tidal creek and substrate effects on oyster reef
TRANSCRIPT
Coastal Carolina UniversityCCU Digital Commons
Electronic Theses and Dissertations College of Graduate Studies and Research
4-30-2018
Tidal Creek and Substrate Effects on Oyster ReefAssociated Fish and Decapod Condition andDensityThomas S. FunkCoastal Carolina University
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Recommended CitationFunk, Thomas S., "Tidal Creek and Substrate Effects on Oyster Reef Associated Fish and Decapod Condition and Density" (2018).Electronic Theses and Dissertations. 18.https://digitalcommons.coastal.edu/etd/18
TIDAL CREEK AND SUBSTRATE EFFECTS ON OYSTER REEF ASSOCIATED
FISH AND DECAPOD CONDITION AND DENSITY
By
Thomas S. Funk
________________________________________________________________________
Submitted in Partial Fulfillment of the
Requirements for the Degree of Master of Science in
Coastal Marine and Wetland Studies in the
School of Coastal and Marine Systems Science
Coastal Carolina University
2018
_________________________
Dr. Keith Walters
Major Advisor
_________________________
Dr. Charles Martin
Committee Member
_________________________
Dr. Richard Viso
Graduate Director
_________________________
Dr. Dennis Allen
Committee Member
_________________________
Dr. Robert Young
Committee Member
_________________________
Dr. Michael Roberts
Dean of Science
ii
Acknowledgements
First and foremost, I would like to thank my major adviser Dr. Keith Walters for
believing in me enough to take me on as a research assistant and guiding me through the
ups and downs of this research project. Working with CORRI is one of the best
experiences I’ve had, and I’m hopeful the reefs built during my research will continue to
benefit the community. I never knew I was capable of building PVC traps, running
complex statistical analyses, staying composed during public presentations, or pulling
together a document with as much depth as this Master’s thesis. I’m sure I’ll continue to
learn as we push toward a publication this summer.
Thanks also to my committee (Drs. Dennis Allen, Charles Martin, and Robert
Young) for their continuous stream of constructive criticism and analytical thinking that
shaped not only my writing but also several of my expeditions into the field. This project
could not have been as robust or well developed without your collective inputs and
different angles on numerous topics.
I would also like to thank Dr. Richard Viso, Dr. James Luken, Julie Quinn, and
Karen Fuss for continuously pointing me in the right direction for all matters
administrative. While I spent all my energy focusing on my research project, you made
sure the files, forms, and last minute details never fell through the cracks.
To Dr. Edgar Newman, thanks for the opportunity to serve as a teaching assistant,
which provided me the financial support I needed to focus on my studies and research.
My tenure also made me a stronger speaker and mentor while providing me a glimpse
into one possible career path.
iii
Thanks to the FishAmerica foundation for their generous funding during my reef
building exploits and Coastal Carolina University for providing an ideal workspace and
occasional funding for sampling equipment. I’d also like to acknowledge Kira Roff’s
contribution of recycled oyster shell from local restaurants, such as Noizy Oyster and
Flying Fish, which made this project possible.
Without my graduate assistants Kathryn O’Shaughnessy, Matt Davin, and Chris
Williams, many of the field protocols laid out in my methods section would have been
impossible, and I’d probably still be looking at samples in the lab. But above all else,
thanks for sharing my burden and keeping me sane during the most rigorous months of
the field season.
Finally, thanks to all my friends and family for supporting, loving, and never
giving up on me. Being able to talk about the day to day challenges of the thesis track
with other graduate students made the hardest times more bearable. To Amy Grogan, I
never felt cooler than when I was allowed into the exclusivity of Surfside Fridays. To
Matt Carter, I owe you a bag of Cheetos. To Matt Helms, Andre Dominguez, and Elise
Agosto, may we forever feel good on Wednesdays. To Maddy Gillis, thanks for letting
me vent to you throughout the process and picking me up if I got down. Thanks to
Jenkins for forcing me to plan healthy work breaks at the beach and dog park. To Atreyu,
thanks for sitting on my lap while I worked at my computer, preventing me from getting
up and forcing me to stay focused on work. Most of all, thanks to my parents for
supporting me in so many ways and helping me reach the series of goals that has gotten
me where I am today.
iv
Abstract
Oyster reefs are essential fish habitat and a worldwide loss of reefs has the
potential to negatively affect reef-associated nekton populations. Along the 100 km
Myrtle Beach, SC shoreline, oyster reefs ostensibly have disappeared within swash tidal
creeks, which are anthropogenically altered estuarine systems that drain into the coastal
ocean directly over shoreline beaches. To address oyster reef losses, a series of shell bag
reefs were constructed within multiple swash tidal creeks. Reefs also were constructed in
tidal creeks associated with estuaries directly connected to the ocean by an inlet. The
purpose of this study was to compare nekton usage of newly constructed restored oyster
reefs, prior to significant spat settlement, between the two creek systems. Baited minnow
traps, gill nets, and pull traps were used to sample nekton on or next to restored reefs.
Limited natural reefs and variable water quality within swash creeks suggested that reef
associated nekton would be different in swash than in inlet creeks. Small fishes (< 120
mm standard length, ≈ 97% Pinfish and Mummichog) were captured in greater number
and condition from restored swash reefs than inlet reefs. Decapod abundance and grass
shrimp condition trended inversely with small fish abundance and condition, suggesting
inlet reefs may provide decapods with higher quality habitat. Larger, piscivorous fishes
were captured more frequently from inlet than swash creeks, but they were likely not
reliant on restored reef substrate. Nekton numbers were also dependent on tidal elevation
(intertidal/subtidal) and diel stage (day/night). Fishes captured by intertidal traps directly
over restored reefs were smaller on average than those captured by subtidal traps just off
the reefs, suggesting shell bags provided nursery habitat even before the restoration site
became a more established reef with the settlement of spat. Rarefied richness of both
v
fishes and decapods was greatest at inlet creeks. A habitat abundance study was
conducted after a residency period of one year to test trap efficacy and attraction of
nekton to restored reefs compared to natural reefs and mudflat substrate. Catches of small
fishes at restored reefs failed to converge with natural reef catches, but were still greater
than from mudflats. Decapods were most abundant at restored reefs, suggesting new
reefs provide them with excellent short-term habitat. Piscivore density was greatest over
mudflat substrate, indicating their indifference to early reefs. Nekton colonization of
constructed oyster reefs occurred rapidly (<6 months) in both inlet and swash tidal creeks
before any appreciable new recruitment and growth of oysters. The abundance and
richness of reef nekton reflected overall differences in the sizes of nekton populations
between inlet and swash creeks. Even during early stages of an oyster reef restoration
project, constructed shell bag reefs provide attractive habitat for many fishes and
decapods critical to estuarine food webs.
vi
Table of Contents
Acknowledgements…………….…………………………………………………....….…ii
Abstract…………………………………………………………………………………...iv
Table of Contents……………………………………………………..……………...…..vi
List of Tables…………………………………………………………...………………..vii
List of Figures………………………………………………………………..……….…viii
Introduction………………………………………………………………………..………1
Methods………………………………………………………………..………………..…7
Results…………………………………………………………………..…………..……16
Discussion……………………………………………………………………………..…23
Tables…………………………………………………………………....……………….35
Figures……………………………………………………………….……………..…….41
Literature Cited…………………………………………………………………..………55
Appendix………………………………………………………………………...……….71
vii
List of Tables
Table 1: Latitude and longitude, estimated low and high tide channel width (m), height
of deployment of intertidal and subtidal minnow traps and pull traps relative to mean low
and high tide, intra-site elevation difference between intertidal and subtidal trap minnow
trap deployment, and tidal range at all 8 locations used during all sampling approaches
from May 2013 to July 2014.
Table 2: Sampling schedule for trap deployment and retrieval for approaches 1 and 2 and
the habitat abundance study.
Table 3: Actual richness, rarefied average richness, rarefied median richness, variance,
and upper and lower confidence intervals for fish captured with all trap methods at 3 inlet
and 3 swash tidal creeks along the Grand Strand of SC from May to September 2013.
Table 4: Actual richness, rarefied average richness, rarefied median richness, variance,
and upper and lower confidence intervals for decapods captured with all trap methods at
3 inlet and 3 swash tidal creeks along the Grand Strand of SC from May to September
2013.
Table 5: Count and total dry mass, in grams, of all fish and decapods collected with
all trap methods at 3 swash and 3 inlet sites in South Carolina from May to
September in 2013.
Table 6: Actual richness, rarefied average richness, rarefied median richness, variance,
and upper and lower confidence intervals for fish and decapods captured with all trap
methods over natural reef, restored reef, and mudflat substrate along the Grand Strand
of SC in July 2014.
Table 7: Count and total dry mass, in grams, of all fish and decapods collected with
minnow traps and gill nets over 3 substrate types at North Inlet, SC in July 2014.
Table 8: Total catch of each species with all trap methods at each of the 6 tidal
creeks sampled in this study in 2013.
Table 9: Total catch of each species with all trap methods over three substrate types
at Bly Creek in North Inlet in July 2014.
viii
List of Figures
Figure 1: Map of the 6 tidal creeks sampled in the study. Orange pins represent inlet sites
and red pins represent swash sites.
Figure 2: Photographs of Hog Inlet (a), Murrells Inlet (b), North Inlet (Oyster Landing)
(c), North Inlet (Bly Creek) (d), Whitepoint Swash (e), Singleton Swash (f), Withers
Swash (g) and Briarcliffe Acres (h) taken during initial site surveys.
Figure 3: A 100 bag oyster shell restored reef immediately after deployment at Hog Inlet,
SC.
Figure 4: A pull trap tethered to a pier at Oyster Landing in North Inlet, SC. Three
dimensional oyster shell structure and stabilizing PVC stake are visible.
Figure 5: A gill net placed over restored reef substrate at Bly Creek in North Inlet, SC.
Figure 6: A minnow trap over shell substrate at North Inlet during low tide.
Figure 7: Mean salinity (ppt) and temperature (˚C) at each of the inlet/swash creek
pairings sampled along the Grand Strand of SC from May to September 2013. An
asterisk indicates a significant (p ≤ 0.05) difference between paired bars.
Figure 8: Sampling approach 1 minnow trap fish and decapod CPUE at three inlet and
three swash tidal creeks from May to September 2013, separated by tidal elevation. An
asterisk indicates a significant (p ≤ 0.05) difference between creek type.
Figure 9: Pull trap CPUE of fish (left) and decapods (right) taken from May to
September 2013 from 3 inlet and 3 swash creeks along the Grand Strand of SC. An
asterisk indicates a significant (p ≤ 0.05) difference between paired bars.
Figure 10: Sampling approach 2 minnow trap fish and decapod CPUE from (A) Hog
inlet and Whitepoint swash, (B) Murrells Inlet and Singleton Swash, and (C) North Inlet
and Withers Swash from July to September 2013, separated by day/night stage. An
asterisk indicates a significant (p ≤ 0.05) difference between creek type.
Figure 11: Gill net fish CPUE from (A) Hog inlet and Whitepoint swash and (B)
Murrells Inlet and Singleton Swash from July to September 2013, separated by day/night
stage. An asterisk indicates a significant (p ≤ 0.05) difference between creek type.
Figure 12: Mean standard length (mm), dry weight (g), and Fulton’s condition factor (K)
of all Pinfish collected from the intertidal and subtidal zones of both inlet and swash
creeks in 2013 along the Grand Strand in SC. An asterisk indicates a significant (p ≤ 0.05)
difference between creek type.
ix
Figure 13: Mean standard length (mm), dry weight (g), and Fulton’s condition factor (K)
of all Mummichog collected from the intertidal and subtidal zones of both inlet and
swash creeks in 2013 along the Grand Strand in SC. An asterisk indicates a significant (p
≤ 0.05) difference between creek type.
