsh substrate due
to heightened shading (Fig. 2). Decking set > 5 ft (1.5 m)
above the marsh platform had the least impact on above-
ground production in Massachusetts estuaries (Logan et al.
2018a, b). Current U.S. state guidelines include minimum
dock height recommendations ranging from 3 ft (0.9 m)
(NCDEQ 2021) to 6 ft (1.8 m) (ACOE 2017; Table 1).
Several U.S. state BMP guidelines recommend a minimum
dock H:W ratio of 1:1 (e.g., NHDES 2020; Table 1), but a
controlled study found that docks with the 1:1 design had
only approximately 50% of the Spartina biomass found in
unshaded controls (Logan et al. 2018b). A 1.5:1 design (i.e.,
6 ft (1.8 m) above the marsh platform for a 4 ft (1.2 m) wide
dock) had significantly greater marsh production relative to
1:1 docks (Logan et al. 2018b).
Decking width and orientation effects on underlying salt
marsh production varied across different regions of the U.S.
east coast. Width was the most important predictor of relative
marsh production under mid-Atlantic U.S. docks but was not
a significant predictor for New England docks (Kearney et al.
1983; Colligan and Collins 1995; Logan et al. 2018a). Most
U.S. state guidelines recommend a four ft (1.2 m) maximum
dock width (Table 1), and most docks in Massachusetts con-
structed over salt marsh have a four ft (1.2 m) width (Logan
et al. 2018a), so uniformity in dock width may have lim-
ited detection of width effects on marsh production in New
England studies. Many U.S. state guidelines recommend a
north–south orientation for docks constructed over vegetation
(Table 1) as a strategy to decrease shading since light avail-
ability should be enhanced as the sun rises in the east and sets
in the west. Light reduction from decking designs will likely
vary with latitude since the elevation and angle of the sun
are related to latitude. Consequently, designs that promote
light penetration in field settings at a given latitude should be
applicable to that geographic region as well as more southern
latitudes, but may not impart the same benefits to docks con-
structed at more northerly latitudes. Similarly, designs that
promote light penetration in U.S. estuaries in spring and sum-
mer seasons may not produce the same improvements during
Fig. 2 Typical salt marsh loss
under docks constructed closely
over the marsh platform. (A)
Private dock in MA, USA
and (B) experimental dock
constructed at < 1:1 H:W ratio
(Logan et al. 2018b)
Estuaries and Coasts
1 3
fall or winter (Alexander 2012). Dock orientation affected
light availability under northeast and southeast U.S. docks
(Alexander 2012; Logan et al. 2018a). In New England, rela-
tive decreases in aboveground production declined as dock
decking shifted towards a north–south orientation with the
greatest marsh loss observed for docks set east–west (Logan
et al. 2018a). However, dock orientation did not influence salt
marsh production under docks in the mid-Atlantic (Vasilas
et al. 2011b) or southeast U.S. regions (Sanger and Holland
2002; Alexander and Robinson 2004; Sanger et al. 2004).
Relative impacts to salt marsh varied among decking
designs (Kearney et al. 1983; Colligan and Collins 1995;
Sanger and Holland 2002; Alexander and Robinson 2004;
Sanger et al. 2004; Vasilas et al. 2011b; Alexander 2012;
Logan et al. 2018a, b). Several U.S. state BMP guidelines
for dock construction recommend alternative decking to pro-
mote light penetration (Table 1). Decking type was not a sig-
nificant predictor of salt marsh stem density under northeast
U.S. docks (Logan et al. 2018a) while stem density reduc-
tions under southeast U.S. docks were higher for grated than
traditional decking (Alexander 2012). Controlled studies of
light penetration through alternative decking in the southeast
U.S. demonstrated that during fall and winter when the sun’s
elevation was low, light was not able to penetrate the grated
decking any more than traditional plank decking. During
spring and summer, when the sun’s elevation was higher,
alternative decking did enhance light penetration, but only
produced a small (< 10%) increase in photosynthetically
active radiation (PAR) relative to traditional plank decking
(Alexander 2012). For docks in the northeast U.S., where the
sun’s elevation is lower, total light availability did not sig-
nificantly differ between decking types (Logan et al. 2018a).
SAV is also vulnerable to shading, and light obstruction
from decking and floats can result in eelgrass mortality
and sub-lethal impacts (Burdick and Short 1999; Beal and
Schmit 2000; Eriander et al. 2017; Sagerman et al. 2020).
Extent and magnitude of impacts vary greatly depending on
seagrass species, orientation, and other environmental con-
ditions (Burdick and Short 1999; Gladstone and Courtenay
2014; Eriander et al. 2017). A recent review by Sager-
man et al. (2020) found that SAV coverage under decking
was only 18% of nearby control areas, but effects varied
across studies and species with a range of 9 to 36%. Even
structures that cause partial shading can largely eliminate
existing eelgrass with minimal chance for regrowth (Pen-
tilla and Doty 1990). Using experimental docks, Shafer and
Robinson (2001) found Florida seagrass shoot densities to
be 52 and 58% of control densities under docks placed at 5
ft (1.5 m) and 4 ft (1.2 m) above mean high water (MHW),
respectively. In Australia, Gladstone and Courtenay (2014)
found that Zostera muelleri subsp. capricorni is reduced
by 75% under docks compared to controls.
