FACTORS AFFECTING INVERTEBRATE AND FISH COMMUNITIES IN
COASTAL WETLANDS OF THE GREAT LAKES
A dissertation submitted
to Kent State University in partial
fulfillment for the requirements of the
degree of Doctorate of Philosophy
by
Douglas J. Kapusinski
December 2012
i
Dissertation written by
Douglas J Kapusinski
B.S., Bradley University, 2004
Ph.D., Kent State University, 2012
Approved by
_______________________________, Chair Doctoral Dissertation Committee
Ferenc A. de Szalay
_______________________________, Members, Doctoral Dissertation Committee
Mark W. Kershner
_______________________________
Robert E. Carlson
_______________________________
Yoram Eckstein
_______________________________
Accepted by
_______________________________, Chair, Department of Biological Sciences
James L. Blank
_______________________________, Dean, College of Arts and Sciences
ii
TABLE OF CONTENTS
LIST OF FIGURES........................................................................................................... v
LIST OF TABLES............................................................................................................ vii
ACKNOWLEDGEMENTS............................................................................................... ix
CHAPTER
I.
Introduction................................................................................................ 1
Great Lake Coastal Wetlands......................................................... 1
Invertebrates and Fish.................................................................... 4
Shorebirds...................................................................................... 6
Invasive Species............................................................................. 7
Hypotheses..................................................................................... 9
Study Site Description.................................................................. 10
Literature cited.............................................................................. 15
II.
Comparing fish and invertebrate communities in Great Lake coastal
wetlands and impounded wetlands........................................................... 30
Abstract......................................................................................... 30
Introduction................................................................................... 31
Methods......................................................................................... 34
Study site description........................................................ 34
Experimental design.......................................................... 35
Data analysis..................................................................... 37
Results........................................................................................... 38
Fish.................................................................................... 38
Benthic invertebrates........................................................ 44
Discussion..................................................................................... 59
Acknowledgements....................................................................... 64
Literature cited.............................................................................. 65
III.
Effects of fish predation on benthic invertebrate communities in a Great
Lake coastal wetland................................................................................. 74
Abstract......................................................................................... 74
Introduction................................................................................... 75
Methods......................................................................................... 77
Study site description........................................................ 77
Fish exclosures.................................................................. 78
Benthic invertebrates........................................................ 79
Fish community................................................................ 80
Data analysis..................................................................... 81
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Results........................................................................................... 82
Environmental data........................................................... 82
Fish taxa and YOY diets................................................... 83
Benthic Invertebrates........................................................ 87
Discussion................................................................................... 101
Management implications............................................... 106
Acknowledgements..................................................................... 108
Literature cited............................................................................ 109
IV.
Predation of epizoic and benthic invertebrates by fish including common
carp Cyprinus carpio, in a Great Lakes coastal wetland........................ 115
Abstract....................................................................................... 115
Introduction................................................................................. 117
Methods....................................................................................... 120
Study site description...................................................... 120
Experimental design........................................................ 121
Invertebrate sampling...................................................... 122
Data analysis................................................................... 123
Results......................................................................................... 124
Benthic invertebrates...................................................... 127
Epizoic invertebrates....................................................... 136
Discussion................................................................................... 146
Acknowledgements..................................................................... 151
Literature cited............................................................................ 152
V.
Effects of fish and shorebird predation on benthic invertebrates in a Great
Lake coastal wetland............................................................................... 162
Abstract....................................................................................... 162
Introduction................................................................................. 163
Methods....................................................................................... 165
Study site description...................................................... 165
Experimental design........................................................ 166
Invertebrate sampling...................................................... 168
Shorebird and fish populations....................................... 169
Data analysis................................................................... 169
Results......................................................................................... 170
Discussion................................................................................... 185
Acknowledgements..................................................................... 189
Literature cited............................................................................ 190
VI.
Synthesis................................................................................................. 197
Introduction................................................................................ 197
Literature cited............................................................................ 206
iv
LIST OF FIGURES
CHAPTER 1: Introduction
Figure 1:
Map of the Ottawa NWR in June 4th, 2009....................................... 12
CHAPTER 2: Comparing fish and invertebrate communities in Great Lake coastal
wetlands and impounded wetlands
Figure 1:
Two-dimensional NMS ordinations of invertebrate communities in
June, July and August 2006. Samples were grouped by water depth
(Shallow, medium, deep) and wetland type (Imp., impoundment;
Open, Open coastal).......................................................................... 53
CHAPTER 3: Effects of fish predation on benthic invertebrate communities in a Great
Lake coastal wetland
Figure 1:
Sizes of predatory fish in Crane Creek Marsh.................................. 86
Figure 2:
Total invertebrate densities collected in each treatment from June to
October............................................................................................. 93
Figure 3:
Densities of dominant taxa collected in each treatment from June to
October............................................................................................ 95
Figure 4:
Two dimensional NMS ordinations of invertebrate communities on
each sampling date........................................................................... 98
CHAPTER 4: Predation of epizoic and benthic invertebrates by fish including common
carp Cyprinus carpio, in a Great Lakes coastal wetland
Figure 1:
Mean richness in benthic sediments (taxa/sample) from May to
September 2008.............................................................................. 128
Figure 2:
Mean total invertebrate densities in benthic sediments in May to
September 2008............................................................................. 130
Figure 3:
Mean densities of the four common taxa in benthic sediments in May
to September 2008........................................................................ 132
Figure 4:
Two dimensional NMS ordinations of benthic invertebrate
communities in May to September 2008....................................... 134
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Figure 5:
Mean invertebrate richness per Q. quadrula unionid..................... 137
Figure 6:
Mean (± SE) total number of invertebrates per Q. quadrula unionid in
September 2008............................................................................. 139
Figure 7:
Mean (± SE) number of common taxa per Q. quadrula unionid in
September 2008............................................................................. 141
Figure 8:
Two dimensional NMS ordinations of invertebrate communities on
the live or dead Q. quadrula unionids in treatments (Fish, Fishless,
Carp)............................................................................................... 143
CHAPTER 5: Effects of fish and shorebird predation on benthic invertebrates in a Great
Lake coastal wetland
Figure 1:
Water depths in treatment area from July 1, 2009 to November 31,
2009............................................................................................... 172
Figure 2:
Mean Shannon diversity (±1 SE) in treatments from July to
November....................................................................................... 176
Figure 3:
Total invertebrate numbers (±1 SE) in each treatment from July to
November....................................................................................... 179
Figure 4:
Densities of common taxa (±1 SE) in treatments from July to
November....................................................................................... 181
Figure 5:
Two dimensional NMS ordinations of invertebrate communities Two
dimensional NMS ordinations of invertebrate communities in
September and November.............................................................. 184
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LIST OF TABLES
CHAPTER 2: Comparing fish and invertebrate communities in Great Lake coastal
wetlands and impounded wetlands
Table 1:
Fish collected in the open coastal wetland (Crane Creek Marsh) in
June, July and August 2006................................................................ 41
Table 2:
Fish collected in three impounded wetlands in June, July and August
2006..................................................................................................... 43
Table 3:
Invertebrate taxa in shallow, medium and deep depths in impounded
and open coastal wetlands at Ottawa National Wildlife Refuge......... 46
Table 4:
Two way ANOVAs results comparing densities of total invertebrates
and dominant taxa between wetland type (impounded, open coastal)
and depth (shallow, medium deep) in June, July, and August 2006... 48
Table 5:
Average density (number per m2) of common benthic invertebrates
between wetland type (impounded [Imp.], open coastal [Open]) and
depth (shallow, medium deep) in June, July, and August 2006......... 51
Table 6:
MRPP pairwise comparison p – values of invertebrate communities in
June, July, and August 2006.............................................................. 55
Table 7:
Indicator taxa for different habitats.................................................... 58
CHAPTER 3: Effects of fish predation on benthic invertebrate communities in a Great
Lake coastal wetland
Table 1:
Total number (n) and size, mean length of fish collected in June
to September 2007 in small mesh and large mesh fyke nets............ 85
Table 2:
Diets of common YOY fish................................................................ 88
Table 3:
Invertebrate taxa collected in exclosures at Crane Creek Marsh....... 90
Table 4:
Statistic results of ANOVAs comparing diversity, total invertebrates
and densities of the three dominant taxa among treatments on each
sampling date...................................................................................... 91
Table 5:
MRPP pairwise comparisons of invertebrate communities in
treatments......................................................................................... 100
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CHAPTER 4: Predation of epizoic and benthic invertebrates by fish including common
carp Cyprinus carpio, in a Great Lakes coastal wetland
Table 1:
Percent of each invertebrate taxa collected in benthic sediments and on
live and dead Q. quadrula unionids.................................................. 126
Table 2:
ANOVAs comparing densities of total invertebrates and common taxa
and richness in benthic sediments in Fish, Fishless, and Carp treatment
on each sampling date....................................................................... 129
Table 3:
MRPP pairwise comparisons of invertebrate communities in benthic
sediments........................................................................................... 135
Table 4:
ANOVAs comparing densities of total invertebrates and common taxa
and richness on live and dead Q. quadrula unionids in Fish, Fishless,
and Carp treatments.......................................................................... 138
Table 5:
MRPP pairwise comparisons of invertebrate communities on the live
or dead Q. quadrula unionids in treatments...................................... 144
Table 6:
Indicator taxa the taxa collected on the live or dead Q. quadrula
unionids in treatments....................................................................... 145
CHAPTER 5: Effects of fish and shorebird predation on benthic invertebrates in a Great
Lake coastal wetland
Table 1:
Shorebirds counts in CCM during the 2009 fall migration season... 173
Table 2:
Benthic invertebrate taxa in treatment areas. Numbers are the percent
that each taxa comprised of total invertebrates in a treatment across all
dates.................................................................................................. 175
Table 3:
Results of ANOVA tests comparing total invertebrates, diversity, and
numbers of the common taxa across treatments on each sampling
date.................................................................................................... 177
Table 4:
MRPP pairwise comparisons of invertebrate communities in
treatments......................................................................................... 183
viii
ACKNOWLEDGEMENTS
First, I would like to thank my wife Rita, my parents Mark and Nancy, and all the
rest of my Ohio and Illinois family and friends for all their support, help and advice
through this long process of research, classes, testing, teaching and writing. I would also
like to thank my advisor, Ferenc de Szalay for helping shape my interests in wetland
ecology. Without his direction, ideas and analysis I would not have been able to complete
these projects. I would also like to thank the other members of my committee, Robert
Carlson and Yoram Eckstein for all their advice and guidance; Mark Kershner, for his
direction, troubleshooting, statistics advice and fantasy football discussions. I would like
to recognize all the office and stock room personnel for their help and understanding in
filling out the reimbursement, ordering and a myriad of other forms. I would especially
like to thank the graduate student secretaries, Donna and Pat for their help with deadlines,
forms, and helping me sign up for classes every semester I forgot (which was most of
them).
I would like to thank the Art and Margaret Herrick Research grant and the Ohio
Division of Natural Resources Wildlife Diversity program grant for funding. Without
these and deals from home improvement stores, my research would not have been
possible. I would also like to thank the all personnel at the Ottawa National Wildlife
Refuge and specifically Doug Brewer, Ron and Kathy Huffman, Eddie Pausch and Jason
Lewis.
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Special thanks goes to my former lab mate Rick Bowers for all his help with field
research and designing projects. The long hot days we spent in the middle of a wetland
were made much easier with humorous stories or in-depth conversations on science
fiction books, TV shows, films and comic books. His help and counsel was invaluable
my first few years here. He passed away way too soon and his great ideas, hard work,
good attitude, musical talents, and understanding of ecology will be missed by all.
I would like to thank all the graduate students and undergraduates in the de Szalay
and Kershner labs for the long hours they put in helping me complete my research. The
days of leaky waders, knee deep sediment, wetsuits, inflatable pool floats, pool noodles,
mussel collecting, dead fish, late November sapling, sieving samples, 25 cent wings,
sunburn, heatstroke, Mexican food, Ralphies, camping, and construction did not go
unappreciated. Thank you: Jenn Clark, Justin Montemarano, Maureen Drinkard, Jackie
Johnston, Emma Kennedy, Allison Brager, Dan Sprockett, Larissa Rybus, Katy Gee,
Mauri Hickin, Emily Faulkner, Adam Custer, Brendan Morgan, Matt Begley, Kristyn
Shreve, John Reiner, Connie Hausman, Matt Eggert, Nathan Yaussy, Steve Robbins, Neil
Drinkard, Nolan Howard, Doug LaVigne, John Carney, and the Zanatta Lab at Central
Michigan University.
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CHAPTER 1
INTRODUCTION
Great Lake Coastal Wetlands
In this dissertation, I will describe the results of several experiments that
examined invertebrate and fish community ecology in Great Lakes coastal wetlands
(GLCW). In the first experiment, I tested if fish and invertebrate biodiversity and
abundance were different between unrestricted coastal wetlands and nearby impounded
wetlands. In a later experiment, I measured the effect that fish predation had on
invertebrate community structure, including zebra mussels attached on shells of native
unionid clams. Because common carp, Cyprinus carpio, are a common invertebrate
predator in these wetlands, I also examined the effect they had on invertebrates. Finally,
I used exclosure studies to determine if fish and shorebirds were competing for
invertebrate food resources in GLCW.
Great Lakes coastal wetlands occur along the Laurentian Great Lake coastlines, in
areas that are intermittently or permanently connected to the lake. Although much
research has examined invertebrate ecology in inland wetlands (Batzer & Wissinger
1996), little is know about GLCW communities. Coastal wetlands provide valuable
functions such as flood storage, sediment control, and shoreline erosion protection
1
2
(Cardinale et al., 1998), and they are areas of high productivity and biodiversity
(Herdendorf, 1987; Krieger, 1992; Randall et al., 1996; Brazner & Beals, 1997; Cardinale
et al., 1998). They are important as habitats for fish, birds, amphibians, reptiles,
invertebrates and mammals (Becker, 1983; Harris et al., 1983; Herdendorf, 1987; Jude &
Papas, 1992; Prince et al., 1992; Maynard & Wilcox, 1997; Cardinale et al., 1998;
Bowers & de Szalay, 2004; Uzarski et al., 2005; Bouchard, 2007). In these systems,
benthic invertebrates and zooplankton are key components of trophic webs, and they
drive ecosystem-level processes such as nutrient cycling (Herdendorf, 1987; Carney &
Elser, 1990; Arnott & Vanni, 1996; Covich et al., 1999; Vaughn & Hakenkamp, 2001;
DeVine & Vanni, 2002; Vadeboncoeur et al., 2002).
The GLCW are unique because their hydrology is very different than inland
wetlands. Water levels in these wetlands are affected by lake water level changes,
including annual, seasonal, and short-term variations (Maynard & Wilcox, 1997). Wind
set-up are short term changes that occur when water levels increase in the direction of
strong prevailing winds. For example, a storm with winds of 80 km/hr caused a 2-m
change in lake water levels (Herdendorf, 1987). Seiches are short term oscillations that
occur when a large wind set-up causes lake levels to “rock” back and forth across the
width of the lake basin. Seasonal variations occur due to changes in water inputs from
tributaries to the Great Lake. Water level in the lakes decrease throughout winter as
snow accumulates in the watershed, and they rise in late spring from snow melt.
Therefore, peak water levels in Lake Erie wetlands occur in May - June and the lowest
levels usually occur during January - February (Maynard & Wilcox, 1997). Water levels
3
also change on a year to year basis due to a-cyclic changes in weather patterns
(Herdendorf, 1987).
GLCW are also affected by physical disturbances such as wave action and ice
scour. These processes disturb sediments and increase turbidity, which uproot aquatic
plants and alter macrophyte diversity. Disturbances in concert with water level changes
affect the location of plant communities across long-term temporal scales (Herdendorf,
1987). These disturbances reduce competitive dominant plant taxa such as trees and
shrubs that would eventually crowd out herbaceous plants. Thus, disturbance creates
vacant habitat patches that are rapidly colonized by annuals and diversify the wetland
community (Herdendorf, 1987).
There are currently 1,200 km2 of wetlands along the Laurentian Great Lakes
(Herdendorf, 1987). Over 60% of historic GLCW have been drained for agriculture or
urban development since European settlement of the Great Lakes region (Comer et al.,
1995), and many remaining wetlands are highly degraded (Krieger, 1992). Furthermore, a
large portion of remaining wetlands are impounded with dikes to manage their water
levels, and therefore they do not experience natural water level fluctuations (Maynard &
Wilcox, 1997). Impounded wetlands provide some important ecosystem functions (e.g.
wildlife habitat), but other functions that are dependent on a connection to the lake (e.g.,
nutrient transformation or uptake from lake water) are lost (Herdendorf, 1987).
Many coastal wetlands are degraded when they are colonized by invasive animals
(common carp, Cyprinus carpio; zebra mussels, Dreissenia polymorpha) and plants
(purple loose strife, Lythrum salicaria; common reed, Phragmites australis). Feeding by
4
carp degrade wetlands because they increase water turbidity by feeding on benthic
invertebrates (Lougheed et al., 2004). This is especially common in wetlands in the
western basin of Lake Erie because they have fine benthic sediments that are easily
suspended in the water column (Herdendorf, 1987). Turbid GLCW lose their emergent
macrophyte communities and become plankton dominated systems (Chow-Fraser, 1998).
Invertebrate and Fish Communities
Understanding the community ecology of coastal wetlands is important because
aquatic invertebrates provide a valuable food resource for fish and wildlife (e.g.
shorebirds) (Herdendorf, 1987; Skagen & Oman, 1996), drive processes such as nutrient
cycling (Carney & Elser, 1990; Arnott & Vanni, 1996; Covich et al., 1999; Vaughn &
Hakenkamp, 2001; DeVine & Vanni, 2002) and are key components in detritus
processing (Konishi et al., 2001; Mancinelli et al., 2002; Ruetz & Newman, 2002).
Common benthic and epiphytic invertebrates taxa include: insects (e.g., Diptera,
Ephemeroptera), oligochaetes, crustaceans (e.g., amphipods, copepods, cladocerans,
ostracods), and molluscs (e.g., snails, unionid mussels, zebra mussels and sphaeriid
clams) (Herdendorf, 1987; Cardinale et al., 1998; Bowers et al., 2005). Invertebrate
community structure is affected by many abiotic factors such as sediment depth, water
level fluxuations (Cooper et al., 2007; Baumgärtner et al., 2008) and distance from the
shoreline (Cardinale et al., 1997; 1998). Biotic factors such as the physical structure of
plant stands also affect invertebrate density and diversity (Gilinsky, 1984; de Szalay &
5
Cassidy, 2001; Brown et al., 1988). Furthermore, plant detritus provide food
(Rasmussen, 1985; Murkin, 1989; Neill & Cornwell, 1992; Bunn & Boon, 1993) and
habitat (Campeau et al., 1994, de Szalay & Cassidy, 2001).
Fish predation is another biotic factor affecting aquatic invertebrates. There are
few studies of fish predation on benthic invertebrates in GLCW, but fish have
pronounced effect on invertebrates in other habitats such as lakes and streams (Crowder
& Cooper, 1982; Flecker 1984; Gilinsky, 1984; Morin, 1984; Mittlebach, 1988; Power,
1990; Diehl, 1992; Hanson & Riggs 1995; Batzer, 1998; Haas et al., 2007; but see Thorp
& Bergey, 1981; Schilling et al., 2009; 2009b). While changes in invertebrate numbers
are usually caused by impacts of fish predation, fish may indirectly affect invertebrate
communities by reducing competitively dominant invertebrate taxa (Batzer et al., 2000),
removing submergent macrophytes (Crivelli, 1983; Lodge, 1991; Sidorkewicj et al.,
1996; Mitchell & Perrow, 1998; Haas et al., 2007) and increasing turbidity (Gido, 2001;
2003).
Impounded wetlands usually have different fish communities and overall lower
fish diversity than open GLCW (Jude & Pappas, 1992; Johnson et al., 1997; Markham et
al., 1997). Great Lakes coastal wetlands are important habitat for many economically
valuable fish species, including channel catfish, yellow perch, crappie, bullhead catfish,
bluegill and gizzard shad (Jude & Pappas, 1992). Some of these species inhabit coastal
wetlands year round (e.g. bluegill and bullhead catfish), while others use wetlands to
breed (e.g. gizzard shad and yellow perch) (Herdendorf, 1987; Jude & Papas, 1992;
Maynard & Wilcox, 1997). Most studies have found a positive correlation between fish
6
diversity and macrophyte density (Minns et al., 1994; Randall et al., 1996; Brazner &
Beals, 1997; Weaver et al., 1997; Hook et al., 2001; Lougheed et al., 2001; Cvetkovic et
al., 2010).
Shorebirds
Many GLCW are also important habitats for shorebird that migrate through the
region. Common shorebirds species found in these wetlands include Dunlin, Killdeer,
Sandpipers, Greater Yellowlegs and Lesser Yellowlegs (Shieldcastle, 2010). Coastal
wetlands along western Lake Erie, including Crane Creek Marsh, are part of the Western
Hemisphere Shorebird Reserve Network (WHSRN), which includes 60 sites in 8
countries. WHSRN’s mission is to “conserve shorebird species and habitat across the
Americas through a network of key sites”.
Shorebirds use Lake Erie wetlands to feed on benthic invertebrates (Myers et al.,
1987; Helmers, 1992), although few studies have examining their effects in these
habitats. In some ocean coastal areas, shorebird predation reduces invertebrate densities
(Schneider, 1978; Quammen, 1981; Schneider & Harrington, 1981; Wilson, 1989, 1991;
Székely & Bamberger, 1992; Mercier & McNeil, 1994; Botto et al., 1998). However, in
other ocean coastal areas and riverine wetlands, shorebirds had little effect on benthic
invertebrate communities (Bay, 1974; Vienstein, 1978; Quammen, 1984; Schneider,
1985; Ashley et al., 2000; Mitchell & Grubaugh, 2005; Hamer et al., 2006). Shorebirds
are opportunistic predators that feed on the most abundant and easy to catch invertebrate
7
species (Hamer et al., 2006; Skagen & Oman, 1996), and birds alter their dietary
preference to local invertebrate communities (Skagen & Oman, 1996). Their diets
overlap with many common GLCW fish (Skagen and Oman, 1996; Pothoven et al., 2009;
Olson et al., 2003; Diehl, 1992; McNeely, 1977), but to my knowledge, there are no
studies that examined if fish and shorebirds compete for food resources in GLCW.
Invasive Species
Currently, a major ecological problem in GLCW are the introduction of invasive
plant, invertebrate, and vertebrate species (Jude & Papas, 1992; Glassner-Shwayder ,
2000; Lougheed et al., 2004; Bowers et al., 2005; Bowers & de Szalay, 2007). One such
species is the zebra mussel (Family Dreissenidae, Dreissena polymorpha), which was
introduced into Lake St. Clair in 1988 from Eurasia. They have spread throughout the
Great Lakes and Mississippi River drainage and have become a dominant species in the
lower Great Lakes (Ricciardi et al., 1995; 1998; Strayer & Smith, 1999). The adult
mussels use byssal threads to attach to hard substrates, including the shells of native
unionid mussels (Chase & Bailey, 1999, 1999b; Toczlowski et al., 1999; Bowers & de
Szalay, 2004; 2007). Zebra mussels and unionids are filter feeders, and thus they
compete for food (Strayer & Smith, 1996; Barker & Levinton, 2003). As a result, native
unionids have been largely extirpated from the lower Great Lakes, except for a few
remnant populations (Gillis & Mackie, 1994; Schloesser & Nalepa, 1994; Schloesser et
al., 1996; Crail et al., 2011). A few unionid mussel populations have recently been found
8
in coastal wetlands in the lower Great Lakes (Zanetta et al., 2002; Bowers & de Szalay,
2004), including a community in Crane Creek Marsh on Lake Erie. This wetland is the
location of my research (see Study Site Description below). For example, the unionid
community in this wetland includes 15 species and is dominated by the native Mapleleaf,
Quadrula quadrula (Bowers & de Szalay, 2004). Some research has examined why
unionids still persist in Great Lake Coastal wetlands (Nichols & Wilcox, 1997; Bowers &
de Szalay, 2007), but no clear patterns have emerged.