Figure 14: Mean standard length (mm), dry weight (g), and Fulton’s condition factor (K)
of all grass shrimp collected with all trap types from inlet and swash creeks along the
Grand Strand in SC in 2013. An asterisk indicates a significant (p ≤ 0.05) difference
between creek type.
Figure 15: Habitat abundance study (A) daytime and (B) nighttime minnow trap fish and
decapod CPUE, separated by tidal elevation, over three substrate types at North Inlet, SC
in July 2014.
Figure 16: Gill Net fish CPUE over three substrate types, separated by day/night stage, at
North Inlet, SC in July 2014.
1
Introduction
Tidal creeks, coastal systems affected by lunar tidal cycles, support an array of
transient and resident fauna (Kneib 1997) that engage in complex energy transformations
(Posey et al. 2002) and generate tremendous system-wide productivity. Many species of
fish rely on estuaries during their early stages of development (Beck et al. 2001). The
proximity of estuarine tidal creeks to the ocean enables exchanges (Perillo 2009). Pinfish
(Lagodon rhomboides) are transient creek inhabitants that travel to the coastal ocean to
spawn upon reaching maturity. Larval Pinfish enter inlets and settle into creeks in winter
(Potthoff and Allen 2003). Resident species such as the Marsh Mummichog (Fundulus
heteroclitus) maintain small home ranges (Lotrich 1975) and high site fidelity (Teo and
Able 2003, Vastano et al. 2017) to tidal creeks as adults, where they spawn during
summer months in the intertidal zone (Kneib 1986). Mummichog are thought to provide
an important energy link between the marsh surface, where they feed at high tide, and
subtidal food chains (Weisberg and Lotrich 1982). Both species consume decapods (e.g.
grass shrimp of genus Palaemonetes) which serve as an energy link to the decomposer,
producer, and primary consumer trophic levels (Odum and Heald 1972, Morgan 1980,
Anderson 1985, McPhee et al. 2015). Small fishes are in turn consumed by piscivores
including birds, large fishes such as Atlantic Sharpnose Sharks (Rhizoprionodon
terraenovae), and large decapods such as blue crabs (Callinectes sapidus). These
piscivores frequently travel and transfer energy between tidal creeks and the adjacent
marine environment (McPhee et al. 2015). Estuarine marshes and tidal creeks rank
among the world’s most productive ecosystems (Odum 1961). The North Inlet estuary, an
unpolluted system in South Carolina, has primary productivity values ranging from 800
2
to 2,100 g per m2 yr-1 (Dame et al. 2000), and other estuaries along the Southeast coast of
the United States have primary productivity values ranging from 400-4,400 g per m2 yr-1
(Dame et al. 2000). High secondary productivity is also characteristic of estuaries.
Mummichog boast one of the greatest reported minimum productivity values of all fishes:
40.7 g-2 yr-1 (Meredith 1979).
Tidal creeks provide a significant amount of habitat complexity, which has a
corroborating influence on fish and decapod abundance and diversity (Walters et al.
2012). Juvenile fishes congregate in the shallow, vegetated edges of marsh systems for
protection (Nevins et al. 2014) where high productivity levels contribute a steady supply
of nourishment. Grass shrimp often reside near macrophytes, where their coloration helps
them avoid predators (Faxon 1879, Coen et al. 1981, Heck and Thoman 1981), or
congregate over oyster shell substrate, which provides food and shelter (Thorp 1976,
Posey et al. 1999). Oyster reefs in tidal creeks are increasingly recognized as essential
fish habitat (Coen et al. 1999b, Minello 1999, Coen and Grizzle 2007) and are a primary
source of intertidal structure. Many smaller fishes, including the skilletfish, oyster
toadfish, naked goby, and several blenny species require oyster shell substrate for
spawning, making reefs indispensable for their survival (Lehnert and Allen 2002). Larger
species such as Red Drum (Sciaenops ocellatus) and Spotted Seatrout (Cynoscion
nebulosus), which do not directly benefit from reef structure as adults, have seen their
populations decline due to reduction of oyster reef available for spawning and nursery
habitat (Swingle et al. 1984, Grabowski et al. 2005).
Oyster reefs contribute to healthy estuarine conditions. Live oysters are filter feeders;
therefore, a healthy population of native oysters also has a significant positive influence
3
on water quality (Dame 1996). Downstream concentrations of TSS and chlorophyll α can
be reduced significantly by small transplanted oyster reefs (Nelson et al. 2004). Feces
and pseudofeces from oysters create denitrification-inducing biodeposits (Newell et al.
2012). Oyster reefs, through the introduction of hard substrate, are hypothesized to
reduce erosion in the intertidal zone and protect salt marsh habitat (Meyer et al. 1997,
Piazza et al. 2005). Finally, vertical relief offered by reefs encourages recruitment of
oyster spat (Knights et al. 2012) and certain larval fish species (e.g. the Naked Goby
(Gobiosoma bosc)) by reducing flow rates and turbulence (Breitburg et al. 1995).
Globally, >85% of coastal oyster populations have disappeared (Beck et al. 2011). In
the U.S., oyster reef restoration programs have become increasingly popular over the past
decade (Peterson et al. 2003, Nelson et al. 2004, Luckenbach et al. 2005, Nestlerode et al.
2007, Geraldi et al. 2009). Funding provided by agencies such as FishAmerica and
NOAA’s Community Restoration Program has helped initiate the construction of
numerous reefs in tidal creeks for use in scientific studies (Peterson et al. 2003, Nelson et
al. 2004, Luckenbach et al. 2005 Nestlerode et al. 2007, Geraldi et al. 2009). Agencies
such as the South Carolina Department of Health and Environmental Control carry out
reef building projects and non-governmental, non-profit, or volunteer organizations such
as Coastal Oyster Recycling and Restoration Initiative (CORRI), Tampa Bay Watch,
Chesapeake Bay Foundation, North Carolina Coastal Federation, and many more actively
build reefs for the benefit of the community (Luckenbach et al. 2005). Increased oyster
shell density and height within experimental, restored reef habitat has been found to be
positively correlated with key community metrics, such as finfish, crabs (e.g.
Eurypanopeus depressus), marsh mussels (e.g. Geukensia demissa) and mud crabs
4
(Panopeus herbstii) (Luckenbach et al. 2005). Introduction of only 10 m2 of restored
oyster reef habitat could increase production by fish and motile crustaceans at an average
Southeastern U.S. reef habitat by 2.6 kg yr-1 for the duration of the reef’s functional
lifespan by providing spatial refuge and additional habitat for recruitment limited species
(Grabowski 2002). In many cases, colonization and faunal use of restored reefs has been
shown to occur quickly (Coen et al. 1999b, Lenihan et al. 2001, Peterson et al. 2003,
Grabowski et al. 2005, Humphries and La Peyre 2015), but few studies have been
conducted to show whether restored reef communities further converge with natural reef
communities as residence time increases (Grabowski et al. 2005). Walters and Coen
(2007) found that even 7 years was not enough for complete convergence of taxa use
between certain natural and restored reefs. Unfortunately, many studies have neglected to
monitor reefs, resulting in a dearth of data and unanswered questions regarding reef
restoration efficacy (Kennedy et al. 2011). In the past, lack of success criteria for reef
restoration projects confounded monitoring efforts (Luckenbach et al. 2005, Walters and
Coen 2007), but recently there have been attempts to establish metrics for the success of
reef restoration projects (Baggett et al. 2015).
Locally, along the northern South Carolina coastline between Georgetown and Little
River, tidal creeks either are part of inlet systems that connect to the ocean through
channels or swash systems that empty directly into the ocean over beach shorelines. Inlet
creeks typically have not been modified to collect and transport stormwater runoff, but all
swash systems have been adjusted by local municipalities to improve coastal drainage
and move rainfall offshore. In 1936, all direct connections between swash and swamp
systems were altered by the construction of a 38.6 km drainage ditch from Little River to
5
the Waccamaw River. After World War II swashes were incorporated into city and
county drainage systems for the management of stormwater, necessary largely due to
increases in residential and commercial development in the Waccamaw area (Wessel and
Walters, in prep). Urbanization commonly associated with swash creeks further increases
inputs from runoff due to a greater concentration of impervious surfaces compared to
forested creeks (Walsh et al. 2009), resulting in salinity fluctuations (e.g., Ahn 2005).
Oysters are unable to survive low (<7 ppt) salinities for more than four days (Wells 1961)
and therefore can be challenged by frequent stormwater input. Runoff can also increase
concentrations of non-point source pollution (Ahn 2005) such as pesticides (Sanger 2004),
polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs) (Van
Dolah et al. 2008). PCBs and PAHs remain in creek systems for particularly long periods
of time, resulting in continuous exposure to organisms (Teal et al. 1992). Coliform or
Enterococci bacteria can also enter creek systems as a result stormwater input. The EPA
in conjunction with the Clean Water Act of 1972 classifies water bodies as either
impaired, containing greater than allowable concentrations of bacteria, or unimpaired. All
local swashes currently are included on the EPA’s 303(d) list of impaired waters.
Excessive nutrient loading from runoff in swash systems can lead to spikes in primary
productivity caused by the appearance of algal blooms, which in turn can result in
hypoxic or anoxic conditions (Carpenter et al. 1998, Paerl et al. 1998, Anderson et al.
2002). Past studies have shown that nitrogen is of particular concern, as the primary
productivity of most estuarine systems is nitrogen limited (Mallin et al. 1993). Estuarine
systems typically act as filters for nutrients traveling downstream, gradually reducing
loads through biological processes (Jaworski 1981). However, if nutrient loads exceed
6
manageable levels for estuarine systems, fish and bivalve kills are known to occur
(Carpenter et al. 1998, Paerl et al. 1998)
More than 90% of commercially important fish species are dependent on estuaries
at some point in their life histories (Chambers 1992), and fish are considered important
indicators of aquatic ecosystem health (Adams et al. 1993, Sheaves et al. 2012). Fish
must expend energy which would otherwise be used for reproduction and growth to deal
with stress (Barton and Schreck 1987, Wedemeyer et al. 1990, Adams et al. 1993). Water
bodies of substandard condition are unlikely to support fish populations as successful as
clean water bodies (Horn et al. 1999, Wedge et al. 2005). Differences in water quality
(Wessel & Walters in prep) and oyster reef density (pers. obs.) exist between swash and
inlet systems, but comparisons of constructed reefs or communities at inlet and swash
creeks have not been conducted. The relationship between anthropogenic impacts and
nekton abundance is not well understood (Partyka and Peterson 2008).
The purpose of this study was to determine whether significant density, size,
condition, and diversity differences exist between assemblages of fish and decapods that
utilize newly restored oyster shell structure at three inlet and three swash type 3 tidal
creek systems along the Grand Strand (coastal Georgetown and Horry counties in South
Carolina). Faunal density and size metrics are listed as appropriate data to quantify
success of habitat enhancement in recent restoration guidelines (Baggett et al. 2015). Fish
and decapod abundances were compared at two tidal elevations and at day and night. A
habitat abundance study tested density and diversity differences of nekton utilizing
natural reef, restored reef (with 1 year residency), and mudflat substrates at Bly Creek in
North Inlet and effectiveness of different trap types at capturing reef associated species.