Eutrophication effects (e.g., decreases in dissolved oxygen
and water clarity, increases in sediment sulfides) impart addi-
tional stresses on eelgrass that can compound decking shad-
ing impacts (Goodman et al. 1995). Shading can also alter
light availability towards levels more favorable for eelgrass
competitors such as macroalgae (Markager and Sand-Jensen
1992). Once algae becomes established, it can outcompete
eelgrass for space and light (Short et al. 1995; Hauxwell et al.
2003), as well as lead to hypoxia and eutrophication upon
decomposition (Han and Liu 2014). In Massachusetts, this
displacement is particularly concerning in inner-harbor areas
subject to eutrophication. Benthic communities in Waquoit
Bay, Massachusetts shifted from eelgrass to macroalgae-
dominated due to deteriorating water quality, leading to shad-
ing impacts to the remaining eelgrass, decreases in dissolved
oxygen, and increases in sulfide in the sediments, which are
toxic to eelgrass (Hauxwell et al. 2003).
Similar to studies of docks constructed over salt marsh
in New England (Kearney et al. 1983; Logan et al. 2018a,
b), Burdick and Short (1999) and Short et al. (2009) found
dock height to be the most important variable determining
underlying light availability and by association, eelgrass bed
quality. Shoot density of Halodule wrightii beneath docks
in Perdido Bay, Alabama was reduced by 40 to 50% at light
levels of 16 to 19% surface irradiance (SI) (Shafer 1999),
with no seagrass found under docks with less than 14% SI.
Mesocosm studies found 30% SI would support 50% of the
normal eelgrass production (Short et al. 1995) and a signifi-
cant reduction in photosynthetic response after 40 days of
shading at ≤ 34% surface irradiance (Ochieng et al. 2010).
Burdick and Short (1999) used mesocosm information to
model dock heights that would achieve adequate surface
irradiance (30%) for eelgrass growth around and beneath the
structure in a microtidal lagoon in southern Massachusetts.
They determined a 3 ft wide (1 m) dock requires a minimum
height of approximately 10 ft (3 m) off the seafloor.
Dock height above the water surface has been shown to
be positively correlated with underlying eelgrass coverage
(Eriander et al. 2017). Short et al. (2009) created a “Dock
Eelgrass Calculator” (DEC) that includes dock height
above mean sea level (MSL), orientation and width vari-
ables that users can manipulate to estimate how impacts
to eelgrass vary with different designs. For a 4 ft wide
(1.2 m) dock typical of Massachusetts, the DEC estimates
a north-oriented dock would need to be approximately
5.7 ft (1.7 m) above MSL to avoid impacts to underly-
ing bed quality. For an east-facing dock, the required
height to avoid impacts would increase to approximately
9.7 ft (3 m). Seagrass presence under docks in Florida
was positively correlated with height based on a sampled
height range of 0 to 5.6 ft (1.7 m) above MHW (based on
Florida Department of Environmental Protection survey
Estuaries and Coasts
1 3
described in Beal and Schmit 2000). Experimental docks
set at heights of 3 ft (0.9 m) and 4.9 ft (1.5 m) above MHW
over seagrass beds in Florida also caused a decline in sea-
grass (percent cover and shoot density) although only shoot
density effects scaled with dock height (Beal and Schmit
2000). Shafer and Robinson (2001) found that seagrass
could persist under docks in Florida that were elevated 4
to 5 ft (1.2–1.5 m) above mean sea level and used fiber-
glass grid for the entire dock. Shading effects may be less
where larger tidal ranges exist, due to the greater separation
between the decking base and the water surface during low
tides and potential for more light infiltration, but to our
knowledge there are no studies that confirm this theory.
Orientation and width influence shading impacts on
SAV (Burdick and Short 1999). East–west oriented docks
cause all-day shading resulting in either poor quality beds or
the complete loss of vegetation (Burdick and Short 1999).
According to the Burdick and Short (1999) and Short et al.
(2009) models, a dock constructed in an east–west orienta-
tion would require approximately twice the height to allow
a given light penetration relative to a north–south oriented
dock in the same location. Light levels are greater under the
centers of narrow docks than under wide docks (Burdick
and Short 1999). In Australia, seagrass (Zostera muelleri
subsp. Capricorni) bed biomass was only approximately
25% of unshaded controls but the losses under docks in
Australia were independent of dock orientation, possibly
due to the narrow width (3–3.6 ft (0.9–1.1 m)) of sampled
docks (Gladstone and Courtenay 2014). Tropic and sub-
tropic seagrass species may also respond to dock shading
differently than temperate seagrasses.