Another important invasive species in GLCW is the common carp (F. Cyprinidae,
Cyprinus carpio). This species was introduced into the United States from Eurasia in the
1800’s (there are varying accounts for their introduction from the early to late 1800’s)
(USGS, 2012). They have spread throughout the United States, and they are abundant
throughout all of the Great Lakes (USGS, 2012). Carp impact the habitat in GLCW
because they disturb the sediments and uproot plants when they feed on benthic
invertebrates. High numbers reduce macrophyte density, increase turbidity and and
reduce benthic invertebrates that are food for native species (Riera et al., 1991; Lougheed
et al., 1998; Parkos et al., 2003; Lougheed et al., 2004; Pinto et al., 2005; Haas et al.,
2007; Weber & Brown, 2009). Carp can reduce native fish populations through
competition for food resources and habitat disturbance (McNeely & Pearson, 1977;
French, 1988; Savino & Stein, 1989; Diehl, 1992; Jude & Papas, 1992; Hook et al., 2001;
Lougheed et al., 2001; Olson et al., 2003; Pothoven et al., 2009). However, carp also
feed on invasive zebra mussels, which may have positive impacts on native species, such
9
as unionid mussels, that compete with these exotic species (Tucker et al., 1996; Thorp et
al., 1998; Magoulick & Lewis, 2002).
Hypotheses
There is little research examining interactions among benthic invertebrates, fish
and shorebird in GLCW, which is the focus of this dissertation. For my first study in
2006, I compared fish and benthic invertebrate communities in a Great Lakes coastal
wetland, Crane Creek Marsh, with nearby impounded wetlands. In 2007 and 2008, I
studied the effects of fish predation on benthic invertebrate communities and exotic
dreissenid mussels. This included an enclosure study that tested if common carp were an
important predator of invertebrate communities. Finally in 2009, I tested if fish and
shorebird compete for invertebrate prey in Crane Creek Marsh.
The four main hypotheses I tested are:
H1: There will be lower fish density and diversity in impounded wetlands than
the unimpounded coastal wetland because impounded wetlands have restricted access to
the Lake. Invertebrate density and diversity will be higher in impoundments because
plant stands are more dense and fish populations are lower.
H2: Fish predation will alter the benthic invertebrate community structure in the
GLCW by changing taxa dominance and reducing overall abundance and diversity.
Small fish will disproportionately affect invertebrates because the coastal wetlands
provides fish breeding habitat and there will be very high numbers of offspring at times.
10
H3: Carp are an important predator and they greatly reduce benthic invertebrate
densities and diversity. Carp also reduce densities of zebra mussels and other epizoic
invertebrates that are attached on the shells of native unionid mussels.
H4: Both fish and shorebird impact the benthic invertebrate communities in
coastal wetlands. However, fish have a greater impact on invertebrate numbers in
summer when water levels are high and they are breeding, and shorebirds have a greater
impact during their fall migration period.
Study Site Description
My dissertation research was conducted at Ottawa National Wildlife Refuge
(ONWR), located at 14,000 West State Route 2, Oak Harbor, Ohio 43449 (Ottawa
County, Ohio). The ONWR was established in 1961, and consists of over 9,000 acres,
including an open coastal wetland, Crane Creek Marsh, and 12 impounded wetlands
(Figure 1). ONWR is managed by the US Fish and Wildlife Service to provide habitat
for birds (e.g. migratory shorebirds and waterfowl), other wildlife (muskrats, mink,
amphibians, and reptiles) and fish (USFWS, 2008). Common benthic invertebrates in
Crane Creek Marsh include: chironomids, oligochaetes, sphaeriids and amphipods.
Common unionid mussels include: Quadrula quadrula, Leptodea fragilis, Amblema
plicata and Pyganodon grandis (Bowers and de Szalay, 2004).
Crane Creek Marsh is a 312 ha wetland where Crane Creek flows into Lake Erie.
The creek has a 144 km2 watershed, which consists mostly row crop agriculture such as
11
Figure 1: Aerial image of the Ottawa NWR in June 4th, 2009 (Google Earth, 2012). The
Ottawa NWR is outlined in orange. The locations of Crane Creek Marsh and the
impounded wetlands that I sampled are labeled. The sites of my fish and shorebird
exclosure experiments are shown with colored squares.
12
13
corn, soybeans and wheat. Benthic sediments are 10–50 cm of soft inorganic silt-clay
over hard packed clay and sand (Bowers et al., 2005). The wave action present in the
marsh along with the soft silt-clay sediments and access by carp leads to high turbidity.
Water clarity is generally less than a few cm (D. Kapusinski, pers. observ.)
Crane Creek Marsh is open to Lake Erie via a 4-m wide opening in a rip-rapped
levee that runs along the northern border of the wetland. The levee protects the marsh
from storm waves, but the opening allows marsh water levels to fluctuate with changes in
Lake Erie (Bowers & de Szalay, 2004). Lake Erie water levels are generally highest in
June and lowest in February, which leads to high water levels in wetland in the early
summer. Water levels throughout the wetland are generally < 50 cm in depth, but they are
up to 2 m in some areas along the southern and eastern edge of CCM (Bowers and de
Szalay, 2004). Short-term water level changes also affect CCM, with daily water levels
depending on the strength and direction of prevailing winds that cause wind-set ups and
seiches. Typical water level fluctuations are < 1 m but can be larger with strong winds.
Winds out of the north and east increase water levels in the wetland, while winds out of
the south and west will decrease water levels.
The typical habitat in CCM is open water with interspersed patches of emergent
vegetation. Emergent vegetation occurs in scattered patches, and dominant plant taxa
include American water lotus (Nelumbo lutea), yellow pond lily (Nuphar advena),
common reed (Phragmites australis), cattail (Typha sp.), and several rush species (Juncus
spp.). Most stands of dense emergent vegetation are found along in the shallow edge of
14
the wetland where seedlings get enough light to grow in the turbid water and frequent
water level changes expose the sediments.
Water levels in the many impounded wetlands at ONWR are manually adjusted
with water control structures (i.e. stop log risers). Management efforts are designed to
promote dense stands of macrophytes that provide food and cover for migratory
waterfowl. The impounded wetlands I sampled at Ottawa NWR were flooded up to 2 m
deep via channels connecting to Crane Creek Marsh. These were semipermanentlypermanently flooded. Most wetlands at ONWR are drained every few years in spring or
summer to control the spread of invasive plants with tractor disking or spraying. The
impounded wetlands I sampled were dominated by emergent herbaceous vegetation such
as American water lotus, yellow pond lily, common reed, cattail, rush and sedge species
(Cyperacea).
15
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CHAPTER 2
COMPARING FISH AND INVERTEBRATE COMMUNITIES IN GREAT LAKE
COASTAL WETLANDS AND IMPOUNDED WETLANDS
Abstract
I sampled benthic invertebrate and fish in a Lake Erie open coastal wetland and
five nearby impounded wetlands. Five sampling sites were selected in the open coastal
marsh, and one site was chosen in each of the five impounded wetlands. In June, July and
August 2006, benthic invertebrates were sampled with an Ekman dredge in shallow
depths (<18 cm), medium depths (18 – 34 cm) and deep depths (>34 cm), and fish were
sampled nearby with fyke nets. I collected 26 fish species that are common in Lake Erie
marshes including bluegill, gizzard shad, emerald shiner, yellow perch and bullhead.
Fish were significantly larger in open coastal wetlands in June. However, I did not detect
any other differences in fish community structure between the wetland types. The
invertebrate community was dominated by Amphipoda, Chironomidae, Ceratopogonidae,
Cladocera, Copepoda, Ostracoda, Nematoda, and Oligochaeta. Total numbers and
densities of cladocerans, ostracods, nematodes and oligochaetes were higher in the open
coastal marsh, but ceratopogonids and copepods were more abundant in impounded
30
31
wetlands. Non-metric Multidimensional Scaling (NMS) analyses showed that
invertebrate community structure differed significantly between open coastal and
impounded wetlands. Indicator taxa analysis showed that Cladocera, Oligochaeta,
Ostracoda were consistently abundant in open coastal wetlands, and Amphipoda,
Ceratopogonidae, Copepoda, Hirudinea, Physidae, Corixidae, and Caenidae were
indicators for impoundment wetlands. Although water depth did not affect community
structure in impoundments, communities in shallow depths were different than in deep
depths in open coastal wetlands. Also, Corixidae, Lymnaeidae, Physidae were indicator
taxa for shallow depths. Differences in the fish and invertebrate communities in wetland
types are probably due to a combination of abiotic factors (i.e., water level changes,
dissolved oxygen) and biotic factors (i.e., plant structure, predation).
Introduction
Great Lake Coastal Wetlands (GLCW) are habitats along the Laurentian Great
Lakes that are hydrologically connected to the lake. In the western Lake Erie, extensive
coastal wetlands are important habitats for many species of fish and wildlife (e.g. wading
birds, waterfowl, and shorebirds). Daily, seasonal, and storm-related lake water levels
fluctuations cause water depth changes in the wetlands of up to several meters. Water
fluctuation transports nutrients and organic matter in and out of the wetland and also
causes scouring of exposed areas (Herdendorf, 1987; Bouchard, 2007). Typical daily
water level changes in Lake Erie are about 20 cm per day, which exposes benthic
32
sediments in shallow areas (Bowers & deSzalay, 2004). This can promote feeding by
shorebirds, but fish feed in the same areas when water levels increase (Skagen & Oman,
1996).
Aquatic invertebrates are key components of the biotic community in wetlands.
They are an important link in food webs because they consume wetland plants and
detritus and are food for fish and wildlife (Eadie & Keast, 1982; Diehl, 1992; Hanson &
Riggs, 1995; Skagen & Oman, 1996; Haas et al., 2007). They also impact ecosystem
processes such as nutrient cycling when they shred detritus and consume fine particulate
organic matter (Devine & Vanni, 2002; Vaughn & Hakenkamp, 2001; Arnott & Vanni,
1996). In other wetland habitats, invertebrate community structure is affected by habitat
characteristics such as sediment depth (Cooper et al., 2007), hydrology (Baumgärtner et
al., 2008), and fish predation (Schilling et al., 2009a, 2009b). Macrophytes also affect
water mixing, turbidity, dissolved oxygen and pH (Cardinale et al., 1997), which in turn
affect invertebrate communities. Macrophyte and benthic invertebrate biomass are often
positively correlated (Brazner & Beals, 1997; Weaver et al., 1997; Lougheed et al., 2001)
because dense vegetation can provide shelter from fish predation (Crowder and Cooper,
1982; Mittelbach, 1988; Schriver et al., 1995).
Fish are another key component of wetland communities, and fish predation can
have a top-down impact on invertebrate community structure (Brönmark, 1994;
Brönmark et al., 1992; Drenner, 1996). Wetland fish and macrophyte densities are often
positively correlated (Cardinale et al., 1998; Brazner & Beals, 1997; Lougheed et al.,
2001), but fish and invertebrate densities are often negatively correlated (Schilling et al.,
33
2009; 2009b, Dorn et al., 2006; Carpenter & Kitchell, 1993). Furthermore, bottom
feeding fish (e.g. common carp) affect macrophyte communities by uprooting plants
(Crivelli, 1983; Sidorkewicj, 1996; Zambrano, 1999), and increasing turbidity and water
column nutrients (Lamarra, 1975; Shorman & Cotner, 1997; Schaus et al., 1997). Thus
patterns of wetland macrophytes, invertebrate and fish populations are complex and will
vary with local habitat characteristics.
In the Great Lakes region, over 60% of the coastal wetlands have been drained,
and many remaining wetlands are degraded by pollution (Maynard & Wilcox, 1997;
Krieger, 1992). Other coastal wetlands have been impounded by dikes, and their
hydroperiods are manually controlled. Most of these impounded wetlands are managed
to provide dense emergent macrophyte stands used by migratory waterfowl. Although
these wetlands still provide valuable wildlife habitat, they are no longer affected by shortterm lake water level fluctuations and the water control structures block access by lake
fish. Previous studies found that fish diversity and abundance are lower in
impoundments than open coastal marshes (Jude & Pappas, 1992; Johnson et al., 1997).
Although many Great Lakes coastal wetlands are now impounded to provide wildlife
habitat, there are few studies that compared invertebrates and fish communities in open
coastal and impounded wetlands.
I compared macroinvertebrate and fish communities in an open coastal wetland in
western Lake Erie to those in nearby impounded wetlands. I also tested if hydrological
differences in these habitats affected invertebrate communities by sampling
macroinvertebrates along a water depth gradient in each habitat. My hypotheses were:
34
H1: Fish communities will be different in the open coastal wetland and
impounded wetlands. The open coastal wetland will have higher fish diversity than the
impounded wetlands, and lake species will be more common in the open coastal wetland.
H2: Species assemblages of invertebrate communities will also vary between
impounded wetlands and the open coastal wetland. Invertebrate density will be lower in
open coastal wetlands than impounded wetlands due to diverse and abundant fish
predators. However, invertebrate species richness will be greater in impounded wetlands
because of the physically complex plant community. There will be pronounced
differences among water depths in open coastal wetlands due to the fluctuating water
levels, although, communities will be similar among water depths in impoundments.
Methods
Study Site Description
This study was conducted at the Ottawa National Wildlife Refuge (Ottawa NWR)
in Oak Harbor, Ohio (Ottawa and Lucas Co.). Ottawa NWR includes many impounded
wetlands where the water levels are manually controlled with water control structures.
Impounded wetlands range in size from 25 to 120 hectares, and they are flooded canals
from Lake Erie. The impounded wetlands are shallow and occasionally drawn down in
summer to produce dense stands of emergent plant vegetation. Dominant plant species in
these wetlands are cattail (Typha spp.), common reed (Phragmites australis), American
35
water lotus (Nelumbo lutea), arrowhead (Sagittaria latifolia), pickerelweed (Pontederia
cordata), smartweed (Polygonum spp.) and bur-reed (Sparganium eurycarpum).
Crane Creek Marsh is a 312 ha open coastal wetland at Ottawa NWR, which is
permanently connected to the adjacent lake through a 4-m wide opening in an earthen
dike. Crane Creek Marsh is mostly shallow (< 2 m depth) open water with interspersed
patches of emergent vegetation such as cattail common reed, American water lotus and
soft stem bulrush (Scirpus acutus). For a complete habitat description, please see the
study site description in the Introduction chapter.
Experimental Design
In summer 2006, I sampled five impounded wetlands at the Ottawa National
Wildlife Refuge: MS3, MS4, MS6, MS8a and MS8b (see Figure 1 in Chapter 1). I
randomly chose one sample site in each wetland. I also chose sampling sites in Crane
Creek Marsh at 5 different locations in the wetland. All sampling sites had stands of
sparsely vegetated emergent plants and were at least 80 cm deep.
On 12 June, 20 July and 20 August 2006, fish communities were sampled with
fyke nets at each sample site. On each date, I randomly selected one impoundment and
one sampling site in Crane Creek Marsh. Thus, the three impoundments were each
sampled once, but Crane Creek was sampled on all three dates at different locations.
I set a large mesh fyke net (1.3 cm mesh) and a fine mesh fyke net (0.5 cm mesh)
in 70 – 80 cm water depth in sparse plant cover. Each net had a 15-m lead net orientated
36
perpendicular from the shoreline to the catch net, and two 3-m wing walls. Fish were
sampled for 24 hours, and all trapped fish were identified, counted and measured to
determine total length. I collected voucher specimens of all species, which were
preserved and identified in the laboratory. All other fish were released back into the
wetland. When there were more than 100 individuals of a species per net, I only
measured the length of the first 100 individuals. The remaining individuals were
identified, counted and released.
Invertebrates were sampled with an Ekman dredge (15 cm x 15 cm) on 11-13
June, 19-21 July, and 21-23 August, 2006. One impoundment was not sampled in
August because it had been drained. To sample benthic invertebrates, I established two
permanent transects at each sampling site (10 transects in the open coastal wetlands and
10 transects in the impounded wetland). Transects were perpendicular to the shoreline,
and ran from shoreline to water that was 50 cm deep; transect length varied depending on
the wetland morphology. Areas with shallow (<18 cm), medium (18-34 cm), and deep
(>34 cm) water depths were marked on each transect. On each sampling date, I collected
a sample in each water depth from each transect. To avoid pseudo replication (Hurlbert,
1984), I combined the samples taken from the same water depth in the two transects at
each sample site. All samples were collected 1 m away from the transect line, to avoid
disturbing the sediments. The collected material was drained in a 300-micron mesh
screen and preserved in 90% ethanol in Ziploc bags. In the lab, samples were rinsed a
through a 300-micron mesh screen to remove silt, and invertebrates were removed under
a dissecting microscope. Invertebrates were identified to the lowest practical taxonomic
37
level (usually family or order) using Peckarsky et al. (1990) and Merritt et al., (2008) and
enumerated.
Data Analysis
I classified invertebrates as dominant taxa if they comprised >3% of all
individuals collected in either impounded wetlands or open coastal wetlands. I used
Tanner et al. (2004), Jude and Pappas (1992) and Herdendorf (1987) to classify fish taxa
as Lake species if they are commonly found in the Great Lakes and only use wetlands as
breeding habitat. I also classified fish as invertebrate predators if they fed mostly on
invertebrates (Pothoven et al., 2009; Olson et al., 2003; Diehl, 1992; McNeely, 1977;
Pearse, 1921; Lindeman, 2006; Morrison et al., 1997; Serrouya et al., 1995; Haas, 2007;
Ellison, 1984, Gido, 2001; Gido, 2003). All data were tested for normality and
transformed log (x+1) for count data, arc sin (square root (x) for percent data) if needed.
I then tested if the transformed data was normally distributed, and used the type of data
(raw or transformed) that best approximated normality.
On each sampling date, I used 2-way ANOVAs to test if total macroinvertebrate
densities and dominant taxa densities differed by water depth and wetland type. When
ANOVAs were significant (P < 0.05), I made pair-wise comparisons among the means
with Tukey’s HSD tests. Because the fish were only sampled once per wetland type on
each sampling date, I used the data collected on the three sampling dates as replicates. I
compared total numbers, richness, percent lake species and percent invertebrate feeders
between wetland types with paired sample T-tests (data were paired by sampling date).
38
Fish sizes were compared between wetland types and dates with a 2-way ANOVA
followed by Tukey’s HSD tests. All univariate statistics were run on JMP software
(JMP 7.0.1, 2007, Cary, NC).
On each sampling date, a Non-metric Multidimensional Scaling procedure (NMS)
was used to graphically compare invertebrate community structure in the six habitats
defined by wetland type and water depth. Fish communities were compared between
impounded wetlands and the open coastal wetland using the data collected on the three
sampling dates. The NMS ordinations were run with Bray-Curtis scaling procedures with
Sorenson distance measures. I used a random starting point with 50 runs and 500
iterations. I also compared the invertebrate communities with a Multi-Response
Permutation Procedure (MRPP) to determine if there were community-level differences
between the three water depths within the two wetland type. The MRPP was based on
Bray-Curtis scaling procedures with Sorenson distance measures using a standard
n/sum(n) weighting of groups. I also tested if there were invertebrate indicator taxa date
using the methods of Dufrene and Legendre (1997). Significance of indicator association
with a wetland type was determined with a Monte Carlo test using 50 runs of randomized
data. The multivariate analysis were performed using PC-ORD version 5.1 (McCune and
Mefford, 2006).
Results
Fish
39
I caught a total of 3,520 fish in 26 species at Ottawa NWR (Tables 1, 2). Fish
were larger in June (8.71 cm [± 0.42 cm]) than in July (5.12 [± 0.25 cm]) and August
(7.39 [± 0.59 cm]), and all dates were significantly different from each other (F 2, 1717 =
27.16, p < 0.0001). The June catch was mostly adult fish, the July catch was mostly
young of year (YOY) fish, and the August catch was a mix of YOY individuals and adult
fish (Tables 1 and 2). Mean fish size was larger in the open coastal wetland (6.42 [± 0.41
cm]), than impounded wetlands (6.12 cm [± 0.21 cm]), however this difference was not
significant (F1, 1719 = 0.49, p = 0.262). The Date by Wetland type interaction was
significant (F5, 1714 = 42.22, p < 0.0001). In June, mean fish size was the higher in the
open coastal wetland (19.7 cm [± 1.2 cm]) than impounded wetlands (7.1 cm [± 0.5 cm]).
In July, mean fish size was slightly higher in the open coastal wetland (5.3 cm [± 0.4
cm]) than impounded wetlands (4.9 cm [± 0.3 cm]). In August, mean fish size in the
open coastal wetland (6.3 cm [± 0.8 cm]) was lower than in impounded wetlands (7.9 cm
[± 0.6 cm]).
Although slightly more species were collected in the open coastal wetland (24)
than the impounded wetlands (17), the difference was not significant (T 1, 5 = 1.84, p =
0.140). The total number of fish collected in impounded wetlands (1,998) and open
coastal wetlands (1,561) were not different between the wetland types (T 1, 5 = 0.24, p =
0.821), Fish total numbers peaked in July (2,890) and were the lowest in August (254).
Fish species assemblages were usually similar between wetland types. Numbers
of lake fish peaked in July, and gizzard shad, emerald shiner and yellow perch were the
40
Table 1: Fish collected in the open coastal wetland (Crane Creek Marsh) in June, July and
August 2006. Fish were trapped with small mesh (SF) and large mesh (LF) fyke nets.
Total is total number collected, Size is mean (SE) Snout-Tail length (cm) of the first 100
individuals of that species collected in the net. Lake Fish are labeled with an *, which are
species found mostly in Lake Erie but use wetlands as breeding habitat. Predominant
feeding type are labeled as: 1Invertebrate predators; 2Detritus feeders; 3Piscivores.
