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HOW THE FEEDING ECOLOGY OF NATIVE AND EXOTIC MUSSELS AFFECTS
FRESHWATER ECOSYSTEMS

By

David Francis Raikow

A DISSERTATION

Submitted to
Michigan State University
in partial fillfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Department of Zoology
W. K. Kellogg Biological Station
and
Program in Ecology, Evolutionary Biology, and Behavior

2002

ABSTRACT

HOW THE FEEDING ECOLOGY OF NATIVE AND EXOTIC MUSSELS AFFECTS
FRESHWATER ECOSYSTEMS

By
David Francis Raikow

As zebra mussels expand their range in North America new ecosystems become
invaded. Beginning with the Great Lakes and continuing through the Mississippi River
drainage network, the zebra mussel invasion is currently spreading among small inland
lakes. Due to the swifiness and magnitude of this invasion and the effects zebra mussels
have had on previously invaded ecosystems, it is imperative that the effects of zebra
mussels on inland lakes be investigated. In addition, the zebra mussel invasion represents
the functional replacement of the multispecies assemblage of native bivalves which
typically disappears after invasion. This dissertation examined the ecology of native
unionid bivalves (Unionoidae) and exotic zebra mussels (Dreissena polmogpha) in order
to learn what North American freshwater ecosystems are losing and gaining.

I examined a community of unionids in a stream by using both natural and
experimentally enriched stable isotopes of nitrogen as a spin-off project of the Lotic
Intersite Nitrogen eXperiment (LINX). This is the first and only isotopic enrichment of
this taxon. The evidence suggests, contrary to conventional wisdom, that native
freshwater mussels they may not be exclusively suspension feeders as adults.

The potential effects of zebra mussels on fish are poorly understood. I examined
how larval bluegill growth and survival could be affected by zebra mussels in an
experimental mesocosm setting. The hypothesized mechanism of competition between

mussels and bluegill for food in the form of microzooplankton was supported.

Competition between zebra mussels and obligate planktivores may contribute to
ecological harm done to small inland lakes as zebra mussels spread throughout North
America. The effects of zebra mussels on microzooplankton reproduction was studied as
part of a large mesosoem ZEbra mussel EXperiment (ZMEX), in order to evaluate the
relative importance of direct predation on microzooplankton by zebra mussels and
indirect competition for resources (phytoplankton as food) causing reduced fecundity.

Another anticipated impact of zebra mussels is the alteration of phytoplankton
community structure, including the promotion of the toxic alga Microcystis. I surveyed
60 lakes in the lower peninsula of Michigan to see whether the phytoplankton community
has changed in the presence of zebra mussels. Standing stocks of phytoplankton have
clearly been reduced in the presence of zebra mussels. While the relative abundance of
Microcystis appears to have increased in the presence of zebra mussels there has been
little effect on overall phytoplankton community structure.

Nutrient regeneration and benthic-pelagic coupling are important issues

concerning zebra mussels. I measured ammonium (NH4), soluble reactive phosphorus

(SRP), and dissolved organic carbon (DOC) concentrations, and the composition of
biodeposits in the mesocosm experiment and survey. While increased SRP was detected
in the mesocosm experiment, no evidence of increased SRP was found in natural
systems. Zebra mussels have, however, evidently reduced the concentrations of DOC in
lakes. This is important because DOC attenuates UV-B radiation, and thus invaded

systems may be more susceptible.

For my parents, who introduced me to the family business.

iv

ACKNOWLEDGEMENTS

To earn a scientific degree is to be an apprentice, and I have had the privilege to
learn from masters of the crafi. This opinion is in no small part due to my advisor
Stephen K. Hamilton. Thanks Steve for allowing me the freedom to find a topic that
interested me and for the opportunities to learn techniques. I have tried to take in all I can
from your mentoring, but your dedication and discipline are tough acts to follow. Thanks
also to Gary Mittelbach and Alan Tessier of my committee for allowing me to knock on
their doors and ask questions. Special thanks to Orlando Samelle who allowed me to
learn phytoplankton identification in his lab and for help interpreting data.

I'm honored to have been able to call the Kellogg Biological Station an academic
home. Everyone who attended KBS brown bag and EEBB colloquium seminars provided
great feedback to embryonic versions of this dissertation. I'd like to thank Jessie Davis,
the ZMEX crew Bethany, Tomoko, Justin, Elizabeth, Carrie and Allison, and my fellow
students Alan Wilson, Chris Stiener, Jessica Rettig, Laura Broughton, Tara Darcy, and
Natalie Dubois for help in the field, methodological and statistical advice, and for
listening to my general complaints about, well, pretty much everything. This dissertation
would not have been possible but for the invaluable logistic support with supplies,
computers and laboratory analyses provided by Nina Consolatti, John Gorentz, and
Madhav Machavaram, and David Weed. Special thanks to lab mate Eric Thobaben,
fellow Cutco enthusiast, for resurrecting my interest in The Game (long live Benthos and

Stefan!) thus providing a well-needed balance to work that helped maintain my sanity. I'd

especially like to thank my lab mate Stefanie Whitmire for all her support during some
tough times. You’re a true fiiend Stef.

I received additional help from D. Graf and R. Mulcrone of the University of
Michigan, Ethan Nedeau, and Nate Dorn who identified mussels, Jennifer Tank who
provided valuable insights into data, R. J. Stevenson who discussed data analysis with
me, and C. Kulpa and D. Birdsell of the Center for Environmental Science and

Technology at the University of Notre Dame who kindly allowed me to use their

analytical facility for 815N analyses. Financial support was provided by the Zoology
Department, Program in Ecology, Evolutionary Biology and Behavior, Lauff Research
Award, Sigma-Xi Grants in aid of Research, Michigan Sea Grant, Steve Hamilton's
career grant fimded by the US. National Science Foundation (DEB-9701714), the KBS
Graduate Research Training Group (RTG) funded by the US. National Science
Foundation (DBI-9602252), the overall LINX grants funded by the US. National Science
Foundation to Virginia Polytechnic University (DEB-9628860) and to Michigan State
University (DEB-9810220 and DEB-9701714). Special thanks to the Environment Now
Fund of the Kalamazoo Community Foundation for generous financial support.

Lastly, I must thank my wonderful wife and emergency field assistant Sue. You
listen to me with enthusiasm, support me when I'm unsure, and inspire me to continue.
Thank you for putting up with late nights, lost weekends, mosquitoes and leaky hip

waders. Nothing I do is possible without you.

vi

TABLE OF CONTENTS

LIST OF TABLES ........................................................................... ix

LIST OF FIGURES .......................................................................... x

CHAPTER 1

INTRODUCTION: BIOLOGICAL INVASIONS, ZEBRA MUSSELS, AND THE

REPLACEMENT OF NATIVE UNIONIDS .............................................. 1
Overview .............................................................................. 1
Life history and physiology ......................................................... 3
Ecological impacts observed and hypothesized .................................. 8
Past research efforts .................................................................. 15
Further motivations for this dissertation ........................................... 15
Literature cited ........................................................................ 17

CHAPTER 2

BIVALVE DIETS IN A MIDWESTERN U.S. STREAM: A STABLE ISOTOPE

ENRICHMENT STUDY .................................................................... 22
Abstract ............................................................................... 22
Introduction ........................................................................... 23
Methods ............................................................................... 26
Results ................................................................................. 30
Discussion ............................................................................. 34
Acknowledgements .................................................................. 43
Literature cited ........................................................................ 44

CHAPTER 3

FOOD WEB INTERACTIONS BETWEEN ZEBRA MUSSELS, ZOOPLANKTON,

AND LARVAL FISH ........................................................................ 63
Abstract ............................................................................... 63
Introduction ........................................................................... 64
Methods ............................................................................... 66
Results ................................................................................. 69
Discussion ............................................................................. 71
Acknowledgements .................................................................. 75
Literature cited ........................................................................ 76

CHAPTER 4

ALTERATION OF NUTRIENT CYCLING BY ZEBRA MUSSELS .............. 87
Abstract ............................................................................... 87
Introduction ........................................................................... 88
Methods ................................................................................ 91
Results ................................................................................. 94

vii

Discussion ............................................................................. 97

Acknowledgements .................................................................. 101
Literature cited ........................................................................ 102
CHAPTER 5
ZEBRA MUSSEL IMPACTS ON THE PHYTOPLANKTON COMMUNITIES
AND WATER CHEMISTRY OF INLAND LAKES .................................... 116
Abstract ............................................................................... 116
Introduction ........................................................................... 1 17
Methods ............................................................................... 122
Results ................................................................................. 125
Discussion ............................................................................. 128
Acknowledgements .................................................................. 136
Literature cited ........................................................................ 137
ZEBRA MUSSEL EFFECTS ON MICROZOOPLANKTON FECUNDITY ...... 159
Abstract ............................................................................... 159
Introduction .......................................................................... T 160
Methods ............................................................................... 162
Results ................................................................................. 164
Discussion ............................................................................. 165
Acknowledgements .................................................................. 167
Literature cited ........................................................................ 168
APPENDICES ................................................................................ 173
Appendix A: Methods for the design and construction of limnocorrals ...... 174
Appendix B: Methods for the collection and rearing of zebra mussels ....... 181
Appendix C: Methods for the collection and rearing of larval bluegill ....... 183
Appendix D: SAS program codes ................................................... 190
Appendix E: Complete references with additions ................................ 195

viii

LIST OF TABLES

CHAPTER 2
Table 1. Hydraulic and geomorphological features of Eagle Creek, MI ............... 49
Table 2. Biomass and nitrogen content of consumers in Eagle Creek, MI ............ 50

Table 3: Chlorophyll a and relative abundance of planktonic and benthic algae in
Eagle Creek, M1, on 20 Aug 1998. The wetlands occur between the lake outflow

and the dripper ................................................................................. 51
Table 4: Food resource partitioning for bivalves in Eagle Creek, MI, based on
observed 15N enrichment on day 42 of the experiment. Stomach gland 515N was
used for bivalves unless otherwise noted in text ......................................... 52
CHAPTER 3

Table 1. Pooled-variance t test of zebra mussel (ZM) effects on larval bluegill
(BG) growth and survival ..................................................................... 80

Table 2. Univariate repeated measures AN OVA test for effects of zebra mussels
and larval bluegill on macrozooplankton biomass over the course of the
experiment (days 3 to 13) ....................................................................... 81

Table 3. Univariate repeated measures ANOVA and multivariate repeated measures
ANOVA (MANOVAR) testing for effects of zebra mussels and larval bluegill on

microzooplankton abundance over the course of the experiment (days 3 to 13) ....... 82
CHAPTER 5

Table 1. Dominant phytoplankton genera encountered during surveys of Michigan
Lakes ............................................................................................ 142
CHAPTER 6

Table l. Experiment-wide response of variables .......................................... 170
APPENDICES

Table A1: Materials needed to build limnocorrals ........................................ 178

ix

LIST OF FIGURES

CHAPTER 2

Figure 1. Natural abundance of 15N in food resources and bivalves of Eagle Creek, MI.
EPI = Epilithon, EPS = Epipsammon, FBOM = Fine Benthic Organic Matter, SPOM =
Suspended Particulate Organic Matter, SPH STR = Sfinhaerium striatinum. EPS and
SPOM are presumed to be the potential food resources for bivalves. Data for food
resources are means for 8 collection dates at the reference station during the course of the
enrichment experiment. Gland and muscle are isolated tissues from Pleurobema sintoxia
+ Fusconaia m. The datum for gahaerium striatinum is a composite of the whole soft
tissue of several individuals. Error bars are d: 1 standard deviation ......................... 54

Figure 2. Enrichment of unionid tissues over the course of the 15N tracer experiment,

with 815N values corrected for background 815N. Error bars are i 1 standard
deviation ............................................................................................ 56

Figure 3. A) Enrichment of potential food resources of bivalves over the course of the

15N tracer experiment, with 615N values corrected for background 615N. The last
SPOM datum was calculated by extrapolating the decay rate. Error bars are i 1 standard

deviation. B) Calculated 15N enrichment of dissolved NH4+ during the experiment, at the

15N H 4+ was
assimilated or nitrified, and at the main bivalve sampling site the N of N H 4+ was 70-

75% that of the dripper (Hamilton et al. 2001) ................................................ 58

dripper. The enrichment decreased across the study reach as the
51 5

Figure 4. Comparison of 15N enrichment in unionid stomach glands and gut contents on

day 42 (515N not corrected for background levels of enrichment). PLE + FUS =
Pleurobemg sintoxia + Fusconaia m; ELL DIL = Elliptic dilitata; STR UND =
Strophitus undulatus; LAM CAR = Lampsilis cardium; VEN ELL = Venustaconcha
ellipsiformis; ACT L1G = Actinonais ligamentina; LAS COS = Lasmigona costata; VIL
IRI = Villosa m ................................................................................... 60

 

Figure 5. Longitudinal patterns of 15N enrichment for unionid stomach-gland samples,
Sphaerium striatinum whole-body samples, and potential food resources across the study

reach on day 42 of the 15

815

N enrichment experiment (values corrected for background

N). Error bars are d: 1 standard deviation ................................................. 62

CHAPTER 3

Figure 1. A) Total macrozooplankton biomass response over the course of the experiment
in the presence of zebra mussels and larval bluegill. Shown are means i: 1 standard
deviation. B) Total micr0200plankton abundance response over the course of the
experiment in the presence of zebra mussels and larval bluegill. Shown are means i 1
standard deviation ................................................................................... 84

Figure 2. A) Chlorophyll-a response over the course of the experiment in the presence of
zebra mussels and larval bluegill. Shown are means :I: 1 standard deviation. B)
Transparency response over the course of the experiment in the presence of zebra
mussels and larval bluegill. Transparency decreases with increasing coefficient of light
extinction. Shown are means i: 1 standard deviation .......................................... 86

CHAPTER 4

Figure 1. Nutrient responses over the course of the experiment in the presence of zebra
mussels and larval bluegill. Shown are means i: 1 standard deviation. A) Sestonic carbon.
B) Sestonic nitrogen............ ................................................................... 107

Figure 2. A) D1N+sestonic N:Tota1 P molar ratio, uncontaminated replicates. B)
D1N+sestonic N:Tota1 P molar ratio, all days. Some replicates on days 6, 9 and 13 were
slightly underestimated due to phosphorus contamination and were excluded from
statistical analysis (see text) ...................................................................... 109

Figure 3. Sedimentation response over the course of the experiment in the presence of
zebra mussels and larval bluegill. In both graphs the sampling dates represent
sedimentation over the preceding time interval. Shown are means :1: 1 standard deviation.
A) Sedimentation rate. B) Sedimentation rate of chlorophyll a .............................. 111

Figure 4. Nutrient ratio in sediment response over the course of the experiment in the
presence of zebra mussels and larval bluegill. Shown are means 3: 1 standard deviation.
A) N:P molar ratio of sediment. B) C:N molar ratio of sediment. C) C:Chlorophyll a mass
ratio of sediment ................................................................................... 113

CHAPTER 5

Figure l. Lakes surveyed in the lower peninsula of the State of Michigan. Black dots
indicate lakes invaded by zebra mussels. Grey dots indicate lakes free of zebra
mussels at the time of surveying ................................................................ 144

Figure 2. Total phosphorus (TP) concentrations in lakes surveyed in 1998 and 1999. Gray
bars are lakes without zebra mussels at the time of the survey and black bars are lakes
with zebra mussels ................................................................................ 146

xi

Figure 3. Effect of zebra mussels on phytoplankton abundance and transparency in inland
lakes of Michigan. Note log scales. Open circles and dashed lines represent lakes without
zebra mussels, closed circles and solid lines represent lakes with zebra mussels. A)
Chlorophyll a concentrations (n = 61). Lakes without zebra mussels: y = 1.148x — 0.762,
B = 0.0003; lakes with zebra mussels: y = 1.407x — 1.352, E < 0.0001. B) Total algal
biovolume (n = 40). Lakes without zebra mussels: y = 2.329x + 3.161, B = 0.0021; lakes
with zebra mussels: y = 2.554x + 2.361, E = 0.0142. C) Secchi depth (n = 61). Lakes
without zebra mussels: y = -0.925 + 1.681, I: < 0.0001; lakes with zebra mussels: y = -
0.893x + 1.693, P < 0.0001 ....................................................................... 148

Figure 4. Effect of zebra mussels on Microcystis relative abundance in inland lakes of
Michigan. Note log scale. Open circles represent lakes without zebra mussels, closed
circles represent lakes with zebra mussels ...................................................... 152

Figure 5. Effect of zebra mussels on the relative abundance of Microcystis in lakes
grouped into low TP (< 25 ug/L) and high TP (> 25 ug/L) classes (Two-way AN OVA,
ZM, F = 5.42, E = 0.0255). Error bars are standard errors ................................... 154

Figure 6. Simple Linear Discriminant Analysis of lakes with and without zebra mussels
using log-transformed biovolumes at the genus level (eigenvalue = 4.323, canonical
correlation = 0.901) ............................................................................... 156

Figure 7. Efi‘ect of zebra mussels on Dissolved Organic Carbon (DOC) concentrations in
lakes surveyed in 1999. Open circles and dashed line represent lakes without zebra
mussels. Closed circles represent lakes with zebra mussels. Note log scale. Lakes without
zebra mussels: y = 0.508x + 0.228, B = 0.0008. Lakes with zebra mussels: B =

0.5086 ............................................................................................... 158
CHAPTER 6

Figure 1. Zebra mussel mortality during the experiment ..................................... 172
APPENDICES

Figure Al. Design of limnocorral frames ...................................................... 180
Figure A2: Location of some bluegill nest colonies in local lakes .......................... 187
Figure A3. Design of brine shrimp rearing tank .............................................. 189

xii

Chapter 1
INTRODUCTION: BIOLOGICAL INVASIONS, ZEBRA MUSSELS, AND THE
REPLACEMENT OF NATIVE UNIONIDS

Overview

One of the most profound scientific advances of the twentieth century was the
recognition that the biosphere is increasingly a human-dominated ecosystem. While
ozone depletion, habitat destruction, chemical pollution, large-animal extinction and
serious risk of global climate change are perhaps the most well known environmental
outcomes of this domination, numerous others abound. One such issue is biological
invasion, defined as the introduction of a species to an area previously uninhabited by
that species. Man is directly and indirectly responsible for thousands of such
introductions. Some of those introduced species become established as viable
populations, and some of those viable populations bloom into ecological and economic
pests. The general concept of invasive species is perhaps less well known than specific
case studies, but represents a problem as old as global human migration itself.