Applications of alternative decking materials have had
variable success in limiting impacts to underlying SAV.
Docks with grated decking in Australian coastal waters
caused seagrass loss, although bed loss was less than that
caused by traditional wooden decking (Gladstone and
Courtenay 2014). Similarly, the installation of prisms
along docks was found to have only minor benefits to eel-
grass growth, unless used in great quantity (Blanton et al.
2002). The same study investigated several other products
(e.g., SunTunnels, halide lights), and while some improved
light conditions, they did not completely eliminate impacts.
Beal et al. (1999) tested fiberglass and alternating wood-
and-fiberglass decking and found only minor differences
in irradiance. Landry et al. (2008) found that grated docks
improved light irradiance under the structure relative to
traditional float designs, but Halophila johnsonii densi-
ties were still reduced relative to unshaded reference areas.
Gayaldo et al. (2001) tested the use of reflective alumino-
metallic film attached to the structure’s underside and piles,
and found an increase in reflected light under the struc-
ture with light levels elevated from 1–3 to 9–11%, which
allowed eelgrass to recolonize beneath the structure.
While impacts of small docks on vegetation have been
well studied, studies of dock shading impacts on fish and
invertebrate communities to date are mostly limited to large
structures (i.e., piers, marinas, and bridges) and results may
not scale to smaller private structures (Able et al. 2013). For
example, most studies of piers have found negative shading
effects on fish and invertebrates (Able and Duffy-Anderson
2005) while a study of small docks in the southeast U.S.
showed that decking shading provided a beneficial thermal
refuge to mangrove tree crabs (Cannizzo et al. 2018). Previ-
ous studies of large commercial piers have shown a variety
of impacts including alteration of fish migratory behavior,
habitat use, community composition, and growth (Munsch
et al. 2017). Light levels under large urban piers decrease
towards the pier interior (Able et al. 2013; Munsch et al.
2014), and shading appears to be the main driver of nega-
tive fish responses (Munsch et al. 2017). The large munici-
pal and commercial piers included in previous studies all
had widths > 60 m (200 ft) with a maximum width of 837
ft (255 m) for Pier 40 in the Hudson River estuary (Able
and Duffy-Anderson 2005; Munsch et al. 2014). Bridges
assessed for benthic invertebrate impacts were > 29.5 ft
(9 m) wide (Struck et al. 2004). By contrast, the width of
private docks in Massachusetts is typically < 5 ft (1.5 m)
(Bliven and Pearlman 2003; Logan et al. 2018a) but situ-
ated closer to the water surface than larger piers (Munsch
et al. 2014) and bridges (Struck et al. 2004).
Shading from decking can impact the communities that
colonize underlying piles and nearby benthic habitat. For
example, the percent cover of serpulid polychaetes, sponges,
and solitary ascidians was often greater on piles in Aus-
tralian marinas than rocks while spirorbid polychaetes,
foliose and filamentous algae were more prevalent on rocks
(Glasby 1999a). The differences between piles and rocks
were attributed to shading, rather than the pile structure
itself, as unshaded piles supported similar communities
as rocks while shaded piles had significantly different spe-
cies composition (Glasby 1999b). Therefore, shading from
overlying dock decking may counteract any positive habitat
contributions of the support piles (See review by Munsch
et al. 2017). Shading from piers and other urban infrastruc-
ture in Brazil affected rocky intertidal community composi-
tion by reducing macroalgal biomass and cover as well as
grazer size (Pardal-Souza et al. 2017). Relatedly, shading
from decking combined with increased filter-feeder colo-
nization of piles could potentially result in dock structures
becoming carbon dioxide sources and energy sinks due to
the resulting imbalance of primary producers and consumers
(Malerba et al. 2019). Relationships between total productiv-
ity and increased filter-feeder colonization of shaded piles
are complex, so indirect effects of net carbon uptake and
carbon dioxide generation are not entirely clear (Riascos
et al. 2020) but nonetheless represent a further example of
Estuaries and Coasts
1 3
potential unintended effects of shading. Additional research
on fish and invertebrate communities associated with docks
and associated environmental conditions would be useful in
furthering our understanding of scaling effects from what is
currently known for larger commercial piers.
Recommendations
• Avoid salt marsh and SAV whenever feasible: Dock
construction should be avoided in areas containing salt
marsh or SAV given the ecosystem services provided by
these vegetation types and their susceptibility to shading.
• Focus on dock height, width, and orientation over salt
marsh and SAV: When avoidance is not feasible, a mini-
mum 1.5:1 H:W ratio is recommended for docks con-
structed over salt marsh and a 10 ft (3 m) height above
the seafloor (or 5.7 ft (1.7 m) above MSL) for docks over
SAV. Orientation should be within 10° of north–south to
increase light penetration.