41
June
July
LF
SF
Species Name
Common Name
Total
Avg size
Ameiurus natalis
Yellow bullhead* 1
Ameiurus nebulosus
Brown bullhead* 1
Amia calva
Bowfin3
Aplodinotus grunniens
Fresh water drum
Carassius auratus
Goldfish1
Cyprinus carpio
Common carp1
Dorosoma cepedianum
Gizzard shad2
Ictalurus punctatus
Channel catfish1
Lepisosteus osseus
Longnose gar3
Lepomis gibbosus
Pumpkinseed sunfish* 1
Lepomis gulosus
Warmouth* 1
Lepomis humilis
Orange spotted sunfish* 1
Lepomis macrochirus
Bluegill* 1
8
11.35 (1.0)
Micropterus salmoides
Large mouth bass* 3
1
34.9
Morone chrysops
3
2
1
62.0
42
6
1
16.8
1
3.7
1
52.5
1
56.0
1
6.1
LF
Avg size Total
29.0
55.7 (2.7)
50.2 (8.8)
SF
Total Avg size Total
1
August
LF
2
58 (2.0)
65.5 (6.5)
SF
Avg size Total
Avg size Total Avg size
3.2 (0.1)
3
3.6 (0.4)
8
56.4 (2.2)
520
4.3 (0.1)
1
62.0
2
1
55.8 (7.3)
1
7.2
17
5.7 (0.3)
1
5.4
34.2
2
7.8 (0.8)
2
4 (0.6)
300
2.8 (0.1)
10
4.4 (0.6)
White perch
1
8.0
Morone saxatilis
Striped bass*
1
9.9
Notemigonus crysoleucas
Golden shiner* 1
7
2.9 (1.1)
Notropis atherinoides
Emerald shiner1
290
3.6 (0.1)
28
4.0 (0.2)
Notropis hudsonius
Spottail shiner1
39
3.8 (0.1)
6
5 (0.2)
Noturus gyrinus
Tadpole madtom*
2
7 (0.5)
Perca flavescens
Yellow perch1
135
3.4 (0.1)
18
5.1 (0.2)
Pimephales promelas
Fathead minnow*
Pomoxis annularis
White crappie* 1
26
3.5 (0.1)
7
6.5 (0.2)
Pomoxis nigromaculatus
Black crappie* 1
1
12.1
13
11.8 (1.0)
5
7.2 (0.4)
1
5.4
2
21.6 (0.2)
1
18.8
42
Table 2: Fish collected in three impounded wetlands in June, July and August 2006 with
small mesh (SF) and large mesh (LF) fyke nets. Total is total number of fish collected.
Size is mean (SE) Snout-Tail length (cm) of the first 100 individuals of that species
collected in the net. Lake Fish are labeled with an*, which are species found mostly in
Lake Erie but use wetlands as breeding habitat. Predominant feeding type are labeled as:
1
Invertebrate predators; 2Detritus feeders; 3Piscivores.
43
June, MS8b
LF
July, MS5
SF
Total
August, MS3
SF
Ameiurus melas
Black bullhead* 1
Ameiurus natalis
Yellow bullhead* 1
Ameiurus nubulosus
Brown bullhead* 1
1
25.1
229
5.0 (0.2)
2
19.0 (6.5)
6
9.0 (2.1)
Amia calva
Bowfin3
1
55.0
2
55.0 (1.0)
6
50.2 (5.0)
3
57.9 (0.9)
Carassius auratus
Goldfish1
1
4.9
1
6.9
50
5.0 (0.1)
Cyprinus carpio
Common carp1
2
51.1 (2.1)
106
5.6 (0.9)
Dorosoma cepedianum
Gizzard shad2
941
3.6 (0.1)
Lepomis cyanellus
Green sunfish1
1
9.0
29
4.6 (0.4)
Lepomis gibbosus
Pumpkinseed sunfish* 1
Lepomis gulosus
Warmouth* 1
6
8.7 (0.4)
7
7.9 (0.2)
Lepomis humilis
Orange spotted sunfish* 1 8
8.0 (0.2)
5
6.7 (0.5)
2
6.2 (0.3)
Lepomis macrochirus
Bluegill* 1
8.0 (1.0)
16
5.4 (0.2)
28
4.6 (0.4)
Micropterus salmoides
Large mouth bass* 3
Notropis atherinoides
Emerald shiner1
25
4.5 (0.1)
Noturus gyrinus
Tadpole madtom*
5
7.4 (0.2)
Pomoxis annularis
White crappie* 1
1
18.0
Pomoxis nigromaculatus
Black crappie* 1
14
Avg size Total
SF
Common Name
16.9 (1.9)
Total Avg size Total
LF
Species Name
7
Avg size
LF
Avg size Total Avg size
10.4 (1.2)
6
114
3
12.0 (1.8)
5
12.6 (1.1)
21
8.3 (0.2)
221
6.3 (0.1)
2
2
Avg size Total
4.8 (0.5)
4.2 (0.1)
7.0 (0.2)
59
3.3 (0.1)
1
16.0
38
3.3 (0.1)
1
8.8
5
5.3 (0.6)
9
9.8 (0.6)
1
7.4
5
5.7 (0.3)
44
most abundant taxa. The most common invertebrate predators were sunfish and catfish
species. Lake fish were slightly more common in open coastal wetlands (50% of all fish)
than impoundments (30% of all fish), but this difference was not statistically significant
(T1,5 = 2.77, p = 0.346). The percent of fish that were invertebrate predators was also not
different between open coastal wetlands (78%) and impounded wetlands (70%) (T1,5 =
2.77, p = 0.734). Multivariate analyses did not detect differences in fish communities in
the wetland types (NMS 2-dimensional solution, F1,5 = 3.39, p = 0.1176; MRPP, p =
0.486). No indicator taxa were detected for either wetland type.
Benthic Invertebrates
I collected over 60,000 invertebrates in 35 taxa in our benthic samples (18,007 in
the impounded wetlands and 44,707 in the open coastal wetland) (Table 3). Dominant
taxa comprising over 3% of all individuals in either wetland type were Chironomidae,
Oligochaeta, Cladocera, Ostracoda, Nematoda, Amphipods, Copepoda and
Ceratopogonidae.
Invertebrate total numbers changed throughout the course of the experiment.
Total numbers were highest in June in all wetlands. Total numbers were significantly
higher in the open coastal wetland than the impounded wetlands on all dates (Table 4).
For example, total numbers in June were in open coastal wetlands (99,661 /m2 [± 16,660
/m2]) were three times higher than in impounded wetlands (30,867 /m2 [± 14,422 /m2]).
45
Table 3: Invertebrate taxa in shallow, medium and deep depths in impounded and open
coastal wetlands at Ottawa National Wildlife Refuge. Values are percent of total for each
taxa across all sample dates. Total are the total number collected in each habitat type.
46
Impounded
Taxa
Shallow
Open
Medium
Deep
Shallow Medium
Deep
Insects
Coleoptera
Chrysomelidae (Donacia)
Haliplidae (Haliplus)
<1
<1
Halipidae (Peltodytes)
<1
<1
Hydrophiloidea (Berosus)
<1
<1
<1
<1
<1
<1
<1
Collembola
Agrenia
<1
<1
Diptera
Ceratopogonidae
Chironomidae
Empididae
Pelecorhynchidae (Glutops)
5.9
4.1
5.9
<1
<1
<1
18.4
32.2
34.0
23.1
21.0
19.8
<1
<1
<1
<1
<1
<1
1.2
<1
<1
<1
<1
<1
Simulidae
Stratiomyidae (Stratiomys)
<1
Tabanidae
<1
<1
<1
1.1
3.2
1.8
Ephemeroptera
Caenidae (Caenis)
Ephemeridae (Ephemera)
<1
<1
<1
Hemiptera
Corixidae
Mesoveliidae (Mesovelia)
1.3
Naucoridae (Pelocoris)
<1
Pleidae (Neoplea)
<1
Odonata
Coenagrionidae (Argia)
Coenagrionidae (Nehalennia)
Libellulidae (Erythemis)
<1
<1
<1
<1
<1
<1
<1
<1
<1
<1
<1
<1
<1
<1
1.0
<1
<1
<1
7.9
10.4
3.1
<1
<1
Crustaceans
Amphipoda
Asellidae (Caecidotea )
Cladocera
5.0
1.8
2.0
9.7
6.1
12.8
Copepoda
15.1
11.3
20.8
<1
<1
<1
14.5
8.3
11.4
15.8
35.2
17.0
1.4
1.5
<1
<1
<1
<1
Lymnaeidae
<1
<1
<1
<1
<1
Physidae
2.6
<1
<1
2.8
<1
<1
Planorbidae
2.4
1.0
1.3
<1
<1
Sphaeriidae
<1
<1
<1
<1
<1
<1
<1
<1
2.4
<1
1.9
16.7
12.3
13.6
Cambaridae
Ostracoda
<1
Leeches
Hirudinea
Molluscs
Dreissenidae
Roundworms
Nematoda
1.7
1.9
Segmented worms
Oligochaeta
Totals
19.2
20.2
14.2
27.8
21.4
31.6
4701
7303
6003
19575
17731
7401
47
Table 4: Two way ANOVAs results comparing densities of total invertebrates and
dominant taxa between wetland type (impounded, open coastal) and depth (shallow,
medium deep) in June, July, and August 2006. Significant differences (p ≤ 0.05) are
bold. Degrees of freedom for June and July are 1, 29; 2, 28 and 5, 25 for wetland type,
water depth and the interaction term respectively. Degrees of freedom for August are 1,
26; 2, 25 and 5, 22 for wetland type, water depth and the interaction term respectively.
48
Taxa
Date
Wetland
Type
Depth
Interaction
Wetland
Ceratopogonidae Type
Depth
Interaction
Wetland
Type
Chironomidae
Depth
Interaction
Wetland
Type
Amphipod
Depth
Interaction
Wetland
Type
Cladocera
Depth
Interaction
Wetland
Type
Copepoda
Depth
Interaction
Wetland
Type
Ostracoda
Depth
Interaction
Wetland
Type
Nematoda
Depth
Interaction
Wetland
Type
Oligochaeta
Depth
Interaction
Total number
June
F; p value
July
F; p value
August
F; p value
11.38; 0.004
1.88; 0.188
2.24; 0.141
7.02; 0.014
0.84; 0.445
1.78; 0.189
6.25; 0.021
1.62; 0.221
2.40; 0.115
6.19; 0.025
0.34; 0.716
3.91; 0.059
0.39; 0.685
0.88; 0.358
0.37; 0.699
0.37; 0.696
0.34; 0.718
0.17; 0.849
3.53; 0.087
1.19; 0.330
0.97; 0.401
0.23; 0.636
0.07; 0.935
3.29; 0.054
1.40; 0.249
0.72; 0.499
3.02; 0.070
2.30; 0.151
0.72; 0.502
0.71; 0.509
2.92; 0.100
1.86; 0.178
0.29; 0.751
2.73; 0.113
1.64; 0.217
1.15; 0.335
17.43; <0.001
0.57; 0.576
0.37; 0.699
0.84; 0.368
2.69; 0.088
1.89; 0.174
1.83; 0.191
1.32; 0.289
1.51; 0.245
2.19; 0.159
7.62; 0.011
6.57; 0.018
0.17; 0.847
0.46; 0.639
0.07; 0.932
0.31; 0.739
0.39; 0.681
0.33; 0.724
4.05; 0.061
2.48; 0.128
8.21; 0.009
1.83; 0.195
0.98; 0.389
1.60; 0.226
1.91; 0.182
0.74; 0.489
1.52; 0.242
18.23; <0.001
12.84; 0.002
16.60;<0.001
2.29; 0.136
1.10; 0.349
2.74; 0.087
2.39; 0.125
1.09; 0.351
2.72; 0.089
15.01; 0.002
1.28; 0.306
13.70; 0.001
0.95; 0.402
7.17; 0.014
1.86; 0.180
1.03; 0.382
2.20; 0.133
3.82; 0.039
49
Total numbers were never different among water depths, and there were no significant
water depth X wetland type interactions (Table 4).
Population of most dominant taxa was highest in June, and declined in summer
(Table 5). On several sampling dates, numbers of many taxa differed between wetland
types but did not differ among water depths (Table 4). Cladocera, Ostracoda, and
Nematoda had higher numbers in open coastal wetlands than impounded wetland depths.
Oligochaeta numbers were also higher in open coastal wetland on all dates, and they had
a wetland type by water depth interaction in August. On this date, numbers in shallow
water were the highest in open coastal wetlands but the lowest in impounded wetlands.
Numbers of Ceratopogonidae and Copepoda were higher in impounded wetlands. The
two other dominant taxa, Chironomidae and Amphipoda, were not different between the
wetland types.
Species composition of invertebrate communities was different among the
habitats we sampled. On each sampling date, there were significant NMS 2 dimensional
ordinations that explained from 73 to 86% of the sample’s variation (Figure 1). Clear
differences were found between invertebrate communities in impounded wetlands and
open coastal wetlands on all dates, but differences among water depths were more
complex. No clear differences among water depths emerged in impounded wetlands. In
open coastal wetlands, communities in shallow and medium water depths usually
grouped together, but those in deep water depths grouped separately on the ordination.
These results were supported by the MRPP pair-wise comparison (Table 6). On all three
sample dates, most pair-wise comparisons between impounded wetlands and open coastal
50
Table 5: Average density (number per m2 ± SE) of common benthic invertebrates
between wetland type (impounded [Imp.], open coastal [Open]) and depth (shallow,
medium deep) in June, July, and August 2006.
51
June
July
Imp.
Mean
Open
1 SE
Mean
August
Imp.
1 SE
Mean
Open
1 SE
Mean
Imp.
1 SE
Mean
Open
1 SE
Mean
1 SE
Total Number
Shallow
32729
10985
120196
38201
9583
4928
52332
7033
5887
2157
44092
11712
Medium
26949
4674
136784
57338
23936
7832
42637
19026
21729
13447
27957
7407
Deep
32966
9662
42089
11409
18529
7878
23221
7980
8481
3286
15248
4078
Shallow
2206
1715
0
0
1145
512
146
65
344
172
0
0
Medium
1335
878
86
50
4210
1883
1403
628
1550
775
121
54
0
0
0
0
1524
682
121
54
108
54
0
0
Shallow
1614
1164
86
86
723
451
215
64
452
313
189
150
Medium
1324
1069
0
0
1231
730
207
85
377
173
344
281
Deep
2798
1089
0
0
585
445
172
151
280
124
112
42
Shallow
4843
1469
21410
7080
2428
1441
11064
2894
1421
672
14956
5900
Medium
7114
2127
32187
25412
6440
1921
5855
1787
10181
5468
6879
1647
Deep
5510
3067
5281
1591
9574
4918
4684
1417
4499
2147
4744
1542
Shallow
1754
1597
22185
11127
560
451
2893
1846
86
86
69
69
Medium
420
264
15541
6377
448
224
34
34
474
389
17
17
1055
625
13589
4368
164
95
0
0
22
22
9
9
0
Amphipoda
Deep
Ceratopogonidae
Chironomidae
Cladocera
Deep
Copepoda
Shallow
2389
976
1521
1478
585
337
146
72
431
317
0
Medium
1614
1270
861
199
938
451
17
11
204
81
0
0
Deep
3024
1698
158
14
775
526
0
0
484
363
17
17
Shallow
721
465
18052
7236
103
83
8722
3831
22
22
8524
2712
Medium
1281
844
18268
7375
146
117
4116
1466
11
11
3737
1310
Deep
1162
627
4635
1590
52
25
3797
1980
11
11
2075
707
Shallow
6350
3098
26461
6172
1774
947
14387
3709
1130
596
16652
3916
Medium
4176
1841
29288
13067
4253
1412
7319
2983
6350
5293
7740
2396
Deep
4101
1913
13173
6211
2118
1382
7783
2409
2400
921
4434
549
Shallow
6544
8583
27925
18562
465
657
7534
5935
215
422
2755
2321
Medium
4509
8201
35732
38343
1946
2974
20242
31397
366
443
7826
6225
Deep
5037
8066
4133
4123
1817
2042
5243
5147
172
338
3143
3370
Nematoda
Oligochaeta
Ostracoda
52
Figure 1: Two-dimensional NMS ordinations of invertebrate communities in June, July
and August 2006. Samples were grouped by water depth (Shallow, medium, deep) and
wetland type (Imp., impoundment; Open, Open coastal). The percent of observed
variation explained by each axis are indicated on the figure. Stress (S) and probability (p)
for the two dimensional ordinations are: June (S = 12.48, p = 0.019); July (S = 12.94, p =
0.019); August (S = 9.84, p = 0.019).
53
54
Table 6: MRPP pairwise comparison p – values of invertebrate communities in June,
July, and August 2006. Habitat labels are Wetland Type/Water depth: IS (impounded
shallow), IM (impounded medium), ID (impounded deep), OS (open coastal shallow),
OM (open coastal medium), and OD (open coastal deep). Significant differences (p ≤
0.05) are bold.
55
Habitat
June
IS
IM
ID
0.884
0.021
0.908
0.020
0.016
0.134
0.029
0.043
0.016
IS
IM
IM
0.959
ID
OS
OM
OD
July
OS
OM
0.044
0.012
0.330
0.033
0.396
ID
OS
OM
IM
0.318
ID
0.900
0.661
OS
0.004
0.004
0.008
OM
OD
0.052
0.026
0.045
0.006
0.137
0.062
0.285
0.093
0.881
ID
OS
OM
August
IS
IM
IM
ID
0.276
0.531
0.824
OS
OM
0.004
0.010
0.052
0.061
0.008
0.061
0.215
OD
0.004
0.007
0.097
0.017
0.352
56
wetlands were significant. However, there were no differences among water depths
within impounded wetlands on any sample date (Table 6). In contrast, open coastal
wetland communities in shallow water were different than those in deep water in June
and August.
Indicator taxa analysis identified several taxa that were associated with each
wetland type (Table 7). Five taxa were indicators of open coastal wetland communities:
Ostracoda, Oligochaeta, Nematoda, Chironomidae and Lymnaeidae. Six taxa others were
indicators of impoundment wetlands: Ceratopogonidae, Copepoda, Corixidae, Caenidae,
Amphipoda and Hirudinea. Cladocera were an indicator of both wetland types on
different dates. The only indicators of water depth were Corixidae, Lymnaeidae, and
Physidae, which were common and abundant in shallow water. Nematodes and
Oligochaetes were both indicators of the shallow depth in open coastal wetlands in July
and August, while Caenidae was an indicator taxa of the medium depth in impoundment
wetlands in August.
57
Table 7: Indicator taxa for different habitats. Habitat labels are Wetland type (Imp.,
impounded wetland; Open, open coastal wetland) and water depth (S, Shallow; M,
Medium; D, Deep).
58
June
July
August
Taxa
P value
Ind.
Taxa
P value
Ind.
Taxa
P value
Ind.
Wetland
Amphipoda
Ceratopogonidae
0.028
0.003
Imp.
Imp.
Wetland
Copepoda
Corixidae
0.043
0.001
Imp.
Imp.
Wetland
Amphipoda
Caenidae
0.007
0.004
Imp.
Imp.
Copepoda
Hirudinea
Physidae
Chironomidae
Cladocera
Oligochaeta
0.008
< 0.001
0.009
0.041
0.012
0.003
Imp.
Imp.
Imp.
Open
Open
Open
Dreissenidae
Lymnaeidae
Nematoda
Oligochaeta
Ostracoda
0.006
0.004
0.028
< 0.001
0.008
Open
Open
Open
Open
Open
Cladocera
Copepoda
Nematoda
Oligochaeta
Ostracoda
0.005
< 0.001
< 0.001
0.015
0.002
Imp.
Imp.
Open
Open
Open
Ostracoda
0.022
Open
Depth
Depth
None
Corixidae
0.023
S
Corixidae
0.045
S
Lymnaeidae
Physidae
0.039
0.038
S
S
Lymnaeidae
Physidae
0.022
0.034
S
S
Caenidae
Nematoda
0.006
0.002
Imp. M
Open S
Oligochaeta
0.010
Open S
Wetland X
Depth
Wetland X
Depth
None
Nematoda
Oligochaeta
Depth
Wetland X
Depth
0.023
0.012
Open S
Open S
59
Discussion
A majority of coastal wetlands along Great Lake shorelines have been impounded
to control their water levels (Comer et al., 1995). This loss of a hydrological connection
to the Great Lakes impacts ecosystem processes as well as plant and animal communities.
For example, lake fish cannot access impounded wetlands except when their water
control structures are manually opened. Johnson et al. (1997) and Markham et al. (1997)
found that impounded wetlands have different fish communities than nearby open coastal
wetlands, but I found few differences in fish communities in Crane Creek Marsh and the
adjacent impounded wetlands. However, invertebrate communities differed between the
Open coastal wetlands and impounded wetlands at Ottawa National Wildlife Refuge,
suggesting that diking wetlands altered key environmental conditions.
More than 100 fish species are found in Lake Erie (Leach & Nepszy, 1976), and
40 species use coastal wetlands (Herdendorf, 1987; Jude and Pappas, 1992). I collected
many of these (25) species at Ottawa National Wildlife Refuge, although I did not find
some species that are common in other Great Lake coastal wetlands (e.g., bigmouth
buffalo, grass pickerel, logperch, redhorse) (Jude & Pappas, 1992). Numbers of fish I
collected with the Fyke net were comparable to surveys conducted in other wetlands
(Herdendorf, 1987; Cardinale et al., 1998). Thus, my results show that this wetland
complex supports a diverse fish community.
I hypothesized that fish communities would have lower species richness and
numbers in the impounded wetlands than the open coastal wetland, which has been found
60
in other Great Lakes wetlands (Jude & Papas, 1992; Johnson et al., 1997; Cardinale et al.,
1998). Changes in the fish assemblages are important because they can affect ecological
factors such as invertebrate and macrophyte community structure (Cirivelli, 1983;
Sidorkewicj et al., 1996; Zambrano & Hinojosa, 1999; Gido, 2003; Olson et al., 2003;
Haas, 2007; Pothoven et al., 2009), detrital breakdown (Short & Holomuzki, 1992) and
abiotic conditions (Lamarra, 1975; Shorman & Cotner, 1997). However, fish numbers
and richness were not different in the open coastal wetland and the impounded wetlands,
which suggest that impoundment does not always lead to a loss of Great Lakes fish
diversity. This was surprising because impounded wetlands at Ottawa NWR are drawn
down every few years to enhance habitat for migratory shorebirds and waterfowl, which
periodically eliminates fish populations. I only sampled three sites of each wetland type,
which would have reduced statistical power. Although my experiment design would find
major differences in community structure, subtle patterns would be harder to detect. For
example, lake-fish species were less common (30% of total collected) in the impounded
wetlands than the open coastal wetland (50% of total), but this was not statistically
significant. Although the multivariate analysis did not detect differences in fish
community structure, examination of the data shows that shiners, shad and yellow perch
were dominant in the open coastal wetland, and catfish and sunfish were dominant in
impounded wetlands. Furthermore, two abundant species were only found in one
wetland type: yellow perch in open coastal wetlands and green sunfish in impounded
wetlands. This supports the idea that subtle differences in fish communities did occur
between the habitat types. Differences would be caused by impacts of impoundment on
61
fish food resources, breeding behavior, or other factors such as macrophyte complexity
(Crowder & Cooper, 1982; Minns et al., 1994; Randall et al., 1996; Brazner & Beals,
1997; Weaver et al., 1997; Hook et al., 2001; Lougheed et al., 2001), dissolved oxygen
levels (Stuber et al., 1982: Johnson, 1989; Stuckey, 1989), turbidity (Brazner & Beals,
1997), and draw downs. Further research may find additional evidence of the effect of
impoundment on fish communities at this site.
Fish size varied among dates, with the largest mean size occurring in June. Adult
fish numbers were fairly consistent throughout the year, but the July and August samples
had large numbers of YOY individuals. I collected many juvenile game fish such as
bluegill, yellow perch, white perch, channel catfish, and crappie and bait fish such as
gizzard shad and emerald shiner. My data is in agreement with other studies that show
that Great Lakes coastal wetlands provide important fish breeding habitat (Tanner et al.,
2004).