7
Methods
The study was conducted between the cities of Georgetown and Little River,
South Carolina USA along the 77 km coastline known as the Grand Strand (Fig. 1). Reef
construction and subsequent nekton sampling occurred within multiple haphazardly
selected inlet and swash tidal creek systems (Fig. 2). Inlet (n = 3) creeks ranged from
minimally (North Inlet) to extensively (Murrells and Hog Inlets) impacted by human
activities. Each inlet is ocean dominated with tidal creek salinities typically ranging
above 30 psu. Swash (n = 3) creeks such as Whitepoint, Singleton, and Withers Swashes
are part of less aerially expansive estuarine systems, all of which are modified
extensively as part of local stormwater management systems. Salinities can drop below
10 psu during storm events (Walters unpubl.). All inlet and swash creeks contained
resident oyster populations, but the extent of reefs both in terms of coverage area and
population densities is several orders of magnitude greater within inlet creeks (Walters
unpubl.). Spartina marsh predominated within the intertidal areas of both inlet and swash
creeks.
Two intertidal sites were selected within each of the inlet and swash tidal creeks for
oyster reef construction. Reefs were constructed by the Coastal Oyster Recycling and
Restoration Initiative (CORRI, www.corri.org). Each reef consisted of 100 bags (~ 19 L
ea.) of recycled oyster shell from local restaurants laid out in a 10 x 10 formation
centered at or just landward of mean low water (Fig. 3). A 2005 study by Grabowski et al.
found reefs constructed over mudflat substrate received more recruitment and ultimately
produced more living oysters than reefs constructed over other sediment types. All reefs
were placed on gradually sloped, depositional banks which were not adjacent to steeper
8
erosional banks. Evidence suggests depositional banks generally provide greater
availability of benthic invertebrate prey (McIvor and Odum 1988, Miltner et al. 1995)
that tends to attract fish moving onto intertidal habitats with the incoming tide (McIvor
and Odum 1988).
Physical characteristics for each tidal creek and constructed reef site were
collected. Geographic coordinates for constructed reefs were determined using a
handheld geographical positioning system (GPS) device (Table 1). Channel width at the
specific sampling location within each tidal creek during high and low tides was
estimated to the nearest meter using Google Earth©. Depths of trap inundation relative to
mean high and low tides were measured from the water’s surface to the top of each trap
using a level.
Several nekton sampling approaches were employed to more completely describe
catch diversity and numbers to address potential sampling biases (Table 2). Selected
sampling approaches were passive (e.g., traps, fixed nets) compared to the more active
and commonly employed towed sampling approaches (e.g., seines, trawls). Passive
sampling devices frequently result in substantially greater small nekton catch efficiencies
in estuarine environments (Rozas and Minello 1997). Difficulties with passive devices
include species-specific behaviors that often result in gear-related attraction or avoidance
(Rozas and Minello 1997, Lehnert and Allen 2002, Hagan and Able 2003). For example,
Pinfish are captured ineffectively by pull traps but easily by minnow traps (Lehnert and
Allen 2002). All sampling took place in summer months when fish densities and diversity
increase as the conditions become more suitable for transient fish species to occupy tidal
creeks (Lehnert and Allen 2002). In winter, fishes tend to occupy deep waters which have
9
high salinity and are more thermally stable than shallow waters. In summer, juveniles of
transient species occupy shallow waters and intertidal traps capture more fish (Lehnert
and Allen 2002). Pull traps, gill nets, and minnow traps were chosen to collect oyster reef
associated nekton during two different sampling approaches.
During the first sampling approach, pull traps (Fig. 4) consisting of a basket filled
with oyster shell, similar in design to Lehnert and Allen (2002), were constructed to
sample fauna that colonized the constructed reefs in each tidal creek system. Natural
faunal abundance, diversity, and biomass typically are greater on oyster reefs compared
to other intertidal habitats (Taylor and Bushek 2008, Grabowski et al. 2005, Manley et al.
2010, Kingsley-Smith et al. 2012), and pull traps incorporating shell are an effective
approach for sampling fauna associated with oyster shell substrate (Lehnert and Allen
2002). Traps consisted of 1 m x 1 m x 0.2 m of 2.5 cm PVC pipe covered in two layers of
mesh (2.5 and 0.64 cm). They were initially filled with 12 kg of loose oyster shell to
attract organisms with structure, but loss of cultch during trap collection necessitated
development of an alternative method to retain shell in traps. Mats consisting of an 80 cm
x 35 cm section of 2.5 cm mesh were fabricated and a 38 cm x 20 cm mesh cylinder was
attached centrally. Oyster shell, ~4kg, was cable-tied onto the mat and cylinder and two
mats were attached to the bottom of each trap before each sampling effort. Between May
25 and October 13, 2013 one pull trap was placed every four weeks within each of the
three selected inlet and swash tidal creeks on a low tide. Pull traps were deployed as close
to constructed reefs as possible. At Whitepoint Swash, recognition that current velocities
were affecting sampling results necessitated relocation further upstream within the swash
from the North Myrtle Beach 48th Ave. Bridge to near Briarcliffe Acres. At North Inlet,
10
lack of available structure to which pull traps could be tied and pulled vertically required
relocation from Bly Creek to Oyster Landing pier (Table 1, Fig. 2). Based on results from
subtidal sampling (Lehnert and Allen 2002), pull traps were left to soak up to 7 days
before retrieval and pulled at high tide. The total distance between sampling locations (77
km) and extended effort to retrieve each pull trap necessitated staggering trap collection.
Pull traps at the 3 southernmost sites (Withers Swash, Murrells Inlet, and North Inlet)
were collected after 6 days and traps at the 3 northernmost sites (Hog Inlet, White Point
and Singleton Swash) were collected after 7 days.
In the second sampling approach, gill nets (Fig. 5) 1.5 by 9 m with a mesh size of 2.5
cm were used to sample the larger, demersal nekton swimming over constructed reefs.
The selection of gill nets was based on an ability to sample an entire constructed reef
(Lenihan et al. 2001) and capture large adult nekton associated with oyster reefs
(Grabowski et al. 2005, Peterson et al. 2003). Between June 27 and September 15, 2013
gill nets were erected over constructed reefs within a single inlet and swash tidal creek
site every two weeks and fished for a total of 4 days during each deployment. Distances
between tidal creeks and sampling efforts required by each net limited deployment to
only a single tidal creek pair every sampling period. Pairing inlet and swash sites enabled
catch comparisons between the different types of tidal creeks. Nets were assembled in a
V-shape over reefs with the opening of the V facing the middle of the tidal creek, a
design reported to result in greater catch rates (Grabowski et al. 2005). Nets were held in
place by three 2.7 m long, 2.54 cm thick PVC poles. Deployment and retrieval of nets
were planned around tidal stage and sunset-sunrise to permit sampling one daylight and
one nighttime high tide each day. Each set of paired sites (Hog Inlet-Whitepoint Swash,
11
Murrells Inlet-Singleton Swash, and North Inlet-Withers Swash) was visited twice over
the course of the summer.
Minnow traps (Fig. 6) of 0.64 cm galvanized mesh with a 1.9 cm opening on each
end were also deployed during both sampling efforts. Each trap was baited with 0.11 L of
commercial dog food (fish attractant) prior to deployment (Able et al. 2015). Baited
minnow traps are effective at capturing baitfish in South Carolina tidal systems.
Mummichog and Pinfish catch per unit effort (CPUE) from traps provided reasonable
estimates of actual field abundances (Meyer 2006). At low tide, baited traps were
positioned intertidally on constructed reefs or subtidally just off constructed reefs. All
traps were collected on the next low tide. Each sample encompassed a single high tide.
Traps were deployed in conjunction with pull trap (6 nighttime samples) and gill net
deployment (6 daytime and 8 nighttime samples).
All fish and decapods collected in samples were placed in plastic bags, labeled, and
transported on ice back to the lab where samples were transferred to a -20 °C freezer and
stored until processed. Sampled fish, polychaetes, and decapods were identified to
species where possible. Standard length, fork length, and total length for all fish species
and standard carapace length and width for decapods were measured with a digital caliper
to the nearest 0.1 mm. Only total length was measured for other taxa. Dry mass was
determined to the nearest 0.001 g for each specimen after placing it in a drying oven at 60
ºC for between 24 to 240 hrs. Species with visible differences between genders (e.g. F.
heteroclitus, C. sapidus) were sexed; species requiring dissection to determine gender
were not sexed.
12
Data were analyzed with general linear model (GLM), generalized linear model
(GZLM), or expected species richness (ESR) approaches depending on design and data
characteristics. A GLM approach was used to analyze physical data collected from tidal
creeks (e.g., temperature, salinity). Differences between inlet or swash creeks (treatment)
in tidal channel width and depth, and in trap tidal elevation were analyzed with a one-way
ANOVA. Inlet and swash temperature and salinity differences were analyzed with a
blocked ANOVA, sampling date being the blocking factor. Temperature and salinity
differences were analyzed according to site pair to account for seasonal variation in
temperature and rainfall induced salinity fluctuation. All data were tested for violation of
critical ANOVA assumptions and, where necessary, appropriately transformed (e.g.,
Keene 1995, Tabachnick and Fidel 1996, Quinn and Keough 2002).
All catch per unit effort (CPUE) data from minnow traps, pull traps, and gill nets
were analyzed with a GZLM to account for over-dispersed count data, a prevalence of 0’s,
and a failure to satisfy basic GLM (e.g., ANOVA) assumptions. Maximum likelihood
estimation was employed for each GZLM fitting a negative binomial distribution using a
log link function (Garson 2012). A negative binomial model was chosen since observed
0’s likely represent true 0’s (Garson 2012) while the log link was applied for general ease
of interpretation and comparison to commonly seen poisson distributions (Garson 2012).
Model fit was determined from the omnibus likelihood ratio χ2 and consideration of the
deviance ratio (Garson 2012).
During sampling approach 1 (pull traps, minnow traps), creek type was the treatment
level of interest with individual sites serving as replicates and dates of collection as a
blocking factor. Pull trap dependent variables fish and decapod CPUE were separately
13
analyzed by the factor creek type. Minnow trap dependent variables fish and decapod
CPUE were separately analyzed by factors creek type, tidal elevation, and interactions
between factors.
During sampling approach 2 (gill nets, minnow traps), diel stage was the treatment
level of interest with different days serving as replicates and sampling set as a blocking
factor. Gill net dependent variable fish CPUE was analyzed by factors creek type, diel
stage, and interactions between factors. Minnow trap dependent variables fish and
decapod CPUE were separately analyzed by factors creek type, diel stage, tidal elevation,
and all possible interactions between two or three factors.
Standard length, dry mass, and Fulton’s condition index for Pinfish and
Mummichog collected in minnow traps and grass shrimp collected in all samples were
analyzed after generating size frequency distributions. Fulton’s condition factor is a
metric for describing “condition” that compares a fish’s actual weight/length relationship
to an expected weight/length relationship, which consists of the formula K = a(W/L3).
Generally, a scaling factor (a) is applied to bring the expected value of K to 1 (Nash et al.
2006); in this study, the scaling factor is set to 100,000. Since Fulton’s K does not
account for differences in body shape, K ranges will differ between species. Therefore,
comparisons of fish condition are only informative within the same species (Barrett et al.
2014). Greater values of Fulton’s condition factor (K) are frequently associated with
greater reproductive success and fecundity and less skipped spawning events (Morgan
2004). It may also indicate increased tolerance of environmental stressors and ability to
capture prey (Barrett 2014). Additionally, overall ecosystem health may be affected by
the condition of small intertidal species such as Pinfish and Mummichog due to the
14
presence of predators such as bluefish, small sharks, and commercially significant species
like flounder (Potthoff and Allen 2003) which may themselves achieve greater condition
by consuming prey of greater condition. Similarly, poor prey condition may force
predators to expand their diets, which may variably affect an ecosystem as a whole
(Barrett 2014). Dry weight is used for W to eliminate unnecessary variability in fish
body mass due to water retention, while standard length is chosen for L so fish with
damaged fins would not erroneously receive increased fish condition scores.