• Maintain light transmission even in unvegetated habitats:
A 1:1 H:W ratio is recommended for docks constructed
over non-vegetated intertidal and subtidal habitats to
reduce shading impacts.
• Use alternative materials in poorly flushed areas: In areas
with low flushing, the decking component should use
alternative materials such as composite, fiberglass, or
local hardwood.
• Use construction BMPs to minimize pressure-treated
wood impacts: As with pile installation, on-site cutting
should be avoided or proper containment used to prevent
sawdust and shavings from entering the aquatic environ-
ment when using pressure-treated wood.
Floats
Floats are located at the seaward or channelward extent of a
dock structure (Fig. 1) and can serve as a platform for swim-
ming or fishing as well as a location for vessel tie-up. Floats
are typically secured with piles or chain and anchor struc-
tures. Some dock structures do not include floats, but instead
have pile-supported decking that terminates over the water.
In some areas, floats are installed on a seasonal basis, and in
other areas they are present year-round. Seasonal floats only
impact the resource area for part of the year, but can have
additional negative impacts if stored in sensitive resource
areas (e.g., salt marsh). Often, floats rise and fall freely with
the tide, secured in place either with piles or anchor systems.
Data on float impacts on benthic habitats are limited
relative to other dock structural effects, but preliminary
research has shown impacts do occur (e.g., Alexander et al.
2006). Depending on water depth and underlying marine
resources, dock floats can have both direct and indirect
impacts. Direct impacts can occur when floats rest on the
substrate during part of the tidal cycle (Fig. 3; Kelty and
Bliven 2003; Alexander et al. 2006) while indirect impacts
are mainly associated with shading (Burdick and Short
1999). Grounded floats can lead to crushing, smothering,
Fig. 3 Examples of potential float impacts on marine resources
including (A) a float grounding on a mudflat at low tide, (B)
increased shading impacts from a large vessel tied to a float, and (C)
propeller scouring of eelgrass in shallow water that can occur when
vessels transit to and from dock floats. All images taken in MA, USA
Estuaries and Coasts
1 3
burial, damage to vegetation, and disturbance to substrates
(Nightengale and Simenstad 2001). As floats rise and fall
with the tide they can lift sediments (i.e., “pumping”) and
impact water clarity (Kelty and Bliven 2003). Float moor-
ing systems common with New England docks also impact
the surrounding seafloor since the anchors and mooring
lines come in contact with the seafloor. The placement of
a float over shellfish habitat can lead to direct habitat dis-
placement by piles or anchors. U.S. state guidelines gener-
ally recommend float water depths be a minimum of 1 to
4 ft (0.3 to 1.2 m) at MLW to minimize habitat impacts
depending on the resources present (Table 1).
Floats installed in intertidal habitat can directly impact
benthic habitat by grounding at low tide or indirectly impact
nearby salt marsh through erosion or circulation alteration.
In addition to the direct conversion of mudflat to artificial
structure, there is often a “dead zone” of soft sediment sur-
rounding float piles or footings (MacFarlane 1996). In such
areas, impacts can be reduced through alternative means of
water access including the use of a dinghy to reach a ves-
sel moored in subtidal habitat, use of a nearby public boat
launch, or creation of a shared, community dock. When
alternatives are not feasible, pile-supported float stops are
sometimes used. Float stop systems incorporate brackets
near the bottom of the support piles to stop the float from
falling below a certain height at low tide, suspending the
float at that height until the next incoming tide. The area of
direct impact can also be reduced through the installation
of legs at the corners or skids running along the float edges.
Benthic Impacts
Assessments of benthic habitat under and adjacent to floats
in Georgia identified impacts to sediment grain size and by
association carbon and nitrogen content. Biological impacts
included reductions in benthic algal production, macrofaunal
numbers, and biomass (Alexander et al. 2006). The floats
included in the Alexander et al. (2006) study all rested on the
substrate at low tide and porosity profiles provided evidence
that the floats were compressing the upper < 1 in (1–2 cm)
surface sediment layer of the seafloor (Alexander et al.
2006). Many of the detected impacts also occurred under the
float sections that did not contact the substrate. The authors
attributed these impacts to flow acceleration and removal of
finer, higher organic content sediment for the substrate areas
under the overall float structure and adjacent to the sub-
floats that rested on the substrate. Alexander et al. (2006)
also found sediment grain size alterations downstream of
the float, which suggests that float impacts may have a larger
spatial footprint in systems with higher tidal velocities.
Shading
In addition to the smothering and crushing impacts of
floats, SAV can be adversely impacted by shading, which
has been documented in both U.S. (Fresh et al. 1995, 2006;
Burdick and Short 1999) and European (Eriander et al.