People have been moving other species around the planet since prehistoric times
(Elton 1958, Williamson 1996). This “aboriginal phase”, as I call it, consisted of
transporting animals and plants as food necessary for survival. Polynesians, for example,
intentionally brought pigs and rats with them to hunt as they colonized Pacific islands.
Meso-americans brought plants with them as they migrated across North America. A
later “European colonial phase” occurred when explorers of the Pacific intentionally

released goats and pigs to islands so future sailors would have something to eat and

 

unintentionally introduced animals including rats and mosquitoes. Plants introduced to
the New World often spread faster than people did, such that later colonizers thought the
plants were native. In the 19th century a “romantic phase” of species introduction
occurred through the work of acclimatization societies established for the purpose of
species introduction. Their motivations were both quaint and practical, spanning the
desire to introduce all the birds mentioned by Shakespeare to wanting something familiar
to hunt. Today we live in the “modern economic phase” of species introductions, where
organisms are transported and released unintentionally through the activities of global
commerce such as air traffic, commercial container traffic, and ballast water exchange
(V itousek et al. 1996).

One of the most famous biological invasions that has occurred in recent times the
arrival of the zebra mussel (Dreissena polmorpha) to North America. The first
discovery of the zebra mussel in North America occurred in 1988, when Herbert et al.
(1989) found large populations in Lake St. Clair and western Lake Erie. A release of
larval mussels during the ballast exchange of a single commercial cargo ship travelling
from the north shore of the Black Sea has been deduced as the likely vector of
introduction (McMahon 1996). The zebra mussel has since dispersed throughout the
Great Lakes, Mississippi River drainage network and other eastern rivers of North
America, and inland lakes including those of Michigan.

Zebra mussels are famous for their biofouling effects on industrial sites (Mackie et

al. 1989, Kovalak et al. 1993). Water intakes, heat exchangers, holding tanks and other

structures of hydroelectric and other plants have been dramatically clogged, producing

densities as high as 750,000 ind/m2. As of the mid 1990’s, industrial mitigation of zebra

mussels cost over $100 million annually in the United States (ZMIS 1996). Biological
invasion differs from many environmental problems in that direct monetary costs can be
obvious and large.

This dissertation examined several effects of zebra mussels on freshwater
ecosystems. Among the impacts zebra mussels have had is the near extirpation of native
unionid mussels. Thus the invasion of zebra mussels to North America is not simply the
addition of a new species, but the functional replacement of one type of fieshwater
mussel with another. By also examining the ecology of native unionid bivalves in this
dissertation I sought to address what North American freshwater ecosystems are gaining
and losing. Much less is known about the basic ecology of unionids than zebra mussels
(see chapter 2 for a review of the state of knowledge concerning unionids). Thus while
this dissertation advances our knowledge of both native and exotic freshwater mussels,
my contribution to the understanding of unionids represents information of a much more
fundamental nature. The rest of this introduction describes some of the basic ecology of
zebra mussels and unionids, impacts that zebra mussels have had on previously invaded

systems, and further motivations for studying these topics.

Life history and physiology

The life history of zebra mussels differs greatly from most endemic Great-Lakes
region bivalves (Pennak 1989, Mackie and Schlosser 1996). Exotic dreissenids are
dioecious, with fertilization occurring in the water column. Endemic bivalves are
mOnoecious, dioecious or herrnaphroditic, some internally fertilized by filtering sperm

fi'om the water column. Under natural thermal regimes, zebra mussel oogenesis occurs in

 

 

 

autumn, with eggs developing until release and fertilization in spring. In thermally
polluted areas, reproduction can occur continually through the year.

The quagga mussel (Dreissena W), a second invader named after an extinct
species of zebra, was discovered in 1991. Large populations of this mussel exist in Lake
Erie and Lake Ontario, and invasion of other areas such as the Mississippi River has
begun (Mills et al. 1996). One endemic genera of dreissenidae exists, Mflilopsis, which
occupies coastal estuaries from New York to the Gulf of Mexico (Pathy and Mackie
1993). It is not known whether Mflilopsis occurs in North America due to early
European introduction (Morton 1993). The false mussel, Mfiilopsis leucophaeata, has
been found in the upper Mississippi, and may pose a similar threat as does the zebra
mussel (Mackie et al. 1989).

Most endemic freshwater bivalves have an obligatory parasitic larval stage called
a glochidium which temporarily infests the gills of fish. The larval stage of Dreissenidae
is a free-swimming pelagic veliger that does not rely on any host. The only other North
American freshwater bivalves with pelagic veliger larvae are fingernail clams,
Sphaeridae, which burrow into sediment, and the Asiatic clam, Corbicula, another exotic
invader which favors running waters causing its own host of ecological and industrial
problems (Morton 1997). The zebra mussel displays high fecundity, annually producing
30,000-40,000 eggs. Veligers typically begin to appear in the water column in June at
13°C, and remain pelagic for 8-15 days. Veligers can occupy the water column for the

entire summer and even into winter depending on differential gamete production and

release.

Once the veliger undergoes morphological changes including development of the
siphon, foot, organ systems and blood, it is known as a postveliger. Further subdivision of
the larval stage has been delineated: (veliger) preshell, straight-hinged, umbonal,
(postveliger) pediveliger, plantigrade, and (juvenile) settling stage (ZMIS 1996). The
settling stage attaches to a substrate via protienaceous threads secreted from the byssal
gland. The vast majority of veliger mortality (99%) occurs at this stage due to settlement
onto unsuitable substrates. Sensitivity to changes in temperature and oxygen are also
greatest at this stage. Once attached, the life span of g polmorpha is variable, but can
range fi'om 3-9 years. Maximum growth rates can reach 0.5 mm/day and 1.5-2.0 cm/year.
Adults are sexually mature at 8-9 mm in shell length (i.e. within one year).

The rapid invasion of North American waterways has been facilitated by the zebra
mussel’s ability to disperse during all life stages. Passive drift of large numbers of pelagic
larval veligers allows invasion downstream. Yearlings are able to detach and drift for
short distances. Adults routinely attach to boat hulls and floating objects and are thus
anthropogenically transported to new locations. Transporting recreational boats disperses
zebra mussels between inland lakes. In addition, speculation exists that waterfowl can
disperse zebra mussels, but this has yet to be conclusively demonstrated.

While byssal threads develop in the larvae of some non-dresissenid endemic
bivalves and are used to attach to fish gills, there are no endemic freshwater bivalves with
byssal adult stages. This adaptation has been important to the zebra mussel’s success in

invading North America. Zebra mussels attach to any stable substrate in the water column
or benthos: rock, macrophytes, artificial surfaces (cement, steel, rope, etc.), crayfish,

unionid clams, and each other, forming dense colonies called druses.

 

Long-term stability of substrate affects population density and age distributions
on those substrates. Within Polish lakes, perennial plants maintained larger populations
than did annuals (Stanczykowska and Lewandowski 1993). Populations on plants also
were dominated by mussels less than a year old, as compared with benthic p0pulations.
These populations of small individuals allow higher densities on plants. In areas where

hard substrates are lacking, such as a mud or sand, zebra mussels cluster on any hard
surface available. Given a choice of hard substrates, zebra mussels do not show a
preference, indicating that veligers cannot discriminate between substrates (with the
exception of substrate rejection due to contaminants). Research on danish lakes shows
that factors exist, however, that cause substrate to be unsuitable for both initial and long
term colonization: extensive siltation, some sessile benthic macroinvertebrates,
macroalgae, and fluctuating water levels exposing mussels to desiccation (Smit et al.
1993)

The dispersion of zebra mussels within a lake is controlled by physical conditions
including wind strength, lake/ shore morphometry, and current patterns (Stanczykowska
and Lewandowski 1993). These conditions affect both spatial patterns of pelagic veliger
density and benthic adult dispersion. Population density of benthic adults has been

observed to vary as widely as two orders of magnitude (e. g. <100 to >1500

individuals/m2) within individual Polish lakes due to these physical conditions.
Tolerance limits of physical and chemical parameters are well known (Sprung
I993, Vinogradov et al. 1993, McMahon 1996). Although discrepancy exists when
comparing temperature tolerance limits of North American and European populations,

this is probably due to the American population being founded by mussels from the

 

southern limit of the European population’s range. Most work in Europe has been done in
the northern range.

North American populations are generally adapted to warmer temperature
regimes than their European counterparts. Although shell growth has been reported to
occur at temperatures as low as 3° C, Lake St. Clair populations and some European
populations display shell growth at 6-8° C. Eggs are released when the environmental
temperature reaches 13° C and release rate is maximized over 17° C. The optimal
temperature range for adults extends to 20-25° C, but _I_)_. polmomha can persist in
temperatures up to 30° C. Short term tolerance of temperatures up to 35° C is possible if
the mussels were previously acclimated to high temperatures. Rapid warming of shallow
lakes has been hypothesized to detrimentally affect reproductive rates in Danish
populations (Smit et al. 1993).

Oxygen demands are similar to those of other freshwater bivalves including
unionids. Tolerance of “anaerobic” conditions has been reported for short time periods
under certain temperatures and sizes, but zebra mussels cannot persist in hypoxic

conditions. The lower limit of p02 tolerance is 32-40 Torr at 25° C. Zebra mussels have

been found in the hypolimnetic zone of lakes with oxygen levels of 01-112 mg/l, and in
the epilimnetic zone with oxygen levels of 4.2-13.3 mg/l. Zebra mussels are described as

poor 02 regulators, possibly explaining their low success rate in colonizing eutrophic

lakes and the hypolimnion.
Zebra mussels can tolerate only slight salinity. Although some populations of
EurOpean zebra mussels can be found in estuaries, their persistence has been

Speculatively attributed to reduced tidal fluctuation. Upper limits of freshwater bivalve

 

 

salinity tolerance reach 8-10 960, and populations of European zebra mussels have been
found to tolerate and range of salinities, from 0.6 %o (Rhine River) to 10.2 %o (Caspian

‘ Sea). North American populations generally tolerate salinity up to 4 %o.

In European populations, calcium concentrations of 24 mg Ca2+/l allow only
10% larval survival due to inhibition of shell development. Optimal calcium

2

concentrations ranges from 40-55 mg Ca +/l, but North American populations have been

found in lakes with lower concentrations. North American populations require 10 mg

Ca2+/l to initiate shell growth and 25 Ca2+ mg/l to maintain shell growth. Larval

development is inhibited at pH < 7.4. Higher rates of adult survival occur at a pH of 7.0-
7.5, but populations have been found in the hypolimnetic zone of lakes with a pH of 6.6-
8.0, and in the epilimnetic zone with a pH of 7.7-8.5. Optimal larval survival occurs at a

pH of 8.4, and optimal adult grth occurs at pH 7.4-8.0.

Ecological Impacts Observed and Hypothesized

Dreissenids are gregarious, epibenthic, byssal filter feeders. As such, zebra
mussels primarily consume phytoplankton, but other suspended material is filtered from
the water column including bacteria, protozoans, zebra mussel veligers, other
microzooplankton and silt. Large populations of zebra mussels in the Great Lakes and

Hudson River reduced the biomass of phyt0plankton significantly following invasion.
Diatom abundance declined 82-91% and transparency as measured by Secchi depth
increased by 100% during the first years of the invasion in Lake Erie (Holland 1993). As

the invasion spread eastward during 1988 to 1990, successive sampling stations recorded

declines in total algae abundance from 90% at the most western station to 62% at the
most eastern (N icholls and Hopkins 1993). In Saginaw Bay, sampling stations with high
zebra mussel populations experienced a 60-7 0% drop in chlorophyll-a and doubling of
Secchi depth (F ahnenstiel et al. 1993). Phytoplankton biomass declined 85% following
mussel invasion in the Hudson River (Caraco et al. 1997). Whether the species
composition of the phytoplankton community can be altered by zebra mussels in
unresolved (see chapter 4).

Increased water clarity allows light to penetrate further, potentially promoting
macrophyte populations. As macrophytes can be colonized by veligers, the macrophyte
community may be altered if such colonization proves detrimental. Indeed, this may be
occurring in Spring Lake, 0 ttowa County Michigan, USA with Vallisneria colonized to
a lesser extent than other plants (Theresa Lauber personal communication). Increased
light penetration may also cause water temperatures to rise and thennoclines to become
deeper, but these effects have not yet been documented. As Chl-a drops as a result of the
phytoplankton consumption, the DOC concentration may drop. Macrophytes could
eventually compensate for this since they are also a source of DOC, but there may be a
lag period between the time when phytoplankton biomass is down and macrophytes
proliferate. This could produce a period of time when UV-B light penetrates deeper into
the water column.

Zebra mussels are able to filter particles smaller than lum in diameter, although

they preferentially select larger particles (Sprung and Rose 1988) . Thus bacteria may
1' epl'esent an important food source (Cotner et al. 1995, Silverman et al. 1996). At a 90%

efficiency rate, zebra mussels are much more efficient at filtration of such small particles

than are Unionids and Asiatic clams. Filtering rate is highly variable, depending on
temperature, concentration of suspended matter, phytoplankton abundance, and mussel
size (reviewed by Noordhuis et al. 1992). Although European zebra mussels are less
active in winter, this seasonal pattern is temperature driven. No diel patterns of filtration
rate have been found. During spring, filtration rates rise dramatically between 5 and 10°
C, then level off with respect to temperature, and may be inhibited at temperatures over
20° C. Increased suspended matter can reduce filtration activity to a minimum required to
maintain oxygen demand. A sigrrroidal relationship exists with filtration rate and size, but
this may be an affect of aging.

Material filtered by zebra mussels is either ingested or expelled as feces or
pseudofeces, which is mucus covered. True fecal pellets are chemically altered, larger
and more dense. Pseudofeces production increases with increasing suspended solid
concentration, as well as increasing temperature, albeit to a much lesser extent

(Noordhuis et al 1992, MacIsaac and Rocha 1995). The rate of biosedimentation through
pseudofeces production was very high (28mg/cm2 day at a density of 1180

individuals/m2) under turbid conditions in Lake Erie, lending support to the hypothesis
that zebra mussels are responsible for increased water clarity observed since mussel
introduction (Klerks et al. 1996). Filtration rate was not related to seston composition
(POC2TSS, chl:TSS) in Saginaw Bay (Fanslow et al. 1995).

Veligers also filter material, but their impact is far less than that of sessile adults.
Settled mussels exerted 103 times the grazing rate of veligers in western Lake Erie, for

example (MacIsaac et al. 1992). Microzooplankton, e. g. rotifers and veligers, are ingested

10

by zebra mussels, but larger zooplankton are not eaten (MacIsaac et al. 1991, MacIsaac et
al. 1995).

It has been speculated that benthic deposition of feces and pseudofeces may aid
bacterial productivity, thus producing a source culture that zebra mussels can feed upon
(Silverman et al. 1996). It has also been speculated that biodeposition of feces and
pseudofeces might cause observed increases in benthic macroinvertebrate populations
(Stewart and Haynes 1994). Biomagnification of PCBs was observed in Gammarus
associated with zebra mussels, indicating concentration of pollutants in zebra mussel
feces/pseudofeces and transfer to other trophic levels (Bruner et al. 1994). In an
experimental study, however, Botts et al. (1996) found greater abundances of
macroinvertebrates associated with both living and non-living (i.e. empty shell) zebra
mussel druses compared with their no~druse treatment. Thus the increased physical
habitat complexity of a mussel colony may benefit macro-invertebrates rather than
deposition of feces and pseudofeces.

Zebra mussels can reduce filtration rates (more frequent interruption of filtering
or slower pumping rates) and/or produce pseudofeces above an incipient limiting
concentration (ILC) of algae to maintain a constant consumption rate (Sprung and Rose
1988, Fanslow et al. 1995 MacMahon 1996,). Feeding activity can be described by the
clearance rate (percentage of algal biomass removed from the water column over time),
biomass of cleared algae (BCA), feces production and pseudofeces production (ug F or
P/BCA). For example, Berg et al. (1996) examined the effects of zebra mussel size and
algae species and concentration on zebra mussel feeding activity. Clearance rates were

constant over varying concentrations of pure cultures of Chlamydomonas reinhardtii, a

ll

gphc‘
tor
box"
col
T:

zel

spherical unicellular species of 7.42 pm (: 0.13pm) in diameter. This indicates that the
concentrations used in experiments were below the ILC. Clearance rates decreased,
however, with increasing concentrations of Pandorina morum. a species made up of
colonies with varying numbers of cells that are individually as large as _C_. reinhardtii.
This indicates that the concentrations used in experiments were above the ILC. Large
zebra mussels (20-25 mm in length) displayed a higher clearance rate across all
concentrations of _C_. reinhardtii than did small mussels (IO-15 mm). ILC differed in this
study fi‘om previous studies done with European populations. Thus zebra mussel size,
phytoplankton species and regional population differences affect clearance rates, ILC and
feces/pseudofeces production.

Zebra mussels produce pseudofeces to avoid ingesting non-food material (e. g.
clay), as a mechanism to deal with overabundance of food (e. g. algal concentrations
above the ILC), and possibly as a way to reject unpalatable algae. Zebra mussels readily
reject blue-green algae, such as Microcystis, as pseudofeces (V anderploeg et al. 2001).
The presence of this cyanobacterium does not inhibit filtering, except in mass abundances
such as a bloom (N oorhuis et al. 1992, Lavrentyev et al. 1995). Zebra mussels can select
material for rejection through pseudofeces production internally, perhaps identifying
cyanobacteria by chemical cues (Ten Winkel and Davids 1982).

Understanding of the fate of pseudofeces once it expelled is poor. Zebra mussels
removed metals from the water column of Lake Erie and deposited it to the bottom at
high rates (Klerks et al. 1997). Pseudofeces production is a key feature in the still
experimental use of zebra mussels as biofilters of undesired pollutants, including

mitigation of cyanobacterial blooms (N oordhuis et al. 1992, Reeders and Bij de Vaate

12

1992). Roditi et al. (1997) found that the biodeposits of zebra mussel were organically
enriched, including 3.9% live algae by weight. Resuspension of this material occurred in
their system, a tidal estuary, reducing the potential impact of biodeposition to the
benthos.

Less well known is the fate of live algae bound into pseudofeces. Bastviken et al.
(1998) speculate that phytoplankton that survives the pseudofeces process must be
resuspended in order for long term survival, a process less likely to occur in inland lakes
than in tidal estuaries. If survivorship following filtration is equal between phytoplankton
species then community species composition can remain unchanged. Other factors may
affect the phytoplankton community, however, including increased light.

The zooplankton community has also been affected by the invasion of zebra
mussels (see also chapters 3, 5). Zooplankton abundance dropped 55-71% following
mussel invasion in Lake Erie, with microzooplankton more heavily impacted (MacIsaac
et al. 1995). Mean summer biomass of zooplankton decreased from 130 to 78 mg dry wt.

rrr'3 between 1991 and 1992 in the inner portion of Saginaw Bay. The total biomass of
zmplankton in the Hudson River declined 70% following mussel invasion, due both to a
reduction in large zooplankton body size and reduction in microzooplankton abundance.
These effects can be attributed to reduction of available food (phytoplankton) and direct
predation on microzooplankton. Increased competition in the zooplankton community for
newly limited food should result from zebra mussel infestation. The size of individual

zooplankters might decrease. Hypotheses can be formulated specifying which species

will prevail based on knowledge of competitive ability.

l3

Effects should continue through the food web to fish (see chapter 3). Reductions
in zooplankton biomass should cause increased competition, decreased survival and
decreased biomass of planktivorous fish. Alternatively, because microzooplankton are
more heavily impacted by zebra mussels the larval fish population may be more greatly
affected than later life stages. This may be especially important to inland lakes with
populations of pelagic larval fish such as bluegills. Benthic feeding fish may benefit as
opposed to planktivorous fish, or behavioral shifts from pelagic to benthic-feeding may

occur. In addition, proliferation of macrophytes may alter fish habitat.