Small sample sizes that were collected from some sampling approaches yielded
unreliable results. Because multiple comparisons increase the chance of a false positive
(Curran-Everett 2000, McDonald 2014), collections from sampling approaches 1 and 2
were analyzed together to produce a sufficiently large sample set. The number of size
classes for each taxon was determined by Sturge’s rule (D = 1 + 3.3 x log[(n(n-1))/2]).
Mean and median size, mode size class, and size range were calculated. A GZLM with
gamma distribution and log link function was applied to test for size differences between
inlet and swash creeks. The gamma distribution is a common choice for continuous data
that are positively skewed (e.g. size frequency distributions) (Garson 2012). Dependent
variables standard length, dry mass, and condition of Pinfish and Mummichog were
analyzed by factors creek type, tidal elevation, and interactions between factors. Total
length, dry mass, and condition of grass shrimp were analyzed by creek type. Individual
sites served as replicates within the treatment level of interest, creek type.
Species richness differences between inlet and swash creeks were analyzed after
correcting for sample size effects (Gotelli and Colwell 2001). ECOSIM (Gotelli and
Entsminger 2000) was used to calculate expected species numbers and 95% confidence
15
intervals. Rarefied average fish and decapod species richness were compared between the
six tidal creeks sampled in 2013.
To evaluate the effectiveness of restored reefs to attract organisms compared to
natural reefs and mudflat substrate at Bly Creek in North Inlet, a separate habitat
abundance study was undertaken in July 2014, one year after initial sampling. Another
goal of the study was to test whether minnow traps and gill nets accurately sample
restored reef associated nekton. Gill nets were initially set up over each of the different
habitats on a late afternoon low tide. Two minnow traps (one intertidal, one subtidal)
were placed in the mouth of each gill net. All traps were sampled and redeployed on 7
consecutive low tides over the next 4 days resulting in 4 nighttime and 3 daytime
collections. Data were analyzed similarly to approach 2, but substrate type replaced creek
type as the treatment level of interest and the blocking factor was removed. Rarefied
average fish and decapod species richness was compared between the three substrate
types at Bly Creek in 2014.
16
Results
Physical Creek Characteristics and Trap Elevations (May – September 2013)
Channel width and tidal range differed by creek type for sites visited during
approaches 1 and 2. Inlet channels were on average ≈ 6x wider than swash channels at
low tide (Table 1). Inlet channels also had a tidal range ≈ 3x greater than swash creeks,
which resulted in minnow and pull traps resting deeper in the water column at high tide.
In sampling approach 1, intertidal minnow, pull, and subtidal minnow traps at inlet creeks
rested approximately 47 cm, 59 cm, and 89 cm deeper on average at high tide,
respectively, than their swash counterparts (Table 1). Similar trends were observed
during approach 2; intertidal and subtidal traps rested nearly 31 cm and 49 cm deeper at
high tide, respectively, at inlet creeks than at swash creeks (Table 1). Creek type exerted
no significant influence on intertidal or subtidal trap elevation at low tide.
Mean channel width and salinity varied by creek type and occasionally month in
sampling approach 2 while water temperature varied only by month at one site pair.
Creek type affected salinity at the HI/WP pair, MI/SS pair, and NI/WS pair. Mean
salinity at swash creeks was typically 5-10 psu lower than at inlet creeks (Fig. 7). Salinity
also fluctuated much further from the mean at swash creeks; standard deviation at swash
creeks was more than twice that of inlet creeks. Salinity dropped as low as 5 psu at swash
creeks without ever falling below 20 psu at inlet creeks. Creek type did not affect
temperature at the HI/WP, MI/SS, or NI/WS pairs (Fig. 7). Monthly temperature
variation was significant only in the MI/SS pair.
17
Sampling Approach 1 (May - September 2013)
Numbers of fishes collected from minnow traps in sampling approach 1 differed
by creek type, which also significantly interacted with tidal elevation. On average nearly
4x more fish were captured by swash compared to inlet minnow traps (Fig. 8). Tidal
elevation did not affect fish catch on its own, but the interaction between elevation and
creek type did influence fish density. At swash creeks, subtidal traps returned slightly
more fish than intertidal traps, but at inlet creeks, intertidal traps returned nearly 4x as
many fish as subtidal traps (Fig. 8). Pinfish and Mummichog comprised 97% of total
catch.
As with fishes, minnow trap returns of decapods differed by creek type, which
significantly interacted with tidal elevation. Decapod catch per unit effort (CPUE)
trended inversely with fish CPUE; inlet creeks returned nearly 9x as many decapods as
swash creeks on average (Fig. 8). Despite a lack of variation by elevation within overall
decapod catch, decapod CPUE was affected by the interaction between creek type and
elevation. Mean CPUE was more than 2.5x greater from subtidal than intertidal traps at
inlet creeks, but CPUE was 10x greater from intertidal than subtidal traps at swash creeks
(Fig. 8); however, elevation differences in swash creeks may be exaggerated due to low
decapod swash captures of < 1 per trap. Grass shrimp were the most abundant captured
decapods, making up 82% of total minnow trap decapod catch.
Numbers of fishes returned by pull traps significantly varied by creek type but
numbers of decapods did not. Swash traps returned, on average, > 3x as many fishes as
inlet traps. (Fig. 9). Naked Gobies, Pinfish, and Atlantic Silversides were the most
commonly captured species, together constituting ≈ 78% of overall pull trap fish catch.
18
Overall decapod CPUE was much greater on average than fish CPUE, and as with
minnow traps, CPUE was greater at inlet than in swash creeks. Grass shrimp constituted
nearly 90% of overall decapod catch from pull traps.
Sampling Approach 2 (June – September 2013)
Mean fish CPUE in sampling approach 2 varied by creek type at all site pairs
(which significantly interacted with tidal elevation at most pairs), sample set at most pairs,
and diel stage at only one site pair. Creek type affected fish density in the HI/WP, MI/SS,
and NI/WS pairs. Fish catch was 13-14x greater at the swash than inlet site in the HI/WP
and NI/WS pairs, but in the MI/SS creek pair, inlet catch was about 3x greater than swash
catch (Fig. 10). Elevation on its own was not a strong indicator of fish density at the
HI/WP, MI/SS, or NI/WS creek pairs, but its interaction with creek type had an impact at
the HI/WP and MI/SS creek pairs. At Hog Inlet, intertidal traps captured more than 2x as
many fish as subtidal traps but at Whitepoint Swash, subtidal traps returned nearly 5x as
many fish as intertidal traps. In the MI/SS creek pair this trend was reversed; Intertidal
traps returned 5x more fish than subtidal traps at the swash creek and 5x fewer at the inlet
creek (Fig. 10). The interaction between creek type and trap elevation was not a strong
predictor of fish catch at the NI/WS site pair. Fish numbers differed with diel stage at the
MI/SS creek pair (where 1.2x more fish were trapped at night than at day) but not the
HI/WP or NI/WS pairs. As seen in sampling approach 1, Pinfish and Mummichog
comprised 97% of total minnow trap catch.
Numbers of decapods captured by minnow traps were affected by creek type at
most pairs, diel stage at all pairs, and elevation and sampling set at one pair each. Creek
19
type impacted decapod minnow trap CPUE at the HI/WP and NI/WS pairs, where inlet
catch was approximately 4x and 100x greater, respectively, at inlet sites (Fig. 10). No
CPUE difference was detected between Murrells Inlet and Singleton Swash. Decapod
CPUE varied with diel stage at the HI/WP, MI/SS, and NI/WS with night catch
exceeding day catch by about 3-6x at all creek pairs. Elevation affected CPUE at the
HI/WP pair (where it was about 2.5x greater from intertidal than subtidal traps (Fig. 10)),
but elevation’s interaction with creek type did not. At the MI/SS creek pair, the
interaction between creek type and elevation affected CPUE but elevation alone did not.
At Murrells Inlet, subtidal traps returned approximately 2x as many decapods as intertidal
traps; at Singleton swash, this trend was reversed (Fig. 10). Neither elevation nor its
interaction with creek type affected decapod CPUE at the NI/WS creek pair. Grass
shrimp were the most numerous species, comprising 65% of captured decapods. White
shrimp (≈25%) and brown shrimp (≈5%) were the next most abundant decapods.
Numbers of fishes trapped by gill nets were affected by creek type but not
sampling set or diel stage. At the NI/WS creek pair, fish CPUE varied with creek type;
more than 7x as many fish were captured from inlet nets than swash nets (Fig. 11).
Despite a similar trend at the HI/WP creek pair, variation was not statistically significant,
likely obscured by low (<1) fish CPUE. Fish catch was noticeably greater during the
night than the day at most creeks, but these differences were not statistically significant.
The MI/SS creek pair returned a gill net CPUE of zero and therefore no statistical
analysis was administered. Atlantic Sharpnose Sharks were the most commonly captured
species, constituting 45% of gill net catch. Spot (Leiostomus xanthurus), large tooth
flounders (Paralichthys sp.), Ladyfish (Elops saurus), Silver Perch (Bairdiella
20
chrysoura), Flathead Grey Mullet (Mugil cephalus), and Pinfish comprised an additional
49% of total gill net catch.
Standard Length, Dry Weight, Condition, Species Richness (May – September 2013)
Creek type and tidal elevation affected mean standard length, dry weight, and
condition (K) for Pinfish collected from minnow traps during sampling approaches 1 and
2. At inlet creeks, trapped Pinfish had a greater standard length by about 3.3 mm and a
greater dry mass by about 0.1 g than at swash creeks. However, mean Pinfish condition
was greater at swash creeks compared to inlet creeks by nearly 10% (Fig. 12). Tidal
elevation also impacted Pinfish mean standard length and dry mass; Pinfish collected
from subtidal traps were approximately 1.7 mm longer and 0.09 g heaver than from
intertidal traps. Though trap elevation alone did not affect mean Pinfish condition, the
interaction between creek type and elevation did have an effect. Pinfish averaged 2%
greater condition when trapped by intertidal traps compared to subtidal traps at inlet
creeks while specimens averaged 5% greater condition when collected from subtidal
traps than intertidal traps at swash creeks (Fig. 12).
Creek type and tidal elevation affected average standard length, dry mass, and
condition for Mummichog. Mummichog captured from swash creeks presented a greater
average standard length by about 1.7 mm, dry mass by about 0.14 g, and condition by
about 1% than Mummichog from inlet creeks (Fig. 13). Elevation also had an impact on
mean Mummichog standard length and dry mass. Mummichog taken from subtidal traps
typically were 1.7 mm longer and 0.03 g heavier than Mummichog from intertidal traps
(Fig. 13). While condition did not differ by trap elevation, it was affected by the
21
interaction between elevation and creek type. At inlet creeks, Mummichog captured from
intertidal traps exhibited approximately 12% greater condition when compared to subtidal
traps, while at swash creeks, Mummichog collected from subtidal traps exhibited nearly
17% greater condition than from intertidal traps (Fig. 13).
Mean total length, dry mass, and condition of grass shrimp varied according to
creek type. Inlet creeks housed larger grass shrimp of greater total length and dry mass.
On average, specimens from inlet creeks were about 2.5 mm longer and 0.02 g heavier.
Inlet creeks also boasted better shrimp condition compared to swash creeks by almost
11% (Fig. 14).
Rarefied average richness and total trapped biomass of fishes and decapods were
greater at inlet than swash creeks. Inlet traps returned a 90% greater fish richness and a
28% greater decapod richness than swash traps (Table 3). Fish biomass collected from
inlet traps was nearly 1.5x greater than from swash creeks, while decapod biomass was
more than 2x greater at inlet than swash creeks (Table 4).