2017) waters. As with decking, the terminal float, if posi-
tioned over existing or historic eelgrass areas, can reduce,
limit, or entirely prevent eelgrass from vegetating the area
directly under the float as well as a halo area, which can
be as much as twice the size of the float (Fresh et al. 1995;
Smith and Mezich 1999). Overall, shading from floats
has greater impacts on SAV than pile-supported decking
(Burdick and Short 1999; Eriander et al. 2017). The float
impacts are likely a result of a combination of shading
and direct impact, since floats may directly come in con-
tact with eelgrass. Unlike pile-supported decking, floats
are directly on the water surface and so impart a greater
shading impact. Burdick and Short (1999) found that no
eelgrass could be supported under floats where tidal range
was < 3.3 ft (1 m). Impacts were greater under floats com-
pared to fixed-height docks, and eelgrass was completely
absent under three of the four floats that they examined.
Fresh et al. (1995) observed that mean eelgrass density
beneath floats in Puget Sound, Washington was 24% of
the mean undisturbed control density, and eelgrass density
one structure width away from the dock was 60% of the
control. In Sweden, eelgrass was completely absent under
all sampled floats while habitats under and adjacent to pile-
supported docks had an average reduction in coverage of
approximately 70% (Eriander et al. 2017). Shading impacts
are not constrained to the area directly under floats, but
instead can extend as far as 20 to 26 ft (6–8 m) from the
float edge (Eriander et al. 2017). This broader area of shad-
ing impact can be a result of side-shading from the float
itself as well as additional shading created by vessels tied
up to the float (Fig. 3B; Eriander et al. 2017).
Several studies have investigated the effectiveness of
alternative construction materials to enhance light pen-
etration under floats with most studies documenting con-
tinued seagrass loss. Steinmetz et al. (2004) found that the
use of glass prisms as part of the float did not prevent loss
of underlying seagrass in Florida. Fresh et al. (2006) used
both steel and fiberglass grated decking over Z. marina,
and found that floats with as much as 50% grated space
still impacted eelgrass in Puget Sound and were of no
ecological benefit compared to traditional planked floats.
Shafer and Robinson (2001) studied alternative materi-
als such as acrylic, lexan, aluminum grating, and fiber-
glass grating, and found significant declines in turtle grass
Estuaries and Coasts
1 3
(Thalassia testudinum) in Florida in both spring and sum-
mer in nearly all treatments. Only the fiberglass grating
showed promise for continued turtle grass growth. Vulner-
ability to shade stress varies among SAV species (Collier
et al. 2016; Nelson 2017) and light transmission through
grating varies with latitude so design benefits may also
depend on the species present and geographic location of
a given dock site.
Anchor System Impacts
The anchor system that secures the float to the sea-
floor can also result in direct impacts, whether concrete
block anchors with chain, or helical anchors and flex-
ible rodes commonly known as conservation moorings.
While common in New England, anchor systems are not
typically used in the southeast U.S. and so present more
of a regional problem (C. Alexander, Pers. Comm). A
study investigating the efficacy of conservation mooring
systems in five Massachusetts harbors found that these
systems can minimize impacts to eelgrass when they
are entirely floating above the sediment, which requires
proper installation and regular maintenance (MA DMF
2019). In some cases, once the eelgrass is uprooted, eel-
grass detritus can linger and prohibit new growth (MA
DMF Pers. Obs.; Walker et al. 1989). Impacts to the rhi-
zome system can lead to reduced habitat and destabilized
sediments. Furthermore, moorings in some embayments
are prone to heavy fouling by mussels, tunicates, and
other organisms that weigh the mooring down causing
additional scour. If conservation mooring rodes are left
in the water to over-winter, unlike chain dropped to the
bottom in one place, the floating rode will partially sink
to the bottom, resulting in a larger scour scar in the sur-
rounding eelgrass (MA DMF 2019). While these systems
can result in lower impacts to the seafloor and SAV than
traditional block and chain systems with proper design,
maintenance, and adequate water depth, in practice they
often still cause damage to eelgrass meadows (MA DMF
2019) and should not be viewed as a mitigating meas-
ure to permit float construction in eelgrass beds. Further
study is also required to assess their efficacy in minimiz-
ing impacts to seagrass when used as float anchors.
Invasive Species
Like piles, floats can serve as habitat for invasive species
(Lambert and Lambert 1998, 2003; Lambert 2002; Arenas
et al. 2006; Ashton et al. 2006; Minchin 2007) by provid-
ing colonizing substrate. Floats can potentially facilitate
the further spread of invasives to natural hard substrate if
present in the surrounding waterbody. Potential invasive
species interactions are described in greater detail in the
“Invasive Species” sub-section of the “Pile Design and
Materials” section and not repeated here as impacts from
floats are similar.
Boating Impacts
Boats are commonly moored to dock floats, and boating
activity associated with floats can also negatively affect the
benthic environment and SAV in particular (Orth et al. 2006;
Sagerman et al. 2020) if the float is located in shallow water.