Invertebrate communities differed between open coastal wetlands and impounded
wetlands. The NMS analysis showed that species assemblages were distinguishable on
each date, and there were several indicator taxa in each wetland type. Abundance patterns
of common taxa indicated that the invertebrate communities were probably strongly
influenced by the aquatic vegetation. Many taxa that were abundant in impounded
wetlands live as clingers on plant stems (i.e., Physidae, Ceratopogonidae, Amphipoda,
Caenidae, Corixidae, and Hirudinea) (Merritt et al., 2008). In contrast, taxa that were
more common in the open coastal wetland are benthic burrowers in unconsolidated
sediments (i.e., Oligochaeta, Nematoda, Ostracoda) (Merritt et al., 2008; Thorp &
62
Covich, 2010). Diking also impacts hydrology, which is another important factor that
affected the invertebrate community. For example, adult dreissenids (zebra and quagga
mussels) were more abundant in the open coastal wetlands. These species are abundant
in Lake Erie but don’t survive winter freezing in the wetlands. These are affected by
hydrology because they enter the open coastal wetlands when seiches bring in juvenile
mussels (i.e. veliger larvae) (Bowers & deSzalay, 2005).
The open coastal wetland had much higher invertebrate densities than the
impounded wetlands. This was unexpected because the impounded wetlands had dense
stands of macrophytes, which enhance food resources and provide cover from predators
(de Szalay & Resh, 1997, 2000; de Szalay & Cassidy, 2001). This suggests that other
unexamined factors controlled invertebrate abundance. Perhaps the stagnant water in the
impounded wetland caused harsh environmental conditions such as anoxia that reduce
invertebrate numbers (USEPA, 1993). Although I did not detect major changes in fish
communities, fish predators may have also affected invertebrate densities. For example,
green sunfish are important macroinvertebrate predators (Stuber et al., 1982), and they
were more common in the impoundments. Further study is needed to examine this
pattern in more detail.
As I hypothesized, water depths are an important factor affecting invertebrate
communities in the open coastal wetland but not the impounded wetlands. The
multivariate analysis showed that species assemblages differed between the shallow and
deep areas in open coastal wetland. The 2-way ANOVA’s Depth by Wetland Type
interaction also tested if patterns among depths differed between impounded wetlands
63
and the open coastal wetland. Oligochaeta had a significant interaction in August, and
others (Chironomidae in July and August; Nematoda in August) had nearly significant (P
<0.10) interactions. Densities of these taxa were markedly higher in the shallow depths
than the deep depths in the Open coastal wetland. Furthermore, Nematoda and
Oligochaeta were indicator taxa in shallow depths in Open coastal wetlands. The Shallow
depths were <18 cm, and they were frequently exposed during Lake Erie seiches in the
open coastal wetland. In contrast, the Deep depths (>34 cm) would rarely be exposed.
Others have found that draw downs impact littoral invertebrate communities in lakes
(Baumgärtner et al., 2008), which can alter ecosystem-level properties (Wantzen et al.,
2008). However, my study is one of the first to show that water level changes affect
benthic invertebrate communities in Great Lakes coastal wetlands.
These results have important management implications. For example, many
invertebrate taxa that were common in open coastal wetland (Cladocera, Copepoda,
Oligochaeta) are important food for juvenile fish (including bluegill, bullhead catfish,
carp, channel catfish, gizzard shad, white perch and yellow perch) (Pothoven et al.,
2009; Olson et al., 2003; Diehl, 1992; McNeely, 1977; Pearse, 1921; Lindeman, 2006;
Morrison et al., 1997; Serrouya et al., 1995; Gido, 2001, 2003; Haas et al., 2007; Ellison,
1984). I collected high numbers of YOY of economically important game fish
(bullhead, bluegill, yellow perch, and crappie) and forage fish (gizzard shad) in the open
coastal wetland. Providing breeding habitats with abundant food resources will benefit
fisheries management goals in the Great Lakes region. Further research is needed to
64
understand how the community-level changes we found affect other ecological properties
in these valuable ecosystems.
Acknowledgements
I would like to thank the personnel at Ottawa National Wildlife Refuge for their
valuable assistance with this project. I also would like to thank J. Montemarano, J. Clark,
E. Kennedy, J. Johnston, E. Faulkner and R. Bowers.
65
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CHAPTER 3
EFFECTS OF FISH PREDATION ON BENTHIC INVERTEBRATE
COMMUNITIES IN A GREAT LAKE COASTAL WETLAND
Abstract
Many fish species that breed in Great Lakes coastal wetlands are dependent on
invertebrates as a key food resource. In June 2007, I built wire mesh exclosures to
examine how fish predation affected benthic invertebrate density and diversity in a Lake
Erie coastal wetland. I built large mesh (2.54 cm mesh), and small mesh (0.64 cm mesh)
exclosures that prevented access by different sized fish. I also built large and small mesh
control treatments with 1-m openings cut into the sides, and sham treatments marked off
by posts that allowed access by all fish. Benthic invertebrates were sampled in June,
July, September and October 2007 with core samplers, and fish were captured with fyke
nets adjacent to exclosures. The most common benthic invertebrate taxa were
chironomid midges, sphaeriid clams and oligochaete worms, which are all collectorgatherers. In June, large fish were the dominant size class, but small Young of the Year
(YOY) species and medium sized fish were the most common sizes later. Diets of the
five most common (bluegill, yellow perch, emerald shiner, gizzard shad and channel
74
75
catfish) included macroinvertebrates (chironomid, corixids) and zooplankton (copepods,
cladocerans, ostracods). In June, there were few differences of invertebrate communities
among treatments. In July to October, invertebrate densities were over twice as high in
exclosures than control or sham treatments indicating that fish predation reduced
invertebrate numbers throughout the remainder of the study. Large fish had the greatest
overall impact because total invertebrate densities were not different between small vs.
large mesh exclosures. Multivariate analyses showed that composition of invertebrate
communities differed in areas with fish access than exclosures without fish predators. In
July, sphaeriid, oligochaete, and midge densities were 2 - 8 times higher in exclosures
than in control and sham treatments. However, multivariate analyses also detected
different communities in small mesh and large mesh exclosures, and sphaeriid,
oligochaete, and midge densities were greater in small mesh exclosures than large mesh
exclosures. Therefore, medium-sized fish are probably important predators of some
benthic taxa. In October, numbers of sphaeriids and oligochaete were highest but midge
densities were lowest in small mesh exclosures. This suggests that midges may indirectly
benefit from fish predation if populations of competing species are reduced. Overall,
these results indicate that fish predation strongly impacts invertebrate density and
community structure in coastal marshes.
Introduction
76
In Great Lakes ecosystems, benthic macroinvertebrates and zooplankton are key
components of trophic webs, and they are drivers of ecosystem processes such as nutrient
cycling (Carney & Elser, 1990; Arnott & Vanni, 1996; Vaughn & Hakenkamp, 2001;
Devine & Vanni, 2002). In Great Lakes coastal wetlands (i.e. wetlands with a
hydrological connection to the lake), many fish species enter the wetlands from the
adjacent lake to breed (Jude & Papas, 1992). For example, channel catfish (Ictalurus
punctatus), northern pike (Esox lucius), yellow perch (Perca flavescens), common carp
(Cyprinus carpio), and freshwater drum (Aplodinotus grunniens) are lake fish, but they
enter coastal wetlands to spawn in the spring. Other invertivorous fish, including crappie
(Pomoxis sp.), brown bullhead (Ameriurus nebulosus), and green sunfish (Lepomis
cyanellus) live in coastal wetlands year round (Jude & Papas, 1992; Maynard & Wilcox,
1997; Batzer et al., 2000). Most species of juvenile fish in these habitats feed on
abundant invertebrate food resources. Fish affect the biomass, diversity, and abundance
of benthic macroinvertebrate prey (Morin, 1984; Diehl, 1992; Haas et al., 2007).
Predatory fish often decrease invertebrate biodiversity (Carpenter & Kitchell, 1993;
Persson, 1999; Dorn et al., 2006) by eliminating some prey species, but predation may
increase diversity if competitively dominant taxa are reduced (Batzer et al., 2000).
Although top down effects of fish on aquatic invertebrates in lakes are wellknown (Carpenter et al., 1987), there are fewer studies on top down effects in wetlands
(Batzer, 1998; Batzer et al., 2000). Some have proposed that cascading effects of
predation are more likely to be important in lakes habitats with a simple physical
structure (Pierce & Heinrichs, 1997). In contrast, wetlands are complex habitats due to
77
the physical architecture of interspersed open water and stands of emergent and
submersed plants, and spatial variability of environmental characteristics (e.g., dissolved
oxygen levels, water temperature). Some recent studies of the impact of fish predation in
wetlands have found mixed effects. Batzer (1998) found small fish decreased numbers of
midges in impounded wetlands, but others found few invertebrate taxa were affected by
predation (Diehl 1992; Batzer et al., 2000). Although many economically important fish
species use Great Lake coastal wetlands, studies of top-down effects in these habitats are
lacking.
In this study, I examined interactions of fish and macroinvertebrates in a Lake
Erie coastal wetland. I tested the effects of fish predation by comparing benthic
invertebrate communities in exclosures that excluded large (body depth > 3.6 cm) or
medium sized fish (body depth 0.9 – 6.3 cm) and nearby open areas. I also trapped fish
to examine temporal changes in fish size-classes and examined gut contents of young of
the year (YOY) to determine key diet items.
H1: Fish will affect density, diversity and community structure of the benthic
invertebrate community. Small fish will have the largest impact, due to their high
numbers, followed by the medium then large fish.
Methods
Study Site Description
78
In summer 2007, I sampled fish and benthic invertebrates in Crane Creek Marsh
(CCM) at the Ottawa National Wildlife Refuge (Latitude / longitude: 41°37′44″N /
83°12′31″W). This marsh is a 166 ha Lake Erie coastal wetland located on the southwest
shoreline in Ohio (Ottawa co.). The CCM watershed is 145 km2 and is mostly
agricultural fields. CCM is mostly open water with scattered beds of emergent aquatic
vegetation. The water is often turbid due to the presence of common carp and waves.
The marsh is permanently connected to Lake Erie via a 4-m wide channel, and water
levels in the marsh fluctuate with levels in the adjacent lake. Water levels in the marsh
are dynamic because Lake Erie water levels are highest in June and lowest during winter.
The deepest water depths in CCM are ~2 m in depth, and wind driven seiches in the Lake
Erie cause marsh water levels to vary by about 20 cm on most days.
Fish Exclosures
I used wire mesh exclosures to limit access by different size classes of fish. The
first set of 5-m diameter circular exclosures were built with large mesh galvanized wire
fencing (2.5 cm mesh size). The mesh was held upright with steel fence posts driven into
the sediments. The large mesh exclosures excluded large fish but allowed access by
medium sized and small fish. I had three treatments: Large Mesh Exclosure was built
with the lower edge of the mesh buried 10 cm into the sediment; Large Mesh Control was
built similarly, but each exclosure had two 1 m X 1 m holes cut into the sides to allow
fish access; and Large Mesh Open was an unfenced 5-m area marked with wooden posts.
79
The Large Mesh Open treatment tested if the structure of the exclosure affected fish
feeding behavior (e.g., if fish would not enter the openings of the Control treatments). A
second set of 1 m X 1 m square exclosures were built with small galvanized wire mesh
(0.64-cm mesh size) held with steel fence posts. The small mesh size was used to
exclude all large and medium fish, but small fish (e.g. fry) could pass through the mesh.
These included three treatments: Small Mesh Exclosure was built the bottom edge of the
mesh buried 10 cm into the sediments; Small Mesh Control was open on two sides, and
Small Mesh Open was an unfenced 1 m x 1 m area marked off with wooden posts. I did
not place tops on the exclosure treatments, however they were build 0.5 m above the
water level. I constructed eight replicates of each of these six treatments. I checked if any
fish had accidentally entered my exclosures by electro-fishing and seining every two
weeks during the study.
All treatments were installed at CCM in late May 2007, and the large mesh and
small mesh treatments were located ~30 m apart. I measured water levels with a meter
stick in June after all treatments were installed. I calibrated these measurements to data
from a water level logger in CCM operated by the USGS and a NOAA gauging station
near Toledo, OH (NOAA station #9063085). I compared dissolved oxygen, temperature
and conductivity at each location in July 2007 with a handheld meter (Model 57, YSI,
Yellow Springs, OH).
Benthic Invertebrates
80
Benthic invertebrates were sampled immediately after treatments were installed in
June 2007 and again in July, September and October 2007. I sampled aquatic
invertebrates with a 5-cm diameter PVC corer driven about 10 cm into the benthic
sediments. On each date, three randomized subsamples were collected from each
treatment area and combined into one sample. Because I re-sampled several times during
the study, I minimized sediment disturbance by standing outside the treatment area
during sampling. Furthermore, I did not sample any previously sampled locations.
Samples were drained in a 300 micron mesh sieve in the field and preserved with 70%
ethanol. In the laboratory, samples were rinsed in a 300 micron mesh sieve to remove
fine silt, and samples were sorted under a dissecting microscope. Invertebrates were
identified to the lowest practical taxonomic level (usually family or order) using
dichotomous keys (Peckarsky et al., 1990; Merritt et al., 2008).
Fish Community
Fish at CCM were sampled with un-baited paired (one of each) large and small
mesh fyke nets (1.3 cm and 0.5 cm mesh, respectively) in June, July, September 2007. I
could not sample in October because water levels were too low to set the nets. Fyke nets
were located ~15 m from the treatment areas. The two 3-m wings of the nets were set
parallel to the shoreline alongside the catch net, and a 15-m lead net ran from the
shoreline to the catch net. Nets were set for 24 h, and afterwards all fish were identified,
81
counted, and their snout-to-tail length was measured. In order to avoid stressing fish held
in the fyke nets, I measured only the first 100 of each species in each net.
I also estimated the number of fish that could enter the large mesh or small mesh
exclosures by counting fish in different size classes. I measured the diagonal of each
mesh opening (small mesh = 0.9 cm, large mesh = 3.6 cm) to determine the maximum
body depth that could enter the exclosures. Because I measured fish length but not body
depth, we estimated the length: depth ratio using drawings of each species in Page & Burr
(1991). On each date, we counted the number of predatory fish in three size classes: 1)
small fish (body depth < 0.9 cm) could pass through large and small mesh, 2) medium
fish (body depth = 0.9-3.6 cm) could pass through the large mesh but not the small mesh,
and 3) large fish (body depth >3.6 cm) could not pass through either mesh. I did not
count fish species that do not feed on invertebrates.
Diets of YOY species were sampled to determine their potential impact on aquatic
invertebrates. In July, I collected 20 individuals of each of the five most abundant YOY
species collected in fyke nets. All fish were euthanized and preserved in 70% ethanol.
They were dissected in the lab, their fore-gut contents were examined under a dissection
microscope, and I counted and identified all invertebrate prey items.
Data Analysis
I calculated Shannon diversity indices (Hʹ) to compare general invertebrate
diversity among treatments (Zar, 1999). I also compared overall benthic invertebrate
82
densities between the treatments. I identified my dominant taxa as those invertebrates
that together comprised >75% of all individuals collected in all samples. I compared
dominant taxa densities, total invertebrate densities, and Shannon’s diversity among
treatments on each sampling date with ANOVAs (JMP v. 7.0.1, 2007, Cary, NC). When
ANOVAs were significant (p < 0.05), I made pair-wise comparisons among treatments
with Tukey’s HSD tests.
Non-metric Multidimensional Scaling (NMS) was used to compare invertebrate
community structure among the six treatments. NMS analyses were run on each
sampling date using the Sorensen (Bray-Curtis) distance measures. I used a random
starting point with 50 runs and 500 iterations. Significance was determined by using a
Monte Carlo test using 50 runs of randomized data. I tested if there were communitylevel differences using Multi-Response Permutation Procedures (MRPP) to compare
treatments on each sample date. The MRPP tests were run using the Sorensen (BrayCurtis) distance measure with groups being defined by treatment type. I also examined if
there were indicator taxa in treatments using the methods of Dufrene and Legendre
(1997). Significance of indicator taxa was tested by using a Monte Carlo Test with 500
permutations. MRPP, NMS, and indicator taxa analysis were run on PC-Ord version 5.1
software (McCune & Mefford, 2006).
Results
Environmental Data
83
Water depths in treatments were 54–59 cm when I initiated the experiment in
June, and decreased to 15-20 cm by the end of the experiment in October. The treatment
areas were never fully dewatered during the experiment, except for a 2-h period during a
large seiche in September. In July 2007, dissolved oxygen levels were often saturated to
super-saturated in the treatment areas. Conductivity was 430 µS/cm – 434 µS/cm and
temperatures ranged from 26.5 ˚C to 26.7 ˚C. There were no differences in dissolved
oxygen (F5,47 = 0.99, p = 0.434), conductivity (F5,47 = 0.06, p = 0.990), or temperature
(F5,47 = 0.29, p = 0.913) among treatments.
Fish Taxa and YOY Diets
I collected 22 fish species in the large and small mesh fyke nets (Table 1). The
most commonly collected fish were bluegill, yellow perch, emerald shiner, gizzard shad
and channel catfish. I also trapped 1-3 adult map turtles (Graptemys geographica) in
fyke nets on each sampling session, and these were released without being measured.
Fish in fyke nets ranged from 2 cm to 61 cm in length. Therefore, fyke net data
probably underestimated the number of fry in this wetland, due to the very small fry
being able to fit through the fyke net. I examined the proportions of small, medium and
large predatory fish on each date. In June, most predatory fish were large (body depth >
3.6 cm), which included adult common carp that enter coastal wetlands to spawn in
spring (Figure 1). On the following two dates, to proportion of juvenile fish increased.
84
Table 1. Total number (n) and size, mean length (± SE) of fish collected in June to
September 2007 in small mesh (SF) and large mesh (LF) fyke nets. Snout-Tail length
(cm) was measured on the first 100 individuals of each species caught. When more than
one individual was collected, variance of the size is shown as the 95% confidence
intervals.
85
June
July
SF
Common Name
Yellow bullhead
Total Avg size
1
September
LF
SF
LF
SF
Total Avg size
Total Avg size
Total Avg size
Total Avg size Total Avg size
29.0
Brown bullhead
Bowfin
Fresh water drum
3
1
55.7 (5.3)
1
3.7
Channel catfish
1
52.5
Longnose gar
1
56.0
Pumpkinseed sunfish
1
6.1
13
11.8 (2.0)
2
52.8 (3.2)
6
6.7 (5.2)
1532
5.6 (0.3)
34
5 (2.5)
3
51.3 (4.2)
1
46.2
11.4 (2.0)
1
62.0
Orange spotted sunfish
Large mouth bass
White perch
8
11.4 (1.9)
1
34.9
2
69 (11.8)
3
11.5 (1.1)
2
Emerald shiner
Spottail shiner
5
1
2
188
8.5 (0.5)
1
6.1
6
5.7 (2.3)
1
8.5
6
3.8 (0.2)
3.1 (0.3)
3
10 (2.8)
19
4.4 (0.6)
1.0
34.0
4
12.7 (14.0)
1
33.0
43
5.1 (0.2)
1
5.7
5
3.8 (0.7)
20
4.5 (0.8)
67
4.3 (0.2)
5
3.8 (0.3)
2
14 (18.6)
174
4.2 (0.1)
3
7.7 (0.6)
5.4
White crappie
Black crappie
61.5 (0.7)
1
60.5
1
11.0
7.2 (0.8)
Yellow perch
Fathead minnow
4
96
Round goby
Tadpole madtom
11.5
50.2 (17.3)
Gizzard shad
Bluegill
5
1
16.8
Goldfish
Common carp
LF
21.6 (0.3)
5
11 (6.6)
2
22.4 (8.0)
3
3.4 (0.5)
5
17.3 (5.9)
86
Figure 1. Sizes of predatory fish in Crane Creek Marsh. Numbers are percent of total
collected in fyke nets on each sampling date. Size classes are: Small (body depth < 0.9
cm) Medium (body depth = 0.9 – 3.6 cm) Large (body depth >3.6 cm).
87
In July and September, medium sized fish (body depth = 0.9 – 3.6 cm) were dominant,
and small fish (body depth < 0.9 cm) were also more abundant.
The five most common YOY species were bluegill, yellow perch, white perch,
gizzard shad and channel catfish. Gut contents analysis showed that their diets included
chironomids, copepods, cladocerans, corixids and ostracods (Table 2). Chironomids
were the most abundant diet item, and these were eaten by all species. Other
invertebrates were less abundant in fish diets, except Corixidae were an important prey
item for yellow perch. Bluegill had the largest range of prey items in their diet that
included chironomids, copepods, cladocerans, corixids and ostracods. Yellow perch,
white perch and channel catfish diets were comprised of only 2 - 3 invertebrate taxa.
Although gizzard shad are mostly detritivores, some had consumed a few invertebrates.
Benthic Invertebrates
I collected 24 invertebrate taxa in the treatment areas (Table 3). Chironomidae,
Sphaeriidae, and Oligochaeta were the three dominant taxa, which comprised over 75%
of all invertebrates collected during the experiment. Species assemblages changed during
the study. Microcrustaceans (Copepods and Cladocera) were found at moderate numbers
at the beginning of the experiment in June, but their numbers were low on other dates.
Chironomidae, Oligochaeta and Sphaeriidae had low densities in June, and they became
increasingly abundant in later dates.
Shannon diversity (Hʹ) of invertebrates ranged from 0.723 to 1.637 during the
study. Shannon’s diversity values were not different among treatments on any sampling
date (Table 4).
88
Table 2. Diets of common YOY fish. Sizes of fish were from 2.6 cm to 7.5 cm. Values are mean number of prey items / fish
± SE “Other” includes amphipods, lepidopterans and nematodes.
.
Fish
Bluegill
Yellow Perch
White Perch
Gizzard Shad
Channel Catfish
Chironomidae
1.2 ±2.2
0.1 ±0.2
0.2 ±0.5
0.1 ±0.2
1.4 ±1.5
Copepoda
0.9 ±2.2
0
0
0
0
Diet Contents
Cladocera
0.3 ±0.7
0
0
0.1 ± 0.2
0
Corixidae
0.2 ±0.5
2.2 ±4.1
0.4 ±1.1
0
0
Ostracoda
0.1 ±0.2
0
0
0.1 ±0.3
0
Other
0.4 ±0.8
0
0
0
0.1 ±0.2
89
Table 3. Invertebrate taxa collected in exclosures at Crane Creek Marsh. Numbers are
the percent that each taxa were of total invertebrates collected in treatments over all
sampling dates.
90
Taxa
Insects
Chironomidae
Ceratopogonidae
Corixidae
Sminthuridae
Ephemera sp.
Simulidae
Caenis sp.
Unidentified Ephemeroptera
Crustaceans
Ostracoda
Cladocera
Copepoda
Amphipoda
Water Mites
Hydrachnidia
Molluscs
Sphaeriidae
Corbicula
Unionidae
Dreissenidae
Physidae
Planorbidae
Lymnaeidae
Open
43.5
2.4
1.3
<1
Treatment
Small Mesh
Control
Exclosure
31.4
1.6
<1
<1
<1
<1
Open
11.7
1.3
40.0
3.4
<1
<1
<1
Large Mesh
Control
Exclosure
37.9
2.8
<1
<1
<1
<1
<1
20.8
1.3
2.5
1.8
3.7
2.6
1.3
1.5
<1
11.4
1.1
<1
<1
<1
<1
8.8
<1
<1
7.8
32.0
<1
<1
<1
<1
<1
<1
1.7
<1
4.2
2.4
10.8
9.0
5.3
<1
1.6
<1
<1
10.1
7.7
<1
<1
<1
<1
32.7
1.3
<1
<1
<1
<1
<1
<1
<1
<1
91
Table 4. Statistic results of ANOVAs comparing diversity, total invertebrates and
densities of the three dominant taxa among treatments on each sampling date. Significant
differences (p ≤ 0.05) are bold.