 

Other effects include the extirpation of native unionid clams through epizootic
colonization (Schloesser et al. 1996, Baker and Hombach 1997). Zebra mussels restrict
valve operation, cause shell deformity, smother siphons, compete for food, impair
movement and deposit metabolic waste onto unionid clams. To date, unionids have been
extirpated from Lake St. Clair and nearly so in western Lake Erie. Many species of bird
known to be predators of zebra mussels also occur in the inland lakes of Michigan. While
a new food source may benefit such predators, biomagnification of toxins into both fish
and birds is possible.

Some effects have been hypothesized as worst—case scenarios. For example, zebra
mussels may cause a shift from pelagically to benthically-based food webs in inland
lakes. Zebra mussels may also shifl lakes from a turbid and phytoplankton-dominated
state to clear and macrophyte-dominated state, i.e. between alternative stable equilibria

(Scheffer et al. 1993).

14

Past Research Efforts

A long tradition of zebra mussel study exists in Europe and the former Soviet Union,
where the zebra mussel has been present for 150 years (see Mackie et al. 1989 for an
annotated bibliography of European references). Work includes spatial distribution
patterns, demography, tolerance limits for physical and chemical parameters, and
physiology. Studies of food web interactions and nutrient regeneration are conspicuously
lacking. Extensive ecological work in the United States began soon as the zebra mussel
was discovered and peaked in the early 1990’s. The literature on ecosystem and
community-level effects of zebra mussels consists primarily of work investigating Lake
Erie, Saginaw Bay, the Hudson River, and Oneida Lake (e.g. Fahnenstiel 1993, Holland

1993, Pace et al. 1998, Idrisi et al., 2001).

Further motivations for this dissertation

Comprehensive investigation of the effects of zebra mussels on North American
inland lakes is currently lacking. Despite the quantification of ecosystem effects in the
Great Lakes and Hudson River ecosystems, predicted effects of zebra mussels cannot be
simply transplanted to inland lakes. Important differences between the Great Lakes and
Hudson River exist that should affect ecosystem response. For example, in both Lake
Erie and Saginaw bay, ecosystem effects caused by the rapid expansion of zebra mussel
populations were localized. In Saginaw Bay, the outer bay with low zebra mussel
populations behaved independently of the inner bay with high zebra mussel populations

(Fahnenstiel et al. 1993). In lake Eric, the eastward expansion of the invasion could be

15

 

charted in the successive reduction of phytoplankton at stations along a transect (N icholls
and Hopkins 1993). Due to their sheer size, the Great Lakes do not respond as whole
units to ecosystem perturbation. Inland lakes are much smaller and may respond to zebra
mussel invasion both more quickly and as whole units. The morphology of inland lakes,
with a larger proportion of littoral zone, should affect ecosystem response, as should the
greater macrophyte populations relative to lake size.

Lastly, Native freshwater bivalves were originally ubiquitous in North American
rivers and streams, but have increasingly fallen victim to anthropogenic pressures such as
over-harvesting, waterway impoundment, pollution and exotic species infestation
(Ricciardi and Rasmussen 1999). Bivalve conservation efforts have included
translocation, but mortality rates have averaged ~50% (Cope and Waller 1995). Captive
rearing programs show highly variable mussel survival rates (Dunn and Layzer 1997).
Incomplete understanding of the feeding ecology of freshwater bivalves impedes
successful conservation efforts. Thus investigation of the basic ecology of these

organisms is sorely needed.

16

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Klerks, P. L., P. C. Fraleigh, and J. E. Lawniczak, 1997, Effects of the exotic zebra
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18

 

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19

Nichols, K. H., and G. J. Hopkins, 1993, Recent changes in Lake Erie (north shore)
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Zebra Mussel Dreissena polmomha, Gustav Fischer Verlag, New York.

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20

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21

Chapter 2

BIVALVE DIETS IN A MIDWESTERN U.S. STREAM: A STABLE ISOTOPE
ENRICHMENT STUDY

with Stephen K. Hamilton

This chapter was published as: David F. Raikow and Stephen K. Hamilton, 2001, Bivalve
diets in a Midwestern U.S. stream: A stable isotope enrichment study, Limnology and

Oceanography 46: 514-522.

Abstract

This study examined a community of stream bivalves (unionids and fingemail clams) in a
second-order woodland stream in southern Michigan, USA using both the natural

abundance of 15N and a six-week whole-stream 15

N enrichment experiment, as part of
the Lotic Intersite Nitrogen eXperiment (LINX). Objectives included addressing what
made up the diet of these bivalves and whether suspended algae consumed by bivalves
were derived fi‘om pelagic phytoplankton imported from an upstream lake or attached
algae sloughed from instream surfaces. Within the examination of bivalve diets we

considered whether suspension and/or deposit feeding modes were employed and

whether bivalves selectively assimilated the algal and microbial portions of bulk material

they ingested. All 12 unionid species reached a level of 15N enrichment greater than the
bulk suspended organic matter. Sphaerium striatinum (Sphaeriidae) were enriched to

levels greater than all presumed food sources. Suspended algae were derived both from

22

sloughed epilithon and pelagic phytoplankton originating from lentic waters upstream. A
mixing model suggested that unionids were consuming 80% deposited and 20%
suspended material. Alternatively, these bivalves were preferentially assimilating the
highly enriched living component of suspended and/or benthic organic matter rather than
assimilating the bulk material. These results advance our understanding of freshwater
bivalve feeding ecology, which is necessary if conservation efforts of these increasingly

threatened organisms are to succeed.
Introduction

Freshwater bivalves were originally ubiquitous in North American rivers and
streams, but have increasingly fallen victim to anthropogenic pressures such as over-
harvesting, waterway impoundment, pollution and exotic species infestation (Ricciardi
and Rasmussen 1999). Bivalve conservation efforts have included translocation, but
mortality rates have averaged ~50% (Cope and Waller 1995). Captive rearing programs
show highly variable mussel survival rates (Dunn and Layzer 1997). Incomplete
understanding of the feeding ecology of freshwater bivalves impedes successful
conservation efforts.

There are several ways to better understand the diet of organisms including direct
observation of feeding behavior, gut contents analysis, and examination of chemical
constituents such as stable isotopes or nutrients within the tissues of organisms compared
with their potential food sources. Direct observation of the feeding behavior of bivalves

has primarily involved either measurement of filtering rates of artificial seston by adults

23

(e.g. Kryger and Riisgard 1988; Silverman et al. 1997) or examination of filtering
morphology (e.g. Ward et al. 1993). Gut contents analyses of freshwater bivalves are rare
and may be misleading because the method cannot distinguish ingested material that is
not assimilated. Algae, for example, can survive passage through the digestive tract, and
fecal material can have a high nitrogen content due to ingested material by-passing the
stomach gland and moving directly to the intestinal tract (Hawkins et al. 1983; Miura and
Yamashiro 1990). Also rare are energy budgets for freshwater bivalves. An energy
budget for the fingernail clam Sphaerium striatinum was reported by Hornbach et al.
(1984), who estimated that 35% of its energy was derived from suspension feeding, and
the rest possibly from deposit feeding.

Most stable isotope studies of food webs can be classed as either surveys of

natural isotope abundance or experimental isotope enrichments. Natural abundances of

stable carbon and nitrogen isotope ratios (513C and 815N) have increasingly been used to
examine the relative importance of various potential autotrophic sources in supporting

food webs that include mollusks (Incze et a1. 1982; Peterson et al. 1985; Thorp et al.

1998). Consumers tend to be enriched with the heavier stable isotope of nitrogen (1 5N)
relative to their diet (Minagawa and Wada 1984). Accounting for this trophic enrichment,
unionids have been used to indicate the isotopic composition of the base of the food web
in studies of lakes because they live long, have slow metabolism, and presumably utilize
primary producers (phytoplankton) as their dominant food source. Unionids should thus
integrate potential short-term fluctuations in primary producer isotope signatures over

time (Cabana and Rasmussen 1996). This concept has been used to study the trophic

24

position of fish in Canadian lakes and effects of exotic species on food webs (Vander
Zanden et al. 1997; Vander Zanden et al. 1999).
In an isotope enrichment experiment, the abundance of the normally rare isotope

is increased in the ecosystem by the introduction of a nutrient enriched in the rare isotope,

allowing a more detailed analysis of the food web. Tracer 15N experiments have been
conducted in the Kuparuk River, AK, USA, Walker Branch, TN, USA and Hugh White
Creek, NC, USA to elucidate food web relationships in streams (Peterson et al. 1997;
Hall et al. 1998; Mulholland et al. 2000a, b). The present study and the latter two cited
above were part of the Lotic Intersite Nitrogen eXperiment (LINX), a multi-site
examination of nitrogen cycling within stream ecosystems across North America. The
study reported here is the first experimental enrichment of stable isotopes in an
ecosystem containing a high abundance and diversity of freshwater bivalves.

We examined two questions of bivalve feeding ecology: 1) What is the diet of
stream bivalves? and 2) To what extent is suspended material available to filter feeders
derived from benthic algae suspended within the stream or from a subsidy of pelagic
phytoplankton derived from upstream? Within the context of bivalve diets we considered
whether bivalves employed deposit and/or suspension feeding modes and whether
bivalves selectively assimilated the algal and microbial portions of bulk material
ingested. To evaluate the potential food sources for bivalves in a woodland stream of the

Midwestern US, we examined unionids, sphaeriids, and their potential food sources

using both natural N isotopic abundances and a nitrogen isotope (1 5N) enrichment

experiment.

25

Methods

Slim _S_i_t§- This study examines a 500m reach of Eagle Creek, Fort Custer State
Recreation Area, Michigan, USA near the city of Battle Creek. Originating as an outflow
from an artificial impoundment 1200m upstream known as Eagle Lake, the stream flows
through secondary forest and wetland upstream of the study reach, where there are
several beaver dams. Below the study reach the stream enters the Kalamazoo River. The
study reach has little wetland and is largely shaded by deciduous forest. Hydraulic and

geomorphological features are summarized in Table l.

_I_SN Tracer Experimen - Experimental design and sampling methods that were
standardized among LINX sites are cursorily described here; details are provided in
Mulholland et al. (20000, b) and Hamilton et al. (2001). Isotope measurements are

expressed as 6 values in units of %o :

615N= [(R /R (1)-11x 1000 (1)

sample standar

where R = 15N: 14N ratio and R stan dar d (N2 in air) = 0.003663. Ammonium chloride

enriched with 15N was dripped into the stream for 42 days beginning on 16 June 1998

815

with the intent of raising the N of ammonium at least +250 %o above natural

515

background levels. Enrichment experiment data were corrected for background N by

subtracting natural delta values measured either just before the experiment or during the

experiment at an upstream reference site (see below).

26

Sampling- Sampling stations were established at 7 locations downstream of the dripper as
well as at a reference reach just above the dripper. Potential food sources for bivalves
included epipsammon (EPS, or detritus and possibly algae mixed with sand) and
suspended particulate organic matter (SPOM). EPS was collected using a turkey baster to
suck organic matter from the surface of sand deposits, and EPS samples likely contained
a substantial portion of mobile fine benthic organic matter (FBOM). SPOM was sampled
by filtering water pumped from the middle of the water column. Samples of EPS and
SPOM were collected on Whatrnan GF/F filters (nominal pore size = 0.7 pm). Additional
ecosystem compartments that served as primary food sources for other organisms in the
stream included Coarse Particulate Organic Matter (CPOM, collected as leaf fragments),
small pieces of rotten wood, and Epilithon (EPI, scraped from rocks and collected on
GF/F filters). Filamentous macroalgae and macrophytes were present at very low
abundances and were thus unlikely to be significant to food webs. Composite samples
were made from several locations within sampling stations. Samples for natural isotope
abundance were obtained from the reference station. Samples were collected weekly
beginning on 15 June 1998 at all or at selected sample stations, including food sources
and various macroinvertebrates. Bivalves were sampled under a different schedule.

To monitor temporal patterns of 15

N enrichment in bivalves, we collected three
individuals of Pleurobema sintoxia from a single station 396 m downstream from the

dripper on the seventh day of the experiment (day 7), as well as on days 14, 28, 35, and

42. We also collected 3 individuals of _13. sintoxia from the same station for temporal

 

analysis of 15N turnover rate the day after the dripper was turned off (post-day l), as well

as on post-days 7, l4, and 28. We studied spatial patterns of unionid isotopic enrichment

27

by collecting 3 B. sintoxia from each station on the day of maximum enrichment (day 42)

 

and spatial patterns of Sphaerium striatinum enrichment by collecting several individuals
from various stations on day 42. We collected unionids of other species from two

adjacent stations on day 42 for interspecific comparisons. To test for seasonal variation in

natural 15

N abundance, we collected bivalves from the reference station on day 42 for
comparison with bivalves collected the day before the dripper was turned on (day -1). We
evaluated natural isotope abundance by comparing mussels collected from the reference
station on day 42 to values for food sources collected at that station averaged over the
course of the enrichment experiment. We froze unionids for storage and preserved S.
striatinum in ethanol until analysis (Peterson et al. 1993).

We surveyed unionids visible on the stream bottom to estimate abundance and
biomass. A more invasive subsurface survey would have disturbed the substrate, altering
seston composition and invertebrate drifl. The unionid population at the surface is also
that portion most directly interacting with the water column, and therefore the most
relevant to food web interactions discussed in this study.

Replicate samples of suspended chlorophyll a (Chl a) were collected on GF/F
filters from several points along the stream on two dates for measurement by flourometry
(W elschmeyer 1994). Suspended algae were sampled on 20 August 1998 at 4 locations:
at the Eagle Lake outlet, the outflow of wetlands 400 m upstream of the study reach, 20
m downstream of the dripper and 460 m downstream of the dripper. Samples were
concentrated by sedimentation and preserved in Lugol's solution for enumeration. Algae

were identified to species and ranked as either more prevalent in benthic habitats (scored

as l), more prevalent in pelagic habitats (scored as -l), or equally prevalent in both

28

pelagic or benthic habitats (scored as 0). The rankings were used to calculate a weighted
average index from -1 to 1 (pelagic to benthic, respectively; Zelinka and Marvan 1961).

These rankings were made by Dr. R. J. Stevenson of Michigan State University.

Tissue analysis- We selected individual unionid mussels for 51 5N analysis and did not

 

pool any unionid samples. Due to the long nitrogen turnover times in bivalves, we
analyzed three "tissues" rather than whole-body samples (Hawkins 1985). The stomach
gland was sampled to reflect short-term assimilation of nutrients because it is a major site
of nutrient absorption (Pechenik 1985). Foot muscle was sampled to integrate long-term
nutrient sources. Gut contents from the intestinal tract were sampled because they contain
partially undigested material. We dissected tissues from frozen specimens and dried and
pulverized the samples for isotopic analysis. Shells were used for species identification.
The entirety of S. striatinum soft tissue was dried and up to five individuals pooled per
sample station for isotopic analysis due to the limited amounts of soft tissue per

15

individual. The N turnover rates for bivalve tissues and food sources were estimated

815

from the slope of the ln( N) over time after the dripper was turned off (i.e. from the

depletion of tracer 15N).

To estimate ash-free dry mass (AFDM) and %N, we dried and ground the soft
tissues (gut contents removed) of 10 whole unionids of various species. We measured dry
weight and % ash gravimetrically and %N using a Carlo Erba elemental analyzer.
IsotOpic analysis of bivalves was done by continuous-flow isotope-ratio mass
Spectrometry with a Carlo Erba Elemental analyzer connected to a Finnigan Delta Plus

SpeCtrometer at the Center for Environmental Science and Technology, University of

29

Notre Dame, Notre Dame, Indiana, USA. The remaining Eagle Creek LINX 51 5N
measurements were made at that facility or using similar equipment at the Ecosystems

Center, Marine Biological Laboratory, Woods Hole, Massachusetts, USA.

Results

Unionids dominated consumer biomass in Eagle Creek (Table 2). A mean

abundance of 1.9 individuals m'2 was present at the surface. Although the collection
effort was not designed to systematically quantify species richness, we found 12 unionid

species: AnQdQntoides ferussacianes Actinonais ligamentina, Elliptio dilitata, Fusconaia

 

 

 

flava, Lampsilis cardium, Lasmigona compressa, Lasmigona costata Pleurobema

sintoxia, flganodon grandis, Strophitus undulatus Venustaconcha ellipsiformis, and

 

 

Villosa irr_s. Sphaeriids were serendipitously sampled, usually fiom underneath rocks, and

thus the biomass of sphaeriids is not known.
FBOM was abundant in the study reach, with 164 g AFDM m'2 in the upper cm

of the stream bottom and 922 g AFDM m"2 in the upper 5 cm (Hamilton et al. 2001).

The mass ratio of organic carbon to Chl a in surficial FBOM was 841, and the ON mass

ratio was 14. The stream carried a mean SPOM concentration of 3.3 mg AFDM L'1 in
the study reach; the C:Chl a ratio of this material was 571. The results of the longitudinal
survey of suspended matter above and within the study reach showed that the Chl 0
concentrations increased from the lake outflow (1200m upstream) to the wetland outflow
(immediately upstream), then decreased through the study reach (Table 3). A similar

decrease in Chl a concentrations through the study reach was observed on 14 July 1998,

30

falling from 2.85 (i 0.19 SD) to 2.18 (d: 0.08 SD) pg L'1 between the dripper and the
461-m station. The species composition of algae in the suspended matter collected in the
longitudinal survey showed that benthic algae predominated in all samples, with an
increase in the relative abundance of benthic algae along the course of the stream (Table
3).

The unionid Fusconaia flava was discovered during processing of B. slum
specimens; these two species are extremely similar in external appearance. Although we

believe the majority of bivalves we had originally classified as B. sintoxia were correctly

 

identified, we cannot rule out the possibility that a few were actually E. flava; we thus

describe this group hereafter as P. sintoxia + E. flava.

Natural isotope abundances- Natural 615

N values for the unionid community and food
resources, as measured just upstream of the enriched reach, are shown in Figure 1. EPS
and FBOM had identical mean values (3.6%o) and were enriched relative to SPOM
(1.3%0). Unionid muscle tissue (4.8%) was 1.6 %o enriched with respect to stomach gland
tissue (3.3%), 1.3%o enriched with respect to EPS and FBOM, and 3.5%o enriched

relative to SPOM. S. striatinum was 3.1 %o enriched relative to SPOM.