SAV can be indirectly impacted by boats tied up to floats
through shading (Eriander et al. 2017). Floats in Sweden
had a distinct border between low and full eelgrass coverage
that was 7 to 8 m (~ 23 to 26 ft) from the float edge. This
extended area of shading impact outside of the float footprint
was likely influenced by additional shading from associ-
ated vessels (Eriander et al. 2017). Boating activity around
floats can also result in direct impacts. Propeller dredging
occurs when a boat’s propeller or water jets suspend sedi-
ments. This may result in the direct loss of vegetation or
loss through burial by sediments (Fig. 3; Burdick and Short
1999; Sagerman et al. 2020). Propeller scarring of SAV has
been well documented in shallow water habitats worldwide
(e.g., Gonzalez-Liboy 1979; Loflin 1995; Burdick and Short
1999; Martin et al. 2008; West 2011; Hallac et al. 2012).
Propeller scarring in Florida Bay was most prevalent in
shallow waters ≤ 2 m (6.6 ft) (Sargent et al. 1995; Hallac
et al. 2012). The disturbed area may become unsuitable for
seagrass growth due to changes in depth or substrate type
and become permanently void of growth or take many years
to heal. For example, eelgrass scars from mussel dragging
in Maine, USA were estimated to require from 6 to 20 or
more years to repopulate depending on growth conditions
(Neckles et al. 2005). Other potential adverse effects of pro-
peller scarring include destabilization of sediments as well
as creation of habitat for algae and invasive species to take
hold. Scar areas also support lower macrofaunal abundance
and diversity relative to SAV (Uhrin and Holmquist 2003).
In addition to SAV impacts, turbidity generated by propellers
can disrupt or even bury newly settled shellfish (MacFarlane
1996) and finfish eggs (Whitfield and Becker 2014).
Recommendations
• Minimize float size: Float size should be the minimum
length and width necessary to berth the associated vessel.
Minimizing the float footprint will reduce the scale of
potential impacts (e.g., shading, turbidity, benthic scour).
• Place floats in the deepest portion of the site: Floats
should be placed in the deepest available water and in
areas where the structure will not ground at low tides.
We support the continued use of existing guidelines rec-
Estuaries and Coasts
1 3
ommended for the protection of shellfish habitat (2.5 ft
(0.8 m) at MLW) and all other locations (1.5 ft (0.5 m))
to protect other benthic fauna in Massachusetts as these
buffers will prevent float grounding and propeller scour
from smaller vessels. These depths should be measured
from the substrate to the lowest structural float compo-
nent.
• Use alternative designs or means of water access when
the float cannot reach subtidal habitat: For locations
where float placement in subtidal habitat is not feasible,
alternatives to individual dock and float systems should
first be explored. In the absence of viable means of alter-
native water access, we recommend pile-supported float
stops to keep the entire float from contacting the seafloor
at low tide. If piles and float stops are not feasible, legs or
skids should be used to minimize the area of direct float
impact.
• Avoid SAV: Given the vulnerability of SAV to direct and
indirect impacts from float installation, MA DMF recom-
mends avoidance as the best strategy for floats and any
associated anchors or piles. Terminal floats should be
located at least 7.5 m (~ 25 ft) from the edge of existing
or historically mapped eelgrass.
• Maintain buffer from salt marsh: Maintain at least 7.5 m
(25 ft) buffer from the edge of the nearest salt marsh to
avoid circulation and/or erosion impacts.
• Seasonal floats should be stored outside of aquatic
resource areas: For cases where floats are removed sea-
sonally, storage locations should avoid sensitive resource
areas. Floats should be stored above the high tide line
and outside of any salt marsh or other sensitive aquatic
habitats.
Cumulative Impacts
While individual small docks may have only limited impacts
on any given aquatic resource, dense build-out can collec-
tively result in greater overall fragmentation, alteration, and
loss of habitat that should be considered in coastal planning
(MacFarlane 1996; Peterson and Lowe 2009). For example,
in Massachusetts, > 2500 docks occupied over 6 ha of salt
marsh (Logan et al. 2018a). Docks occupied > 28 ha and
60 ha of coastal habitat, in Georgia and South Carolina,
respectively (Sanger et al. 2004; Alexander and Robinson
2006). In addition to causing direct habitat loss, docks and
other artificial structures can disrupt ecological connectiv-
ity by creating barriers to the movement of certain species,
providing new structure that facilitates the movement of
other species, and altering trophic connectivity (Bishop et al.
2017). While cumulative impacts are important to quantify
and account for in coastal management, quantifying broader
impacts remains a challenge due to the current lack of con-
sensus on an approach.
System-level assessments and planning may be achieved
by considering individual dock applications in the con-
text of relative ecosystem-level impacts and quantitative
comparisons of benefits and negative impacts. MacFar-
lane et al. (2000) described an index to identify areas of
greater resource sensitivity in Pleasant Bay, Massachu-
setts, which could then be avoided during construction
activities. The sensitivity index used a variety of abiotic,
biological, and anthropogenic factors to assign a ranking
to a given region of the study system of 0, 0.5, or 1, for
low, moderate, and high resource sensitivity, respectively.