June
F 5,47; P value
Total Invertebrates
1.96; 0.104
Shannon's Diversity
1.32; 0.276
Chironomidae
0.7; 0.623
Oligochaeta
0.57; 0.723
Sphaeriidae
1.02; 0.417
July
September
F 5,47; P value
F 5,47; P value
13.29; < 0.001 23.59; < 0.001
1.83; 0.127
4.43; 0.003
4.04; 0.004
19.78; < 0.001
0.79; 0.562
3.18; 0.016
October
F 5,47; P value
9.64; < 0.001
1.86; 0.123
4.98; 0.001
9.54; < 0.001 12.67; < 0.001
25.41; < 0.001 11.83; < 0.001
92
Invertebrate numbers increased during the experiment, especially in the exclosure
treatments (Figure 2). Total densities were not different among treatments on the first
sampling date in June, but they were on all other dates (Table 4). In July to October, the
Small Mesh and Large Mesh Exclosure treatments had higher densities than the Open
and Control treatments. For example, October total invertebrate densities were between
10,000 – 20,000 invertebrates/m2 in exclosures, and only between 2,000 – 7,000
invertebrates /m2 in Open and Control treatments. Densities were rarely different
between the Open and Control treatments or the Large Mesh Exclosure and Small Mesh
Exclosure treatments (Figure 2).
Patterns of abundance of the three dominant taxa were more complex. Numbers
of all dominant taxa were low in June and did not differ among treatments (Figure 3,
Table 4). Numbers increased on later sampling dates, especially in the exclosures.
Chironomids were more abundant in exclosures than Open or Control treatments in July
and September. However in October, chironomid numbers increased markedly in most
treatments but remained low in Small Mesh Exclosure (Figure 3). Sphaeriid clams
peaked in September, and oligochaetes peaked in October (Figure 3). In September, both
taxa were higher in Small and Large Mesh Exclosure than the Open or Control
treatments. However in October, both taxa were higher in Small Mesh Exclosure than all
other treatments. Both taxa were also more abundant in Large Mesh Exclosures than in
Open treatments.
Multivariate analyses showed temporal changes in patterns between invertebrate
community structure in the different treatments. Community structure in each treatment
93
Figure 2. Total invertebrate densities (± SE) collected in each treatment from June to
October. Sm and Lm indicate Small mesh and Large mesh treatments, respectively.
One-way ANOVAs comparing all 6 treatments were run on each sample date. Letters
over bars indicate that treatments are different (p ≤0.05) on that sampling date.
94
Figure 3. Densities of dominant taxa (± SE) collected in each treatment from June to
October 2007. Sm and Lm indicate Small mesh and Large mesh treatments, respectively.
One-way ANOVAs comparing the 6 treatments were run on each sample date. Letters
over bars indicate that treatments are different (p ≤0.05) on that sampling date.
95
96
was not distinguishable on the NMS ordination in June, but community structure was
clearly divergent in July, September, and October (Figure 4). In July, September, and
October, Open and Control treatments usually overlapped on the NMS ordination and
were not significantly different in the MRPP analyses (Table 5). However on the same
dates, Small Mesh Exclosure and Large Mesh Exclosure treatments were usually
different than the Control and Open treatments. Furthermore, invertebrate communities
in Large Mesh Exclosure and Small Mesh Exclosure treatments were different from each
other in July and October but not in September (Table 5).
Indicator taxa analysis found that different taxa were associated with the
treatments on each date. In June when the communities were similar in all treatment
types, there were no indicator taxa. In July, chironomids (p = 0.047) and ostracods (p =
0.002) were indicators of the Small Mesh Exclosure treatment, while oligochaetes (p =
0.036) and sphaeriids (p < 0.001) were indicators species for the Large Mesh Exclosure
treatments. In September, oligochaetes (p = 0.003) and sphaeriids (p < 0.001) were
indicators of the Small Mesh Exclosure treatment. In October, oligochaetes (p < 0.001),
sphaeriids (p < 0.005) and Corbicula clams (p < 0.001) were indicators of the Small
Mesh Exclosure treatment, while chironomids were indicators of the Large Mesh Control
treatment (p = 0.034).
97
Figure 4. Two dimensional NMS ordinations of invertebrate community structure on
each sampling date. Sm and Lm indicate Small mesh and Large mesh treatments,
respectively. The percentages of observed variation explained by each axis are indicated
on the figures. Stress (S) and probability (p) values for the two dimensional solutions
are: June S =15.27, p = 0.020; July S = 12.94, p = 0.020; September S = 11.75, p = 0.039;
October S = 6.45, p = 0.020.
98
99
Table 5. MRPP pairwise comparisons of invertebrate communities in treatments. Sm
and Lm indicate Small mesh and Large mesh treatments, respectively. Significant
differences (p ≤ 0.05) are bold.
100
June
Sm. Open
Sm. Control
Sm. Exclosure
Sm. Control
0.282
Sm. Exclosure
0.225
0.417
Lm. Open
0.278
0.243
0.480
Lm. Control
0.048
0.034
0.266
Lm. Exclosure
July
Sm. Control
Sm. Exclosure
Lm. Open
Lm. Control
0.601
0.130
0.112
0.141
0.954
0.829
Sm. Open
Sm. Control
Sm. Exclosure
Lm. Open
Lm. Control
0.821
< 0.001
< 0.001
Lm. Open
0.383
0.030
< 0.001
Lm. Control
0.137
0.011
< 0.001
0.211
< 0.001
< 0.001
0.017
< 0.001
0.001
Sm. Open
Sm. Control
Sm. Exclosure
Lm. Open
Lm. Control
Lm. Exclosure
September
Sm. Control
Sm. Exclosure
0.645
< 0.001
< 0.001
Lm. Open
0.634
0.073
< 0.001
Lm. Control
0.182
0.006
< 0.001
Lm. Exclosure
October
Sm. Control
Sm. Exclosure
0.184
< 0.001
< 0.001
0.323
< 0.001
< 0.001
Sm. Open
Sm. Control
Sm. Exclosure
Lm. Open
Lm. Control
0.989
< 0.001
< 0.001
Lm. Open
0.081
0.161
< 0.001
Lm. Control
0.008
0.020
< 0.001
0.072
Lm. Exclosure
0.002
0.005
< 0.001
0.008
0.498
101
Discussion
Top-down controls of fish on macroinvertebrates have been documented in some
freshwater wetlands (Batzer, 1998; Batzer et al., 2000; Hentges & Stewart, 2010), and I
also found that fish predators also have a strong impact on benthic invertebrate
communities in this Great Lakes coastal wetland. Although invertebrate numbers were
not different among treatments in June, they have high reproductive rates (Batzer &
Wissinger, 1996) and increased rapidly the absence of predators after the exclosures were
installed. As a result, total densities in July were 3-4 times higher in exclosures than in
open areas with predator access. Invertebrate densities remained low in the Open
treatments from July to October indicating that fish predation occurred throughout the
rest of the study. Invertebrate communities were also different in areas with fish
predation. NMS ordinations showed clear differences between areas with fish (Large
Mesh Exclosure and Small Mesh Exclosure treatments) vs. without fish (Open and
Control treatments). Indicator taxa analysis showed that most taxa (sphaeriids,
chironomids, oligochaetes, ostracods, and Corbicula clams) were associated with areas
that lacked predation effects. These taxa are often important in diets of species such as
carp, bluegill, channel catfish that were found in CCM (McNelly & Pearson, 1977; Thorp
& Bergey, 1981; Haas et al., 2007). Thus, excluding fish in this coastal wetland allowed
distinct benthic invertebrate communities to develop that were dominated by taxa that
were controlled by top down predation pressure.
102
The wire mesh exclosures I used were an effective way to test the impacts of fish
feeding on benthic invertebrates. I electro-shocked the exclosures every two weeks
throughout the experiment, and I only found four fish in the exclosures. Cage effects
(i.e., if the wire mesh increased sedimentation, shaded algae, or altered fish behavior) are
a potential problem of this study design (Virnstein, 1978; Hulberg & Oliver, 1980).
However, this was not significant because invertebrate communities and abiotic variables
were similar in exclosures with holes cut into the sides (Control treatment) and
unrestricted areas (Open treatment). On the occasions when invertebrate communities in
Control and Open treatments were not the same, the differences between these treatments
were much less pronounced than between these and the exclosure treatments.
Furthermore, invertebrate communities and abiotic variables were similar in the locations
when we installed the Small Mesh and Large Mesh treatments. Therefore, I am confident
that the exclosure design and location did not significantly influence invertebrate
communities by changing impacts of predation or environmental conditions.
Fish were the most important benthic invertebrate predators in this coastal
wetland. Fish species composition changed temporally, but they were abundant
throughout the study. For example, bluegill and yellow perch and gizzard shad peaked in
July, but emerald shiners were most abundant in September. These communities were
similar to those described in other coastal wetlands (Jude & Papas, 1992). I observed
some other potential predators at CCM. Map turtles were collected in fyke nets, which
feed on mollusks and other invertebrates (Serrouya et al., 1995; Lindeman, 2006).
However, their numbers were much lower than fish densities. I also observed shorebird
103
and dabbling duck feeding nearby (D. Kapusinski, pers. observ.), but water depths in the
treatments were too deep for them to feed effectively.
It is difficult to predict which the most important predator species were because
CCM supported a diverse fish community. Bluegill, gizzard shad, yellow perch, white
perch, channel catfish, and emerald shiner were abundant in CCM, and all consume or
indirectly effect benthic invertebrates (McNeely, 1977; Diehl, 1992; Gido, 2001; 2003;
Olson et al., 2003; Pothoven et al., 2009). Common carp were also observed, but they
were not trapped in large numbers. The diet study confirmed that YOY bluegill, yellow
perch, white perch, gizzard shad, and channel catfish consumed pelagic invertebrates
(corixids, copepods and cladocera) and benthic invertebrates (chironomids, ostracods),
which has been reported by others (McNeely, 1977; Diehl, 1992; Olson et al., 2003;
Pothoven et al., 2009). Although I did not find sphaeriid clams in fish diets, these
molluscs are common food for yellow perch, gizzard shad, channel catfish, bullhead, and
fresh water drum that occur in CCM (Pearse, 1921; Serrouya et al., 1995; Morrison et al.,
1997; Lindeman, 2006). Oligochaetes were also not found in fish diets, but they would
be difficult to detect because their soft bodies are quickly digested. Gizzard shad were
the most numerous species in fyke nets, but these mainly feed on deposited organic
detritus. I found some invertebrates in their diets, which they probably inadvertently
consumed when eating detritus. However, gizzard shad can indirectly increase
invertebrate predation when they stir up the benthic sediments (Gido, 2001; 2003).
Fish size classes changed during the study. Many fish spawn in coastal wetlands
in spring (Jude & Papas, 1992), and the fish community in CCM in June was dominated
104
by large fish (body depth >3.6 cm). For example, I collected adult freshwater drum,
common carp, black crappie and channel catfish, which probably entered CCM from
Lake Erie to spawn. Medium sized fish (body depth = 0.9 – 3.6 cm) peaked in July and
September, and small fish (body depth < 0.9 cm) numbers increased through September.
YOY fish comprised most of the increase in medium and small fish. Numbers of fry
probably increased as well, but they were underrepresented in fyke net catches because
they could pass through the net.
My results indicate that large fish had the greatest impact on total invertebrate
density in this coastal marsh. Large Mesh Exclosures excluded large fish that could not
pass through the coarse mesh, and total invertebrate densities were much higher in these
areas than in Open and Control treatments. Furthermore, total invertebrate densities in
Small Mesh Exclosure and Large Mesh Exclosure treatments were similar on all dates,
which suggests that medium sized fish that passed through the Large Mesh did not
greatly reduce total densities. This pattern was found in July and September at a time
when medium fish outnumbered large fish. Therefore, large fish were always present in
ample numbers to reduce benthic invertebrates. However, presence of different fish sizes
had a measurable impact on overall invertebrate community structure. Multivariate
analysis in July and October showed communities were different between Large Mesh
Exclosure and Small Mesh Exclosure treatments. Therefore, medium size fish were
affected abundance of some common invertebrate taxa even if they did not greatly
change total numbers. It is important to note that small predators (e.g., fry, invertebrates)
105
may also be important in this wetland, but my experimental design did not test their
impact because they could access all treatment areas.
Densities of the dominant invertebrate taxa (sphaeriids, oligochaetes,
chironomids) were usually higher in fish exclosures than in open areas. Although
sphaeriid clam and oligochaete worm were not abundant in the first sampling date, they
increased rapidly in the fish exclosures. For example, sphaeriids peaked in September at
8400 / m2 in exclosures vs. only 300 / m2 in open areas. Impacts of fish on sphaeriid
clams and oligochaetes have been reported before (Hendrika et al., 2004; Bowers et al.,
2005). Furthermore, numbers of both taxa were higher in Small Mesh Exclosure than
Large Mesh Exclosure in October, which suggests that their populations were reduced by
medium sized fish that could pass through the large mesh but not the small mesh.
Therefore, feeding by medium sized fish may become more important in Great Lakes
coastal wetlands when their numbers increase in late summer.
Changes in chironomid midge numbers were more complex. In July and
September midges were highest in Small and Large Mesh Exclosure treatments, which
show the effect of excluding large fish. In October, their numbers increased in Open and
Control areas, and they were the same as in Large Mesh exclosures. Chironomids have
high reproductive rates (Coffman & Ferrington, 1996), and midges in October samples
were mostly small instars (D. Kapusinski, pers. observ.). Therefore, gains by midge
reproduction may have been higher than losses due to large fish predation. October
midge densities were lowest in Small Mesh Exclosures, and there are several possible
reasons for this pattern. First, numbers of medium sized fish may have increased inside
106
exclosures where large fish were excluded, and these selectively fed on chironomids.
Alternately, competition for resources may have been higher in these exclosures, which
decreased midge numbers. I did not study competition impacts, but it is interesting to
note that October densities of oligochaetes and sphaeriids were highest in Small Mesh
Exclosures. Oligochaetes and sphaeriids feed on organic matter in sediment (Peckarsky
et al., 1990; Vaughn & Hakenkamp, 2001), and thus they may compete with chironomids
for food resources. Others have suggested that fish predation alters competitive
interactions between invertebrate species (Batzer & Resh, 1991; Diehl, 1995; 1992). For
example, midge numbers increased after snail populations were suppressed by fish
predation (Batzer et al., 2000). Thus, fish at CCM may have altered invertebrate
communities directly via predation and indirectly by altering the outcome of competition
among invertebrate taxa.
Management Implications
This is one of the few studies that have tested impacts of fish predation in Great
Lakes coastal wetlands, and my results have important implications to manage these
ecologically valuable habitats. I trapped many game fish (e.g., yellow perch, bluegill,
channel catfish) in CCM, and my diets study show their YOY feed on benthic
invertebrates. Gizzard shad and emerald shiners were also abundant, and these are
important food for game fish such as walleye (Sander vitreus) (Bur et al., 2008).
Furthermore, several common taxa are primarily lake fish (e.g., gizzard shad, channel
107
catfish, and yellow perch); these were probably using wetland as a nursery for their
offspring. Many coastal wetlands along Lake Erie have been impounded to control their
water levels, but this eliminates fish movement between the lake and the wetland.
Therefore, CCM is an important location because it provides habitat for a diverse fish
community that helps support the economically game fishing industry in Lake Erie.
I found that fish reduced densities of benthic invertebrates, which may reduce
food resources for other wildlife species. For example, corixids, chironomids, and
oligochaetes eaten by YOY fish are also are important in shorebird diets (Skagen &
Oman, 1996), which use coastal wetlands as stopovers during migration (Herdendorf,
1987). Therefore, fish and shorebirds may compete for food resources in these habitats.
However, I also found that fish predation may indirectly increase chironomid numbers,
which can benefit shorebirds. Therefore, more information is needed to understand
competition for food among fish and shorebirds in Great Lake coastal wetlands.
The indicator taxa analysis species showed chironomids, oligochaetes, sphaeriids,
ostracods, and corbiculids were associated with areas without fish access. These taxa
feed mostly on deposited or floating detritus, and thus impact detrital breakdown and
nutrient cycles (Merritt & Cummins, 1996; Thorp & Covich, 2001). It has been shown
that fish predation in streams and can influence detritus processing mediated by aquatic
invertebrates (Short & Holomuzki, 1992) and can influence nutrient cycling (Schaus et
al., 1997). Coastal wetlands are key sites for nutrient uptake, and they export organic
matter to the Great Lakes during seiche induced out-wellings (Bouchard, 2007). Further
108
studies should examine how fish predation of benthic invertebrates affects nutrient cycles
and other important ecosystem processes in coastal wetlands.
Acknowledgements
I would like to thank the personnel at Ottawa National Wildlife Refuge for their
invaluable assistance with this project. I also would like to thank R. Bowers, F. de
Szalay, M. Drinkard, N. Drinkard, K. Gee, B. Morgan, E. Kennedy, J. Montemarano, D.
Sprockett, J. Clark, D. LaVigne, N. Howard and L. Rybus for their help constructing and
sampling the exclosures. Funding for this study was provided by the Ohio Division of
Natural Resources Wildlife Diversity program.
109
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Jersey.
CHAPTER 4
PREDATION OF EPIZOIC AND BENTHIC INVERTEBRATES BY FISH
INCLUDING COMMON CARP, CYPRINUS CARPIO, IN A GREAT LAKES
COASTAL WETLAND
Abstract
Great Lake coastal wetlands are important habitat for fish, including the invasive
species such as common carp (Cyprinus carpio). In May 2008, I constructed wire mesh
cages to test the effects of predation of carp and other fish on benthic invertebrates in a
Lake Erie coastal wetland. I also included native unionid mussels in each 5-m diameter
exclosure to test if predation affected epizoic invertebrates. I used three treatments: 1)
Fishless treatment was an exclosure (2.54 cm mesh) that prevented access by all fish
except small individuals, 2) Carp treatment was an enclosure that held one live large carp
(~0.05 carp/m2 density), and 3) Fish treatment was an open area that was accessible by all
fish. I placed one dead Quadrula quadrula shell and four live Q. quadrula mussels in
each treatment area. Live unionids remained exposed above the soft benthic sediments,
but silt eventually covered the dead shells. Benthic invertebrates were sampled using
sediment cores in May, July, August and September, and mussel shells were collected in
115
116
September. Benthic macroinvertebrate numbers and diversity were lower in both
treatments with fish than Fishless treatments. The effect of carp predation (Carp
treatment) was similar to the impact of access by all fish (Fish treatment). The most
common invertebrates were Chironomidae, Oligochaeta, Ostracoda and Sphaeriidae.
Ostracoda were not affected by predation, but numbers of the other taxa were lower in
Fish and Carp treatments than Fishless treatments. Multivariate analysis detected
differences in the benthic invertebrate communities among all three treatment types. Fish
predation of epizoic invertebrates reduced total numbers and richness on live unionids.
However, predation did not affect total numbers or richness on dead unionid shells that
had settled into the soft benthic sediments. Common epizoic invertebrates were
Chironomidae, Dreissenidae, Nematoda, Oligochaeta and Ostracoda, and predation
reduced dreissenid, oligochaete and ostracod numbers. Similar to the results of the
benthic invertebrates, carp predation was as important as predation by the entire fish
community. Multivariate analysis found invertebrate communities on live unionids were
similar in Fish and Carp treatments but were different in Fishless treatments. However,
there were no community differences among treatments on dead shells. Pairwise
comparisions between treatments using Multi-response permutation procedure showed
that invertebrate communities were not different on live unionids and dead shells except
in the presence of carp. These results show that the shell surface of live unionids
provides a unique microhabitat in the silty sediments in Lake Erie coastal wetlands. The
results also show that fish can control community structure of benthic and epizoic
invertebrates, and the non-native common carp are one of the most important predators in
117
these wetlands. Further research is needed to determine if fish are affecting other wildlife
such as shorebirds that use Great Lake coastal wetlands and rely on invertebrate food
resources.
Introduction
Coastal wetlands along the Laurentian Great Lakes are critical habitat for many
wildlife and fish species. For example, bluegill, yellow perch, white perch, emerald
shiners, gizzard shad, and bullhead catfish all feed on benthic invertebrates in Great Lake
coastal wetlands (Pearse, 1921; McNeely, 1977; Ellison, 1984; Diehl, 1992; Serrouya et
al., 1995; Morrison et al., 1997; Gido, 2003; Olson et al., 2003; Lindeman, 2006;
Pothoven et al., 2009). Fish predation can affect invertebrate (e.g. chironomid) densities
and communities (Batzer, 1998; Batzer et al., 2000; Hentges & Stewart, 2010), and taxa
such as; oligochaetes, sphaeriids, chironomids and ostracods are associated with areas
that lack predation effects (McNelly, 1977; Thorp & Bergey, 1981; Haas et al., 2007).
Some fish species, such as gizzard shad (Dorosoma cepedianum) can indirectly increase
invertebrate predation when they stir up sediment (Gido, 2003).
The common carp, Cyprinus carpio L., was introduced to the United States from
Eurasia in 1831 (Page & Burr, 1991). The fish was widely distributed by the U.S. Fish
Commission after 1877, and they are now found across the continental United States and
Hawaii (Edwards & Twomey, 1982; Nico et al., 2012). Carp are usually found in lakes,
ponds, and rivers with moderate flow, and they are abundant in the Great Lakes. They
118
feed on detritus, benthic invertebrates such as chironomids, oligochaetes, and mollusks,
and attached invertebrates such as zebra and quagga mussels (Dreissenidae) (Riera et al.,
1991; Tucker et al., 1996; Lougheed et al., 1998; Thorp et al., 1998; Zambrano and
Hinojosa, 1999). Carp spawn in shallow habitats with emergent vegetation (Edwards &
Twomey, 1982), and they enter coastal wetlands in large numbers to spawn (Wilcox &
Whillans, 1999)
Carp have several negative impacts in aquatic habitats. They uproot plants and
increase turbidity when they feed on benthic invertebrates (Lougheed et al., 1998, 2004;
Pinto et al., 2005; Haas et al., 2007), which can reduce feeding activity of visual
predators such as sunfish and largemouth bass (Panek, 1987). They also increase
epiphyton and phytoplankton when by re-suspending nutrients in benthic sediments
(Crivelli, 1983; Brabrand et al., 1990; Haas et al., 2007; Weber & Brown, 2009). When
feeding carp are abundant, they decrease invertebrate communities (Riera et al., 1991;
Zambrano & Hinojosa, 1999; Haas et al., 2007), and thus compete with other predators
including waterfowl (Haas et al., 2007) and fish (McNeely & Pearson, 1977; Olson et al.,
2003; Pothoven et al., 2009).
Unionidae is a family of native freshwater bivalves with over 300 species in
North America (Bogan, 1993). They are an important component of freshwater
ecosystems because they affect sediment stability when they burrow, and feed on fine
particulate organic matter (Thorp & Covich, 1991; Zimmerman & de Szalay, 2007) they
also can increase organic matter and chlorophyll a in sediments (Spooner & Vaughn,
2006). Their shells also provide epizoic habitat for epizoic invertebrates, including;
119
invasive zebra and quagga mussels, and algae (Bowers & de Szalay, 2007). Live mussel
shells have also been shown to provide habitat for macroinvertebrates, increasing their
densities as compared to deceased mussels (Vaughn et al., 2008). Live mussels have also
been shown to increase benthic invertebrate abundances and richness in surrounding
sediment (Spooner & Vaughn, 2006).