Habitat abundance study (July 2014)
Minnow trap fish collection numbers were affected by substrate type and tidal
elevation, but not diel stage. Fish density over natural reef substrate was nearly 10x
greater than over restored reef substrate and 40x greater than over mudflat substrate.
Minnow trap returns were also influenced by trap elevation, with intertidal traps
capturing more than 5x as many fish as subtidal traps (Fig. 15). Diel stage, however, did
not affect minnow trap capture rates. As with sampling approaches 1 and 2, Pinfish and
Mummichog comprised 97% of minnow trap fish catch.
22
Decapod minnow trap CPUE significantly differed according to diel stage but not
substrate type or tidal elevation. Nighttime decapod CPUE was approximately 6x greater
than daytime CPUE (Fig. 15). Despite greater decapod catch from restored reefs and the
subtidal zone, small sample size and significant trap by trap catch variability may have
obscured the statistical significance of substrate type’s and tidal elevation’s effects on
decapod density. Grass shrimp were the most abundant species (≈71%), followed by
white shrimp (≈11%) and brown shrimp (≈8%)
Mean gill net CPUE was not affected by substrate type or diel stage. Despite
greater gill net catch at night and over mudflat substrate, low per trap catch numbers (1-2
fish per net average) and a limited sampling period obscured whether differences were
simply random or conveyed actual significance (Fig. 16). Bluefish (Pomatomus saltatrix)
(≈32%) and Atlantic Sharpnose Sharks (≈26%) were captured most frequently, followed
by Atlantic Menhaden (Brevoortia tyrannus) (≈11%) and Ladyfish (≈9%).
23
Discussion
Newly constructed oyster shell bag reefs attracted baitfish in greater numbers and
better physical condition within swash compared to inlet tidal creeks. Fishes and
decapods were collected over constructed reefs consisting primarily of loose shell within
months after placement in creeks and before appreciable recruitment and growth of live
oysters occurred. While both reef presence (Hadley et al. 2010, Humphries et al. 2011)
and development of architectural complexity (Luckenbach et al. 2005, Colden 2015, Karp
et al. 2018) are of value to nekton, the relative role of each may depend on the quantity
and quality of available habitat existing in the study system (Turner et al. 2000). At swash
creeks, presence alone appears to be enough to immediately boost fish productivity by
providing refuge for habitat limited species (Peterson et al. 2003, Beck et al. 2011,
Grabowski et al. 2012, Humphries and La Peyre 2015), increasing longevity and growth
of individuals already residing in creeks (Peterson et al. 2003, Kroeger and Guannel 2014,
Humphries and La Peyre 2015). At inlet creeks, abundant natural reefs seemingly
outcompete newly restored reefs as fish habitat (Geraldi 2009). This should not
discourage restoration efforts; an expansive study by Karp et al. (2018) shows that after a
sufficient residency period (5-21 years), constructed reefs can be restored to an ostensibly
natural state. Restoring for live oyster growth and habitat complexity at creeks that
already have a sufficient quantity of available habitat would likely produce a more
optimal outcome.
Trends found in this study may not necessarily hold true for all fish species. It is
important to remember that minnow traps, which provided the most robust data in this
study, target a very narrow range of nekton. Other studies using various trap types have
24
found many more species to be associated with restored oyster reefs (Coen et al. 1999a,
Lehnert and Allen 2002, Peterson et al. 2003, Humphries et al. 2011, Kingsley-Smith et
al. 2012, Kroeger and Guannel 2014, Humphries and La Peyre 2015). Careful
consideration of local context prior to reef construction is essential to meeting restoration
goals (Grabowski 2005, Humphries and La Peyre 2015).
Irrespective of reefs, the observed greater abundance and better physical condition
of swash fishes may reflect broad scale differences in nekton populations between inlet
and swash creeks. Anthropogenic influences (e.g., increased salinity variations, increased
pollutants) on local swash creeks are linked to decreased fish condition (Horn et al. 1999),
size, liver somatic index, and caloric density (Wedge et al. 2005). However, Pinfish and
Mummichog both exhibit physical traits that help them deal with environmental extremes
(Sogard 1997) at swashes. Juvenile Pinfish experience rapid growth even in estuaries of
dynamic salinity between 15-30 ppt (Shervette et al. 2007), while Fundulus species are
capable of quickly activating or deactivating various ion channels (Hoffmann et al. 2002),
adjusting gill epithelium morphology (Daborn et al. 2001, Katoh and Kaneko 2003), and
regulating genes (Weis 2002, Wood and Laurent 2003, Patterson et al. 2012) to deal with
salinity fluctuations. Mummichog also produce embryos resilient to bioaccumulative
toxins when exposed to poor water quality (Nacci et al. 1999, Weis 2002). Physical
differences between specific swash and inlet creeks selected in this study were likely a
bigger factor than fish adaptability to anthropogenic influences. Inlet creeks were wider
and deeper, while swash creeks were narrower and shallower. Inlet creeks also had a
greater tidal range than swash creeks, meaning traps rested deeper in the water column at
inlet than swash creeks at high tide. Allen et al. (2007) found Mummichog and Pinfish
25
are more commonly associated with shallow creeks perched high in the tidal range.
Finally, fish attraction and aggression toward baited traps may vary between inlet and
swash creeks. Pollutants, which are more prevalent at swash creeks, are known to alter
the behavior of resident organisms (Farr 1977, Hartman 1978, Weis and Weis 1995,
Smith and Weis 1997).
Decapod biomass and density and grass shrimp standard length, weight, and
condition were greater at constructed inlet than swash reefs overall, despite abundant
natural reefs. This suggests newly constructed reefs immediately provide high quality
habitat for decapods at inlet creeks (Turner et al. 2000, Humphries et al. 2011, Humphries
and La Peyre 2015). At swash reefs, variable water quality and spikes in salinity can
result from stormwater running off of impervious surfaces typically associated with
developed watersheds; accordingly, nekton abundance could be reduced in urban creeks
(Sanger et al. 2004, Lawless 2008, Washburn and Sanger 2011). Larval grass shrimp are
known to suffer metabolic stress during salinity fluctuations (Anderson 1985). Reduced
decapod catch at restored swash reefs is also likely correlated with increased presence of
common predators such as Mummichog (Harrington and Harrington 1972, Welsh 1975,
Kneib and Stiven 1982, Anderson 1985) and Pinfish (Potthoff and Allen 1993,
Lukzkovich et al. 1995) at the same reefs. Decapods may spend more energy avoiding
predators at swash creeks, negatively affecting metabolic rate (Barton and Schreck 1987,
Wedemeyer et al. 1990). Predation often affects community structure (Schoener 1986);
Mummichog are known to alter local density and distribution of their common prey
(Vince et al. 1976, Kneib 1980). Bretsch and Allen (2006) noted depth of greatest grass
shrimp density trended inversely with depth of greatest Pinfish density. In some instances,
26
grass shrimp may have used the same minnow traps as their predators but were eaten
within the traps by fish which had also been trapped. Minnow traps that returned large
numbers of fish almost never simultaneously returned many decapods and occasionally
contained pieces of grass shrimp. Therefore, minnow traps might not provide accurate
estimates of grass shrimp when fish are also present.
Greater aggregate fish biomass and CPUE of larger, piscivorous fish species from
inlet than swash reefs may be attributable to physical differences between inlet and swash
creeks. These species, which travel frequently between estuaries and the open ocean,
benefit from a larger tidal cycle (McPhee et al. 2015) and have an easier time navigating
the wider channels found at inlet creeks for reef access as adults (Sklar and Browder
1998). At swash creeks, increased development and urbanization not only causes
metabolic stress (Goto and Wallace 2010) but also reduces habitat connectivity (Kennish
2001, Thrush et al. 2008, Rozas et al. 2013, Lowe and Peterson 2014, Rudershausen et al
2016). Larger species are known to achieve greater size, live longer, and benefit
metabolically in the deeper waters found at these inlet creeks (Rypel et al. 2007, Miltner
et al. 1995, Linehan et al. 2001). Many large fishes, such as adult sharks, are restricted
from ever entering shallow coastal waters to hunt for prey (Heithaus 2004), and those that
can, such as Paralichthys sp. flounders (Packer 1999), often make ontogeny-dependent
habitat migrations from shallow mudflats to deep sandy bottoms (Burke et al. 1991)
where the oldest and largest adults remain year-round (Festa 1977). Greater piscivore
abundance at inlet creeks might also help explain reduced small fish catch at inlet reefs.
Small fishes likely expend more time and energy hiding from predators at inlet creeks
(Barton and Schreck 1987, Wedemeyer et al. 1990) and safely foraging at swash creeks
27
(Jordan et al. 1996). Similarly, larger fishes trapped in gill nets at inlet creeks may have
discouraged smaller fishes from using proximal minnow traps, potentially biasing CPUE
results.
Tidal elevation trends seen in small fishes and decapods may have been
influenced by piscivore abundance. At inlet creeks, increased predation pressure means
small fishes likely find increased metabolic success in maximizing predator avoidance by
utilizing intertidal structure for shelter (Harvey and Stewart 1991, Ruiz et al. 1993,
Miltner et al. 1995, Linehan et al. 2001, Rypel et al. 2007), where we saw greater small
fish abundance and condition. Both Pinfish and Mummichog are well adapted to
intertidal survival; intertidal seagrass beds provide Pinfish with abundant nutrition at a
minimal energy cost (Callaway and Josselyn 1992, Montgomery and Targett 1992, Levin
et al. 1997) while Mummichog typically rely heavily on the marsh surface for food
during diurnal high tides (Weisberg et al. 1981). Decapods likely move into the subtidal
zone as a result of intertidal (and reduced overall) fish presence. Grass shrimp in
particular are known to prefer the subtidal zone during periods of reduced fish encounter
frequency (Bretsch and Allen 2005). At swash creeks, these trends are reversed.
Abundance and condition of small fishes were greater in the subtidal zone, where they
could safely feed due to reduced encounter frequency with piscivorous fishes (Jordan et
al. 1996). Foraging success of these species is greater over flat bottoms when vision and
mobility is not impeded by structure (Vince et al. 1976, Heck and Thoman 1981, Ryer
1988, Jordan et al. 1996, Bretsch and Allen 2005, Harter and Heck 2006). Proximity of
subtidal traps to shell substrate likely allowed Pinfish to quickly retreat (Summerson and
Peterson 1984, Ferrell and Bell 1991) should the need arise. Mummichog are
28
opportunistic predators (Kneib 1986) that can assimilate over 75% of their energy from
subtidal food sources in certain habitats (Weisberg and Lotrich 1982). Additionally,
entering the subtidal zone minimizes energy spent on avoidance of avian species (Power
1984, 1987, Crower et al. 1997, Kneib 1997) which are major predators of Pinfish and
Mummichog (Jenni 1969, Kushlan 1976, Bildstein et al. 1981, 1982). In turn, subtidal
fish presence drives grass shrimp and other decapods further into shallow water to forage
(Kneib and Wagner 1994, Bretsch and Allen 2006) and escape predation (Stoner 1980,
Posey and Hines 1991, Kneib 2000). Divergences from these trends seen at some creek
pairs suggest high variability in fish and decapod elevation preferences, possibly
attributable to the motility of nekton allowing frequent back and forth movement between
the two zones.
At both creek types, Pinfish and Mummichog captured by intertidal traps over
reef substrate were smaller on average than those captured by subtidal traps off the reef,
suggesting shell bags provide nursery habitat. Larger Pinfish likely venture further from
intertidal structure as they achieve size refuge from predation due to gape size limitation
in piscivores (Christensen 1996, Reimchen and Nosil 2002, Price et al. 2015). Juvenile
Mummichog stay in the intertidal zone at all times, inhabiting small puddles that remain
after the tide recedes (Kneib 1986).