Sites with total scores > 5.5 were considered highly sen-
sitive and consequently unsuitable for dock installation
(MacFarlane et al. 2000). This framework was used by the
towns surrounding Pleasant Bay to identify regions within
Pleasant Bay to restrict new dock construction (Pleasant
Bay Resource Management Alliance 2018). Needles et al.
(2015) described a similar matrix approach that provides
managers with a decision support tool to quantify tradeoffs
of a variety of management actions. The matrix consid-
ered cultural, climate regulation, biological services (e.g.,
food), storm protection, water quality, and biodiversity ser-
vices by assigning positive, negative, or neutral effects of
each considered management action on these services. The
Programmatic General Permit (PGP83) of the U.S. Army
Corps of Engineers for private docks in Georgia included a
cumulative impact assessment based on annual dock build-
out and estimates of the percent of different coastal habitat
types being impacted (e.g., marsh, tidal flats). The envi-
ronmental impact assessment estimated a variety of direct
and indirect impacts (e.g., shading, wrack accumulation,
sediment re-suspension, float grounding) and found that the
impacted area was < 1% of the state’s total marsh habitat,
although cumulative impacts were not assessed at the level
of individual systems (ACOE 2012; King and Blair 2012).
Best management practices including the siting, timing,
installation methods, materials, and designs employed in the
construction of docks can minimize environmental impacts
of individual dock structures, but cumulative impacts still
should be accounted for in the permitting process. Most
BMPs will function to minimize rather than completely
avoid environmental impacts and so adoption of these prac-
tices without consideration of the cumulative impacts with
dock proliferation may instill a false sense of environmen-
tal preservation. For example, dock shading and loss of
underlying marsh production can be minimized by BMPs
that increase the dock’s height to width (H:W) ratio, but
even docks with optimized H:W under practical constraints
will induce some level of shading and marsh loss. Docks
set at a 1.5:1 H:W ratio and oriented N-S reduced shading
Estuaries and Coasts
1 3
Table.3 Best management practice (BMP) recommendations for docks in New England estuarine and coastal waters
Component Recommendation
Piles
Design – general •Minimize footprint (diameter and number)
•Maximize pile spacing
Design – over sensitive habitats (salt marsh, SAV)•Avoid where feasible
Installation •Work from upland where feasible
•Use shallow-draft barge that avoids grounding
•Use vibratory rather than impact driving where feasible
•Avoid jetting
•Employ “soft start” to reduce turbidity and noise impacts when using impact
driving
•Use silt-curtains in areas with fine-grained sediment
•Avoid sensitive life history periods (e.g., fish spawning)
Materials •Use CCA or local hardwood instead of ACQ or other high-copper content treated
wood
•Avoid cutting over resource area to reduce leachate release to aquatic environment
Decking
Design – general •Minimize width (≤ 4 ft (1.2 m))
•Maximize H:W ratio (≥ 1:1)
•Orient within 10° of North
•Avoid add-ons that increase shading (e.g., bump-outs, gazebos)
Design – over sensitive habitats (salt marsh, SAV)•Avoid where feasible by micro-siting walkway in area of property outside sensi-
tive habitats
•Follow General Design BMPs with the following modifications:
oSalt Marsh
-Maximize H:W ratio (≥1.5:1)
oSAV
-Minimize width (≤ 3 ft (0.9 m))
-Maximize height above seafloor (> 10 ft (3 m))
Material •Use composite or local hardwood
Floats
Design – general •Place in deepest available water to avoid direct and indirect impacts during low
tide
•Maintain minimum 1.5 ft (0.5 m) depth under float at mean low water (MLW)
•Avoid intertidal habitat where feasible
•If only intertidal habitat available, use pile-supported float stops to avoid or legs/
skids to minimize contact with seafloor
•Orient within 10° of North
•Minimize float area to size required for intended use
Design – over sensitive habitats (salt marsh, SAV, shellfish)•Follow General Design BMPs with following modifications:
•Maintain minimum 25 ft (7.5 m) separation from any bordering salt marsh or
SAV
•Avoid placing float anchors in SAV
•Maintain minimum 2.5 ft (0.8 m) depth under float at mean low water (MLW) in
shellfish habitat
Installation •Avoid grounding of work barges
•Remove seasonally and store outside resource areas
Cumulative impacts
Community planning •Quantify cumulative impacts at ecosystem-level to place individual dock permit-
ting in proper context
•Develop by-laws and tools for minimizing cumulative impacts to sensitive
resources (e.g., salt marsh, SAV) and cultural resources (e.g., shellfishing flats)
•Require mitigation for unavoidable impacts
Alternative access approaches
•Public boat ramps
•Community docks
•Moorings
•Boat lifts
Estuaries and Coasts
1 3
impacts relative to lower H:W designs, but still caused an
approximate 30% decline in low marsh production (Logan
et al. 2018b). When viewed in the context of a single dock,
such impacts may be considered minor and consequently
permittable, but dense dock build-out even following BMPs
(Table 3) could cause collective, cumulative impacts to salt
marsh and other estuarine resources.