Many species of unionids have become imperiled due pollution, over-harvesting,
and introduction of exotic species (Nalepa et al., 1991; Bogan, 1993). For example, adult
zebra mussels attach to unionid shells and compete with them for food (Chase & Bailey,
1999; Bowers & de Szalay, 2004, 2007). As a result, unionids in the Great Lakes
declined dramatically after the introduction of dreissenids in 1988 (Ricciardi et al., 1995,
1998; Strayer & Smith, 1999). However, some remnant populations have been found in a
few shallow littoral areas (Gillis & Mackie, 1994; Schloesser & Nalepa 1994; Crail et al.
2011) and coastal wetlands (Zanetta et al., 2002; Bowers & de Szalay, 2004).
In this study, I tested the impact of fish predators on invertebrates in benthic
sediments and on unionid shells in a Lake Erie coastal wetland. I also studied if the
common carp was a significant predator in these wetlands. I experimentally manipulated
fish predation with wire mesh exclosures to prevent fish access. I also used enclosures
that contained carp to test the impact of this species. My hypotheses were:
H1: Fish predators control invertebrate community structure in coastal wetlands.
Fish predation will reduce their diversity and total numbers, and change the relative
abundance of dominant taxa.
120
H2: Species assemblages of epizoic invertebrates living on unionid shells are
different than benthic invertebrate communities. The burrowing behavior of live
unionids alters the microhabitat conditions and impacts of predation on epizoic
invertebrate communities.
H3: Carp are one of the most important fish predators in this coastal wetland.
Carp predation will alter community structure of benthic invertebrates and epizoic
invertebrates.
Methods
Study Site Description
This study was conducted at the Ottawa National Wildlife Refuge, located in Oak
Harbor, OH. This site which includes a Lake Erie coastal marsh, Crane Creek Marsh, and
a number of impounded wetlands where water levels are artificially controlled. Crane
Creek Marsh is dominated by shallow (< 2 m depth) turbid water with patches of
emergent vegetation, while the impounded wetlands are dominated by emergent and
woody vegetation. For a complete habitat description, please see the study site
description in the Introduction chapter. Fifteen species of unionids are found in Crane
Creek Marsh, and the most common species, Quadrula quadrula, is ~ 40% of all
unionids (Bowers & de Szalay, 2004). Carp are common in Crane Creek Marsh, and they
reach densities of 3500 / ha in coastal wetlands during the breeding season (Lougheed et
al., 1998).
121
Experimental Design
In May 2008, I installed carp enclosures and fish exclosures in Crane Creek
Marsh. The Carp treatment used circular enclosures (5-m diameter) that contained live
carp. The enclosures were built with wire mesh poultry fencing (2.54 cm mesh, 1.5 m
high) that was attached to fence posts embedded in the sediments. The fencing excluded
medium to large sized fish. On May 30th, we added one large carp (30-45 cm length) that
was trapped in Crane Creek Marsh. Therefore, carp density in the enclosure was 509
carp/hectare or 0.05 carp/m2. The Fishless treatment was built with the same design (i.e.
1.5 m high fencing, 2.54 cm mesh, 5-m diameter) but we did not add any fish inside the
exclosures. The Fish treatment was an unfenced area (5-m diameter) marked with 8
fence posts that allowed unrestricted access by all fish.
I also collected live and dead Q. quadrula unionids in Crane Creek Marsh. All
unionid shells were cleaned to remove any attached invertebrates including zebra
mussels. To hold the dead shells in a realistic posture in the sediments, I inserted a flat
wooden stake (20-cm) inside the shell cavity and filled the shell with plaster of paris. The
dead shells were held upright by pressing the wooden stake into the mud until the shell
was halfway embedded. Live mussels were numbered with a permanent marker and
fitted with a metal washer glued to the posterior end of the shell near the umbo. The
metal washer was used to relocate the mussels using an underwater metal detector. On
29 May, I stocked each treatment with one dead and four live Q. quadrula and allowed
122
invertebrates and dreissenids to colonize their shells. All unionids remained in the
treatment areas until they were collected in September 2008.
I installed 6 replicates of each treatment (Carp, Fish, and Fishless). In order to
reduce any impacts of environmental variation (e.g. sediment type, water depth) on
invertebrates or fish, I randomly located the treatments in a 3 X 6 grid where the 6 rows
each contained one replicate of each treatment. The all treatments were located 10 m
away from their neighbors. If carp became stressed in the exclosures, it could affect their
feeding behavior. Therefore, carp were only held between consecutive invertebrate
sampling dates (see below) and then released. I then trapped six new carp that were held
until the following invertebrate sampling date,. The Carp enclosures and the Fishless
exclosures were also seined on each sampling date to ensure that no other fish had
entered these treatments.
Invertebrate Sampling
Benthic invertebrates were sampled with a sediment core sampler (5-cm dia.) in
May, July, August and September 2008. On each date, I randomly sampled each
treatment by collecting four core samples from the upper ~10 cm of sediments. The four
samples were combined, drained in a sieve (300 micron mesh), and preserved in Ziploc
bags with 90% ethanol. In the laboratory, samples were rinsed through a sieve (300
micron mesh) and sorted under a dissection microscope.
Epizoic invertebrates were sampled on the shells of the dead and live unionids in
September, 2008. All mussels were removed carefully from the sediment so as to not
123
disturb the attached invertebrates. We removed epizoic invertebrates from the live
unionids by scraping them off with our fingers and a toothbrush into a Ziploc bag with
ethanol. All live unionids were released into the wetland at the end of the experiment.
The dead unionids were placed them into a Ziploc bag with 90% ethanol. In the
laboratory, the epizoic invertebrates on dead unionids were scraped into a sieve (300
micron mesh) and re-preserved until they were processed.
All invertebrates were identified to the lowest practical taxonomic level (usually
family) using Merritt et al. (2008) and Peckarsky et al. (1990) and then counted. Taxa
that were >3% of all invertebrates in either the benthic samples, and on dead or live
unionids were termed our dominant taxa.
Data Analysis
I expected that invertebrate communities would change throughout the sampling
season. Therefore, we compared invertebrate communities on each sampling date. I used
one-way ANOVAs to compare taxa richness and numbers of total invertebrates and
dominant taxa among treatments. When ANOVAs were significant (P<0.05), I ran pairwise comparisons among means with Tukey’s HSD tests. All univariate statistics were
run on JMP statistical software (JMP v. 7.0.1, Cary, NC).
I also used multivariate statistics to examine if species assemblages changed
among treatments on each date. Non-metric Multidimensional Scaling (NMS) was used
to test if species assemblages changed in response to fish predation. Ordinations were
124
performed using the Sorensen (Bray-Curtis) distance measure using a random starting
point with 50 runs and 500 iterations. I tested if ordinations were significant using a
Monte Carlo test with 50 runs of randomized data. I also tested if there were significant
community-level differences among treatments using Multi-Response Permutation
Procedures (MRPP). The MRPP used the Sorensen (Bray-Curtis) distance technique to
make pairwise comparisons of community dissimilarity between treatments. I also
determined if there were any indicator taxa for treatments using the methods of Dufrene
and Legendre (1997). The significance of indicator taxa was tested by using a Monte
Carlo Test with 500 permutations. The MRPP, NMS, and indicator taxa analysis were
performed using PC-ORD version 5.1 (McCune & Mefford, 2006).
Results
I collected 26 aquatic invertebrate taxa in this study (Table 1). Dominant benthic
taxa (>3% of total) in sediment samples were chironomid midge larvae, oligochaete
worms, sphaeriid clams and ostracods. These comprised over 96% of all invertebrates
collected. Dominant epizoic taxa on Q. quadrula were chironomids, oligochaetes,
ostracods, hydroptilid caddisfly larvae, leeches, nematodes, and dreissenid mussels.
These comprised about 94% of the total community. Population numbers of the
dominant taxa were somewhat different on live and dead unionids. Leeches and
dreissenids comprised a greater proportion of the invertebrate community on live
unionids, and chironomids, ostracods, and hydroptilid caddisflies were more abundant on
dead unionids. Although some taxa were abundant in all habitats (e.g. chironomids,
125
Table 1. Percent of each invertebrate taxa collected in benthic sediments and on live and
dead Q. quadrula unionids. Total is the total number of invertebrates collected.
126
Taxa
Insects
Diptera
Chironomidae
Ceratopogonidae
Ephemeroptera
Ephemeridae
Caenis
Hemiptera
Corixidae
Megaloptera
Sialidae
Trichoptera
Hydroptilidae sp. 1
Hydroptilidae sp. 2
Polycentripodidae
Leptoceridae
Collembola
Sminthuridae
Isotomidae
Crustaceans
Ostracoda
Amphipoda
Water Mites
Hydrachnidia
Molluscs
Dreissenidae
Sphaeriidae
Physidae
Bithyniidae
Unionidae
Corbiculidae
Segmented Worms
Oligochaeta
Leeches
Unidentified spp.
Nematodes
Unidentified spp.
Flatworms
Turbellaria
Freshwater Jellyfish
Hydra sp.
Total
Benthic Sediments
Core Samples
Unionids
Live
Dead
18.8
1.8
10.7
-
32.0
-
<1
<1
-
-
<1
-
-
<1
-
-
-
1.4
<1
<1
<1
5.4
<1
2.2
<1
-
-
<1
<1
4.7
<1
4.8
<1
15.5
-
<1
<1
<1
<1
4.0
<1
<1
<1
60.7
<1
<1
<1
-
32.0
-
68.8
3.9
5.4
<1
6.5
<1
<1
5.1
5.1
-
2.1
<1
3312
2.9
2462
<1
1435
127
ostracodes), burrowers were more important in benthic sediments (oligochaetes,
sphaeriids) and clingers (dreissenids, hydroptilidae, leeches) were more important on
unionid shells.
Benthic Invertebrates
Invertebrate taxa richness in the benthic sediments changed during the
experiment. Mean richness ranged from 1.8 to 5.5 taxa/sample from May to September
2008. Invertebrate richness was different among treatments in May, July and September
(Figure 1, Table 2). Richness was highest in the Carp treatment in May, but it was higher
in the Fishless than the Fish treatment in July and September.
Invertebrate densities varied among dates and treatments. On the first sampling
date in May, densities were generally low and were not different among treatments
(Figure 2, Table 2). Total invertebrate numbers in the Fishless treatment increased
during the experiment and were 4 times greater in August than May. In the Fish and
Carp treatments, numbers stayed approximately the same. In July, August and
September, invertebrate densities were higher in Fishless treatments than Fish and Carp
treatments. In August, invertebrate density in the Carp treatment was also lower than the
Fish treatment (Figure 2, Table 2).
Abundance of the four dominant benthic taxa also varied by date and among
treatments. Numbers of the four dominant taxa were not different among treatments in
May, but they generally increased later (Figure 3, Table 2). In July, chironomid numbers
128
Figure 1. Mean richness in benthic sediments (taxa/sample) from May to September
2008. One way ANOVAS comparing numbers in Fish, Fishless, and Carp treatments
were run on each sampling date. Letters over bars indicate that treatments are different (p
≤0.05) on that date.
129
Table 2. ANOVAs comparing densities of total invertebrates and common taxa and richness in benthic sediments in Fish,
Fishless, and Carp treatment on each sampling date. Significant differences (p ≤ 0.05) are bold. Missing values indicate that
taxa that were not collected.
Density
Species Richness
Chironomidae
Oligochaeta
Ostracoda
Sphaeriidae
May
July
F2,17; p value F2,17; p value
1.32; 0.306
7.87; 0.005
6.66; 0.009
9.21; 0.003
1.54; 0.247
1.88; 0.187
1.02; 0.383
5.37; 0.017
0.96; 0.403
0.73; 0.506
0.87; 0.439
August
September
F2,17; p value F2,17; p value
5.74; 0.014
5.62; 0.015
0.71; 0.507
7.00; 0.007
11.80; < 0.001 7.06; 0.007
4.35; 0.032
3.91; 0.043
0.52; 0.601
1.45; 0.266
3.81; 0.046
0.88; 0.436
130
Figure 2. Mean (± SE) total invertebrate densities in benthic sediments in May to
September 2008. One way ANOVAS comparing numbers in Fish, Fishless, and Carp
treatments were run on each sampling date. Letters over bars indicate that treatments are
different (p ≤0.05) on that date.
131
Figure 3. Mean (± SE) densities of the four common taxa in benthic sediments in May to
September 2008. One way ANOVAS comparing numbers in Fish, Fishless, and Carp
treatments were run on each sampling date. Letters over bars indicate that treatments are
different (p ≤0.05) on that date.
132
133
were not different between treatments, but their numbers were slightly higher in Fishless
treatment than in Fish and Carp treatments. In the last two dates, chironomids were much
more abundant in the Fishless treatment. For example, numbers were about three times
higher in the Fishless treatment than Fish or Carp treatments in August. Oligochaeta
were the most abundant taxa in the sediments, and they reached mean densities of over
10,000 worms/m2. In July and August, the Fishless treatment had the highest oligochaete
densities, Carp treatment had the lowest densities, and the Fish treatment was
intermediate. In September, numbers were highest in the Fishless treatment, lowest in the
Fish treatment, and intermediate in the Carp treatment. Ostracod densities were not
different during any sampling date, however there was a trend of higher densities in the
Fishless treatment. Sphaeriid numbers were more variable, and they were not different in
May, July, or September. In August, numbers were highest in the Fishless treatment,
followed by the Fish treatment, and then the Carp treatment.
Impacts of carp predation on community structure were apparent on the NMS
ordinations. There were significant 2-dimensional ordinations in May, August and
September (Figure 4). Also, MRPP comparisons detected significant differences among
treatments on all dates (Table 3). In May, species assemblages in Fish and Fishless
treatments were not different, but the Carp treatment was different from Fishless and Fish
treatments. There was no significant ordination in July text (S = 7.10, p = 0.314), but the
Carp and Fishless treatments were different in the MRPP analysis. In August, the three
treatment types were all different from each other. In September, the Fish and Fishless
treatments were different from each other (Table 3).
134
Figure 4. Two dimensional NMS ordinations of benthic invertebrate communities in
May to September 2008. The percent of observed variation explained by each axis are
indicated on the figures. Stress (S) and probability (p) values for the two dimensional
solutions are: May S = 4.85, p = 0.019; August S = 4.45, p = 0.019; September S = 7.37,
p = 0.019. There was no statistically significant ordination in July, S = 7.10, p = 0.314.
135
Table 3. MRPP pairwise comparisons of invertebrate communities in benthic sediments.
Significant differences (p ≤ 0.05) are bold.
May
Fishless
Carp
July
Fishless
Carp
August
Fishless
Carp
September
Fishless
Carp
Fish
0.084
0.006
Fish
0.163
0.404
Fish
0.037
0.009
Fish
0.014
0.997
Fishless
0.010
Fishless
0.009
Fishless
0.004
Fishless
0.090
136
Indicator taxa analyses examined which taxa were correlated with each treatment,
and indicator taxa varied among dates. In May, there were no indicator taxa for any
treatment. Oligochaetes were indicators of the Fishless treatment in July (p = 0.024) and
August (p= 0.032). Chironomids were indicator taxa of Fishless treatments in August
(p< 0.001) and September (p = 0.007). Ceratopogonids were indicator taxa of the
Fishless treatment in September (p = 0.009). There were no indicator taxa for Fish or
Carp treatments on any date.
Epizoic Invertebrates
Epizoic invertebrate diversity was generally lowest on live unionids in the
presence of fish. For example, taxa richness on dead unionids in Fish or Carp treatments
was 2-3 species / unionid, but was 4-8 species / unionid in all other treatments (Figure 5,
Table 4). On live unionids, richness was lower in Fish and Carp treatments than the
Fishless treatment (Figure 5, Table 4). On dead unionids, richness was lowest in the Fish
treatment but was not different between Fishless and Carp treatments (Figure 5, Table 4).
Total epizoic densities on unionids had a similar pattern. Invertebrates on live
unionids in the presence of fish predators (Fish and Carp treatments) had lower numbers
than the fishless treatment (Figure 6, Table 4). On dead unionids, treatment densities
were not different (Figure 6, Table 4). There were some differences in densities of
dominant epizoic taxa among treatments. On live unionids, densities of dreissenids,
ostracods and oligochaetes were higher in the Fishless treatment than Fish or Carp
treatments (Figure 7, Table 4). The other three dominant taxa showed similar patterns,
137
Figure 5. Mean (± SE) invertebrate richness per Q. quadrula unionid. One way
ANOVAs compared richness among Fish, Fishless and Carp treatments on live and dead
unionid shells. . Letters over bars indicate that treatments are different (p ≤0.05).
138
Table 4. ANOVAs comparing densities of total invertebrates and common taxa and
richness on live and dead Q. quadrula unionids in Fish, Fishless, and Carp treatments.
Significant differences (p ≤ 0.05) are bold. Missing values indicate that taxa that were
not collected.
Density
Species Richness
Dreissenidae
Chironomidae
Nematoda
Hydroptilidae sp. 1
Ostracoda
Oligochaeta
Polycentropodidae
Hirudinea
Live Mussel
F2,17; p value
8.93; 0.003
11.37; 0.001
6.46; 0.010
2.09; 0.159
1.91; 0.184
6.12; 0.012
6.46; 0.010
2.55; 0.114
Dead Mussel
F2,17; p value
2.28; 0.137
4.31; 0.033
2.15; 0.151
3.47; 0.058
3.08; 0.076
2.74; 0.097
0.72; 0.504
4.29; 0.033
1.17; 0.339
-
139
Figure 6. Mean (± SE) total number of invertebrates per Q. quadrula unionid in
September 2008. One way ANOVAs compared invertebrate densities among Fish,
Fishless and Carp treatments on live and dead unionid shells. . Letters over bars indicate
that treatments are different (p ≤0.05).
140
Figure 7. Mean (± SE) number of common taxa per Q. quadrula unionid in September
2008. One way ANOVAS compared invertebrate densities among Fish, Fishless and
Carp treatments on live and dead unionid shells. Letters over bars indicate that treatments
are different (p ≤0.05). Abbreviations are: Chironomidae (Chiro.), Dreissenidae (Drei.)
Hirudinea (Hiru.), Nematoda (Nema.), Oligochaeta (Oligo.), Ostracoda (Ostra.),
Hydroptilidae (Hydr.) and Polycentropodidae (Poly.).
141
142
but they were not statistically significant. On dead unionids, the differences among
treatments were less pronounced. Oligochaetes had significantly higher densities in the
Fishless and Carp treatments than the Fish treatment (Figure 7, Table 4). Patterns varied
among the other dominant taxa, and no differences were significant.
Multivariate analysis showed that epizoic species assemblages were also affected
by fish predation. There was a significant 2-dimensional NMS ordination in September
(Figure 8). MRPP analysis found the community in the Fishless treatment on the live
unionids was clearly different than communities in Fish and Carp treatments on the live
unionids (Table 5). However, there were no differences among treatments on dead
unionids. When we made pairwise MRPP comparisons of Fish and Fishless treatments
on live and dead unionids (i.e. Fishless treatment on live unionids vs. Fishless treatment
on dead unionids) the communities were not different. However, communities in the
Carp treatment on dead and live unionids were different (Figure 8, Table 5)
Indicator taxa analyses also detected that some species correlated with treatments.
On live unionids, amphipods, dreissenids, leeches and turbellarian flatworms were all
indicators of the Fishless treatment. On dead unionids, chironomids, hydroptilid
caddisflies and oligochaetes were indicators of the Carp treatment (Table 6).
143
Figure 8. Two dimensional NMS ordinations of invertebrate communities on the live or
dead Q. quadrula unionids in treatments (Fish, Fishless, Carp). The percent of observed
variation explained by each axis are indicated on the figure. Stress (S) and probability (p)
values for the two dimensional solutions are: invertebrate community on the living
mussels S = 12.58, p = 0.019.
144
Table 5. MRPP pairwise comparisons of invertebrate communities on the live or dead Q. quadrula unionids in treatments
(Fish, Fishless, Carp). Significant differences (p ≤ 0.05) are bold.
Dead/Fish
Dead/Carp
Live/Fishless
Live/Fish
Live/Carp
Dead/Fishless
0.139
0.222
0.406
0.029
0.021
Dead/Fish
Dead/Carp
Live/Fishless
Live/Fish
0.089
0.017
0.194
0.455
0.003
0.009
0.011
0.003
0.046
0.117
145
Table 6. Indicator taxa the taxa collected on the live or dead Q. quadrula unionids in
treatments (Fish, Fishless, Carp).
Taxa
Amphipoda
Chironomidae
Dreissenidae
Hirudinea
Hydroptilidae sp.1
Hydroptilidae sp.2
Oligochaeta
Turbellaria
P value
0.017
0.002
0.039
0.028
0.022
0.044
0.025
0.012
Indicator
Live/Fishless
Dead/Carp
Live/Fishless
Live/Fishless
Dead/Carp
Dead/Carp
Dead/Carp
Live/Fishless
146
Discussion
Fish predation is known to affect community structure in many freshwater
ecosystems (Thorp et al., 1998; Zambrano & Hinojosa, 1999; Haas et al., 2007). For
example, fish predation decrease invertebrate density and diversity in inland wetlands
(Batzer et al., 2000), which can create ecosystem level effects such as changes in nutrient
availability and primary production (Carney & Elser, 1990; Arnott & Vanni, 1996;
Vaughn & Hakenkamp, 2001; Devine & Vanni, 2002). In this Lake Erie wetland,
benthic invertebrates densities in the Fishless treatment was over 15,000 invertebrates /
m2, but benthic densities in areas with fish access were less than 5,000 invertebrates / m2.
Also, epizoic densities on live unionid shells were 3 to 5 times higher in Fishless
treatments than Fish and Carp treatments. Although benthic invertebrate richness was not
affected, epizoic richness on live unionids was 2 – 3 times higher in Fishless exclosures.
Thus, fish predation had strong top-down impact on benthic macroinvertebrate
community structure in this coastal wetland as predicted in Hypothesis #1.
I did not determine which fish were the dominant predators in this study.
However in 2007, I collected common carp, gizzard shad, yellow perch, white perch,
bluegill and channel catfish in this wetland (Kapusinski et al. In review), and these feed
heavily on invertebrates (Herdendorf, 1987; Jude & Pappas, 1992). Densities of benthic
and epizoic invertebrates in Carp and Fish treatments were the same on most dates. Thus
carp in the enclosures consumed approximately the same number of invertebrates as the
entire fish community. Carp are important predators in other aquatic systems (Riera et al.,
147
1991; Tucker et al., 1996; Zambrano & Hinojosa, 1999; Haas et al., 2007), and my data
suggests that they are one of the most important invertebrate predators in this wetland as
predicted in Hypothesis #3.
The fish selectively preyed on certain invertebrate taxa in the benthic and epizoic
habitats. Multivariate analysis found benthic and epizoic communities were distinctly
different between the Fishless treatment and Fish or Carp treatments on every sampling
date. For example, three dominant taxa (chironomids, oligochaetes and sphaeriids) were
decreased by fish predation, however, ostracods were not. Indicator taxa analysis is used
to determine which species are abundant and common in a treatment, and chironomids,
oligochaetes and ceratopogonids were indicators of the Fishless treatment. These taxa
are important prey for many species of fish, including carp (Riera et al., 1991; Tucker et
al., 1996; Zambrano & Hinojosa, 1999), and other native fish species (McNeely, 1977;
Diehl, 1992; Gido, 2003; Olson et al., 2003; Pothoven et al., 2009), and our results
suggest that fish have strong top-down control of their populations in Great Lakes coastal
wetlands.