Diel use of reefs was consistent with known diel activity patterns for the species
studied; therefore, introduction of reef substrate likely had little to no effect on diel
patterns. Fish diel reef usage expectedly showed no obvious patterns by creek type.
Mummichog are visual predators that feed primarily during daylight hours (Weisberg et
al. 1981) and several studies found Pinfish to be most active during the day (Low 1973,
29
Hobson 1979, Ryan 1981, Sogard et al. 1988). However, Pinfish activity is known to
vary greatly at some creeks, including North Inlet (Shenker and Dean 1979). Seagrass-
using fish often forage over adjacent bare substrata at night when predation risk is
reduced (Summerson and Peterson 1984), and traps may not appear as visually
threatening at night. Piscivorous fishes and certain decapods are predominantly
crepuscular feeders (Antheuisse et al. 1971, Oakley 1979, Shenker and Dean 1979,
Aguzzi et al. 2004, Aguzzi et al. 2005). Grass shrimp are often stimulated into foraging
patterns by optimum light intensity present during twilight hours (Oakley 1979, Aguzzi et
al. 2004) and are thought to remain sedentary during the majority of daylight hours
(Shenker and Dean 1979). Night traps, which were active during both dawn and dusk,
expectedly returned greater catches of decapods than day traps.
Inlet creeks exhibited a greater mean rarefied richness of fish and decapods than
swash creeks. Greater area, habitat heterogeneity, and connectivity to the open ocean
increase the likelihood of encountering a new species at inlet creeks and provide more
ecological niches for these species to fill (Scheiner 2003, Guilhaumon 2008, Nicolas et al.
2010, Guilhaumon et al. 2012, Vasconcelos et al. 2015). Inlet creeks also had a more
tolerable average salinity and fewer extreme salinity drop-offs due to rainfall events.
Creeks with low salinities due to turbulent river flow have resulted in reduced species
richness (Whitfield and Harrison 2003), as have creeks of hyperhaline conditions
(Whitfield et al. 2012). Finally, reduced urbanization at inlet creeks meant a reduction in
impervious surfaces and increases in episodic stormwater runoff events, which have been
linked to decreases in biological diversity (Lerberg et al. 2000, Holland et al. 2004, Krebs
et al. 2014).
30
The observed, within creek variability in minnow trap captures likely reflects the
natural schooling behavior of fishes captured (Shenker and Dean 1979). The number and
density of transient Pinfish schools within a creek at any given time is subject to a great
deal of fluctuation (Shenker and Dean 1979). Mummichog are also schooling fishes
(Symons 1971), though as residents they generally remain in a small home range (Lotrich
1975). Discovery of a minnow trap by a school may help explain the frequency of both
traps that return no catch and traps that return catches several orders of magnitude above
the mean. Evidence suggests faunal use of natural reefs at the same site often differ
significantly over both short and long periods of time (Walters and Coen 2007).
Dog food in the stomachs of organisms captured by minnow traps may have
biased recorded mass and condition. However, all traps were filled with the same
quantity of food, so traps that captured more organisms would have less available food
per organism. Organism condition was always greatest from creek type and tidal
elevation at which they were also most abundant. Therefore, condition differences are
unlikely to suffer from type I error and may be even greater than reported here.
Poor pull trap catches compared to Lehnert and Allen (2002) may be attributable
to intertidal rather than subtidal trap deployment. Compared to subtidal trap catches,
Lehnert and Allen (2002) experienced lower densities of fish in the intertidal zone, even
though a highly efficient blocknetting method was used. Previous studies found subtidal
densities of summer transient fishes approached 10x greater than intertidal densities;
moreover, many fish caught in the subtidal zone are considered uncommon or rare in the
intertidal zone (Wenner et al. 1996, Coen et al. 1999b).
31
Notable catch irregularities at Singleton swash occurred possibly due to a
disruption of the tidal cycle. Sediment which had slowly accumulated at the mouth of the
swash over the sampling period resulted in a reduced height differential between high
tide and low tide. In tidal impoundments, decreased functional species diversity was due
to fragmentation, though resident estuarine specialists that prefer the marsh interior were
not affected significantly (Carswell et al. 2015). Tidal impoundments have also been
associated with reductions in dissolved oxygen and turbidity and increases in
sedimentation and salinity fluctuations (Wenner et al. 1986, Montague et al. 1987). All
intertidal traps were submerged permanently toward the end of the sampling period at
Singleton Swash which may account for the greater density and diversity of fish species
caught from pull traps.
Habitat Abundance Study (July 2014)
Differences in minnow trap fish captures over mudflat, restored and natural reefs
within Bly Creek (North Inlet) suggested reefs were the preferred habitat, but restored
reefs were less attractive than natural reefs. Many studies have indicated architecturally
complex reef habitats provide significant advantages for fishes over less complex
mudflats (Coen et al. 1999b, Minello 1999, Grabowski 2005, Coen and Grizzle 2007,
Humphries and La Peyre 2015), but the difference between natural and restored reef
usage is notable. Few direct comparisons of natural and restored reef nekton assemblages
exist (Grabowski et al. 2005), but in some specific cases populations using restored reefs
have not significantly changed in the first 3 years (Humphries and La Peyre 2015) or
converged completely with natural reefs after 7 years (Walters and Coen 2007), so the
32
one-year residence period of restored reefs at the time of sampling likely influenced data.
Additionally, the abundance of high quality habitat at North Inlet means fish might not
choose to venture away from numerous natural reefs at which they are already
comfortable (Geraldi 2009, Humphries and La Peyre 2015). However, natural reef
minnow trap fish CPUE seen in this study was similar to restored reef minnow trap fish
CPUE at swash creeks in sampling approach 1. Lenihan et al. (2001) also found fish trap
(0.5 cm mesh) Pinfish catch to be similar between restored and natural reefs after only 4
years. Newly restored reefs may, in certain conditions, work similarly to natural reefs to
attract certain species, particularly if those species are limited by quantity of available
habitat (Turner 2000, Humphries et al. 2011).
Decapods seemed to prefer restored reefs to abundant natural reefs, suggesting
early restored reefs supply them with excellent quality habitat (Turner 2000, Humphries
and La Peyre 2015). It is likely restored reefs provide more hiding places than mudflats
(Thorp 1976, Posey et al. 1999) but they are not as heavily cohabitated by predators with
which they share affinity for shell substrate as natural reefs (Thorp 1976, Posey et al.
1999). Organisms are known to be differentially attracted to reefs based on structural
complexity (Sherman 2002, Humphries 2011). Larger organisms reach a threshold after
which increasing habitat complexity leads to decreasing density faster than small
organisms (Karp et al. 2018). In this case, it is possible invertebrates prefer the smaller
crevices created by the mesh bag structure of the restored reefs while fishes favor larger
spaces created by natural reefs. However, usage may continue to converge with natural
reef usage as spat settles and grows on restored reefs.
33
Gill net catch was greatest over mudflat substrate. Grabowski et al. (2005)
obtained a similar result, discovering that piscivore abundance was significantly greater
over mudflat controls than at reefs likely because structure interferes with visibility and
mobility for predators (Vince et al. 1976, Hobson 1979, Heck and Thoman 1981, Ryer
1988, Jordan et al. 1996). Many of the species caught in the gill nets, such as large-tooth
flounders, Atlantic Sharpnose Sharks, Red Drum, Ladyfish, and Spotted Seatrout are
known to be supported by flat substrate (Ryan 1981, Jordan et al. 1996). It is also
possible predators travel at higher speeds in pursuit of prey over mudflat substrate
(Hobson 1979), increasing the likelihood of entanglement in a gill net.
Results from the habitat abundance study imply minnow traps effectively sample
organisms that rely on restored reefs, but gill nets do not. The target range of species
captured by a minnow trap is narrow and fails to encompass the entire spectrum or
species present or reliant on reef substrate, but greater catch of organisms from reef
substrates than mudflat substrate indicates these species are reliant on oyster reef habitat.
Gill net catch was greater from mudflat substrate than either of the reef substrates,
suggesting target species likely use reefs in a very limited capacity or were merely
trapped while passing by.
Conclusion
Newly constructed reefs were heavily used by different assemblages of nekton at
swash and inlet creeks. Small fishes appeared to be limited by availability of structure at
swash creeks, and as a result, colonized new reefs immediately. At inlet reefs, reduced
numbers likely result from redundancy of habitat function and may improve with
34
increased residence time. Decapods colonized restored inlet reefs quickly and in great
numbers despite abundant natural reefs, suggesting early restored reefs provide high
quality habitat. Larger, piscivorous fishes appear to be unaffected by reef construction but
may have influence on creek or tidal elevation preferences seen in smaller nekton. At
inlet creeks, predation pressure likely forces smaller fishes into the intertidal zone,
successively pushing decapods into the subtidal zone. At swash creeks, reduced predation
pressure allows small fishes to spend more time safely foraging in the subtidal zone, in
turn driving decapods into the intertidal zone. Regardless of creek type, fish captured by
intertidal traps were smaller on average than those captured by subtidal traps, suggesting
early restored reefs successfully provide nursery habitat. Finally, after a residence time of
one year, restored reefs at North Inlet were not convergent with natural reefs. Small
fishes preferred longstanding natural reefs to newer bag reefs, but still used restored reefs
more than mudflat substrate. Larger piscivores were most common on mudflats, but
visited reefs on occasion. Decapods preferred restored reefs to natural reefs, at least prior
to significant spat settlement. Convergent with natural reefs or not, restored reefs seem to
be an effective way to increase productivity at both swash and inlet creeks in South
Carolina. However, careful consideration should be given to quantity and quality of
nearby habitat and to local trophic interactions prior to reef construction.
35
Tables
Table 1: Latitude and longitude, estimated low and high tide channel width (m), height
of deployment of intertidal and subtidal minnow traps and pull traps relative to mean low
and high tide, intra-site elevation difference between intertidal and subtidal trap minnow
trap deployment, and tidal range at all 8 locations used during all sampling approaches
from May 2013 to July 2014.
Hog Inlet
(HI)
Murrell’s
Inlet (MI)
North Inlet-
Oyster
Landing
(NI)
North Inlet-
Bly Creek
(NI)
Whitepoint
Swash--48th
Ave Bridge
(WP)
Whitepoint
Swash--
Briarcliffe
Acres (WP)
Singleton
Swash (SS)
Withers
Swash (WS)
Latitude 33.503133˚ 33.344558˚ 33.349667˚ 33.333802˚ 33.792511˚ 33.471421˚ 33.45363˚ 33.405999˚
Longitude -78.370359˚ -79.001136˚ -79.188777˚ -79.203918˚ -78.736775˚ -78.443242˚ -78.47289˚ -78.53329˚
Low Tide Channel
Width (m) 80 112 34 36 12 10 18 11
High Tide Channel
Width (m) 207 130 80 42 16 10 19 16
Intertidal Trap
(Low Tide) (cm) 35.9 82.3 38.1 43.1 27.9 8.2 -34 10
Intertidal Trap
(High Tide) (cm) -99.7 -64.3 -93.9 -74.9 -54.1 -25.9 -45.5 -46
Subtidal Trap
(Low Tide) (cm) -34.7 -35.2 -46.6 -51 -48.5 -38 -75.2 -52
Subtidal Trap
(High Tide) (cm) -170.8 -181.2 -178.6 -153.2 -130.5 -68.1 -86.7 -108
Pull Trap
(Low Tide) (cm) 25 20 63 N/A N/A 3 -42 7
Pull Trap
(High Tide) (cm) -110.6 -126 -69 N/A N/A -27.1 -53.5 -49
Trap Elevation
Difference (cm) -70.6 -116.9 -84.7 -78.3 -71.4 -42.2 -41.2 -62
Tidal Range (cm) 135.6 146 132 118 82 34.1 11.5 56
36
Table 2: Sampling schedule for trap deployment and retrieval for approaches 1 and 2
and the habitat abundance study.