Isotope enrichment experimen - The addition of 15N -labeled NH4+ to the stream

elevated the 615N of food resources and consumers. Unionid stomach gland and gut

15

contents became measurably enriched with N during the addition (P < 0.0001, P <

0.0001, respectively) and then became depleted over time after the dripper was turned off

31

(_B = 0.0007, B = 0.0034, respectively, Fig. 2). Estimated turnover rates for 15N in the
stomach gland and gut contents were similar at 78 and 85 (1, respectively. Unionid
stomach gland tissue was on average 0.8%; (i 0.7 SD) enriched relative to gut contents

during the enrichment experiment. Muscle tissue from the foot responded very slowly to

15

the addition of N -enriched NH 4+ (3 = 0.02, Fig. 2). Although muscle tissue did

become depleted after the dripper was turned off, the depletion curve was not significant
(P = 0.16), making the estimated turnover rate of 357 d questionable. These turnover
rates may be overestimated by the possible consumption of enriched FBOM after the
dripper was turned off. FBOM, however, had a N turnover rate of 7 (1 (Hamilton et al. in
review) and this rapid rate would have minimized the influence of enriched FBOM

following termination of tracer addition. Based on the greater response to experimental

15N enrichment of stomach gland compared with muscle, the stomach gland was used for

seasonal, species and mixing model analyses within the enrichment experiment. Muscle
tissue was used for mixing model analysis of natural abundance data.

15

The potential food sources EPS and SPOM also became enriched during the N

addition (Fig. 3A). The temporal pattern of enrichment for EPS and SPOM followed the
same general pattern as the calculated 81 5N enrichment of the NIL:+ in the stream water
(Fig. 3B). The peak in calculated tracer 51 5N NH4+ between days 25-30 was caused by a
combination of lower discharge and lower ambient NH4+ concentrations, resulting in less

dilution of the added 15N (Fig. 3B; Hamilton et al. 2001).

32

It was not clear whether unionid tissues approached isotopic equilibrium during

15

the N enrichment period. Gut contents may have reached a plateau by the end of the

enrichment experiment (Fig. 2). Enrichment of stomach gland tissue appeared to still be

rising on day 42, although variability in SISN values of this tissue increased near the end

of the experiment.

The stomach gland SISN values fiom various unionid species were similar

between day -l and day 42 at the station above the dripper (mean for community: 3.4%o

:1: 0.6 SD and 3.3%o d: 0.4 SD, respectively). The 515N values of S. striatinum were also
similar between day -1 and day 42 at the station above the dripper (4.4%o and 4.3%,
respectively). Thus no seasonal isotopic change occurred in either the unionid community
or S. striatinum. The various unionid species reached similar levels of 15N enrichment
(Fig. 4).

The S‘SN for EPS, SPOM and unionid stomach gland varied over the length of the
stream on day 42 (Fig. 5). Enrichment of EPS appeared to increase in the downstream
part of the reach, as did the 5’5N of unionid stomach glands. S. striatinum varied widely
between locations with an average enrichment of 13.9%, which exceeds that of EPS and

SPOM, the presumed potential food sources (Fig. 5).

33

Discussion

The simultaneous consideration of natural abundance of stable isotopes with the
results of our isotope enrichment experiment yields several insights into the feeding
ecology of stream bivalves that are not available from other kinds of studies, but also
raises some interesting new questions that challenge our traditional view of freshwater
bivalves as primarily suspension-feeders. In the following discussion we address the
evidence regarding dietary sources of nitrogen for the bivalves, the role of bivalves in the

overall cycle of nitrogen in the stream, and the implications for conservation.

contents and tissues was due to the experiment rather than seasonal changes in natural
isotope abundance. The community was remarkable in its similar response suggesting
that all unionid species were utilizing similar food resources. The small response of
unionid muscle compared with stomach gland during the enrichment experiment
indicated that muscle tissue N integrates food resources over a much longer time period
than does the stomach gland.

Turnover times for muscle calculated in this study agree well with a 333 (1 whole-
body N turnover time calculated for Mytilus edulis by Hawkins (1985). Studies utilizing
unionids as a baseline for comparison of food webs between ecosystems would benefit
from analysis of isolated muscle tissue rather than the whole body because of longer-term
integration of dietary sources within muscle, whereas studies needing shorter-term

dietary information should use the stomach gland.

34

Food resources reached isotopic equilibrium or nearly so during the experiment

(Hamilton et al. in review), as illustrated by the similar patterns of changing 81 5N values
between the tracer, SPOM, and EPS (Fig. 3). This equilibrium reflected the rapid
turnover rates and direct utilization of suspended ammonium by algae and heterotrophic
microbes within SPOM and EPS. Unionids did not clearly reach isotopic equilibrium
during the course of the enrichment experiment. Two explanations of unionid food
resource utilization are thus possible: a diet of both EPS and SPOM (their presumed
potential food resources), and/or differential assimilation of food resource components

within the SPOM rather than uniform assimilation of the bulk organic material.

Sources pf suspended algae— The total amount of algae decreased along the study reach as

 

indicated by Chl a, while the relative abundance of benthic algae increased. This pattern
can be explained by settling and active removal of suspended pelagic algae by bivalves
and macroinvertebrate filtering collectors, coupled with sloughing of benthic algae from
the stream bottom. Lake outlet streams have long been recognized to support a high
density of filter feeders including unionids (Brbnmark and Malmqvist 1982; Richardson
and Mackay 1991). While the study reach was not strictly a lake outlet because there are
also wetlands below the lake, the importation of high quality food fi'om lentic
environments might explain the observed bivalve abundance and diversity. Thus, within
the suspension mode of feeding, the bivalves of Eagle Creek could be supported both by
constant re-suspension of material from deposited pools and an ecological subsidy of

pelagic phytoplankton supplied by the upstream lake and wetland. The concept of an

35

ecological subsidy could also be extended to organic detritus exported from the wetland

and entering the steam (Polis et al. 1996).

Food resource mixing model- The relative abundance of multiple food sources in the diet

of organisms can be quantified using an isotope mixing model:

15 _ 15
consumer _ 2 fi 8 Ni (2)

515

where fi and Ni are the fractional contributions and 51 5N values of each food

resource i, and 2 fi = 1. Use of a mixing model assumes both the organism and potential

food resources to be at isotopic equilibrium (Gannes et al. 1997).

Bulk SPOM could not be the predominant food resource for bivalves in the
stream because we observed a consistent enrichment of stomach gland tissue to levels
greater than that of SPOM (Fig. 5). The simplest explanation of food resources for
unionids then becomes a mixture of SPOM and EPS. In the enrichment experiment each

sampling station was treated as a different test of unionid food resource partitioning. The

15N enrichment for stomach gland slightly exceeded that of EPS at the 251m station

(5.8960 and 5.7960 respectively), and the 461m station (8.4%o and 7.2960 respectively).
Unionids were assumed to have consumed EPS exclusively at these stations for the
purposes of mixing model analysis. The mixing model using stomach gland data from the
enrichment experiment suggested unionids consumed 80% EPS and 20% SPOM on
average (Table 4).

Application of the mixing model to natural muscle isotope abundance data

required correction of consumer 51 5N values due to fractionation that occurs in

36

consumers relative to their food resources. In the enrichment experiment, trophic
fractionation was accounted for by the subtraction of background (natural) 815N values

from the measured SISN values. A 3.4%o enrichment per trophic level is commonly used
in food web studies, but this is an average value for a wide variety of animals (Minigawa

and Wada 1984). Data on marine bivalves from the literature suggest that bivalves tend to

be about 1.7%o enriched with 15N relative to suspended food resources (calculated from

Minagawa and Wada 1984; Fry 1988). Application of the mixing model to natural-

abundance muscle 815N values, assuming a trophic enrichment of 1.7960, indicated that
unionids consumed 81% EPS and 19% SPOM. Using the same trophic correction of
1.7%o with whole-body S. striatinum 5'5N values, the mixing model indicated that S.
striatinum consumed 64% EPS and 36% SPOM. Application of the mixing model to
natural isotope abundance data for S. striatinum agreed well with Hornbach et al.'s (1984)
energy budget for this species that indicated greater reliance on deposit feeding than

suspension feeding.

Suspension y_s_. deposit feeding- Suspension feeding is the removal of suspended particles
including phytoplankton from the water column. Deposit feeding is the consumption of
particles from the sediment. Descriptions of unionid behavior and the role of unionids in
aquatic ecosystems emphasize the suspension mode of feeding (e.g. McMahon 1991;
Strayer et al. 1994; Strayer et al. 1999, Box and Mossa 1999).

Use of the deposit feeding mode by freshwater bivalves has generally been
described within specific contexts in which direct feeding upon sediment by adults is not

usually considered to be typical. Examples include: 1) Illustrations of exceptional

37

deposit-feeding behaviors (Way 1989 cited in MacMahon 1991; see also Reid et al.
1992); 2) Burial of sphaeriids in the substrate (e.g. Pennak 1989; MacMahon 1991); and
3) Pedal transport of sediment by juvenile unionids (Reid et al. 1992; Yeager et al. 1994).
Re-suspended sediment has been examined as a factor affecting suspension-feeding
behavior and clearance rates, but this is not an example of deposit feeding (e. g. Kiorboe
et al. 1980; Bricelj et al. 1984).

In the scientific literature, descriptions of adult unionids as both suspension and
deposit feeders are uncommon (e. g. Singh et al. 1991), while the belief that adult
unionids are exclusively suspension feeders is pervasive yet not supported by empirical
evidence. Ecological studies often examine unionids within the context of suspension
feeding only (Kramer 1979, Brbnmark and Malmqvist 1982, McCall et al. 1995, Parker
et al. 1998). If citations are given in reference to stating that unionids are suspension
feeders, the citations are usually not of studies that quantify freshwater bivalve feeding
ecology (Brbnmark and Malmqvist 1982, Yeager et al. 1994). To state that adult unionids
are exclusively suspension feeders is therefore an inductive conclusion.

The mixing model used in this study suggested a predominance of the deposit
mode of feeding in unionids, with good agreement between model results obtained with
natural abundances and results obtained from the enrichment experiment. Indeed,
application of a mixing model might be considered a standard procedure in a study of this
kind. Use of the model, however, requires satisfaction of several conditions: isotopic

equilibrium and assimilation of bulk material. The 15

N enrichment of gut contents
appeared to reach a plateau over the 42-day addition, but the stomach gland became more

variable at the end of the experiment. These results obscure evaluation of whether

38

unionids reached isotopic equilibrium. If unionids were near equilibrium by the end of

the experiment, then the mixing model is still reliable. If not, the model yields a maximal

estimate of the importance of the relatively 15N -depleted SPOM because the SISN of
unionid tissues would have continued to rise had the experiment been longer.

The duration of the present study was insufficient to definitively evaluate feeding
mode preference. Future in-situ isotope enrichment studies of bivalves will need longer
periods to accommodate slow N turnover rates. Laboratory experiments, however, should
prove more valuable in determining the extent to which adult freshwater bivalves employ
the deposit mode of feeding (Gannes et al. 1997). Laboratory isotope enrichment

experiments could entail use of smaller amounts of 15

N, thereby costing less, and
different levels of isotopic enrichment between suspended and deposited organic matter

could be produced in the laboratory.

Preferential assimilation pf microbes and algae- Pennak (1989) describes the diet of

 

freshwater bivalves as consisting chiefly of fine organic detritus dislodged from the
substrate, with phytoplankton of minor importance. In contrast McMahon (1991) lists
phytoplankton, bacteria and fine detritus as the chief components of the freshwater
bivalve diet, with deposit feeding as a potential, and perhaps underestimated, supplement.
This dichotomy of opinion should not be surprising given the common assertion that the
ecology of freshwater bivalves is poorly understood (e.g. Bogan 1993; Neves 1993, 1997;
Box and Mossa 1999).

In LINX experiments in other streams, macroinvertebrate consumers of epilithon

often became more highly labeled than their food resources (e.g. Tank et al. in press).

39

This could only have occurred if the organisms were assimilating highly labeled algal
components of epilithon, rather than the bulk material that included algae, mucilage and
entrained detritus. Unionids could also have preferentially assimilated the microbial and
algal components of the bulk material they ingested. Nichols and Garling (2000) found
that unionids in a Michigan river and lake fed primarily on bacteria and concluded that
the unionids were selectively assimilating bacteria from ingested detritus. The enrichment
of S. striatinum to levels greater than that of bulk potential food sources supports the
preferential assimilation explanation (Fig. 5). Due to small body size relative to unionids,
S. striatinum likely had higher N turnover rates, resulting in greater enrichment during
the course of the experiment.

If bivalves differentially assimilate living components of bulk fine organic

material, we cannot use the 815N data of bulk samples to distinguish between suspension
and deposit feeding. This is because SPOM represented suspended material of multiple
origins including lentic phytoplankton exported to the stream, attached algae sloughed
from surfaces, and FBOM in constant exchange between suspension and deposition. Lake
phytoplankton, however, passed through the study reach very quickly (~32 min) due to
high current velocities and this short residence time may have prevented much uptake of
labeled ammonium by plankton that entered the study reach already in suspension.
Labeled microbial portions of SPOM were more likely derived from resuspended stream
sediments and benthic algae. In light of the probability of preferential assimilation of
microbial and algal components of food, the mixing model becomes questionable. The
possibility of preferential assimilation does not, however, eliminate the potential for

deposit feeding to exist in this community. The bivalves of Eagle Creek may to a

40

substantial extent feed directly on sediments and likely derive their nutrition from the

living microbial and algal components of fine organic material on the sediment surface.

represented the largest pool of consumer N in the stream ecosystem. The present study
shows that the slow N turnover rate of unionids, however, effectively prevented N bound
in biomass from cycling in the short term. The importance of this pool of N may appear
upon the death of a unionid by releasing large amounts of N to the sediment in a localized
area. Predation upon bivalves by terrestrial mammals such as raccoons represents a loss

of this N from the stream.

Implications Q conservation- Both the mixed diet of EPS + SPOM and differential
assimilation explanations of patterns seen in this study could improve our understanding
of the relationship between stream bivalves and sediment. For example, while
improvements in the success of freshwater bivalve translocation efforts have been made
by refining handling techniques, limited understanding of unionid ecology still impedes
optimal site selection (Dunn and Sietrnan 1997). Poor success of translocation efforts
might be explained in part by differences between locations in the microbial and algal
components of sediment and seston. A radical change in sediment characteristics can be
encountered by unionids during a translocation project, and living components of
sediment may differ in abundance or palatability in the new location. Such differences in
the sediment would affect deposit feeding directly and suspension feeding indirectly by

altering one source of suspended organic matter. Changes in the microbial community of

41

sediment and seston may also explain why greater translocation success has been seen in
projects that minimize the distance mussels are moved, especially to new locations in the
same river (Valovirta 1998). Such alteration might not need to occur in the stream itself
to have an impact if an upstream wetland or lake were providing an ecological subsidy of
suspended organic material. Clearly we should consider how alteration of the sediment
impacts the food sources of unionids within the context of the general ecology and

conservation of these increasingly threatened organisms.

42

Acknowledgements

We thank D. Graf, R. Mulcrone, E. Nedeau and N. Dom for help identifying mussels. J.
Tank provided valuable insights into the data. R. Dufford enumerated algae and R. J.
Stevenson identified the habitat preferences of algae. C. Kulpa and D. Birdsell of the

Center for Environmental Science and Technology at the University of Notre Dame

kindly allowed us to use their analytical facility for 515N analyses. This study was a
spinoff of the Lotic Intersite Nitrogen eXperiment (LINX); P. Mulholland, J. Meyer, J.
Webster, and B. Peterson were the PPS and numerous other investigators contributed to
the experimental design. Support for the lead author was provided by Sigma Xi Grants-
in-aid-of-research and the KBS Graduate Research Training Group (RTG) funded by the
National Science Foundation (NSF, DBI-9602252). The overall LINX experiment was
supported by grants from the Ecosystems Program of NSF to Virginia Polytechnic
University (DEB-9628860) and to Michigan State University (DEB-9810220 and DEB-

9701714). This is contribution 945 of the W. K. Kellogg Biological Station.

43

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Box, J. B., and J. Mossa. 1999. Sediment, land use, and freshwater mussels: prospects
and problems. J. North Am. Bethol. Soc. 18:99-117.

Bricelj, V. M., R. E. Malouf, and C. de Quillfeldt. 1984. Growth of juvenile Mercenaria
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Bronmark, C., and B. Malmqvist. 1982. Resource partitioning between unionid mussels
in a Swedish lake outlet. Holarct. Ecol. 52389-395.

Cabana, G., and J. Rasmussen. 1996. Comparison of aquatic food chains using nitrogen
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47

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48

Table 1. Hydraulic and geomorphological features of Eagle Creek, MI.

 

Slopea

Water velocityb
Water depthb
Channel widtha
Dischargeb

Substratea:

Habitata:

0.6 %

24 cm s'1

19 cm

5m

202 L 8-1

Sand and/or FBOM

Gravel

Cobble

Riffle

Run

54%

35%

10%

90%

10%

 

a: Mean for study reach.

b: Mean during enrichment experiment.

49

Table 2. Biomass and nitrogen content of consumers in Eagle Creek, MI.

 

 

g AF DM m'2 g N m-2

Unionids l .93 0.196
Fishes 0.83 0.074
Macroinvertebrate Functional Feeding Groups:

Shredders 0.1 1 0.010

Scrapers 0.07 0.007

Collectors 0.04 0.005

Filterers 0.02 0.002

Predators 0.08 0.010

 

50

Table 3: Chlorophyll a and relative abundance of planktonic and benthic algae in Eagle
Creek, M1, on 20 Aug 1998. The wetlands occur between the lake outflow and the

 

 

dripper.
1200m
upstream 400m 20m 460m
(Eagle Lake upstream downstream downstream
Outflow) of dripper of dripper of dripper
Chlflmean pg L-l) 2.15 5.72 3.15 2.14
(SD) (0.02) (0.87) (0.34) (0.20)
Cell density (cell 15‘)
Bacillariophyta 1 10 136 60 20
Chlorophyta 1 8 l 122 200 1 70
Chrysophyta 1 144 48 20
Cryptophyta 126 141 122 168
Cyanoprokaryota 10601 3862 2796 4393
Euglenophyta 0 9 3 0
Pyrrophyta 1 1 0 0
Total 10731 4157 2973 4594
Benthic-Pelagic Index -5.4 E-02 -5.3 E-04 -1.9 E-06 -9.3 E-08

 

51

Table 4: Food resource partitioning for bivalves in Eagle Creek, MI, based on observed

15N enrichment on day 42 of the experiment. Stomach gland SISN was used for bivalves
unless otherwise noted in text.