The impacts of fish predation I found may alter various ecosystem processes. For
example, fish reduced numbers of several detritivores including chironomids,
oligochaetes, dreissenids and sphaeriid clams. In streams and lakes, detrital decay rates
dropped by 50% when fish reduced detritivore densities (Konishi et al., 2001; Mancinelli
et al., 2002; Ruetz & Newman, 2002). Changes in burrowing invertebrates such as
chironomids and oligochaetes biomass can also affect water chemistry because they
relocate nutrients from sediments into the water column (Vanni, 2002). Our results
148
suggest that further study is needed to determine if top-down control of invertebrates in
wetlands also affects ecosystem processes in these habitats.
Another unexamined potential impact is that fish may reduce food resources for
other wildlife. For example, many shorebirds feed on aquatic invertebrates in Great
Lakes coastal wetlands when they migrate to overwintering habitat in the south. Fish can
decrease water bird densities if they compete for invertebrate food resources (Haas et al.,
2007), and they may be having a similar impact on shorebirds in these critical migratory
stopovers. Furthermore, the exotic carp was an important predator in this wetland. Many
native fish breed in coastal wetland (Herdendorf, 1987), and our results indicate that carp
may be competing with the young of native species.
Unionids are a key group that affects environmental conditions by stabilizing
sediments, reallocating bethnic nutrients, and bioturbation (Vaughn & Spooner, 2006;
Vaughn et al. 2007; Zimmerman & de Szalay, 2007). In streams they also affect
macroinvertebrate community structure in mussel beds (Gutierrez et al. 2003; Howard &
Cuffey, 2006). For example, unionid shells can be an important substrate for epizoic
invertebrates in streams (Vaughn et al. 2007). Recent studies found that unionids can be
abundant in coastal wetlands (Zanetta et al., 2002; Bowers & de Szalay, 2004; Crail et al.
2011), and thus they may be important to macroinvertebrates in these habitats. I found an
abundant and diverse epizoic fauna, which differed from the benthic community. For
example burrowing organisms including oligochaetes and sphaeriids were abundant in
benthic sediments, but trichopterans, nematodes and dreissenids were more abundant on
the unionid shells. Therefore, unionid shells provided a unique microhabitat in the soft
149
benthic sediments of this coastal wetland as predicted in Hypothesis #2. Furthermore,
invertebrate communities differed between live unionids and dead unionids in carp
enclosures. The impact of predation on epizoic invertebrates was also more pronounced
on live than dead unionids, and numbers were lowest on live unionids in areas with fish.
This probably occurred because the live unionids remained above the sediments but the
dead shells settled into the sediments and were covered by silt. The fish may have grazed
the exposed shells of the live unionids more effectively than the dead shells buried in the
benthic sediments. It is also possible that live unionids scraped off attached invertebrates
when they burrowed into the sediments. Thus, live unionids presumably provide a
different microhabitat for epizoic invertebrates than other solid substrates such as dead
shells, rocks or woody debris.
Exotic Dreissenidae (zebra and quagga mussels) were common taxa on unionid
shells. Others have shown that dreissenids have largely extirpated unionids in the lower
Great Lakes because they compete food (Strayer & Smith, 1996). However, these
molluscs co-exist in some Great Lakes coastal wetlands (Schloesser et al., 1996; Bowers
et al. 2004). A possible reason for their co-existence is that fish predation in wetlands
controls zebra mussel numbers and allows unionids to persist (Tucker et al., 1996;
Molloy et al., 1997; Morrison et al., 1997; Thorp et al., 1998; Magoulick & Lewis, 2002;
Bowers et al., 2005; Bowers & de Szalay, 2007). I found dreissenid densities were 3 to 6
times lower in the Fish and Carp treatments vs. the Fishless treatment. Although this
support the importance of carp and other fish in reducing dreissends, it is still not clear
why fish do not control dreissenids in the adjoining Great Lake. For example carp are
150
abundant in Lake Erie (Jude & Pappas, 1996), so they should feed on dreissenids in the
lake.
Clusters of dreissenid can increase invertebrate density and diversity because they
add to habitat complexity in soft sediments (Stewart & Haynes, 1994; Stewart et al.,
1998) and reduce fish predation (Beekey et al., 2004). Taxa commonly collected on
dreissenid clusters are Trichoptera, Ephemeroptera, Chironomidae, Hirudinea, and
Oligochaeta (Ricciardi et al., 1997). These taxa were common in the epizoic samples,
and therefore, the high number of dreissenids in the Fishless treatment probably affected
the other epizoic invertebrates.
Fish enclosure experiments have sometimes been criticized when they do not test
realistic habitat conditions. If fish are attracted and enter the structures, this will
artificially increase predation rates. Enclosures may also affect sedimentation, shading,
or water flow, which can stress fish and reduce their feeding (Virnstein, 1978; Hulberg &
Oliver, 1980). However, I feel that my enclosure design was an effective method to test
the impacts of fish predation. I did not find other fish inside the exclosures, showing that
they successfully blocked fish access. Furthermore, I used a similar wire mesh enclosure
in past experiments, and I did not detect changes in abiotic variables such as dissolved
oxygen, pH, water conductivity or water flow (Kapusinski et al., In reveiw). When I
checked the enclosures on the sampling dates, all but 3 carp were found alive. Thus,
most carp fed on invertebrates throughout the experiment. Another potential problem is
if abnormally high stocking densities are created inside enclosures. In this experiment,
carp density in the enclosure was 0.05 carp/m2 (509 carp/ha), which is within the range of
151
natural densities. For example, Lougheed et al. (1998) reports 3500 carp/ha during
breeding season and 400 carp/ha during non breeding season in coastal wetlands, and
others found 0.25 – 1.1 carp/m2 in other aquatic systems (Zambrano & Hinojosa, 1999).
However, we acknowledge that there may have been some unexamined impacts of carp
on habitat conditions, (e.g., nutrients, Lougheed et al., 1998; turbidity Lougheed et al.,
1998, 2004) thus, additional studies are needed to further understand fish and carp effects
on invertebrate communities in Great Lakes coastal wetlands.
Acknowlegements
I would like to thank the personnel at Ottawa National Wildlife Refuge for their
invaluable assistance with this project. I also would like to thank J. Montemarano, D.
Sprockett, J. Clark, D. LaVigne, N. Howard and L. Rybus for their help constructing and
sampling the exclosures. Funding for this study was provided by the Ohio Division of
Natural Resources Wildlife Diversity program and the Herrick Grant.
152
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CHAPTER 5
EFFECTS OF FISH AND SHOREBIRD PREDATION ON BENTHIC
INVERTEBRATES IN A GREAT LAKE COASTAL WETLAND
Abstract
Great Lake Coastal Wetlands (GLCW) provide many important ecosystem
functions. Lake and wetland fish use these as breeding habitat and shorebirds use them as
migratory stopovers. Both fish and shorebirds feed on aquatic invertebrates, and thus
they could be in competition for food resources. In July 2009, I constructed floating
mesh exclosures in shallow water in Crane Creek Marsh at Ottawa National Wildlife
Refuge to examine effects of fish and shorebird predation on benthic invertebrates in
GLCW. I compared invertebrate numbers in control areas with numbers in exclosures
that excluded fish, shorebirds, and fish and shorebirds. Invertebrates were sampled using
a sediment corer in July, August, September and November. Common benthic
invertebrate taxa included Chironomidae, Ceratopogonidae, Oligochaeta and Sphaeriidae.
Shorebirds were counted during the Fall migration period, and common species were
killdeer, dunlin, long-billed dowitcher and greater and lesser yellowlegs. Migratory
shorebirds were counted from October 20th to November 28th, and numbers in Crane
162
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Creek Marsh peaked at over 3,800 birds in early November. Fish predation greatly
reduced total invertebrate numbers and diversity and numbers of common taxa in this
GLCW, but shorebirds did not have an impact. Multivariate analyses also showed that
fish predation changed overall invertebrate community structure, but there was no effect
of shorebird predation on community structure during the fall migration period. These
results show that fish are important predators in Great Lakes coastal wetlands, and they
may be reducing prey availability in areas where shorebirds are feeding. Shorebirds had
less of an impact on invertebrate communities in this marsh, but further study is needed
to show if higher shorebird numbers have a different impact and if they are affected by
competition with fish in these important migratory bird habitats.
Introduction
The Laurentian Great Lakes have annual and seasonal water level changes due to
precipitation patterns and short term water level changes due to wind set-up and seiches
(Herdendorf, 1987). Great Lakes coastal wetlands (GLCW) are unique class of marshes
because their hydrology is controlled by lake water level changes. Their dynamic
hydrology impacts the wetland biota because the frequent water level changes affect plant
community structure and fish, wildlife and aquatic invertebrates that live or feed in these
areas.
Benthic invertebrates are a key component of GLCW ecosystems. In many
aquatic habitats, invertebrates drive ecosystem-level processes like nutrient cycling
164
(Carney & Elser, 1990; Arnott & Vanni, 1996; Vaughn & Hakenkamp, 2001; Devine &
Vanni, 2002). Invertebrate communities vary within coastal wetlands depending on
sediment depths, hydrology (Cooper et al., 2007; Baumgärtner et al., 2008), and other
abiotic variables (Friday, 1987; Rader & Richardson, 1994; Balla & Davis, 1995). Less is
known about biotic interactions that affect invertebrate community structure, although
benthic invertebrate density and diversity are strongly influenced by macrophyte
communities (Gilinsky, 1984; Brown et al., 1988; de Szalay & Cassidy, 2001).
In GLCW that are open to the adjacent lakes, many fish species enter the wetlands
to feed or breed. These habitats also are important for shorebirds that migrate through the
region to their breeding grounds or overwintering areas. Common shorebirds in GLCW
are killdeer, least and pectoral sandpipers, dunlin, long-billed and short-billed dowitchers
and greater and lesser yellowlegs (Herdendorf, 1987), which overwinter as far south as
South America and breed as far north as Canada (Skagen & Oman, 1996). Therefore
shorebirds need areas with high invertebrate productivity to replenish their energy
reserves during these long-distance migrations (Mihue et al., 1997).
Fish and shorebirds feed on a wide variety of aquatic macroinvertebrates,
including insects, oligochaetes, molluscs, and crustaceans (Diehl, 1992; Helmers, 1992;
Skagen & Oman, 1996; Batzer et al., 2000). Fish and shorebird predation selectively feed
on some benthic or epiphytic taxa, but they can increase overall community diversity if
they reduce numbers of competitively dominant taxa (Batzer et al., 2000). Although fish
impact invertebrate numbers year-round (Mercier & McNeil, 1994), shorebird predation
165
is variable within the year as they migrate through the region (Ashley et al., 2000;
Hammer et al., 2006).
Studies on coastal marine wetlands found shorebirds were more important
invertebrate predators than fish (Quammen, 1984). In freshwater systems, fish are often
important predators (Batzer et al., 2000) while shorebird effects are variable (Ashley et
al., 2000). However, little is known about top-down effects of fish and shorebirds on
macroinvertebrates in GLCW, or if they compete with each other in these habitats.
I tested the impact of fish and shorebird predation on macroinvertebrates in Crane
Creek Marsh, which is a GLCW on Lake Erie. I sampled macroinvertebrate communities
in unrestricted areas and in exclosures that prevented access by fish and/or shorebirds.
H1: I expected that fish predation would have a greater impact in the summer
when water levels are higher, and shorebirds would have a greater impact during the fall
migration period when water levels decrease. Therefore, I compared the impact of fish
and shorebirds through the summer and fall.
Methods
Study Site Description
Crane Creek Marsh is a 166 ha wetland at Ottawa National Wildlife Refuge
(Ottawa Co.; latitude / longitude: 41°37′44″N / 83°12′31″W). The marsh is open to Lake
Erie by a 4-m wide channel. It is affected by Lake Erie water level fluctuations, which
166
include daily seiches and storm surges that can change water levels as much as a meter.
Furthermore, seasonal changes in Lake Erie water levels usually peak in June and
decrease by around 20 cm in winter.
Habitat within Crane Creek Marsh is mostly shallow open water (< 1 m) with
sparse submersed plants (Potamogeton, Elodea, Myriophyllum) and some scattered beds
of emergent aquatic vegetation (e.g., American water lotus, Nelumbo lutea). The benthic
sediments are deep, unconsolidated mineral clays and silts. Large mudflats are found in
the wetland when shallow areas are exposed during seiches.
Fish populations have been well documented at Crane Creek Marsh. Lake fish,
such as yellow perch, enter Crane Creek Marsh to feed and spawn in spring, and other
species (e.g. bluegill) live in the wetland year-round (Kapusinski et al. In review, Ron
Huffman, Ottawa NWR, unpublished data). Feeding by benthic feeding fish such as
common carp and gizzard shad and wave action stirs up the bottom and increase water
turbidity. Shorebirds also feed on the mudflats exposed during seiches. The Lake Erie
Marsh Region that encompasses Crane Creek Marsh and other nearby wetlands has been
recognized by the Western Hemisphere Shorebird Reserve Network
(http://www.whsrn.org) as important shorebird habitat.
Experiment Design
In summer 2009, I used exclosure experiments to test the impact of fish and
shorebirds on benthic invertebrates. I modified an exclosure design used by Quammen
167
(1981, 1984). My exclosures were a 2 m x 2 m x 0.1 m (L x W x H) wood frame with
attached Styrofoam floats. The frame was tethered to stakes to hold it in a location, but it
could float up and down when the water levels changed. I had five treatments: No fish,
No shorebirds, No fish and shorebirds, Control and Open treatments. The exclosure of the
No fish treatment was open on the top, and it had a curtain of nylon mesh (6.4 mm mesh
size) around the floating wooden frame. The bottom of the nylon mesh curtain was fixed
to the sediments with stakes to prevent fish from entering the floating exclosure. When
the water level dropped and exposed the sediments, shorebirds could enter the top of the
exclosure to feed. In the No shorebird treatment, the floating frame did not have the
nylon curtain, but the top was covered with a galvanized wire mesh (5 cm mesh size).
The frame was also extended about ~10 cm to hold the mesh off the water’s surface to
eliminate algal growth on the mesh. This blocked shorebirds from entering, but allowed
fish to enter the sides of the exclosure when it floated in the water. The No
fish/shorebird treatment was a floating exclosure with both the nylon mesh curtain and
the wire mesh top, and this prevented access by both fish and shorebirds. The Control
treatment was a floating wood frame that lacked both the nylon curtain and the wire mesh
top. This allowed entry by both fish and shorebirds, but it tested if the physical structure
of the floating wooden frame affected invertebrate numbers. The Open treatment was a 2
m x 2 m area marked with four posts that tested the effect of unrestricted access by fish
and shorebirds. All treatments were established in early July 2009 and were sampled
until November 2009.
168
Water depth changes in Crane Creek Marsh were monitored each hour with a
water level logger operated by the USGS that was installed <100 m away. In July 2009, I
measured depths in each exclosure with a meter stick to establish their baseline water
level. I calibrated these measurements with the water level logger to estimate changes
throughout the experiment. The USGS meter malfunctioned on 15 September 2009. I
estimated water levels our exclosures after that date using water depth measurements
collected by a NOAA data logger near Toledo, Ohio (field station number 9063085). To
check that these two data sets were temporally aligned, I ran a correlation analysis
comparing the USGS and the NOAA data from July to September.
Invertebrate Sampling
Benthic invertebrate densities were sampled in July, August, September and
November 2009. In each exclosure, I sampled invertebrates at three random locations
with a core sampler (5 cm dia.) embedded in the sediments (10 cm deep). The three
subsamples were combined, drained through a sieve (300 micron mesh) and stored in
bags with 90% ethanol. Because I re-sampled the exclosures several times, I minimized
disturbing the sediments by standing outside the exclosures during sampling. I also did
not re-sample previously sampled locations.
In the lab, samples were rinsed through a sieve (300 micron mesh) to remove fine
silt, and invertebrates were sorted under a dissecting microscope. Invertebrates were
169
identified to the lowest practical taxonomic level (i.e. family or genus) with taxonomic
keys (Merritt et al., 2008; Peckarsky et al., 1990) and counted.
Shorebird and fish populations
Shorebird numbers in Crane Creek Marsh were monitored by personnel from the
Black Swamp Bird Observatory during the fall migration period (late August to
November). Birds were surveyed on 10 dates using binoculars, and shorebirds were
identified to species and counted. On dates that I sampled invertebrates, I also monitored
for the presence of shorebirds and checked for their tracks in the exclosure. Diets of fish
and shorebird species were determined from published literature (e.g. Skagen and Omen,
1996) to compare to invertebrates collected in Crane Creek Marsh.
Data analysis
I compared total invertebrate numbers and numbers of dominant taxa, which were
those that comprised over 3% of all individuals on any date. I also calculated Shannon
Diversity (Hʹ), which is a widely used metric of biodiversity (Zar, 1999). Shannon’s
diversity, numbers of dominant taxa and total invertebrate numbers were compared
among treatments with one-way ANOVAs. I expected that fish, shorebird and
invertebrate populations would change among the sampling dates. Thus, I ran separate
analyses on each date. Significant ANOVAs (P <0.05), were followed by pair-wise
170
comparisons using Tukey’s HSD tests. All univariate analyses were run on JMP (version
7.0.1, 2007) statistical software.
I also tested if species assemblages differed among treatments with multivariate
statistics. Non-metric Multidimensional Scaling (NMS) ordinations were run on each
date with a Sorensen (Bray-Curtis) distance measure. I used a random starting point with
50 runs and 500 iterations with a maximum amount of 6 axes stepping down in
dimensionality. Significance was determined with a Monte Carlo test using 50 runs of
randomized data. The number of dimensions retained in the final ordination was
determined by including all axis that reduced stress (i.e. increased the model’s goodness
of fit) by at least 5 (on a scale of 0 to100) and yielded a significant model (p<0.05). I also
checked if there were indicator taxa of the treatments using the methods of Dufrene and
Legendre (1997). The significance of indicator taxa was tested with a Monte Carlo Test
with 500 permutations. I also tested if there were community-level differences among
treatments using Multi-Response Permutation Procedures (MRPP) on each sampling
date. The MRPP tests were run using the Sorensen (Bray-Curtis) distance measure with
groups being defined by treatment type. All MRPP, NMS, and indicator taxa analysis
were performed using PC-ORD (version 5.1, McCune and Mefford, 2006).
Results
Water depths were 33 – 40 cm in the exclosures on 1 July. Water level changes
measured with the USGS water level logger in Crane Creek Marsh correlated very well
171
with data collected at the NOAA field station at Toledo, OH (r 2 = 0.89). Therefore, we
used these data sets to estimate water level changes in the exclosures from July through
November. Water levels were highest in June, and declined gradually through November
(Figure 1). In June through August, exclosures were intermittently exposed during
pronounced seiches. By September, seasonal changes had decreased water levels ~10
cm, and mudflats were often exposed by seiches during the fall shorebird migration
period.
About 10,000 shorebirds in 12 species were observed in Crane Creek Marsh on
the 10 survey dates from 21 August to 28 November. The most common species were
dunlin, killdeer, long-billed dowitchers, lesser yellowlegs and greater yellowlegs (Table
1). Other species were present in lower numbers including least sandpiper and
semipalmated plovers.
I collected 16 benthic invertebrate taxa in our samples (Table 2). Chironomidae,
Sphaeriidae, Oligochaeta and Ceratopogonidae were the most common taxa, which
comprised over 3% of all invertebrates. All of the common taxa are considered important
in fish or shorebird diets.
Invertebrate biodiversity changed during the experiment. Shannon’s diversity
was low (Hʹ = 0.48-0.73) in all treatments in July (Figure 2). Diversity increased in
August and September, but declined slightly in November. Diversity was not different
among treatments in July and in November (Table 3). In August and September,
diversity was highest in treatments that prevented fish access (No Fish, No
Fish/Shorebird), but remained low in the other treatments (No Shorebird, Control, Open).
172
Figure 1: Water depths in treatment area from July 1 2009 to November 31 2009. Mean
(±1 SE) exclosure depth was set at 0 ±3 cm. Exclosures were dewatered when water
depths were 0 or less.
173
Table 1: Shorebirds counts in CCM during the 2009 fall migration season. Data provided by M. Shieldcastle (Black Swamp
Bird Observatory).
Common Name
Aug
Aug
Sep
Oct
Oct
Oct
Nov
Nov
Nov
Nov
Total
21
26
23
1
20
22
3
9
19
28
Semipalmated Plover
Killdeer
Black-bellied Plover
Pectoral Sandpiper
Least Sandpiper
0
1
0
0
0
4
8
0
0
0
0
33
0
0
0
0
21
0
0
0
45
85
2
0
0
4
78
2
15
54
0
0
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
53
227
4
15
54
Semipalmated Sandpiper
0
3
0
0
0
0
0
0
0
0
3
Dunlin
0
0
0
0
1800
1350
3800
1200
29
356
8535
Long-billed Dowitcher
Greater Yellowlegs
Lesser Yellowlegs
Hudsonian Godwit
0
7
2
0
0
6
9
0
0
3
24
0
0
0
2
0
18
6
4
0
59
26
52
0
11
0
2
1
19
5
0
0
0
0
1
0
0
0
0
0
107
53
96
1
Snipe
0
0
0
0
1
0
0
0
0
0
1
174
Table 2: Benthic invertebrate taxa in treatment areas. Numbers are the percent that each
taxa comprised of total invertebrates in a treatment across all dates. Total are the total
number of invertebrates in a treatment. Taxa with an asterisk* are common taxa.
175
Taxa
Treatment
Open
Insects
Ephemeridae
Ephemera
Caenidae Caenis
Corixidae
Chironomidae*
Ceratopogonidae
Crustaceans
Ostracoda
Amphipoda
Water mites
Hydrachnidia
Molluscs
Sphaeriidae*
Unionidae
Dreissenidae
Corbiculidae
Segmented
worms
Oligochaeta*
Leeches
Hirudinea
Total Numbers
<1
46.6
<1
Control
No Fish
No
Shorebird
No
Fish/Shorebird
<1
<1
<1
51.6
1.2
<1
38.2
7.1
<1
51.5
1.3
<1
<1
31.2
4.5
<1
<1
4.8
<1
5.8
41.6
<1
1.7
<1
36.4
18.4
291
<1
706
<1
7.1
7.9
<1
1.2
<1
23.6
<1
1.1
<1
44.3
37
27.7
253
<1
254
647
176
Figure 2: Mean Shannon diversity (±1 SE) in treatments from July to November. Oneway ANOVAs comparing treatments were run on each sample date. Different letters
over bars indicate that treatments are different (P ≤ 0.05) on that sampling date.
177
Table 3: Results of ANOVA tests comparing total invertebrates, diversity, and numbers of the common taxa across treatments
on each sampling date. Bold P values indicates there were significant differences among treatments.
Total Invertebrates
Shannon's
Diversity
Chironomidae
Oligochaeta
Sphaeriidae
Ceratopogonidae
July
F4,39; P value
0.87; 0.491
August
F4,39; P value
5.58; 0.001
September
F4,39; P value
24.77; < 0.001
November
F4,39; P value
4.41; 0.006
0.96; 0.443
1.12; 0.361
0.81; 0.530
0.97; 0.439
Not collected
9.17; < 0.001
0.91; 0.472
2.25; 0.084
4.60; 0.004
7.66; 0.002
6.36; < 0.001
10.01; < 0.001
1.85; 0.141
5.51; 0.002
11.36; < 0.001
1.73; 0.165
0.78; 0.545
2.34; 0.074
5.74; 0.001
3.35; 0.0200
178
Total benthic invertebrate densities were not different among treatments on the
first sampling date in July (Figure 3, Table 3). Later, invertebrate densities declined in
treatments with fish access but increased in treatments that prevented fish access. In
August and September, the No fish and No fish/shorebird treatments had higher densities
than other three treatment types (Figure 3, Table 3). In November, densities in the No
fish and No fish/shorebird treatments were again higher than the Open and Control
treatments, and densities in the No shorebird treatment were intermediate.