Approach 1
(Minnow traps,
pull traps) 2013
Trap
deployment
Minnow
trap
retrieval
Pull trap
retrieval
(WS, MI, NI)
Pull trap
retrieval
(HI, WP, SS)
Set 1 5/28 5/29 6/2 6/3
Set 2 6/23 6/24 6/29 6/30
Set 3 7/21 7/22 7/27 7/28
Set 4 8/19 8/20 8/25 8/26
Set 5 9/16 9/17 9/22 9/23
Approach 2
(Minnow traps, gill nets) 2013 Start date End date
NI/WS
Set 1 6/28 7/1
Set 2 8/13 8/16
MI/SS
Set 1 7/14 7/17
Set 2 8/27 8/30
HI/WP
Set 1 7/28 7/31
Set 2 9/12 9/15
Habitat abundance study
(Minnow traps, gill nets) 2014 Start date End date
Set 1 7/16 7/19
37
Table 3: Actual richness, rarefied average and median richness, variance, and upper and
lower confidence intervals for fish captured with all trap methods at 3 inlet and 3 swash
tidal creeks along the Grand Strand of SC from May to September 2013.
Site
Actual
Richness
Rarefied
Average
Richness
Rarefied
Median
Variance
Diversity
95%
confidence
interval low
95%
confidence
interval high
Hog Inlet 10 10 10 0 10 10
Murrell’s Inlet 4 2.8 3 0.3 2 4
North Inlet 12 11.1 11 0.7 9 12
Inlet (avg.) 8.7 8.0 8 0.3 7 8.7
Whitepoint Swash 8 2.9 3 0.7 2 5
Singleton Swash 10 5.5 5 1.1 4 8
Withers Swash 9 4.3 4 1.3 2 7
Swash (avg.) 9 4.2 4 1.0 2.7 6.7
Table 4: Actual richness, rarefied average and median richness, variance, and upper and
lower confidence intervals for decapods captured with all trap methods at 3 inlet and 3
swash tidal creeks along the Grand Strand of SC from May to September 2013.
Site
Actual
Richness
Rarefied
Average
Richness
Rarefied
Median
Variance
Diversity
95%
confidence
interval low
95%
confidence
interval high
Hog Inlet 10 4.4 4 1.3 2 7
Murrells Inlet 6 3.2 3 0.8 2 5
North Inlet 11 5.9 6 1.1 4 8
Inlet (avg.) 9 4.5 4.3 1.1 2.7 6.7
Whitepoint Swash 6 3.0 3 0.9 1 5
Singleton Swash 7 4.5 4 0.9 3 6
Withers Swash 3 3 3 0 3 3
Swash (avg.) 5.3 3.5 3.3 0.6 2.3 4.7
38
Table 5: Count and total dry mass, in grams, of all fish and decapods collected with
all trap methods at 3 swash and 3 inlet sites in South Carolina from May to
September in 2013.
Site Fish
Catch
Total Fish
Biomass
Avg. Fish
Biomass
Invert.
Catch
Total Invert.
Biomass
Avg. Invert.
Biomass
Hog Inlet 87 461.9 5.3 421 246.0 0.6
Murrells Inlet 1164 701.2 0.6 221 87.6 0.4
North Inlet 110 1675.4 15.2 270 137.9 0.5
Inlet (avg.) 453.6 946.1 2.1 304 157.2 0.5
Whitepoint Swash 1031 731.9 0.7 229 68.5 0.3
Singleton Swash 448 290.9 0.6 184 86.2 0.5
Withers Swash 876 912.9 1.0 35 61.8 1.8
Swash (avg.) 785 645.3 0.8 149.3 72.1 0.5
Table 6: Actual richness, rarefied average richness, rarefied median richness, variance,
and upper and lower confidence intervals for fish captured with all trap methods over
natural reef, restored reef, and mudflat substrate along the Grand Strand of SC in July
2014.
Site
Actual
Richness
Rarefied
Average
Richness
Rarefied
Median
Variance
Diversity
95%
confidence
interval low
95%
confidence
interval high
Fish
Mudflat 6 6 6 0 6 6
Natural Reef 9 2.9 3 0.7 2 5
Restored Reef 9 6.1 6 1.3 4 8
Decapods
Mudflat 8 7.6 8 0.4 6 8
Natural Reef 5 5 5 0 5 5
Restored Reef 8 5.9 6 1.6 4 8
Table 7: Count and total dry mass, in grams, of all fish and decapods collected with
minnow traps and gill nets over 3 substrate types at North Inlet, SC in July 2014.
Substrate Fish
Catch
Total Fish
Biomass
Avg. Fish
Biomass
Invert.
Catch
Total Invert.
Biomass
Avg. Invert.
Biomass
Natural Reef 334 501.4 1.5 31 100.0 3.2
Restored Reef 53 488.0 9.2 65 93.9 1.4
Mudflat 24 915.3 38.1 35 27.8 0.8
39
Table 8: Total catch of each species with all trap methods at each of the 6 tidal
creeks sampled in this study in 2013.
Species Hog
Inlet
Murrells
Inlet
North
Inlet
Whitepoint
Swash
Singleton
Swash
Withers
Swash
Lagodon rhomboides 29 136 10 897 138 666
Fundulus heteroclitus 39 1009 51 123 261 180
Fundulus majalis 0 18 0 1 0 0
Anguilla rostrada 0 0 0 0 0 12
Rhizoprionodon terraenovae 0 0 23 0 0 0
Lutjanus griseus 0 1 2 4 18 6
Mugil cephalus 0 0 3 1 0 3
Evorthodus lyricus 0 0 0 0 1 0
Sciaenops ocellatus 0 0 1 0 0 0
Elops saurus 0 0 4 0 0 0
Gobiosoma bosc 1 0 0 0 23 0
Gobiesox strumosus 0 0 0 0 0 3
Opsanus tau 3 0 0 2 1 0
Bairdiella chrysoura 3 0 1 0 1 0
Paralichthys dentatus 0 0 0 0 0 2
Paralichthys lethostigma 3 0 0 0 0 0
Leiostomus xanthurus 4 0 3 0 0 1
Monocanthus hispidus 2 0 0 0 2 0
Chasmodes bosquianus 0 0 0 2 0 0
Eleotris pisonis 0 0 0 0 1 0
Myrophis punctatus 0 0 0 0 0 3
Selene setapinnis 0 0 0 1 0 0
Cynoscion nebulosus 0 0 1 0 0 0
Chaetodipterus faber 0 0 0 0 2 0
Symphurus plagiusa 0 0 1 0 0 0
Menidia menidia 1 0 10 0 0 0
Orthopristis chrysoptera 2 0 0 0 0 0
Total Fish 87 1164 110 1031 448 876
Palaemonetes Sp. 341 196 179 210 88 30
Litopenaeus setiferus 43 16 26 0 3 0
Farfantepenaeus duorarum 3 0 1 1 0 0
Callinectes sapidus 8 4 12 3 4 2
Eurypanopeus depressus 1 0 4 10 0 0
Uca pugnax 6 0 2 3 12 0
Alpheus heterochaelis 0 1 1 0 1 0
Panopeus herbstii 7 0 1 0 0 0
Armases cinereum 0 0 27 2 0 3
Uca minax 0 0 1 0 0 0
Clibanarius vittatus 1 0 0 0 2 0
Menippe mercenaria 1 1 0 0 0 0
Farfantepenaeus aztecus 10 4 16 0 74 0
Total Decapods 421 221 270 229 184 35
Overall Total 508 1385 380 1260 632 911
40
Table 9: Total catch of each species with all trap methods over three substrate types
at Bly Creek in North Inlet in July 2014.
Species Mudflat Natural Reef Restored Reef
Lagodon rhomboides 2 158 18
Fundulus heteroclitus 7 162 23
Rhizoprionodon terraenovae 5 1 2
Mugil cephalus 0 1 0
Pomatomus saltatrix 6 4 2
Leiostomus xanthurus 1 0 1
Elops saurus 3 0 0
Brevoortia tyrannus 0 0 4
Cynoscion nebulosus 0 2 1
Orthopristis chrysoptera 0 4 1
Paralicthys dentatus 0 1 0
Gobiesox strumosus 0 1 0
Cynoscion regalis 0 0 1
Total Fish 24 334 53
Palaemonetes sp. 16 20 49
Armases cinereum 1 0 0
Callinectes sapidus 3 7 4
Farfantepenaeus aztecus 3 1 4
Farfantepenaeus duorarum 0 0 1
Litopenaeus setiferus 9 2 2
Menippe mercenaria 1 0 0
Panopeus herbstii 1 1 0
Uca pugnax 1 0 3
Cymothoa exigua 0 0 1
Clibanarius vittatus 0 0 1
Total Decapods 35 31 65
Overall Total 59 365 118
41
Figures
Figure 1: Map of the 6 tidal creeks sampled in the study. Orange pins represent inlet sites
and red pins represent swash sites.
42
Figure 2: Photographs of Hog Inlet (a), Murrells Inlet (b), North Inlet (Oyster Landing)
(c), North Inlet (Bly Creek) (d), Whitepoint Swash (48th Ave. Bridge) (e), Singleton
Swash (f), Withers Swash (g) and Whitepoint Swash (Briarcliffe Acres) (h) taken during
site surveys.
43
Figure 3: A 100 bag oyster shell restored reef immediately after deployment at Hog Inlet,
SC.
Figure 4: A pull trap tethered to a pier at Oyster Landing in North Inlet, SC. Three
dimensional oyster shell structure and stabilizing PVC stake are visible.
44
Figure 5: A gill net placed over restored reef substrate at Bly Creek in North Inlet, SC.
Figure 6: A minnow trap over shell substrate at North Inlet during low tide.
45
HI/WP MI/SS NI/WS
Me
an
Sa
lin
ity (
x +
SE
)
0
10
20
30
40HI/WP MI/SS NI/WS
Me
an
Te
mp
era
ture
(x
+ S
E)
25
26
27
28
29
30Inlet
Swash
Figure 7: Mean temperature (˚C) and salinity (ppt) at each of the inlet/swash creek
pairings sampled along the Grand Strand of SC from May to September 2013. An
asterisk indicates a significant (p ≤ 0.05) difference between paired bars.
* * *
46
Inlet Swash
De
ca
po
d C
PU
E (
x +
SE
)
0
5
10
15
20
Fis
h C
PU
E (
x +
SE
)
0
10
20
30
40
Intertidal
Subtidal
Figure 8: Sampling approach 1 total minnow trap fish and decapod CPUE at three inlet
and three swash tidal creeks from May to September 2013, separated by tidal elevation.
An asterisk indicates a significant (p ≤ 0.05) difference between creek type.
*
*
47
Fish Decapod
Pu
ll T
rap
CP
UE
(x
+ S
E)
0
10
20
30
40
Inlet
Swash
Figure 9: Total pull trap CPUE of fish (left) and decapods (right) taken from May to
September 2013 from 3 inlet and 3 swash creeks along the Grand Strand of SC. An
asterisk indicates a significant (p ≤ 0.05) difference between paired bars.