 

 

Location (m) % EPS % SPOM
116 50 50
176 77 23
251 100a

301 48 52
351 96 4
396 86 14
461 100a

Mean 80 20
Unionid Muscleb 81 19
S. striatinumb 64 36

 

a: SISN of tissue slightly exceeded 815N of EPS at this location.

b: Natural abundance 815N with 1.7%o correction for trophic level enrichment.

52

Figure 1. Natural abundance of 15N in food resources and bivalves of Eagle Creek, MI.
EPI = Epilithon, EPS = Epipsammon, FBOM = Fine Benthic Organic Matter, SPOM =
Suspended Particulate Organic Matter, SPH STR = Sphaerium striatinum. EPS and
SPOM are presumed to be the potential food resources for bivalves. Data for food
resources are means for 8 collection dates at the reference station during the course of the

enrichment experiment. Gland and muscle are isolated tissues from Pleurobema sintoxia

 

 

+ Fusconaia flava. The datum for Sphaerium striatinum is a composite of the whole soft

tissue of several individuals. Error bars are :t 1 standard deviation.

53

 

 

Em 3mm

cams: case:

case Econ:

mm><mq

20mm

0003

20mm

mmm

Em

54

15

Figure 2. Enrichment of unionid tissues over the course of the N tracer experiment,

with 615N values corrected for background SISN. Error bars are i: 1 standard deviation.

55

515N enrichment over background

 

 

I I j I 'l"

+Gland

- 0— - Gut contents ‘

- - El- - - Muscle ‘,

6_ ’13‘~:‘

I .

['1 \\
I” \§\\

‘ ‘~ 0
A-;$-'O-E ---------- QI'--$---:&l I .~.§‘--r--'§

O 5101520253035404550556065

Day of experiment

56

Figure 3. A) Enrichment of potential food resources of bivalves over the course of the

15N tracer experiment, with 815N values corrected for background 815N. The last

SPOM datum was calculated by extrapolating the decay rate. Error bars are i 1 standard

15

deviation. B) Calculated N enrichment of dissolved NIL.+ during the experiment, at the

15

dripper. The enrichment decreased across the study reach as the NH 4+ was assimilated

615 15NH +" was 70-75% that

Nof 4

or nitrified, and at the main bivalve sampling site the

of the dripper (Hamilton et al. 2001).

57

147

+EPS A
-'A"- SPOM ‘

l

12

 

10-

815N enrichment over background

 

 

 

0 I I I I T I I I r I Ifi'fi'fl
0 5101520253035404550556065

1000 1
B
900 -

800 -
700 -
600 d
500 -
400 -
300 -
200 -
100 -

 

Calculated 615N of NH4+

 

 

 

0 I 1 I I r I I I I T I I j

0 5 10152025303540455055 6065
Day of experiment

58

Figure 4. Comparison of 15N enrichment in unionid stomach glands and gut contents on

day 42 (515

N not corrected for background levels of enrichment). PLE + FUS =
Pleurobe_m_a sintoxia + Fusconaia flava; ELL DIL = Elliptio dilitata; STR UND =

Strophitus undulatus; LAM CAR = Lampsilis cardium; VEN ELL = Venustaconcha

 

ellipsiformis; ACT L1G = Actinonais ligamentina; LAS COS = Lasmigona costata; VIL

IRI = Villosa iris.

59

O Gland

<> Gut contents

I

:82

5: .d>

mOU was;

05 HU<

15m 75>

M<U $24

T

923 MHm

I

an Adm

mam + mum

 

10-

 

60

Figure 5. Longitudinal patterns of 15N enrichment for unionid stomach-gland samples,
Sphaerium striatinum whole-body samples, and potential food resources across the study

5

reach on day 42 of the l N enrichment experiment (values corrected for background

51 5N). Error bars are i: 1 standard deviation.

61

515N enrichment over background

25'}

+ EPS
O —Q—' Gland
20 J _ 'A ' - SPOM
O S. striatinum
15 - O

 

 

 

110 160 210 260 310 360 410 460

Distance downstream from dripper (m)

62

Chapter 3
FOOD WEB INTERACTIONS BETWEEN ZEBRA MUSSELS, ZOOPLANKTON,
AND LARVAL FISH

Abstract

Food web interactions between exotic invasive zebra mussels (Dreissena polymorpha),
zooplankton, and native larval bluegill (Lepomis macrochirus) were examined with a
mesocosm experiment. Hatchling larval bluegill collected from nests were reared in the
presence of size-structured populations of zebra mussels in 1,500 L limnocorrals
suspended in an artificial pond for two weeks. Chlorophyll a and other limnological
variables, and zooplankton abundance and biomass (including copepod nauplii and
rotifers) were monitored over time. During their first two weeks of life, larval fish reared
in the presence of mussels grew 24% more slowly than fish reared alone. Competition
between mussels and bluegill for food in the form of microzooplankton can explain the
differential growth rates, but also likely was an indirect competition via starvation of the
zooplankton community as zebra mussels consumed the phytoplankton. Either direct or
indirect trophic competition between zebra mussels and obligate planktivores may result

in ecological harm as zebra mussels spread throughout inland lakes of North America.

63

Introduction

As zebra mussels (Dreissena polygporpha) expand their range in North America
new ecosystems become invaded. Beginning with the Great Lakes and continuing
through the Mississippi River drainage network, the zebra mussel invasion is currently
spreading among small inland lakes. The southern peninsula of the State of Michigan is
at the forefront of this invasion due to its proximity to the Great Lakes, its high density of
inland lakes, and its high rate of inter-lake recreational boat traffic (Buchan and Padilla
1999). Over 150 of Michigan’s inland lakes are known to have zebra mussels and over
75% of Michigan’s thousands of inland lakes possess suitable pH and calcium to support
zebra mussels (Ramcharan et al. 1992, Klepinger 2000).

Among the ecological concerns of zebra mussel invasion are food web
interactions. Zooplankton communities, for example, have been negatively affected by
the invasion of zebra mussels in the Great Lakes and Hudson River. Zooplankton
abundance dropped 55-71% following mussel invasion in Lake Erie, with the smallest of
these animals more heavily impacted (MacIsaac et al. 1995). The total biomass of
zooplankton in the Hudson River declined 70% following mussel invasion, due both to a
reduction in large 200plankton body size and reduction in small zooplankton abundance
(Pace et al. 1998). These effects can be attributed to reduction of available food
(phytoplankton) due to competition with zebra mussels and because the mussels feed
directly on microzooplankton such as copepod nauplii and rotifers. Less well understood

are interactions between zebra mussels and higher trophic levels including fish.

Although the feeding habits of the larval stages of many fish species are not well
understood, larvae are most often visually-oriented predators of zooplankton during the
earliest life stage, and they experience high mortality rates (Gerking 1994). Body size is
generally believed to affect survival because larger size confers advantages in starvation
resistance, swimming ability, predation avoidance, foraging ability, and overwinter
survival (Adams et al. 1982, Werner and Gilliam 1984, Blaxter 1986, Crowder et al.
1987, Miller et al. 1988, Pepin 1991, Schindler 1999). Prey availability can also affect
larval size development (Mills et al. 1989, Rettig and Mittelbach in press). It is thus
logical that if zebra mussels can affect the abundance or size structure of zooplankton
communities, larval fish growth or survival can be diminished (Richardson and Bartsch
1997, Thayer et al. 1997, Trometer and Busch 1999). No studies to date have employed
experiments to examine interactions between zebra mussels and larval fish.

One fish species that might compete for food with zebra mussels in inland lakes is
bluegill (Lepomis macrochirus). In early summer, bluegill hatch from eggs in nests built
in the littoral zone, migrate to the limnetic zone, then retum to the littoral zone in the fall
(Faber 1967, Werner 1967, 1969). Larval bluegill feed on microzooplankton such as
copepod nauplii and rotifersas well as small-bodied cladocerans while in the limnetic
zone (Seifert 1972, Barkoh and Modde 1987, Bremigan and Stein 1994, Welker et al.
1994).

The purpose of this study was to test whether zebra mussels negatively impact the
growth and survival of larval bluegill in an experimental setting using limnocorrals
suspended in a pond as mesocosms. A negative impact of zebra mussels on larval bluegill

could occur by several mechanisms: 1) The entire zooplankton community might decline

65

due to starvation if algal biomass [as indicated by chlorophyll a (chl a) concentrations]
decline in the presence of zebra mussels; 2) Microzooplankton abundance might decline
in the presence of zebra mussels independently of macrozooplankton dynamics if zebra
mussels heavily consumed microzooplankton and thus competed directly with larval
bluegill for food and if microzooplankton were less tolerant of lower algal
concentrations; 3) Toxic conditions might be produced by zebra mussels through

excretion of ammonia or consumption of oxygen.

Methods

Twelve 1.5-m deep, l-m2 limnocorrals holding 1,500 L of water each were
installed in an artificial pond at the Experimental Pond Facility of the W. K. Kellogg
Biological Station, Michigan State University, Hickory Corners, MI, USA. The artificial
pond measured 29 m in diameter and was 1.8-m deep, and did not have zebra mussels but
contained redear sunfish (L. microlophus) that maintained a zooplankton community
dominated by small-bodied cladocerans and copepods. A balanced factorial design of
zebra mussels (ZM) x bluegill (BG) with three replicates each of the four treatments
(ZM, BG, ZM+BG, control) was used. Limnocorrals were arranged in two parallel rows
and treatments assigned in a stratified random fashion with one replicate of each
treatment at the end of a row. The limnocorrals were filled with pond water screened
through 100-um mesh and then stocked with natural densities of zooplankton collected
from the pond with a 100-um net and mixed in a large container before removing aliquots

for each replicate.

66

Zebra mussels were collected from nearby Gull Lake, then held in a tank with
continuous flow-through of Gull Lake water and allowed to attach to Plexiglass sheets for
two weeks. Size-structured populations of 200 mussels per treatment were created: 25 of
20-30 mm, 75 of 10-20 mm, and 100 of <10-mm, for a total of approximately 2 g AFDM
/ m2 of the mussel's soft tissue (Young et al. 1996). The Plexiglass sheets with attached
zebra mussels were suspended vertically 50 cm below the water surface in the
limnocorrals. Sediment traps were attached immediately under the Plexiglass sheets.
Control treatments received Plexiglass sheets and sediment traps without any attached
mussels. Dead mussels were removed from the experiment when detected, but any
change in mussel abundance within a replicate was disregarded in the data analysis due to
a very low mortality rate.

Hatchling bluegill were collected directly from nests in nearby Warner Lake
(Barry County) and reared in the lab for several days. When the fish reached the
swimming stage they were fed brine shrimp nauplii. Fish that had fed upon nauplii were
then selected from a single cohort in one tank for use in limnocorrals. The standard length
of larval fish from the cohort used for the experiment was 5.5 mm (SD :1: 0.16) estimated
by measuring 20 fish not used in the experiment but from the same cohort. Twenty larval
fish were used per treatment. Control treatments had no fish. Bluegill and zebra mussels
were added to the mesocosms at the same time three days after zooplankton had been
added.

Samples were taken approximately every three days from 9 July to 22 July 1999.
The experiment was terminated after two weeks in order to focus on the period when

larval fish and zebra mussel diets overlap. Temperature, 02, pH, and specific

67

conductance were measured with a YSI multisensor. Sestonic chl a was measured by
flourometry (W elschmeyer 1994). Measurements and sampling of water were performed
before moving Plexiglass plates and sediment traps to prevent sediment from being
resuspended and thereby contaminating the water column. Transparency was measured
with a light meter. A two-tiered regime was used to sample zooplankton: 2 l-m tows
using a lO-cm diameter lOO-um net for macrozooplankton and the filtration of 4 L of
water through 30-um mesh for microzooplankton. All zooplankton were stored in 95%
ethanol. To collect the larval bluegill at the conclusion of the experiment, a l-m diameter
500-um net was pulled through each limnocorral until three consecutive tows yielded no
fish. Fish were placed into containers, put on ice, and measured the day of collection.
Zooplankton were enumerated with a dissecting microscope. Body size was
measured for macrozooplankton only using a digitizing pad. Entire zooplankton samples
were enumerated, but only up to 50 individuals of a species were measured for
calculation of biomass using published length-mass regressions (McCauley 1984).
Several measures were undertaken to ensure that zebra mussels did not escape
fi'om the experiment (Reid et al. 1993). The tops of mesocosms were covered with gray
fiberglass window-screen to prevent animals from entering the limnocorrals and
transporting water potentially infested with zebra mussel veligers to other experimental
ponds. After the experiment, the limnocorrals were treated with Rotenone and chlorine
before the bags were cut away from their frames. Lastly, the entire pond was drained
(with the water infiltrating nearby soil) and allowed to dry completely before being

refilled.

68

Time series measurements were analyzed by Repeated Measures Analysis of
Variance (rmAN OVA) using SAS for Windows v.8 (von Ende 1993). Only days 3
through 13 were used for statistical analysis because day zero represented initial
conditions, i.e. those measurements were taken before treatment factors were in place. In
rmANOVA terminology, between-subjects effects refer to differences between the means
of treatments for the time period and within-subject effects refer to differences between
treatments in the temporal pattern of response (i.e. the time x treatrrrent interactions). If a
sphericity test on orthogonal components was rejected then Greenhouse-Geisser
corrected rmANOVA results were used for within-subject effects, otherwise Multivariate
Repeated Measures ANOVA (MANOVAR) results were used (SAS Institute 1999).

Larval fish sizes were compared with a pooled-variance t-test.

Results

Larval bluegill reared in the presence of zebra mussels were shorter at the end of
the experiment than larval bluegill reared alone (initial standard lengths were 5.5 m :t
0.16, final standard lengths were 8.4 m i 0.03 SE in ZM+BG treatments and 9.3 m i
0.15 SE in BG treatments, P = 0.03, t = 5.33, Table 1). Thus bluegill growth rates were
about 24% lower in the ZM+BG compared to the BG treatments (0.22 mm/d and 0.29
mm/d, respectively). There was no significant difference in survival among treatments
(initial count was 20 fish, final counts were 10.7 :1: 1.7 SE fish in ZM+BG treatments and
12.0 i 2.6 SE fish in BG respectively, B = 0.69, t: 0.43, Table 1).

Macrozooplankton included calanoid and cyclopoid copepods, as well as the

cladocerans Daphnia ambigua, Bosmina longirostris, Diaphanosoma spp., Scaphloberis

69

spp., Simocephalus spp., Chydorus spp., and Si_da spp. Although significant between-
subjects effects of zebra mussels and bluegill were found for total macrozooplankton
biomass (B = 0.0021, E = 19.82; B = 0.0076, F = 12.53, respectively), greater initial
macrozooplankton densities were present in control treatments due to the presence of
greater numbers of Bosmina and copepods in spite of attempts to equalize densities
across all enclosures (Fig. 1A). This difference in initial conditions confounds
interpretation of zebra mussel effects on macrozooplankton (between-subjects effects).
Total macrozooplankton biomass fell over time during the experiment (within-subject
effect, G-G Time _12 = 0.0103, E = 6.51) but zebra mussels and bluegill had no within-
subject effect (Table 2).

Microzooplankton included copepod nauplii and the rotifers Keratella cochlearis
and Polyarthra Mam. Initial microzooplankton densities were less variable among all
treatments (Fig. 18). Both zebra mussels and bluegill caused significant between-subjects
effects, reducing the abundance of microzooplankton (B = 0.0011, F = 24.91; B = 0.0300,
F = 6.94, respectively, Table 3). Temporal dynamics of the major microzooplankton taxa
were variable. Copepod nauplii dominated the microzooplankton count and drove the
temporal pattern seen in figure 1B, where nauplii increased over time in control
treatments, declined quickly in ZM and ZM + BG treatments, and were affected by fish
late in the experiment in BG treatments. Keratella abundance dropped to below
detectable levels in all treatments by the third day. Polyarthra was reduced to low levels
by day three and held at low constant levels in ZM and ZM+BG treatments while

increasing in abundance in control treatments.

70

Within the control treatment, chl a fell steadily from a mean of 2.70 ug/L to a
mean of 1.21 rig/L (Fig. 2A). Zebra mussels further decreased the concentration of chl a
(between subjects I: < 0.0001, F = 415.39). Zebra mussels increased the transparency of
the water column (between subjects E = 0.0001, F = 46.27, Fig. ZB). Most of the impact
of zebra mussels occurred within the first three days. Transparency decreased in all
treatments between days 3 and 6.

Specific conductance (corrected to 25°C) fell from an initial experiment-wide
mean (i SD) of 315 (d: 1.3) [IS/cm to 311 (i 1.7) uS/cm on day 3 and then remained
level. Dissolved oxygen fell from an initial experiment-wide mean (:1: SD) of 75 (d: 4) %
saturation to 54 (:1: 7) % saturation and then remained level. The pH of all treatments rose
from an experiment-wide mean (i SD) of 7.67 (:1: 0.03) to 7.75 (d: 0.05) by day 9 (day 13
pH data were not available). Temperature did not vary among treatments and ranged

between 23 and 26 °C during the experiment.

Discussion

Zebra mussels substantially reduced the growth of bluegill during the critical
larval stage. The most likely cause of this effect was competition with fish for
consumption of microzooplankton. A reduction in the abundance of algae probably also
impacted the microzooplankton community. Toxic conditions were not produced by
zebra mussels.

The temporal pattern of microzooplankton abundance suggested direct

consumption by zebra mussels. Temporal patterns of rotifer abundance mirrored those of

71

phytoplankton with an initial decline followed by low abundances, although Keratella
abundance was below detectable levels after day 3. Sustained feeding by zebra mussels
appeared to prevent nauplii abundance from increasing as seen in the control treatment.

Although the susceptibility of microzooplankton including the rotifers Keratella
and Polyarthra to predation by zebra mussels has been conclusively demonstrated by
previous researchers (Shevtsova et al. 1986, MacIsaac et al. 1991, MacIsaac et al. 1995),
additional interactions may have existed between zebra mussels and microzooplankton in
the present study. Microzooplankton could have been incorporated into pseudofeces
instead of being consumed by mussels. The net effect of pseudofeces incorporation of
microzooplankton would, however, have been the same for larval fish, i.e. exploitative
competition for resources. Less probable was either overall or selective consumption of
algae by zebra mussels having caused selective suppression of microzooplankton through
exploitative competition for algal food resources. Zebra mussels preferentially feed on
particles 5 to 45 pm in size, a range which overlaps that which rotifers consume (4 to 17
um) and that which copepods consume (<1um to 1mm) (Gilbert 1985, Sprung and Rose
1988, Williamson 1991). Zebra mussels are also able to feed on a larger range of
particles, from <1 11m to 750 pm including bacterioplankton, Ankistrodesmus cells and
cyanobacterial filaments (Ten Winkel and Davids 1982, Cotner et al. 1995, Horgan and
Mills 1997). Zebra mussels were thus unlikely to have selected a size range of particles
that was detrimental to microzooplankton and simultaneously not detrimental to
macrozooplankton.

