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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 +

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)

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10

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20

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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

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2

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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