Temporal patterns of abundance of dominant invertebrate taxa were complex.
Chironomid midges were abundant in all treatments in July but decreased in all
treatments in August (Figure 4, Table 3). In September, Chironomidae numbers
increased and were higher in No fish and No fish/shorebird treatments than the Open,
Control and No shorebird treatments. However, densities in November decreased and
were not different among treatments. Oligochaete worm numbers were relatively
constant, and they did not differ among treatments on any date (Figure 4, Table 3).
Sphaeriid clam numbers were low in July and did not differ among treatments. Their
populations increased in later dates. Sphaeriidae numbers were higher in No fish and No
fish/shorebird treatments in August, September and November than in treatments with
fish access (Figure 4, Table 3). For example, August Sphaeriidae densities were <100
individuals m -2 in Open, Control and No shorebird treatments but were about ten times
higher in No fish and No fish/shorebird treatments.
Multivariate analyses detected different invertebrate species assemblages among
treatments on several dates. There were no significant NMS ordinations in July and
179
Figure 3: Total invertebrate numbers (±1 SE) in each treatment from July to November.
One-way ANOVAs comparing treatments were run on each sample date. Different
letters over bars indicate that treatments are different (P ≤ 0.05) on that sampling date.
180
Figure 4. Densities of common taxa (±1 SE) in treatments from July to November. Oneway ANOVAs comparing treatments were run on each sample date. Different letters
over bars indicate that treatments are different (P ≤ 0.05) on that sampling date.
181
182
August. Statistical results for the two-dimensional solution in July were Stress (a) =
13.45, P = 0.098, and in August were Stress (a) = 29.53, P = 0.706. MRPP statistics did
not detect differences among treatments in July (Table 4). However, in August MRPP
detected differences between the No fish and No fish/shorebird and the other treatment
types. NMS ordinations in September and November were significant (Stress (a) =
14.54, P = 0.039 and Stress (a) = 20.01, P = 0.039 respectively) (Figure 5). In these
months, a two dimensional ordination provided the best resolution of the data. MRPP
tests in September and November found that invertebrate communities in the No fish and
No fish/shorebird treatments usually differed from the Open, Control and No Shorebird
treatments (Table 4). However, the No fish and No fish/shorebird treatments were never
different from each other.
During the initial sampling session in July there were no indicator taxa for any
treatment type. In August, Sphaeriidae were indicators of the No fish/shorebird treatment
(P = 0.002), and Ceratopogonidae were indicators of the No fish treatment (P = 0.003).
In September, Ceratopogonidae were again an indicator of the No fish treatment (P <
0.001), and Chironomidae and Sphaeriidae were indicators of the No fish/shorebird
treatment (P = 0.004, P < 0.001, respectively). In November, Sphaeriidae was again an
indicator taxa of the No fish/shorebird treatments (P = 0.001). The other three treatment
types had no indicator taxa.
183
Table 4: MRPP pairwise comparisons of invertebrate communities in treatments. Bold values are statistically significant.
July
Control
No Fish
No Shorebird
No Fish/Shorebird
Open
0.775
0.116
0.091
0.436
Control
No Fish
No Shorebird
0.288
0.261
0.661
0.765
0.543
0.349
August
Control
No Fish
No Shorebird
No Fish/Shorebird
Open
0.476
0.057
0.158
0.001
Control
No Fish
No Shorebird
0.007
0.266
< 0.001
0.006
0.437
< 0.001
September
Control
No Fish
No Shorebird
No Fish/Shorebird
Open
0.867
< 0.001
0.926
< 0.001
Control
No Fish
No Shorebird
< 0.001
0.602
< 0.001
< 0.001
0.126
< 0.001
November
Control
No Fish
No Shorebird
No Fish/Shorebird
Open
0.978
0.013
0.283
0.003
Control
No Fish
No Shorebird
0.027
0.202
0.016
0.163
0.198
0.012
184
Figure 5: Two dimensional NMS ordinations of invertebrate communities in September and November. The percentages of
observed variation explained by each axis are labeled on the figures.
185
Discussion
Top-down effects of fish on macroinvertebrate community structure have been
well-documented in freshwater systems (Batzer et al., 2000; Gido, 2003; Pothoven, 2009;
Hentges & Stewart 2010;). Shorebirds sometimes affect macroinvertebrate numbers
(Schneider, 1978; Quammen, 1981; Wilson, 1989; Mercier & McNeil, 1994), but other
studies did not detect impacts of shorebird predation (Ashley et al., 2000; Mitchell &
Grubaugh, 2005; Hammer et al., 2006). In this unimpounded GLCW, fish predation
decreased invertebrate densities by about 80% (i.e. total numbers decreased from 5,000
invertebrates m-2 to 1,000 invertebrates m-2). Water depths in these areas were usually 25
cm or less, which means that these changes occurred in shallow areas that were
frequently dewatered. The effect of fish predation was probably even greater in deeper
areas. However, shorebird predation during the fall migration period did not greatly
impact invertebrate numbers. Multivariate analyses found that communities differed
between open areas and fish exclosures (No Fish and No Fish/shorebird treatments) but
not shorebird exclosures (No Shorebird treatment). Therefore, in shallow water areas in
this coastal wetland, fish cause a strong top down pressure but shorebirds are less
important predators.
Exclosures studies have sometimes been criticized because they may cause
unintended changes to the habitat that confound the variables being tested, for example,
the physical structure of exclosures may alter predator behavior or other key
environmental conditions (i.e., increase sedimentation or reduce algae growth; Virnstein,
186
1978; Hulberg & Oliver, 1980). However, I used a similar exclosure design in a previous
study and found that abiotic variables (pH, dissolved oxygen, conductivity) were not
different between exclosures and open areas (Kapusinski et al., In Review). Most
shorebird species prefer open areas, but I don’t believe they were deterred from feeding
in my exclosures because I observed their tracks inside the No Fish treatments on several
dates (D. Kapusinski, pers. observ.). I also believe that fish behavior and environmental
variables were not affected by the physical structure of the exclosures because
invertebrate communities were similar in Open, Control, and No shorebird treatments.
Thus, the exclosure design I used was a realistic method to test the impacts of fish and
shorebird predation.
Fish communities in GLCW are diverse, and I cannot directly determine which
species were the dominant predators in this wetland. However in a 2007 study at Crane
Creek Marsh (Kapusinski et al., In Review), we collected many fish species that are
invertebrate predators. I found gizzard shad, yellow perch, white perch, bluegill and
channel catfish were abundant, and these fish feed mostly on benthic invertebrates
(McNeely, 1977; Diehl, 1992; Gido, 2003; Olson et al., 2003; Pothoven et al., 2009).
The species collected in this wetland are similar to assemblages described in other
GLCW (Jude & Papas, 1992), and thus fish probably have a similar impact on
invertebrate communities in other sites.
I found top-down control by fish had a strong impact on invertebrate community
structure in this Great Lake Coastal Wetland. The NMS and MRPP analysis determined
that fish community structure was clearly different in treatments that excluded fish (No
187
fish and No fish/shorebird treatments) and the other treatment types. Most of the
common taxa (chironomids, ceratopogonids and sphaeriids) had lower densities in areas
accessible by fish, although oligochaete numbers were not significantly affected.
However in a previous study, oligochaete densities were also reduced by fish predation
(Kapusinski et al., In Review). Indicator taxa analysis also found that all indicator taxa
(sphaeriids, chironomids, and ceratopogonids) were associated with areas that lacked fish
predation effects. These taxa are important in diets of carp, bluegill, channel catfish
(McNelly & Pearson, 1977; Thorp & Bergey, 1981; Haas et al., 2007) that were common
in CCM. Therefore, excluding fish in shallow water areas in this coastal wetland,
allowed distinct benthic invertebrate communities to develop that were dominated by
taxa that were reduced in the rest of the wetland by predation pressure.
The changes in invertebrate communites that I found are important because it may
affect ecosystem level processes. For example, invertebrates impact nutrient cycling and
detritus processing (Carney & Elser, 1990; Arnott & Vanni, 1996; Vaughn &
Hakenkamp, 2001; Devine & Vanni, 2002), and loss of detritivores can decrease detritus
processing rates in streams and lake littoral zones (Konishi et al., 2001; Mancinelli et al.,
2002; Ruetz & Newman, 2002). Fish predation can also reduce herbivore densities and
indirectly increase periphyton (Dorn et al., 2006). However, no studies have been done
in GLCW and further research is needed to determine if the community-level changes I
found changed ecosystem processes in these habitats.
I did not detect an effect of shorebird predation on invertebrate community
structure. Wetlands along the Laurentian Great Lakes are important stopovers during
188
migrations for shorebirds, and the exclosures were constructed in intermittently exposed
areas that were shallow enough for shorebirds to feed. Shorebirds were seen feeding in
and around the treatments that were accessible by shorebirds (Open, Control, No Fish).
Common shorebirds observed at Ottawa NWR included dunlin, killdeer, greater and
lesser yellowlegs and long-billed dowitchers. Shorebird numbers at Ottawa were similar
or slightly higher than other inland migration stopovers (Ashley et al., 2000; Mitchell &
Grubaugh, 2005; Hamer et al., 2006). Therefore, I should have detected an affect of
shorebirds on benthic invertebrate numbers, and my results support the conclusion that
fish have a greater impact than shorebirds on benthic invertebrate communities in
GLCW. Studies that detected shorebird impacts on benthic invertebrates are generally in
coastal flyways, where shorebird densities are much higher (Mitchell & Grubaugh, 2005).
Therefore, shorebird may have a greater impact in GLCW if they occur in higher
concentrations.
Although shorebirds did not reduce invertebrate numbers at Ottawa NWR, they
are probably competing with fish for similar abundant food resources. Shorebirds seek
areas with shallow water and abundant invertebrate populations, and they are
opportunistic feeders on the common prey items (Hamer et al., 2006). Fish can decrease
food for water birds in some habitats (Haas et al., 2007). In Crane Creek Marsh, total
invertebrate numbers were much lower in areas accessible to fish. Furthermore, taxa
reduced by fish predation are important in shorebird diets. For example, dunlins
accounted for 93% of all shorebirds in this wetland, and their diets include diptera
including chironomids, oligochaetes, crustaceans, and molluscs (Skagen & Oman, 1996).
189
Further studies are needed to determine if the patterns I observed are common in other
Great Lake coastal wetlands, and how fish predation influences habitat preference by
shorebirds in these important migratory bird habitats.
Acknowlegements
I would like to thank the personnel at Ottawa National Wildlife Refuge for their
assistance with this project. I also would like to thank J. Montemarano, J. Clark and M.
Bagley for their help constructing and sampling the exclosures. Funding for this study
was provided by the Herrick Grant.
190
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CHAPTER 6
SYNTHESIS
Introduction
In this dissertation, I described a series of field-based ecological experiments that
examined community ecology in Great Lakes coastal wetlands (GLCW) along the
Laurentian Great Lakes. In the past, these wetlands were considered economically
worthless, and as a result, 60-80% of GLCW have been drained, filled or impounded
(Maynard & Wilcox, 1997; Comer et al., 1995). More recently, the economic and
ecological values of this habitat type are becoming better understood. For example,
GLCW provide critical habitat for fish, birds, amphibians, reptiles, invertebrates and
mammals (Becker, 1983; Harris et al., 1983; Herdendorf, 1987; Jude & Papas, 1992;
Prince et al., 1992; Maynard & Wilcox, 1997; Cardinale et al., 1998; Bowers & de
Szalay, 2004; Uzarski et al., 2005). However, there is still a lack of understanding about
the factors that drive community structure in these wetlands.
I studied several the impact of several factors that potentially influenced aquatic
invertebrate community structure including: if communities differed in open coastal
wetlands and impounded wetlands, if migratory shorebirds and fish predation controlled
benthic invertebrate populations in open coastal marshes, and if fish predation affected
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epizoic invertebrates on native unionid mussels? These studies provided insight on the
ecology and community dynamics in these wetlands, and also information to better
conserve and manage these habitats. In my initial project, I sampled impounded (i.e.
diked) wetlands and an open (i.e. unimpounded) coastal wetland, Crane Creek Marsh at
Ottawa National Wildlife Refuge, to compare their fish and invertebrate communities. In
subsequent experiments, I used field exclosures to examine impacts of fish and
shorebirds on aquatic invertebrates in the open coastal wetland. Below, I will summarize
the results of my dissertation research and discuss the implications in detail.
Do open Great Lakes coastal wetlands have different invertebrate and fish
communities than other types of wetlands?
Although all wetlands are affected by water level changes, open GLCW are
unique because they are affected by short-term water level fluctuations, seasonal and
inter-annual water level fluctuations through a connection to the Great Lakes.
Predictable seasonal changes in Lake Erie water levels cause levels in Crane Creek
Marsh (CCM) to reach a maximum in early summer. Short-term unpredictable lake
water level changes (e.g. seiches) change water levels by as much as 2 m on a daily basis.
Furthermore, many GLCW are also affected by riparian inputs because they are found at
the mouths of rivers and streams that flow into the lakes (Herdendorf, 1987). My
hydrographs at Crane Creek marsh show both unpredictable daily water level changes
and the expected peak level in June as expected Lake Erie coastal wetlands. In contrast,
inland riparian wetlands generally flood once or twice in a year after spring thaws or
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large rainfall events (i.e. a “flood pulse hydrology”). Water levels in large-order riparian
wetlands remain high for days to weeks and then gradually decline as floodwaters recede
(Middleton, 1999). Thus, the hydrology I described in Crane Creek Marsh is very
different than in other common wetlands in Ohio.
While open GLCW are different than riparian wetlands, open GLCW are also
different from the coastal wetlands that have been impounded by man-made dikes.
Impounded wetlands are hydrological isolated, and they are manually flooded by opening
water control structures. Water levels in impounded GLCWs are often managed to
provide specific management goals (i.e. migratory waterbird habitat), and levels are
stable unless being manually flooded or drained. Impounded wetlands at ONWR are
either seasonally or permanently flooded, but they are usually flooded in Fall when many
shorebirds and waterfowl used these wetlands as stopovers in their winter migration. In
contrast, water levels Crane Creek Marsh had dropped and exposed much of the shallow
areas by fall.
Due to these hydrological differences, open coastal wetlands and other wetland
types often differ in their biota (e.g. species presence/absence, abundance) (Gilinsky,
1984; Campeau et al., 1994; Cardinale et al., 1998), and abiotic factors (water mixing,
dissolved oxygen, turbidity, pH, salinity, sedimentation rates) (USEPA, 1993; Cardinale
et al., 1997). For example, I observed that the impounded wetlands had higher
macrophyte densities than the adjacent Crane Creek Marsh. I also found that the fish
community in impounded wetlands was dominated by species such as green sunfish and
bullhead catfish, while Crane Creek Marsh was dominated by yellow perch, gizzard shad
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and shiners. The benthic invertebrate community also was different, which was probably
affected by differences in abiotic conditions and fish predation. I found that the dominant
invertebrates in impounded wetlands were crustaceans (Amphipoda, Copepoda) biting
midges (Ceratopogonidae), leeches (Hirudinea) and snails (Physidae), but open coastal
wetlands were dominated by midges (Chironomidae), roundworms (Nematoda), worms
(Oligochaeta) and other crustaceans (Cladocera , Ostracoda). Hydrology probably played
a major role in these differences. For example, I detected different invertebrate
communities in three water depth strata (shallow (<18 cm), medium (18-34 cm), and
deep (>34 cm)) water depths in open coastal wetlands, but not in impounded wetlands.
However, I did not gather data to directly test the effect of many potentially important
biotic or abiotic factors. Thus, further research is needed to determine which factors
control the community structure of benthic invertebrates.
What is the importance of fish and shorebird predation in structuring benthic
invertebrate communities in Great Lakes coastal wetlands?
Fish are important predators and have top-down impacts on invertebrate
communities in inland wetlands (Batzer et al., 2000; Haas et al., 2007), lakes (Mc Neely,
1977; Olson et al., 2003), reservoirs (Gido, 2003), streams (Winkelmann et al., 2011) and
within the Great Lakes (Pothoven et al., 2009; Morrison et al., 1997). Past studies have
found that predators often decrease invertebrate abundance (Haas et al., 2007; Diehl,
1992; Morin, 1984). I showed that medium and large-bodied fish such as bluegill,
201
emerald shiners, channel catfish, yellow perch and gizzard shad were some of the key
predators of benthic invertebrates.
Predatory fish can also reduce biodiversity (Dorn et al., 2006; Persson, 1999;
Carpenter & Kitchell, 1993) by greatly reducing or eliminating populations of their prey
taxa. However, I found that fish did not significantly reduce overall invertebrate taxa
richness. Instead, I found numbers of some dominant taxa were reduced (e.g.,
Sphaeriidae and Oligochaeta) while others increased (e.g., Chironomidae) in areas with
fish access. Others have also found that fish predation can have a minimal effect on
diversity if they eliminate competitively dominant taxa. For example, predation by
pumpkinseed sunfish, black crappie, brown bullhead and common carp increased midge
density when they reduced numbers of competitive taxa (Planorbidae and Physidae) and
predators (Corixidae and Glossiphoniidae) (Batzer et al., 2000).
It is not entirely clear how invertebrate numbers increased rapidly inside the
exclosures. Many larval invertebrates may have been physically transported into the
exclosures by water flow during seiches. However, many invertebrates are multivoltive
and have high reproductive rates (e.g. chironomids) (Coffman & Ferrington, 1996).
Adult sphaeriid clams also reproduce rapidly. For example, adult Sphaerium clams
produce several cohorts of offspring each year and their offspring reach maturity within 6
months (Heard, 1977). Therefore, the observed population differences could have been
due to reproduction when eggs were washed into the exclosure, and offspring had higher
survival inside the exclosures.
202
Fish also affect benthic invertebrates indirectly by altering habitat conditions.
Gido (2003) found that foraging detritivorous gizzard shad dislodged benthic
invertebrates (chironomids and ostracods) where they were consumed by other fish
predators. Gizzard shad were common in Crane Creek marsh , and I found gizzard shad
consumed benthic invertebrates (chironomids and ostracods), perhaps by accidentally
ingesting them in the detritus. Thus, these common detritivore fish may also be
important in structuring invertebrate communities in GLCW.
In this dissertation, I showed that fish predation had a pronounced impact on
benthic invertebrate community structure even in shallow areas that were intermittently
exposed. Thus, significant numbers of fish moved into these areas during high water
periods to feed. This horizontal migration of fish into flooded areas in open coastal
wetlands would not occur in the stable water levels of impounded wetlands. This shows
wildlife managers who providing habitat for breeding fish may want to manage exposed
areas adjacent to deeper waters. For example, mowing could be conducted in the
exposed areas to provide abundant plant detritus for aquatic invertebrates (de Szalay and
Resh, 1997)
Migratory shorebirds can also be important predators in some wetlands
(Schneider, 1978; Quammen, 1981; Wilson, 1989; Mercier & McNeil, 1994). Due to the
large populations of shorebirds that stopover in western Lake Erie marshes during the fall
migration, I tested if they have a top-down effect on invertebrate numbers. Surprisingly,
I detected little effect of shorebirds on benthic invertebrate communities. Invertebrate
communities in shorebird exclosures vs. open control areas had similar total numbers,
203
biodiversity, and numbers of individual invertebrate taxa. This occurred even though I
observed shorebirds feeding around the exclosures. I also observed footprints of
shorebirds inside the open exclosures that allowed shorebird access. However,
shorebirds may not have occurred in sufficient numbers to deplete benthic invertebrates
(at least to a level that I could detect in my samples). Most studies that describe an effect
of shorebirds feeding were conducted in marine tidal wetlands, where shorebirds reach
high numbers. In contrast, studies that did not find an effect (Ashley et al., 2000;
Mitchell & Grubaugh, 2005; Hammer et al., 2006) were conducted in freshwater
wetlands. For example, Mitchell et al. (2005) sampled six inland federal wildlife refuges
and only found impacts on chironomid numbers in a single site. Shorebird numbers in
Crane Creek Marsh peaked on November 3rd 2009 at ~4000 individuals. Shorebirds
densities in Crane Creek Marsh were comparable to other studies in freshwater wetlands
(about 12 birds / ha) or 4000 per day (Ashley et al., 2000). However, shorebirds numbers
can be far higher at times in marine tidal marshes. For example, Clark et al. (1993) found
average daily numbers of shorebirds in Delaware Bay were 216,000 with a peak of over
426,000! Therefore, the lack of an impact may be due to the relatively low numbers of
shorebirds in my study area, and they may be more important in other wetlands. Since
few open GLCW remain, all remaining coastal wetlands provide essential habitat for fish
and migratory shorebirds. Competition between these species would lead to a reduction
in fecundity, survival or biomass of these economically important taxa. Therefore,
further research should examine potential competition between shorebirds and fish in
other GLCW.
204
Do fish affect epizoic communities on shell of native unionid mussels?
Few studies have examined epizoic invertebrate communities found on mollusk
shells, and even fewer have examined fish predation of epizoic invertebrates on unionid
shells in Great Lakes coastal wetlands. Mollusc shells are a unique microhabitat in
GLCW because they are a stable substrate in soft benthic sediments. For example,
Burlakova et al., (2012) found higher invertebrate densities associated with zebra mussels
than nearby benthic substrates. Bowers and de Szalay (2004, 2007) also found that
unionid shells provided a substrate colonized settling juvenile zebra mussels.
Understanding the ecology of the invasive zebra mussel is important because their
introduction into the Great Lakes caused drastic declines in native unionid populations
(Ricciardi et al., 1995, 1998; Strayer & Smith, 1999). A few remnant populations have
been found in shallow littoral areas (Gillis & Mackie, 1994; Schloesser & Nalepa, 1994;
Crail et al., 2011) and coastal wetlands (Zanetta et al., 2002; Bowers & de Szalay, 2004).
In lakes and rivers, vertebrate predators such as carp (Tucker et al., 1996), freshwater
drum, pumpkin seed sunfish, rock bass (Watzin et al., 2008) yellow perch (Morrison et
al., 1997; Watzin et al., 2008) and map turtles (Serrouya et al., 1995; Lindeman, 2006)
consume exotic zebra mussels. Predation has been proposed as an important biological
factor that reduces zebra mussels in coastal wetlands and other habitats (Bowers & de
Szalay, 2007; Carlsson et al., 2011; Goote & Bergman, 2012). I also found that fish ate
high numbers of zebra mussels on unionid shells. In my study, common carp were one of
the most important predators. These results are similar to those conducted in other
205
habitats suggesting the importance of carp in controlling zebra mussels (e.g. Tucker et
al., 1996).
Furthermore, fish predation also affected other epizoic invertebrates. Common
epizoic invertebrates were Chironomidae, Dreissenidae, Nematoda, Oligochaeta and
Ostracoda, and predation reduced dreissenid, oligochaete and ostracod numbers. Overall
numbers were lower in the presence of fish. I also found that epizoic communities were
different on shells of live and dead unionids. This observation shows that native unionids
may be valuable because they increase habitat complexity and therefore overall
biodiversity (Beckey et al., 2004).
206
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