Macrozooplankton abundance did not decline to a greater degree in the presence

of zebra mussels. This result is consistent with the idea that zebra mussels directly

72

consumed microzooplankton and did not counsume macrozooplankton. Larger
200plankton, however, may have tolerated possible starvation conditions for longer than
smaller zooplankton, and hence may not have declined on the short time scale of the
experiment. While chl :1 concentrations were reduced by zebra mussel consumption
within the first three days, macrozooplankton biomass fell relative to initial values in all
treatments. In the ZM treatment, macrozooplankton biomass actually rose between days 6
and 13. The difference in initial macrozooplankton biomass densities could be
attributable to uneven stocking despite attempts to add equal amounts. While this
confounds comparison of between-subjects effects, the lack of a statistically significant
within-subjects effect of zebra mussels on macrozooplankton biomass shows that the
decline in macrozooplankton biomass in treatments with zebra mussels occurred with the
same temporal pattern as in other treatments.

It is possible that reduced phytoplankton abundance in treatments with zebra
mussels reduced available energy to copepods, and that as a result copepod reproduction
rates might have been lower in treatments with zebra mussels (Williamson 1991). While
the average biomass of an individual copepod increased in all treatments, it is not known
whether this reflects the growing of organisms or the culling of small individuals from
the population.

No toxic conditions were produced by zebra mussels. Oxygen concentrations
were not reduced in any treatment. While ammonium concentrations changed during the
experiment, rising during the first half then falling during the second half, zebra mussels

had no effect. Moreover, ammonium concentrations never exceeded toxic levels.

73

The potential for zebra mussels to affect bluegill growth in natural systems should
occur during the earliest life stages when bluegill are most restricted in their diet.
Richardson and Bartsch (1997) studied zebra mussel-bluegill food web interactions in
mesocosms using omnivorous juvenile bluegill (34-mm long) that were not limited to
microzooplankton, and found that the growth of bluegill was not affected by the presence
of zebra mussels. Competition with zebra mussels should not occur during later life
stages when bluegill display more omnivorous feeding behavior. Several questions,
however, affect the potential for and biological significance of bluegill-zebra mussel
competition in natural systems.

The obvious question is whether zebra mussels have actually had an effect on
microzooplankton abundance in inland lakes. MacIsaac et al. (1991) speculated that
vertical zonation of zooplankton and poor water column mixing might ameliorate the
impact of microzooplankton predation by benthic zebra mussels in the Great Lakes, and
the same could be true of small inland lakes. The epilimnion of a small inland lake,
however, is arguably in greater contact with the benthos than the epilimnia of the Great
Lakes. An exception would be where the Great Lakes are shallow, like the western Basin
of Lake Erie and Saginaw Bay. If obligatory feeding on microzooplankton is restricted to
a short enough time period then larval fish might survive long enough to outgrow reliance
on the size class of zooplankton fed upon by zebra mussels (Lazzaro 1987). This
ontogenetic niche shift may include feeding on benthic or littoral invertebrates and may
help to explain why detrimental impacts on fish populations have not yet been observed
in inland lakes invaded by zebra mussels (Gopalan et al. 1998, Rutherford et al. 1999,

Trometer and Busch 1999, Mayer et al. 2000, Idrisi et al. 2001).

74

Acknowledgements

I thank Sue Raikow and Stefanie Whitmire for assistance in the field. Alan
Wilson helped build limnocorrals and prepare the pond. Stephen Hamilton, Gary
Mittelbach, Alan Tessier, and Orlando Samelle provided helpful comments about this
manuscript. Everyone who attended KBS brown bag and EEBB colloquium seminars
provided useful feedback. Jessica Rettig, Chris Steiner, and Laura Broughton kindly
answered methodological questions. Madhav Machavaram, Nina Consolatti, and John
Gorentz provided invaluable logistic support. This research was based in part upon work
supported by the US. National Science Foundation (DEB-9701714), the KBS Graduate
Research Training Group (RTG) funded by the US. National Science Foundation (DBI-
9602252), and by grants from the Environment Now Fund of The Kalamazoo

Community Foundation.

75

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79

Table 1. Pooled-variance t test of zebra mussel (ZM) effects on larval bluegill (BG)

 

growth and survival.

Treatment N Mean SE (If t P
BG Length ZM + BG 3 8.5 0.03 2.2 5.330 0.0274
(mm)

BG 3 9.3 0.15
BG Survival ZM + BG 3 10.7 1.7 3.4 0.426 0.6956
(count)

BG 3 12.0 2.6

 

80

Table 2. Univariate repeated measures ANOVA test for effects of zebra mussels and

larval bluegill on macrozooplankton biomass over the course of the experiment (days 3 to

 

13).
Source df MS F P

Between-subjects ZM 1 0.021 19.82 0.0021
BG 1 0.013 12.53 0.0076
ZM X BG 1 0.002 2.54 0.1493
Error 8 0.001

Within-subject Time 3 0.002 6.51 00103"
Time X ZM 3 0.000 2.08 01621"
Time X BG 3 0.001 3.21 00720"
Time X ZM X BG 3 0.000 2.06 01640”
Error 24 0.000

 

* Greenhouse-Geisser corrected value.

81

Table 3. Univariate repeated measures ANOVA and multivariate repeated measures

ANOVA (MANOVAR) testing for effects of zebra mussels and larval bluegill on

microzooplankton abundance over the course of the experiment (days 3 to 13).

 

Source df MS F P

Between-subjects ZM 1 l 16328.5 24.91 0.001 1
BG 1 32396.0 6.94 0.0300
ZM X BG 1 15950.5 3.42 0.1018
Error 8 4670.6

Within-subject* Time 3 5.00 0.0452
Time X ZM 3 18.37 0.0020
Time X BG 3 19.92 0.0016
Time X ZM X BG 3 18.57 0.0019

 

* MANOVAR (sphericity test of orthogonal components: Mauchly's Criterion = 0.07, X

= 17.8, _B = 0.0032).

82

2

Figure 1. A) Total macrozooplankton biomass response over the course of the experiment
in the presence of zebra mussels and larval bluegill. Shown are means i standard
deviation. B) Total microzooplankton abundance response over the course of the

experiment in the presence of zebra mussels and larval bluegill. Shown are means :t 1

standard deviation.

83

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84

Figure 2. A) Chlorophyll-a response over the course of the experiment in the presence of
zebra mussels and larval bluegill. Shown are means 1: standard deviation. B)
Transparency response over the course of the experiment in the presence of zebra
mussels and larval bluegill. Transparency decreases with an increasing coefficient of light

extinction. Shown are means i standard deviation.

85

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86

Chapter 4

ALTERATION OF NUTRIENT CYCLING BY ZEBRA MUSSELS

Abstract

The effects of the exotic zebra mussel (Dreissena polmogpha) on nutrient cycling in
small inland lakes was investigated in a mesocosm experiment. Dissolved and particulate
nutrients and sedimented material were monitored for two weeks in 1,500 L limnocorrals
stocked with zebra mussels and/or larval bluegill sunfish (Lepomis macrochirus). Zebra
mussels reduced concentrations of sestonic carbon, nitrogen and phosphorus and
increased concentrations of total dissolved phosphorus (TDP) and soluble reactive
phosphorus (SRP). Zebra mussels did not affect the relative abundance of nutrients as
measured by dissolved inorganic nitrogen (DIN):TDP, sestonic stestonic P, and
Sestonic C:Sestonic N. The ratio of D1N+sestonic Nztotal P increased in non-zebra
mussel treatments following rain events that contained high levels of nitrogen, reflecting
the flow of introduced nitrogen to periphyton due to removal of phytoplankton by zebra
mussel filtration. Sedimented material in zebra mussel treatments varied over time in
abundance and chemical composition. Alteration of foodweb structure by zebra mussels
had a more important impact on nutrient cycling than nutrient regeneration or

biodeposition.

87

Introduction

Small inland lakes in the lower peninsula of Michigan are currently being invaded
by zebra mussels at an alarming rate due to their proximity to the Laurentian Great Lakes
and high rate of inter-lake recreational boat traffic (Buchan and Padilla 1999). Most
research of North American zebra mussel invasions, however, has been carried out in the
Laurentian Great Lakes, large rivers, and the inland Lake Oneida. Ecosystem effects of
zebra mussel invasions that have been documented in larger water bodies do not
necessarily apply to small inland lakes. Important differences between the Great Lakes,
large rivers and inland lakes exist that should affect ecosystem response. In Saginaw Bay
of Lake Huron, for example, the outer bay hosted low zebra mussel populations and
behaved independently of the inner bay, which hosted high zebra mussel populations
(F ahnenstiel et a1. 1993). In Lake Erie, the eastward expansion of the invasion could be
charted in the successive reduction of algae at stations along a transect (Nichols and
Hopkins 1993). Thus, due to their sheer size, the Great Lakes do not respond as whole
units to ecosystem perturbation. Inland lakes may respond to zebra mussel invasion more
rapidly and homogeneously. Inland lake morphometry, with a larger proportion of littoral
zone, might also affect ecosystem response, as should larger and more diverse
communities of macrophytes in the littoral zone. Large river systems such as the
Mississippi and tidal Hudson River are distinct from inland lakes because of their flow
regimes and inorganic suspended matter. Such differences might produce radically
different responses to zebra mussel invasion in inland lakes compared to previously

observed ecosystems.

88

Zebra mussels remove phytoplankton and other seston from the water column and
biodeposit much of this material to the benthos in altered forms. This benthic-pelagic
coupling, along with excretion of dissolved materials into the water column, has fueled
speculation that zebra mussels might significantly alter nutrient cycling regimes and food
webs in lakes (Ludyanskiy 1993, MacIsaac 1996). Such speculation has ranged from
simply promoting benthic communities to shifting lakes from a stable equilibrium of high
turbidity and phytoplankton domination to another stable equilibrium of clear water and
macrophyte domination (Scheffer et al. 1993).

Studies of zebra mussel biodeposition began with examinations of factors
affecting rates of pumping, filtration, clearance, and pseudofeces production (Morton
1971, Walz 1978). This knowledge has been refined through studies of the effects of
specific algal species and concentrations on biodeposition (Sprung and Rose 1988, Berg
et al. 1996), effects of clay on biodeposition (MacIsaac and Rocha 1995), in situ
biodeposition rates (Klerks et al. 1996), effects of biodeposition on contaminant cycling
and food web transfer of contaminants (Bruner et al. 1994, Klerks et al. 1997), use of
biodeposition by zebra mussels for water quality management (Reeder and Bij de Vaate
1990, 1992, Noordhuis et al. 1992), and effects of biodeposition on benthic macro-
invertebrate abundance (Thayer et al. 1997). Less common are investigations of
biodeposit composition. Ten Winkel and Davids (1982) noted that pseudofeces pellet
composition from individual mussels varied greatly in the relative abundance of algae
and the species present. Roditi et al. (1997) examined in situ zebra mussel biodeposit

composition in depth and found that sediment deposited by zebra mussels contained more

89

organic matter, more live algae, higher rates of bacterial production, and lower C:N ratios
than passively deposited material.

When large populations of zebra mussels become established in small inland
lakes, biodeposit composition might become important to the overall nutrient cycling. For
example, Mellina et al. (1995) speculated that biodeposition played a small role in the
removal of nutrients from the water column of a microcosm, but that assimilation of
nutrients into zebra mussel biomass and resuspension of biodeposits were important
processes for nutrient cycling. Holland et al. (1995) speculated that resuspension of zebra
mussel biodeposits may alter relative amounts of nutrients in the water column during
seasonal turnover events in lakes. Mixing events might differentially alter pelagic nutrient
availability due to episodic or seasonal resuspension of biodeposits. Temporal variation
in biodeposit composition might affect benthic communities differently at different times.

Observations of the alteration of dissolved nutrient concentrations by zebra
mussels include increases in Dissolved Inorganic Nitrogen (DIN, i.e. ammonium and
nitrate) seen in experimental systems, Lake Huron, and Lake Erie (Gardner et al. 1995,
Heath et al. 1995, Holland et a1. 1995, Johengen et al. 1995), and increased Total
Dissolved Phosphorus (TDP) and Soluble Reactive Phosphorus (SRP) concentrations in
experimental systems and Lake Erie (Heath et al. 1995, Holland et al. 1995, James et al.
1997, James et al. 2000). Unresolved is whether such zebra mussels can alter the relative
abundance of nutrients through regeneration.

The purpose of this study was to examine how zebra mussels affect nutrient
cycling. Sedimented material and dissolved nutrients were monitored over time in a

mesocosm experiment stocked with zebra mussels, zooplankton and larval fish. Two

90

primary questions were addressed: 1) Can zebra mussels alter the relative availability of
potentially limiting nutrients as indicated by pelagic nutrient ratios? 2) How does

biodeposit content change as the mussels deplete available food?

Methods

This study was conducted in an artificial pond 29 m in diameter and 1.8-m deep at
the Experimental Pond Facility of the W. K. Kellogg Biological Station, Michigan State
University, Hickory Comers, MI, USA. The pond did not have zebra mussels but
contained redear sunfish (L. microlophus) that maintained a zooplankton community
dominated by small-bodied cladocerans and copepods. Twelve 1.5-m deep, l-m2
limnocorrals holding 1,500 L of water each were installed, filled with pond water
screened through 100-um mesh, and stocked with natural densities of zooplankton
collected from the pond with a lOO-um net and mixed in a large container before
removing aliquots for each replicate. Limnocorrals were arranged in two parallel rows
and the treatments with zebra mussels (ZM), bluegill (BG), ZM+BG, and control
assigned in a stratified random fashion with one replicate of each treatment at the end of a
row.

Zebra mussels collected from Gull Lake were held in a tank with continuous

flow-through of Gull Lake water, placed on Plexiglass sheets, and allowed to attach for

two weeks. Approximately 2 g AFDM (soft tissue) / m2 of mussels represented by size-
structured populations of 200 mussels were used in each treatment: 25 of 20-30 mm, 75

of 10-20 mm, and 100 of <10-mm (Young et al. 1996). Sediment traps were attached to

91

vertically oriented acrylic sheets with attached zebra mussels. The sheets were then
suspended in limnocorrals 50 cm below the water surface. Control treatments received
acrylic sheets without mussels and sediment traps.

Hatchling bluegill collected directly from nests in Warner Lake were reared in the
lab for several days until reaching the swimming stage. Fish were then fed brine shrimp
nauplii. Fish (standard length 5.5 :1: 0.16 mm, estimated by measuring 20 fish not used in
the experiment but from the same cohort) that consumed nauplii were selected from one
tank for use in limnocorrals. Twenty larval fish were used per treatment. Control
treatments received no fish. Bluegill and zebra mussels were added to the mesocosms at
three days after zooplankton had been added. Samples were taken from 9 July to 22 July

1999. Temperature, 02’ pH, and specific conductance were measured with a YSI

multisensor. Measurements and water sampling were performed prior to collecting
sedimented material to prevent accidentally resuspended sedimented material from
contaminating water samples.

Dissolved organic carbon (DOC) was measured with a high-temperature
combustion DOC analyzer. Nitrate was measured with an ion chromatograph (detection

limit of ca. 15 rig/L NO3-N). Ammonium, particulate phosphorus (PP), total dissolved

phosphorus (TDP), soluble reactive phosphorus (SRP), and phosphorus content of
sedimented material were analyzed colorimetrically, after persulfate oxidation in the case
of total P (Langrrer and Hendrix 1982, Aminot et al. 1997). Dissolved calcium was
measured with a flame atomic absorption spectrophotometer.

Seston was collected on Gelman AE glass-fiber filters. Sestonic carbon, carbon

content of sedimented material, sestonic nitrogen and nitrogen content of sedimented

92

material were measured with a CHN analyzer. Inorganic carbonates were removed from
seston samples on filters by acid fumigation prior to CHN analysis. Sestonic CaCO; and
associated P were measured by extraction of the material on AE filters overnight in dilute

I—INO3 (0.5%), followed by analysis of dissolved Ca and SRP in the extract. Total

sestonic phosphorus (TP) was calculated by summing PP and TDP. Sestonic chl a and chl
a content of sediment were measured by flourometry (W elschmeyer 1994).

Measures were taken to prevent the spread of zebra mussels to other ponds at the
facility (Reid et al. 1993). Mesocosms were covered with gray fiberglass window-screen
to prevent animals from entering the limnocorrals and transporting water potentially
infested with zebra mussel veligers. After the experiment, Rotenone and chlorine were
added to limnocorrals before bags were cut away from frames. The entire pond was
drained directly to soil and allowed to dry before being refilled.

Data from days 3 through 13 were used for time series analysis with Repeated
Measures Analysis of Variance (rmANOVA) using SAS for Windows v.8 (von Ende
I993). Between-subjects effects refer to differences between the means of treatments for
the time period and within-subject effects refer to time x treatment interactions.
Greenhouse-Geisser (G-G) corrected rmANOVA results were used for within-subject
effects if a sphericity test on orthogonal components was rejected, otherwise Multivariate

Repeated Measures AN OVA (MANOVAR) results were used (SAS Institute 1999).

93

Results

Algal biomass and nutrients in the water column - Chlorophyll _a fell steadily in the
control treatment from a mean of 2.70 ug/L to a mean of 1.21 rig/L (see Chapter 3). The
concentration of chl _a fell to about 0.5 rig/L and stabilized in zebra mussel treatments
after day 3 (between subjects P < 0.0001, F = 415.39). Transparency of the water column
increased in treaments with zebra mussels (between subjects 2 = 0.0001, F = 46.27).

Prior to the addition of fish and zebra mussels, mean concentrations (i SD) of
sestonic C, N, and P among all limnocorrals were 0.64 (d: 0.17) mg/L, 23.3 (:1: 16.5) rig/L,
and 3.47 (:1: 0.96) rig/L, respectively. Zebra mussels reduced sestonic carbon (between
subjects B = 0.0054, F = 14.29, Fig. 1A) and nitrogen (between subjects P < 0.0001, _F_ =
101.32, Fig 1B). Some sestonic phosphorus samples became contaminated due to the use
of P-contaminated filters and were removed fi'om statistical analysis. Subsequently time
series analysis could not be used for sestonic phosphorus or nutrients ratios including
sestonic phosphorus; instead, AN OVA was used to analyze only Day 3 data. Zebra
mussels reduced mean sestonic phosphorus on Day 3 (ANOVA P < 0.0001, E = 49.34).

Our measurements failed to demonstrate an appreciable increase in acid-soluble
calcium and phosphate isolated from seston samples. Both sestonic and dissolved calcium
dropped by day 3 in all treatments, perhaps indicating calcium precipitation. overall
temporal patterns, however, did not conform to expectations and there was no relation
between acid-soluble Ca and P in the seston.