In some U.S. states, towns can implement BMP conditions
comprehensively through the development of local bylaws
that can be more specific and conservative than state laws
(Patterson 2003; Goetsch 2011; Massachusetts Association of
Conservation Commissions 2016). In Massachusetts, many
towns have taken this approach to maximize protection of
eelgrass and other marine resources (Goetsch 2011; Massa-
chusetts Association of Conservation Commissions 2019). In
Massachusetts, the Waterways Regulations includes language
(310 CMR 9.38(2)(b)) that prohibits state permitting of docks
in areas that have been identified by town-level management
plans as being unsuitable due to competing uses (e.g., shell-
fishing) or presence of sensitive resources (e.g., SAV). The
Massachusetts Association of Conservation Commissions
created a template to help guide towns in developing such
bylaws (Massachusetts Association of Conservation Com-
missions 2016). A bylaw approach would allow a town to
broadly implement BMPs to all future dock projects, which
serves the added benefit of streamlining the permitting pro-
cess by clearly defining required designs and practices. In
Massachusetts, individual town bylaws include specific con-
ditions relating to dock construction methods (e.g., prohibi-
tion of jetting in pile installation), location relative to sensi-
tive marine resources (e.g., floats need to be seaward of salt
marsh and eelgrass), and size (e.g., 4 ft (1.2 m) maximum
width) (Town of Bourne 2000; Town of Hingham 2014;
Town of Yarmouth 2016).
Recommendations
• Map and quantify important resource areas at a system
level: Improved characterization of existing resources
and abiotic conditions of regions within individual estu-
aries can guide BMPs and restriction decisions at a sub-
embayment scale.
• Develop sensitivity indexes to guide the use of BMPs vs.
moratoriums: Sensitivity indexes should be developed for
individual estuaries to identify areas where dock installa-
tion could occur with limited environmental impact with
appropriate BMPs as well as areas where restrictions are
more appropriate due to the presence of more vulner-
able resources (e.g., SAV) and/or important cultural or
recreational resources (e.g., productive shellfish flats).
• Consider BMP guidance in the context of anticipated
dock build-out: A system-wide approach to manage-
ment may identify areas where dock build-out can be
accomplished with limited resource impacts. If restric-
tions are not feasible for a given town or estuary due
to state or federal permitting standards, individual dock
BMPs should be scaled towards the most conservative
parameters to account for cumulative impacts. Decisions
on BMP guidance for individual dock projects can best
be made in the context of anticipated cumulative impacts
from continued build-out of similar structures in a given
system.
• Consider mitigation strategies for unavoidable resource
impacts: When avoidance is not feasible, cumulative
impacts of dock build-out within an estuary can also
be addressed through mitigation approaches. For cases
where BMPs are employed but unavoidable impacts still
occur, mitigation can be required to avoid net loss of
habitat within a system.
Data Gaps and Next Steps
Additional data are still needed to better understand dock
impacts (e.g., float grounding), possible designs to mini-
mize impacts (e.g., effective light-transmitting decking), and
strategies to incorporate cumulative impacts. Data assessing
potential impacts of grounding floats on epi- and infaunal
communities are lacking as is an assessment of how such
impacts may be avoided or minimized across different mini-
mum depth thresholds. While alternative designs intended to
reduce shading have largely been unsuccessful to date, some
designs have shown promise and further research into where
those designs work as well as new designs and materials is
warranted. For example, further development and testing of
different grated decking designs as well as light reflective
pile material or reflective material under decking or floats
is warranted to potentially reduce shading and associated
habitat alteration. Further development of pile materials and
designs that promote colonization by native species is also
warranted as such an approach could allow dock structures
to promote ecosystem services (e.g., biomass of filter feed-
ers). Relatedly, further study of the relationship between
submerged dock structures (piles and floats) and invasive
species colonization is warranted as piles should ideally be
designed and situated in a manner that does not promote
the spread of invasives (Chapman et al. 2018). Continued
development of ecological engineering designs (See review
in O’Shaughnessy et al. 2020) and similar field trials are
needed to identify appropriate designs for different habitat
types and geographical regions. Minimization of dock pro-
liferation combined with selection of designs and materi-
als that limit negative impacts and where possible provide
habitat enhancements are key management strategies for the
preservation of coastal and estuarine ecosystems.
Estuaries and Coasts
1 3
Acknowledgements We would like to thank P. Colarusso, E. Reiner, D.
Janick, J. Burtner, K. Glenn, R. Boeri, C. Alexander and an anonymous
reviewer for helpful comments on earlier drafts of this manuscript. We
also thank S. Kass and C. Cassidy for assistance with figure creation.
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