*
48
0
5
10
15
20
Night
Day
De
ca
po
d C
PU
E (
x +
SE
)
0
5
10
15
20
Inlet Swash
0
5
10
15
20
Fis
h C
PU
E (
x +
SE
)
0
10
20
30
40
50
Inlet Swash
0
10
20
30
40
50
0
10
20
30
40
50
Figure 10: Sampling approach 2 total minnow trap fish and decapod CPUE from (A)
Hog inlet and Whitepoint swash, (B) Murrells Inlet and Singleton Swash, and (C) North
Inlet and Withers Swash from July to September 2013, separated by day/night stage. An
asterisk indicates a significant (p ≤ 0.05) difference between creek type.
A
B
C
*
* *
* *
49
0
1
2
3
4
5
Night
Day
Inlet Swash
Fis
h C
PU
E (
x +
SE
)
0
1
2
3
4
5
Figure 11: Total gill net fish CPUE from (A) Hog inlet and Whitepoint swash and (B)
Murrells Inlet and Singleton Swash from July to September 2013, separated by day/night
stage. An asterisk indicates a significant (p ≤ 0.05) difference between creek type.
A
B *
50
Me
an
Sta
nd
ard
Le
ng
th (
x +
SE
)
0
20
40
60
80
Intertidal
Subtidal
Mea
n D
ry M
as
s (
x +
SE
)
0.0
0.5
1.0
1.5
2.0
Inlet Swash
Me
an
Co
nd
itio
n (
x +
SE
)
0.0
0.2
0.4
0.6
0.8
1.0
Figure 12: Mean standard length (mm), dry mass (g), and Fulton’s condition factor (K)
of all Pinfish collected from the intertidal and subtidal zones of both inlet and swash
creeks in 2013 along the Grand Strand in SC. An asterisk indicates a significant (p ≤ 0.05)
difference between creek type.
*
*
*
51
Inlet Swash
Me
an
Co
nd
itio
n (
x +
SE
)
0.0
0.2
0.4
0.6
0.8
1.0
Me
an
Sta
nd
ard
Len
gth
(x +
SE
)
0
20
40
60
80
Intertidal
Subtidal
Mea
n D
ry M
as
s (
x +
SE
)
0.0
0.5
1.0
1.5
2.0
Figure 13: Mean standard length (mm), dry mass (g), and Fulton’s condition factor (K)
of all Mummichog collected from the intertidal and subtidal zones of both inlet and
swash creeks in 2013 along the Grand Strand in SC. An asterisk indicates a significant (p
≤ 0.05) difference between creek type.
*
*
*
52
Me
an
To
tal
Le
ng
th (
x +
SE
)
0
5
10
15
20
25
30
Inlet
Swash
Me
an
Co
nd
itio
n (
x +
SE
)
0.0
0.2
0.4
0.6
0.8
1.0
Me
an
Dry
Ma
ss
(x
+ S
E)
0.00
0.02
0.04
0.06
0.08
0.10
Figure 14: Mean standard length (mm), dry weight (g), and Fulton’s condition factor (K)
of all grass shrimp collected with all trap types from inlet and swash creeks along the
Grand Strand in SC in 2013. An asterisk indicates a significant (p ≤ 0.05) difference
between creek type.
*
*
*
53
Fis
h C
PU
E (
x +
SE
)
0
20
40
60
80
Nat. Reef Rest. Reef Mudflat
0
20
40
60
80
Nat. Reef Rest. Reef Mudflat
De
ca
po
d C
PU
E (
x +
SE
)
0
5
10
15
20
0
5
10
15
20 Intertidal
Subtidal
Figure 15: Habitat abundance study (A) daytime and (B) nighttime total minnow trap
fish and decapod CPUE, separated by tidal elevation, over three substrate types at North
Inlet, SC in July 2014.
A
B
54
Natural Reef Restored Reef Mudflat
Fis
h C
PU
E (
x +
SE
)
0
1
2
3
4
5
Night
Day
Figure 16: Total gill net fish CPUE over three substrate types, separated by day/night
stage, at North Inlet, SC in July 2014.
55
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Appendix
Table A1: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
negative binomial regressions examining the influence of various factors on trap CPUE in
sampling approach 1. “NS” indicates a factor was not significant (p > 0.05).
Sampling approach 1 Result
Fish minnow trap CPUE
Creek type Wald χ2 = 11.135, df = 1, p = 0.001
Tidal elevation NS
Creek type * tidal elevation Wald χ2 = 3.865, df = 1, p < 0.05
Month NS
Decapod minnow trap CPUE
Creek type Wald χ2 = 18.252, df = 1, p < 0.001
Tidal elevation NS
Creek type * tidal elevation Wald χ2 = 5.455, df = 1, p < 0.05
Month NS
Pull trap CPUE
Fish by creek type Wald χ2 = 4.349, df = 1, p < 0.05
Invertebrates by creek type NS
Month NS
72
Table A2: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
negative binomial regressions examining the influence of various factors on minnow trap
fish CPUE in sampling approach 2. “NS” indicates a factor was not significant (p > 0.05).
Sampling approach 2 Minnow trap fish CPUE Result
Creek type
Hog Inlet/Whitepoint swash Wald χ2 = 47.440, df = 1, p < 0.001
Murrells Inlet/Singleton swash Wald χ2 = 26.364, df = 1, p < 0.001
North Inlet/Withers swash Wald χ2 = 67.306, df = 1, p < 0.001
Tidal elevation
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash Wald χ2 = 15.229, df = 1, p < 0.001
North Inlet/Withers swash NS
Creek type * tidal elevation
Hog Inlet/Whitepoint swash Wald χ2 = 9.980, df = 1, p < 0.01
Murrells Inlet/Singleton swash Wald χ2 = 61.754, df = 1, p < 0.001
North Inlet/Withers swash NS
Creek type * diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Creek type * tidal elevation * diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Sampling set
Hog Inlet/Whitepoint swash Wald χ2 = 12.964, df = 1, p < 0.001
Murrells Inlet/Singleton swash Wald χ2 = 19.614, df = 1, p < 0.001
North Inlet/Withers swash NS
73
Table A3: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
negative binomial regressions examining the influence of various factors on minnow trap
decapod CPUE in sampling approach 2. “NS” indicates a factor was not significant (p >
0.05).
Sampling approach 2 Minnow trap decapod CPUE Result
Creek type
Hog Inlet/Whitepoint swash Wald χ2 = 13.074, df = 1, p < 0.001
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash Wald χ2 = 65.088, df = 1, p < 0.001
Tidal elevation
Hog Inlet/Whitepoint swash Wald χ2 = 6.701, df = 1, p = 0.01
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Diel stage
Hog Inlet/Whitepoint swash Wald χ2 = 15.426, df = 1, p < 0.001
Murrells Inlet/Singleton swash Wald χ2 = 17.375, df = 1, p < 0.001
North Inlet/Withers swash Wald χ2 = 3.930, df = 1, p < 0.05
Creek type * tidal elevation
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash Wald χ2 = 5.077, df = 1, p < 0.05
North Inlet/Withers swash NS
Creek type * diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Creek type * tidal elevation * diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Sampling set
Hog Inlet/Whitepoint swash Wald χ2 = 4.648, df = 1, p < 0.05
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
74
Table A4: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
negative binomial regressions examining the influence of various factors on gill net
CPUE in sampling approach 2. “NS” indicates a factor was not significant (p > 0.05).
Approach 2 gill net CPUE Result
Creek type
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash Wald χ2 = 6.853, df = 1, p < 0.01
Diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Creek type * diel stage
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Sampling set
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
75
Table A5: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
gamma regressions examining the influence of various factors on pinfish, mummichog,
and grass shrimp size metrics in sampling approaches 1 and 2. “NS” indicates a factor
was not significant (p > 0.05).
Size metrics Result
Pinfish standard length
Creek type Wald χ2= 29.118, df = 1, p < 0.001
Tidal elevation Wald χ2 = 5.205, df = 1, p < 0.05
Creek type * tidal elevation NS
Pinfish dry mass
Creek type Wald χ2 = 9.500, df = 1, p = 0.001
Tidal elevation Wald χ2 = 8.169, df = 1, p < 0.01
Creek type * tidal elevation NS
Pinfish condition (K)
Creek type Wald χ2 = 39.796, df = 1, p < 0.001
Tidal elevation NS
Creek type * tidal elevation Wald χ2 = 6.2324, df = 1, p < 0.05
Mummichog standard length
Creek type Wald χ2 = 28.297, df = 1, p < 0.001
Tidal elevation Wald χ2 = 25.474, df = 1, p < 0.001
Creek type * tidal elevation NS
Mummichog dry mass
Creek type Wald χ2 = 53.648, df = 1, p < 0.001
Tidal elevation Wald χ2 = 27.205, df = 1, p < 0.001
Creek type * tidal elevation NS
Mummichog condition (K)
Creek type Wald χ2 = 6.574, df = 1, p = 0.01
Tidal elevation NS
Creek type * tidal elevation Wald χ2 = 75.966, df = 1, p < 0.001
Grass shrimp (by creek type)
Total length Wald χ2 = 47.760, df = 1, p < 0.001
Dry mass Wald χ2 = 79.546, df = 1, p < 0.001
Condition (K) Wald χ2 = 40.102, df = 1, p < 0.001
76
Table A6: Power (F), degrees of freedom (df), and statistical significance (p) of
ANOVAs examining the influence of various factors on trap CPUE in sampling
approaches 1 and 2. “NS” indicates a factor was not significant (p > 0.05).
Physical creek characteristics Result
Approach 1
Mean low tide
Intertidal trap height NS
Subtidal trap height NS
Pull trap height NS
Mean high tide
Intertidal trap height F1,4= 13.071, p < 0.05
Subtidal trap height F1,4 = 55.867, p < 0.01
Pull trap height F1,4 = 9.722, p < 0.05
Channel width F1,4 = 8.369, p < 0.05
Approach 2
Mean low tide
Intertidal trap height NS
Subtidal trap height NS
Mean high tide
Intertidal trap height F1,4= 8.211, p < 0.05
Subtidal trap height F1,4= 15.881, p < 0.05
Channel width F1,4 = 7.874, p < 0.05
Temperature by creek type
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash NS
North Inlet/Withers swash NS
Temperature by month
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash F = 30.295, df = 4, p = 0.001
North Inlet/Withers swash F = 4.984, df = 4, p < 0.01
Salinity by creek type
Hog Inlet/Whitepoint swash F = 11.890, df = 1, p = 0.001
Murrells Inlet/Singleton swash F = 23.638, df = 1, p < 0.001
North Inlet/Withers swash F = 7.660, df = 1, p < 0.01
Salinity by month
Hog Inlet/Whitepoint swash NS
Murrells Inlet/Singleton swash F = 5.377, df = 4, p = 0.001
North Inlet/Withers swash NS
77
Table A7: Power (Wald χ2), degrees of freedom (df), and statistical significance (p) of
negative binomial regressions examining the influence of various factors on trap CPUE in
the habitat abundance study. “NS” indicates a factor was not significant (p > 0.05).
Habitat abundance study Result
Minnow trap fish CPUE
Substrate type Wald χ2 = 59.635, df = 2, p < 0.001
Tidal elevation Wald χ2 = 15.935, df = 1, p < 0.001
Diel stage NS
Substrate type * tidal elevation NS
Substrate type * diel stage NS
Substrate type * tidal elevation * diel stage NS
Minnow trap decapod CPUE
Substrate type NS
Tidal elevation NS
Diel stage Wald χ2 = 18.359, df = 1, p < 0.001
Substrate type * tidal elevation NS
Substrate type * diel stage NS
Substrate type * tidal elevation * diel stage NS
Gill net CPUE
Substrate type NS
Diel stage NS
Substrate type * diel stage NS