Other nutrients showed changes over the course of the experiment but with little

or no effect of zebra mussels. Dissolved inorganic nitrogen (DIN) consisted chiefly of

94

ammonium as nitrate was not found above detectable levels. Ammonium followed a
similar temporal pattern as DOC starting at a mean of 4.5 (d: 0.4) ug/ L, peaking on day 6
at 37.4 (:1: 14.3) rig/L, and falling to 17.8 (:1: 11.9) rig/L on day 13, but zebra mussels had
no effect on ammonium. The concentration of TDP fell in non-zebra mussel treatments
from an initial experiment-wide mean of 6.1 (d: 1.3) rig/L to 5.2 (i 0.4) rig/L by day 13,
while TDP in zebra mussel treatments increased slightly (between subjects, _12 = 0.0049, _F_
= 14.74). The concentration of SRP increased in all treatments from an initial
experiment-wide mean of 1.2 (d: 0.2) rig/L to an experiment-wide mean of 1.8 (d: 0.6)
rig/L; zebra mussel treatments on day 13 had concentrations of SRP that were < 1 rig/L
greater than non-zebra mussel treatments (between subjects 2 = 0.0228, F = 7.70). Zebra
mussels had no effect on TP on Day 3. Treatments with zebra mussels had different DIN
+ Sestonic N:TP molar ratios (between subjects, E = 0.0032, F = 17.26, Fig. 2), but zebra
mussels did not significantly affect relative amounts of nutrients as measured by molar
ratios of DIN:TDP, Particulate N:Particulate P (Day 3 only), and Particulate CzParticulate
N. Dissolved organic carbon (DOC) increased from a mean (SD) of 7.4 (:1: 0.7) rig/L
across all replicates on day 0 to 8.5 (i 0.8) rig/L on day 9 before falling back to 7.5 (i:
0.8) rig/L on day 13 (within-subjects Time G-G P} 0.0091, F = 6.23), but zebra mussels
had no effect on DOC concentrations. Mean (d: SD) concentrations of dissolved calcium
among all limnocorrals prior to the addition of fish and zebra mussels was 51.3 (i 1.3)
mg/L. Mean dissolved calcium levels dropped in the control and BG treatments (by 5%
and 2.5%, respectively) and rose slightly in the ZM and ZM+BG treatments (by 0.2% and
0.3% respectively) between day 0 and day 3, but returned to previous levels by the end of

the experiment.

95

Sedimented Material- Zebra mussels produced copious amounts of sediment in the first

 

three days of the experiment, but then reduced their sedimentation rate to a constant level
slightly above that in control and BG treatments (Fig. 3A). Treatments with zebra
mussels had higher rates of C, N, and P sedimentation (between subjects for carbon _12 <
0.0001, F = 276.2; for nitrogen E < 0.0001, F = 113.9; for phosphorus P < 0.0001, 1;“ =
981.0) following temporal patterns essentially the same as in Figure 3A. Sedimentation
rates of chl a were more complex with zebra mussel x bluegill effects, i.e. there was more
chl a in sediment from ZM+BG treatments (between subjects ZM P < 0.001, E = 241.5;
BG P = 0.0097, F = 11.38; ZM x BG _B = 0.0168, F = 9.07; Fig. 3B).

Analysis of nutrient ratios within sedimented material (C:N, N:P, and C:Chl-a)
using repeated measures ANOVA on the entire time period showed varying influences of
zebra mussels (between subjects for C:N P = 0.0550, F = 5.04; for N:P _E = 0.6658, E=
0.20; for C:Chl a P < 0.001, 1;“ = 61.03). Sediment sampling dates represent collection of
material that sedimented over the time interval since the previous sampling, with the first
sampling representing days 0-3. Temporal patterns indicated that sedimented material
collected on day 3 consisted of material substantially different than sediment collected on
days 6, 9 and 13 (Fig. 4). Sedimented material in zebra mussel treatments had higher C:N
on day 3 (within-subjects effect, time x ZM P = 0.0009, F = 7.75), lower C:Chl a from
day 3 onward (within-subjects effect, time x ZM _P = 0.0041, F = 5.76), but no
discemable differences in N:P throughout the experiment (within-subjects effect,

MANOVA, time x ZM P = 0.0792, F = 3.75).

96

Discussion

Nutrients in the water column: Observed nutrient cycling effects by zebra mussels were
generally consistent with previous studies. Reduction in suspended carbon, nitrogen, and
phosphorus concentrations was attributable to the general filtering activity of zebra
mussels. Increased TDP and SRP concentrations were attributable to zebra mussel
excretion and were consistent with observations in previous experimental systems as well
as Lake Erie (Heath et al. 1991, Holland et al. 1995, James et al. 1997, James et al. 2000).
Fahnenstiel et al. (1995) and Johengen et al. (1995) reported reductions in total
phosphorus in areas of Lake Huron with large zebra mussel populations but no effect on
TP was observed in the present experiment. Incorporation of phosphorus into zebra
mussel biomass and increased SRP availability through remineralization may represent
the most important alterations of nutrient availability by zebra mussels for inland lakes,
considering the high frequency of phosphorus limitation of algal growth in inland lakes
(W etzel 2001).

Zebra mussels evidently prevented a change in DIN +PN:TP ratios that was
observed in non-zebra mussel treatments on days 3 and 6. The temporal patterns of
DIN+PN:TP ratios in all treatments, including the lower ratios seen in treatments with
zebra mussels, and the pattern of changes in transparency might be partially explained by
the precipitation of calcium carbonate from the water column. Through this mechanism,
the reduction of algae and thus photosynthetic activity in zebra mussel treatments reduces

the rate of calcium carbonate (CaCO3) precipitation, the whiting effect (light attenuation)

97

of such precipitation, and the removal of dissolved phosphorus through adsorption (Kelts
and H511 1978). Thus the calcium carbonate explanation was not supported.

A possible explanation for the temporal pattern of DIN+PN:TP exists with rain
entering the enclosures. The temporal pattern of DIN+PN:TP was driven by DIN+PN,
which itself was largely driven by sestonic nitrogen (Fig. 38). Any explanation must then
account for greater nitrogen present in living plankton in treatments without zebra
mussels (BG and control treatments). Rain fell on 9 July 1999 (day 0) after initial

conditions were measured and fish and zebra mussels added (2.82 cm, 0.54 mg/L NH 4,
1.95 mg/L N03, Kellogg Biological Station #M126, National Atmospheric Deposition

Program 2001). Each 1m x 1m limnocorral therefore received an infusion of 24.3 mg of
DIN in 28.2 L of water, raising the concentration of DIN in the entire 1,500 L mesocosm
by approximately 16 rig/L. Rain also fell between days 6 and 9 (0.27 cm, 0.43 mg/L

NH 4, 3.64 mg/L N03) which raised the concentration of DIN by approximately 2 rig/L.

In addition, the mean annual concentration of dissolved organic nitrogen in Kalamazoo
County precipitation is 0.43 mg/L (Rheaume 1990), and this form could add up to 50%
more nitrogen to precipitation. Thus enrichment can at least partially account for the
increase in total N seen in non-zebra mussel treatments on days 3 and 9.

Nitrogen entering the mesocosms through rain could have been taken up primarily
by phytoplankton in non-zebra mussel treatments and periphyton in zebra-mussel
treatments. Reduced phytoplankton biomass in zebra mussel treatments would have
provided less nutrient-uptake competition for periphyton. Moreover, less turbid
conditions in zebra mussel treatments would have provided more light and thus better

conditions for periphyton growth.

98

The present experimental pond system, however, may have been limited by
nitrogen, as indicated by a low initial DIN + Sestonic N:TP ratio (< 10). Although an
increase in ammonium concentration has been observed in the presence of zebra mussels
in experimental systems, Lake Huron, and Lake Erie (Gardner et al. 1995, Heath et al.
1995, Holland et a1. 1995, Johengen et al. 1995), ammonium did not increase in the
presence of zebra mussels in the present study. Rapid utilization of ammonium by
phytoplankton and periphyton within limnocorrals could have masked an increase in
ammonium regeneration produced by zebra mussels. Rapid uptake of nitrogen introduced

in rain would also be consistent with a nitrogen-limited system.

Biodeposition- Temporal patterns in the nutrient ratios of sedimenting material indicate
that sediment produced by zebra mussels in the first 3 days differed substantially from
sediment produced later. The sediment produced by zebra mussels from days 0 to 3 had
higher C:N ratios and similar N:P and C:Chl g ratios compared with the other treatments.
After day 3, nutrient ratios in sedimenting material did not differ among treatments,
except for lower C:Chl a in zebra mussel treatments, yet the zebra-mussel treatments
consistently yielded larger quantities of sediment.

Temporal patterns in the content of sedimenting material in the zebra-mussel
treatments might be explained by two mechanisms: 1) greater production of feces which
changed in composition as the seston changed, or 2) pseudofeces production during the
first three days. The initial concentration of sestonic C in limnocorrals (0.64 i 0.17 mg
C/L) was greater than a previously documented threshold for beginning pseudofeces

production (~O.2 mg/L) but less than the threshold for reduced filtering rates (incipient

99

limiting concentration, ~2.0 mg C/L) (Walz, 1978). While pseudofeces production is
expected to begin at the incipient limiting concentration for pure algal cultures,
pseudofeces are also produced in order to reject inedible inorganic or organic particles
(Sprung and Rose 1988, MacIsaac and Rocha 1995, Berg et a1. 1996). Thus zebra

mussels could have been producing pseudofeces early in the experiment.

Conclusions- Indirect effects on nutrient cycling by zebra mussels through alteration of
foodweb structure were more important than direct effects via dissolved nutrient
regeneration or biodeposition in this experiment. Specifically, the increase in DIN +
SestonichTP ratios in non-zebra mussel treatments was evidence that removal of
phytoplankton by zebra mussels allowed nitrogen naturally introduced by rainfall to be
taken up by periphyton. Biodeposits produced by zebra mussels varied substantially over
short time periods in terms of relative nutrient composition. In natural systems temporal
change in biodeposit composition might occur both due to seston availability and seston
content including phytoplankton abundance or community structure, and seasonal
changes in biodeposit composition may occur in response to phytoplankton succession. A
trophic cascade can explain why greater amounts of chl-a were found in sedimented
material in enclosures with zebra mussels and fish, compared to sedimented material in
enclosures with zebra mussels and no fish. Models of zebra mussel alteration of nutrient
cycling including stoichiometric relationships in natural systems must include flow paths

to benthic communities produced by indirect food-web effects.

100

Acknowledgements

I thank Sue Raikow and Stefanie Whitmire for assistance in the field. Alan
Wilson helped build limnocorrals and prepare the pond. Stephen Hamilton, Gary
Mittelbach, Alan Tessier, and Orlando Samelle provided helpful comments about this
manuscript. Everyone who attended KBS brown bag and EEBB colloquium seminars
provided useful feedback. Jessica Rettig, Chris Steiner, and Laura Broughton kindly
answered methodological questions. Madhav Machavaram, Nina Consolatti, and John
Gorentz provided invaluable logistic support. This research was based in part upon work
supported by the US. National Science Foundation (DEB-9701714), the KBS Graduate
Research Training Group (RTG) funded by the US. National Science Foundation (DBI-
9602252), and by grants from the Environment Now Fund of The Kalamazoo

Community Foundation.

101

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105

Figure 1. Nutrient responses over the course of the experiment in the presence of zebra
mussels and larval bluegill. Shown are means i 1 standard deviation. A) Sestonic carbon.

B) Sestonic nitrogen.

106

Sestonic C (mg/L)

Sestonic N (pg/L)

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107

Figure 2. A) D1N+sestonic N:Tota1 P molar ratio, uncontaminated replicates. B)
DIN+sestonic N:Tota1 P molar ratio, all days. Some replicates on days 6, 9 and 13 were
slightly underestimated due to phosphorus contamination and were excluded from

statistical analysis (see text).

108

 

 

 

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109

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A) Sedimentation rate. B) Sedimentation rate of chlorophyll a.

110

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111

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ratio of sediment.

112

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115

Chapter 5

ZEBRA MUSSEL IMPACTS ON THE PHYTOPLANKTON COMMUNITIES
AND WATER CHEMISTRY OF INLAND LAKES

Abstract
A survey of 61 Michigan inland lakes spanning a wide phosphorus gradient was
conducted in 1998 and 1999 to assess the impact of zebra mussels on phytoplankton
abundance, phytoplankton community structure, lake transparency, and nutrient
concentrations. Lakes with zebra mussels had lower phytoplankton abundances as
measured by chlorophyll a concentration and total algal biovolume, differences in
phytoplankton community structure, and lower dissolved organic carbon concentrations
than lakes without zebra mussels. Lakes with and without zebra mussels were
indistinguishable in their transparencyand concentrations of ammonium and soluble
reactive phosphorus. Total phytoplankton abundance and abundance of cyanobacteria
were positively related to total phosphorus. Most differences in phytoplankton
communities were subtle but there were marked increases in the relative abundances of
the colonial cyanobacteria Microcystis at low total phosphorus concentrations (i.e., < 25
_g P/L). Recent invasion by zebra mussels may have directly contributed to Microcystis
blooms recently observed in several Michigan inland lakes. The reduction of dissolved

organic carbon may render inland lakes more susceptible to UV-B radiation.

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Introduction

Zebra mussels (Dreissena polymorpha) are notorious for their ability to disperse
widely and quickly (Johnson and Carlton 1996). Since their introduction to North
America zebra mussels have invaded the Great Lakes, the Mississippi drainage network
and other rivers, and numerous small inland lakes. Less is known about the ecological
effects of zebra mussels in North American inland lakes compared with other ecosystems
including the Laurentian Great Lakes. While important limnological differences between
inland lakes and the larger lakes and rivers that have been invaded may result in different
ecological impacts, observations in those other systems do provide a basis for hypotheses
about what may take place in smaller lakes.

Effects of the invasion on phytoplankton communities have been of primary
ecological concern due to the intense filtering activity of zebra mussels (MacIsaac 1996,
Ram and McMahon 1996). Some effects of zebra mussels on phytoplankton communities
have occurred in all invaded ecosystems while others have been system-specific.
Chlorophyll a (chl a) concentrations, for example, have been reduced following zebra
mussel invasion in the Great Lakes (MacIsaac et al. 1992, Holland 1993, Nichols and
Hopkins 1993, Fahnenstiel et a1. 1995), the Hudson River (Caraco et al. 1997), and the
inland Oneida and Hargus Lakes (Yu and Culver 2000, Idrisi et a1. 2001). Increases in
transparency have often accompanied these reductions in phytoplankton abundance
(Holland 1993, Skubinna et al. 1995, Caraco et al. 1997, Yu and Culver 2000, Idrisi et al.

2001).

117

The effects of zebra mussels on cyanobacteria and Microcystis in particular have
been of particular interest due to the undesirable nature of cyanobacterial blooms as well
as the potential toxicity of Microcystis to vertebrates including humans if ingested in
large quantities (Carmichael and Falconer 1993). Algal blooms are unslightly,
malodorous and reduce the recreational value of freshwaters. Filamentous and colonial
cyanobacteria (blue-green algae) are the most important taxa causing blooms because
they increase turbidity, enhance oxygen depletion in bottom waters, and can produce
toxins that threatens the health of livestock and humans when ingested (Fogg et al. 1973,
Paerl 1988, Chorus and Bartram, 1999). In addition, cyanobcteria are often avoided or
poorly assimilated by herbivores reducing planktonic food chain efficiency and thus can
negatively impact pelagic-based fisheries (Gilwicz and Pijanowska 1989, DeMott 1986).
Aesthetic degradation and public health threats are greatest when floating scums of algae
accumulate at the shore, and therefore scum-forming species that potentially produce
toxins, such as Microcystis aeruginosa, are the most notorious of the noxious
cyanobacteria.

Smith et al. (1998) documented a striking shift in phytoplankton dominance from
cyanobacteria to diatoms following zebra mussel invasion of the Hudson River. In
contrast, Microcystis aeruginosa blooms were observed soon after zebra mussel invasion
of Saginaw Bay in Lake Huron (Vanderploeg et al. 2001), and Aphanizomenon blooms
were observed soon after zebra mussel invasion in Oneida Lake (Horgan and Mills 1997),
although in Oneida Lake long term trends in phytoplankton community composition have
not shown consistent change following invasion (Idrisi et al. 2001). No change in the

relative abundance of dominant phytoplankton taxa was observed immediately following

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zebra mussel invasion in the nearshore waters of Lake Erie (Nichols and H0pkins 1993),
and phytoplankton community composition in offshore waters of Lake Erie was
minimally impacted during the first years of zebra mussel invasion (1989-1993,
Makarewicz et al. 1999). In 1995, however, a Microcystis bloom occurred in the western
basin of lake Erie, where zebra mussel densities are greatest (Budd et al. 2001).

The long history of research in Europe and the former Soviet Union on zebra
mussels yields little information on shifts in phytoplankton community composition
following zebra mussel invasion (Karatayev et al. 1997). One exception is the exploration
of the use of zebra mussels as bioprocessors to control effects of eutrophication including
cyanobacterial blooms (Reeder and Bij de Vaate 1990). Noordhuis et al. (1992), for
example, demonstrated the suppression of Aphanizomenon and Oscillatoria blooms
following experimental introduction of zebra mussels to a hypereutrophic pond.

Experimentation has yielded conflicting results concerning the ability of zebra
mussels to alter phytoplankton community structure. Diatoms were more heavily reduced
by zebra mussels while cyanobacteria were relatively unaffected in a short mesocosm
experiment in Lake Huron (Heath et al. 1995). Nanoplanktonic algae were selectively
removed while Microcystis was not in a microcosm experiment using Lake Huron
plankton (Lavrentyev et al. 1995). Cyanobacteria were suppressed at high zebra mussel
densities and diatoms promoted at mid-level densities in an in-situ experiment on the
Ohio River (Jack and Thorp 2000). Some experimentation has shown that zebra mussels
can cause change through selective rejection and subsequent survival of algae including
Microcystis in feces and pseudofeces. Bastviken et al. (1998), for example, demonstrated

filtering of all algae from the water column (gross clearance) by zebra mussels in

119

microcosms and alteration of phytoplankton community composition upon resuspension
of biodeposited material (net clearance), including an increase in the relative abundance
of Microcystis colonies. Vanderploeg et al. (2001) showed rejection of Microcystis in
pseudofeces depending on whether the algae strains produced the toxin microcystin. In
contrast, Baker et al. (1998) demonstrated preferential ingestion of Microcystis by zebra
mussels in the Hudson River.

In addition to Microcystis blooms following zebra mussel invasion reported in the
literature, Microcystis blooms have been observed following zebra mussel invasion in
several inland lakes in the vicinity of the Kellogg Biological Station, Michigan State
University. For example, in 1996 one of us (Hamilton) observed a Microcystis bloom in
Gull Lake, where the Kellogg Biological Station is located, two [three?] years after zebra
mussels were first observed in the lake, and similar blooms have since been observed in
nearby Gun Lake; neither lake had experienced such blooms in the recent past.

Most phytoplankton taxa become more abundant with increasing total phosphorus
concentration (Watson et al. 1997). Total phosphorus has been further demonstrated as an
excellent predictor of cyanobacterial biomass (Trimbee and Prepas 1987) and the
potential for bloom formation (Paerl 1998). The response of cyanobacterial biomass to
total phosphorus concentrations is nonlinear, however, dramatically increasing at total
phosphorus concentrations above 20-30 rig/L (Downing et al. 2001). Blooms observed in
Gull and other local inland lakes were thus surprising due to their low total phosphorus
concentrations (<25 ug/L).

Alteration of nutrient cycling by zebra mussels is also only partially understood.

Regeneration of and subsequent increases in concentrations of total dissolved phosphorus

120

(TDP), soluble reactive phosphorus (SRP), and ammonium by zebra mussels have been
demonstrated under laboratory conditions and observed in Lake Erie (Heath et al. 1991,
Holland et al. 1995, James et al. 1997, James et al. 2000). Mellina et al. (1995), however,
observed variable phosphorus trends over time with increasing zebra mussel density in
microcosms. Zebra mussel excretion of ammonium may contribute to altered
biogeochemical cycles in lakes (Lavrentyev et al. 2000). Zebra mussels have also
recently been shown to be able to directly assimilate dissolved organic carbon (DOC,
Roditi et al. 2000). Potential reduction of DOC is of particular concern because DOC
attenuates ultraviolet-b (UV-B) radiation in aquatic ecosystems, and this radiation has
been shown to detrimentally impact freshwater ecosystems by inhibiting photosynthesis
and increasing amphibian mortality (Bothwell et al. 1994, Kiesecker and Blaustein 1995,
Schindler et al. 1996).

The objectives of this study were to evaluate the effects of zebra mussels on
inland lake phytoplankton communities and nutrient availability. We examined several
questions: 1) Do lakes with zebra mussels differ from lakes without zebra mussels in
algal biomass and transparency across a range of phosphorus availability? 2) Do lakes
with zebra mussels differ from lakes without zebra mussels in phytoplankton community
composition, specifically Microcystis and other cyanobacteria? 3) Do lakes with zebra
mussels differ from lakes without zebra mussels in nutrient availability, particularly
ammonium, SRP, and DOC concentrations? A survey of 61 lakes preselected along a
total phosphorus (TP) gradient and classed by zebra mussel presence or absence was

conducted in the lower peninsula of the State of Michigan to address these hypotheses.

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Methods

In Michigan, zebra mussels were known to inhabit 150 inland lakes by the late
1990’s (Kraft and Johnson 2000, Klepinger 2000). A list of lakes infested by zebra
mussels was obtained from the Michigan Sea Grant. The Sea Grant list also included the
year zebra mussels were first sighted, and from this the age of invasion at time of
sampling was estimated. For lakes adjacent to Lake Michigan, the age of invasion was
estimated based on when zebra mussels were sighted along the shore of Lake Michigan
near the lake that was sampled, since numerous boats travel between Lake Michigan and
those lakes. Lakes to be sampled were chosen to represent the range of total phosphorus
concentrations common in Michigan Lakes, but only included lakes with pH and calcium
concentrations suitable for zebra mussels (Ramcharan et al. 1992). Lakes sampled had
similar distributions of TP levels in each category. Historical data on total phosphorus
concentrations in Michigan inland lakes were obtained from the Michigan Department of
Environmental Quality. Data on maximum depth and surface area were used to find lakes
of roughly similar morphometries with and without zebra mussels. Most lakes sampled
were thermally stratified.

Lakes were sampled in late summer (August or early September) of 1998 and
1999. Phytoplankton samples were analyzed from all lakes sampled in 1998 but from
only selected lakes sampled in 1999 with mid-range TP due to time restraints (n = 40
lakes analyzed). Late summer was chosen to maximize the potential of measuring
cyanobacterial dominance of the phytoplankton community (W etzel 2001). Samples were

taken by boat at the deepest area of the lake as indicated by a bathymetric map and depth

122

finder. Transparency was evaluated by Secchi depth. A vertical profile of temperature,
oxygen, pH, and specific conductance was measured to locate the therrnocline using a
YSI multisensor. An integrated water sample of the epilimnion was then taken with a
tube sampler and mixed in an opaque cooler. From this integrated sample, 250 ml of
water was preserved in Lugol’s fixative as a phytoplankton sample and 1 L was put on
ice for later chemical analysis.

Dissolved organic carbon (DOC) was measured with a Pt catalyzed high-
temperature-combustion C analyzer. Ammonium, particulate phosphorus (PP), total
dissolved phosphorus (TDP), and soluble reactive phosphorus (SRP) were analyzed
colorimetrically (W etzel and Likens 1991, Aminot et al. 1997, Langner and Hendrix
1982). Total phosphorus (TP) was calculated by summing PP and TDP. Dissolved
calcium was measured with a flame atomic absorption spectrophotometer. Sestonic chl a
was collected on Gelman AE glass-fiber filters and measured by fluorometry
(W elschmeyer 1994).

Phytoplankton were settled in 10 ml subsamples for several days prior to
enumeration. The circular field of the settling chamber was divided into an inner disk and
outer ring of equal areas. Plankton were examined using an inverted microscope at 160x,
400x, and 1000x. Within each of the inner and outer areas 30 random fields were
examined for a total of 60 fields per sample at each magnification. Phytoplankton were
identified to genus. Relative abundance was estimated by calculating the biovolume using
measurements of cell size at 1000x. Biovolume of Microcystis was estimated by counting
the number of grid squares occupied by colonies and estimating the number of cells per

grid square before measurement of individual cells.

123

Analysis of covariance (ANCOVA) was used to compare the following

characterlstlcs 1n lakes wrth and wrthout zebra mussels usrng TP as a covarlate: NH 4 ,

SRP, DOC, Secchi depth, chl-a, biovolume of phytoplankton, biovolume and relative
abundance (as % of total phytoplankton biovolume) of all cyanobacteria, bloom-forming
cyanobacteria (sum of Anabaena, Aphanizomenon, Microcystis. and Oscillatoria),
colonial coccoid cyanobacteria (sum of Aphanocapsa, Chroococcus. Merismopedia.
Microcystis), filamentous cyanobacteria (sum of Anabaena, Aphanizomenon, ngbya,
Oscillatoria), Microcystis. and other selected genera if correlated with multivariate
functions. Data from 1998 and 1999 were pooled for all analyses and log-transformed

(log10+ 1), including TP. Values for lakes sampled in both years were the average of

both years.

The AN COVA was preceded by a test for homogeneity of slopes (ZM*TP
interaction). If the interaction term was not significant at P_ s 0.05 we ran ANCOVA.
Principal Components Analysis (PCA using the covariance matrix), Correspondence
Analysis (CA), and Multi-dimensional Scaling (MDS using Euclidean distances) of
biovolume and relative abundance were used to evaluate alteration of phytoplankton
community composition by ordination (PCA of relative abundance excluded, Jackson
1997). Simple Linear Discriminant Analysis (LDA) was used to attempt classification of
lakes based on phytoplankton biovolume and relative abundance. Rare genera (<5% of
total biovolume in any lake or present in only one lake) were removed from the data set
used for ordination and classification. Analyses were conducted with SAS for Windows

version 8 (ANCOVA, PCA, CA), Systat version 9 (MDS), and SPSS version 10 (LDA).

124

Results

Selection _o_f I_akg- In 1998, 22 lakes without zebra mussels and 12 lakes with zebra
mussels were surveyed. In 1999, 17 lakes without zebra mussels and 21 lakes with zebra
mussels were surveyed. Ten lakes were surveyed in both years. The final data set
consisted of 33 lakes with zebra mussels and 28 lakes without zebra mussels. In the lower
peninsula of Michigan, lakes with zebra mussels are concentrated in the southeast and
southwest corners and along the western coast, and the distribution of lakes with zebra
mussels chosen in this study reflects those geographic patterns (Fig. 1). Total phosphorus
in surveyed lakes ranged from 12.4 to 112 rig/L and did not differ between lakes with and

without zebra mussels (Kolmogorov-Smirnov, D = 0.135, _13 = 0.9444, Fig. 2).

thplankton abundance m transparency- The AN COVA results showed that lakes
with zebra mussels had lower abundances of phytoplankton, as indicated by chl a
concentrations (ZM P = 0.0028, TP 2 < 0.0001, Fig. 3A). Total algal biovolume was also
lower in lakes with zebra mussels (ZM P = 0.0489, TP 13 < 0.0001, Fig. 3B). Secchi depth
decreased with increasing TP but lakes with zebra mussels did not differ in Secchi depths

from lakes without zebra mussels (ZM P = 0.1722, TP 2 < 0.0001, Fig. 3C).

Cyanobacteria- Lakes with zebra mussels did not differ from lakes without zebra mussels
with respect to biovolume of all cyanobacteria (ZM E = 0.6231, TP B = 0.0024),
biovolume of bloom-forming cyanobacteria (ZM B = 0.9430, TP 2 = 0.0087), or

biovolume of filamentous cyanobacteria (ZM _13 = 0.0680, TP B = 0.0002). Lakes with

125

zebra mussels had greater biovolumes of coccoid colonial cyanobacteria (ZM _P = 0.0403,
TP 13 = 0.4948). Lakes with zebra mussels did not differ from lakes without zebra
mussels with respect to the relative abundance of all cyanobacteria (ZM _B = 0.6889, TP _13
= 0.2734), relative abundance of bloom-forming (ZM P = 0.7399, TP B = 0.0899), or
relative abundance of filamentous cyanobacteria (ZM P = 0.0817, TP _13 = 0.0010). The
relative abundance of coccoid colonial cyanobacteria was greater in lakes with zebra
mussels (ZM B = 0.0009, TP _B = 0.0771).

The relative abundance of Microcystis did not vary linearly with total phosphorus
in lakes with or without zebra mussels precluding the use of ANCOVA (Least squares
regression P = 0.5127, P = 0.3961, respectively Fig. X). Similarly, biomass of
Microcystis did not vary linearly with total phosphorus in lakes with or without zebra
mussels (Least squares regression P = 0.2193, P = 0.7928, respectively). The relative
abundance of Microcystis, however, appeared to be greater in lakes with zebra mussels at
lower TP levels (Kolmogorov-Smirnov, D = 0.437, B = 0.0507), and this result was
clearly seen when lakes were grouped into low TP (< 25 jig/L) and high TP (>25ug/L)
classes (Two-way ANOVA, ZM, F = 5.42, P = 0.0255, Fig. 4). No relationship was
found between relative abundance of Microcystis and the time since initial observations
of zebra mussel invasion (P = 0.1584). While the abundance of phytoplankton as
indicated by total algal biovolume increased with increasing TP concentrations, the
strength of the positive relationship weakened within nested subsets of total algal
biovolume. For example, total algal biovolume, cyanobacteria, colonial cyanobacteria,
and Microcystis responded to TP with P values of < 0.0001, 0.0024, 0.4948, and 0.4835,

respectively.

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Phjfloplankton community §_tr_uc_tm- Enumerated genera are listed in Table 1. Principal
Components Analysis did not efficiently reduce the data. Component 1 of the PCA using
biovolume by genus, for example, explained only 22% of the variance and principal
component 2 explained 12%. Data reduction using Correspondence Analysis was
similarly poor. The first factor in the CA using relative abundance by genus, for example,
explained only 11% of the variance, and the second factor explained 10%. Thus
ordination employing PCA and CA failed to separate lakes with and without zebra

mussels. Multi-Dimensional Scaling also failed to separate lakes with and without zebra
mussels using genera and division biovolume (R2 = 0.76 and R2 = 0.92, respectively)

and relative abundance (R2 = 0.74 and R2 = 0.96, respectively).

Linear Discriminant Analysis using algal division biovolume and relative
abundance failed to classify lakes with and without zebra mussels (Wilk’s Lambda =
0.919, df = 5, P = 0.701, and Wilk’s Lambda = 0.914, df = 5, _I_’ = 0.669, respectively).
Linear Discriminant Analysis using genera biovolume and relative abundance succeeded
in classifying lakes with and without zebra mussels (Wilk’s Lambda = 0.188, df = 23, B =
0.005 and Wilk’s Lambda = .226, df = 23, P = 0.018, respectively, Fig. 5). Canonical
discriminant function 1 of the LDA using genera biovolume was correlated with the
biovolumes of Anabaena (R = 0.43, B = 0.0055), Merismopedia (R = -0.54, B = 0.0003),
and Scenedesmus (R = -0.71, E < 0.0001). Canonical discriminant firnction l of the LDA
using genera relative abundance was correlated with the relative abundances of Anabaena

(R = -0.32, P = 0.044), Coelospherium (R = -0.42, B = 0.0072), Merismopedia (R = 0.40,

127

R = 0.0110), Microcystis (R = 0.36, R = 0.0230) and Scenedesmus (R = -0.0.38, R =

0.0170).

Nutrients - Concentrations of DOC increased with increasing TP in lakes without zebra
mussels, but zebra mussels altered this relationship by decoupling the TP-DOC
relationship (i.e. a non-significant relationship between TP and DOC, P = 0.5086, Fig. 6).
Thus the test for homogeneity of slopes was not passed, and ANCOVA was not used

(ZM*TP P = 0.0227). Lakes with zebra mussels had lower concentrations of DOC
(ANOVA, R = 18.72, R < 0.0001). Concentrations of ammonium (NH 4+) increased with

increasing TP but lakes with zebra mussels did not differ from lakes without zebra
mussels (ZM R = 0.2819, TP R = 0.0160). Concentrations of SRP showed the same

pattern (ZM R = 0.0764, TP R < 0.0001).
Discussion

This study is the first interlake comparison of the effects of zebra mussels on
phytoplankton communities. Because the invasion of inland lakes by zebra mussels is
ongoing at the present time, the results of this study serve as a snapshot of ecological
conditions in ecosystems at or near the beginning of colonization. In the following
discussion we evaluate the impact zebra mussels have had on phytoplankton communities
and nutrient concentrations in inland lakes of Michigan, and relate these impacts to those

seen in previously invaded systems.

128

Phjaoplankton abundance- Zebra mussels have clearly reduced standing stocks of
phytoplankton in inland lakes as evidenced by the reduction of chl a and total algal
biovolume. Reduction of algal abundance thus remains a universal effect of zebra
mussels in all ecosystems they invade. Mellina et al. (1995) hypothesized that upon
invasion of inland lakes by zebra mussels, the Dillon-Rigler relationship (Dillon and
Rigler 1974) of chl _a to TP in inland lakes might be “decoupled” (a term they used
incorrectly to mean changing the shape of the relationship), but found no evidence of
such an effect in European inland lakes. The reduction of algal abundance in Michigan
inland lakes by zebra mussels has changed the Dillon-Rigler relationship, but there was
no discemable difference in the slopes of the response curves of chl a to TP in lakes with

and without zebra mussels.

Transparency- Contrary to our hypothesis we did not observe an effect of zebra mussels
on transparency. This result was surprising given the observed effect on algal abundance,
and could be due in part to the poor sensitivity of the transparency measurement method
(Secchi depth). Directly measuring light extinction with a quantum meter might have
revealed an effect of zebra mussels on transparency. Secchi depth has been used,
however, to demonstrate increased water clarity following zebra mussel invasion in other
systems (e.g. Idrisi et al. 2001). In the Hudson River resuspension of material including
clay and detritus biodeposited by zebra mussels may have offset the large reductions in
phytoplankton abundance, resulting in only a modest increase in transparency (Baker et

al. 1998). Mixing within the epilimnion might also resuspend biodeposited material in

129

inland lakes, thus contributing to turbidity despite reductions in phytoplankton

abundance.

Phytoplankton community structure- With the exception of Microcystis abundance in
low-P lakes, zebra mussels had surprisingly subtle effects on late summer phytoplankton
community composition. Several multivariate ordination techniques failed to separate
lakes into classes of zebra mussel presence and absence, suggesting little overall change
in the phytoplankton community. The magnitude of this result stands in contrast to the
tidal Hudson River, where Smith et al. (1999) demonstrated dramatic changes in the
phytoplankton community following zebra mussel invasion. Poor data reduction by the
ordination techniques employed here likely reflects the multiple controlling factors that
interact to determine differences between phytoplankton communities in lakes. Most of
the genera encountered in surveys were not discemibly different in abundance in lakes
with and without zebra mussels. None of the genera classified as more abundant in
Michigan inland lakes with zebra mussels benefited from zebra mussel invasion in the
tidal Hudson River (Smith et al. 1999). Linear Discriminant Analysis successfully
classified lakes based on genera biovolume and relative abundance. While differences in
the overall phytoplankton communities in lakes with and without zebra mussels were
small, some genera displayed differential abundances.

The relative abundance of the colonial cyanobacterium Microcystis was greater in
low-P lakes with zebra mussels. The observation that absolute abundance (as biovolume)
of Microcystis did not differ while relative abundance did suggests several possibilities.
Microcystis might have avoided or survived ingestion (including encapsulation in

pseudofeces) rather than being directly promoted in some way in the presence of zebra

130

mussels. Alternatively, in lakes with zebra mussels, Microcystis. Merismopedia, and
Scenedesmus might have had rates of production greater than other genera, enabling
them to restock their populations at a higher rate. This possibility is unlikely, however,
given that laboratory determinations of Microcystis growth rates are relatively slow
(Reynolds 1989). In addition, Baker et al. (1998) demonstrated selective feeding in zebra
mussels that included the active rejection of Scenedesmus. Thus rejection of
mucilaginous colonial cyanobacteria and Scenedesmus could account for the changes in
phytoplankton community structure observed here, with the effects being most marked at
lower algal abundances typical of lower-P lakes, where grazing pressure might be more
influential.

The response of phytoplankton community composition in the inland lakes of
Michigan to zebra mussel invasion differed from the responses reported for some other
ecosystems. In the Hudson River, for example, cyanobacteria were strongly suppressed
(Smith et al. 1999). Aphanizomenon blooms were observed shortly after invasion in
Oneida lake (Horgan and Mills 1997).

The conclusions regarding zebra mussel impacts on phytoplankton community
structure in this study are limited to late summer communities. The timing of this study
was selected because late summer is the typical period of cyanobacterial dominance in
stratified lakes, and by sampling then be we were more likely todetect the potential effect
of zebra mussels on cyanobacterial abundance. Zebra mussels could, however, affect
phytoplankton community structure (including the abundance of cyanobacteria) at other
times during the year. For example, Makarewicz et al. (1999) documented seasonal

differences in the relative abundance of cyanobacteria in the western basin of Lake Erie

131

after zebra mussel invasion, where cyanobacteria relative abundance increased in the

spring but decreased in the summer.

Dissolved Opganic Carbon- Zebra mussels appear to have reduced the concentrations of
DOC in Michigan lakes, supporting our hypothesis. This result is consistent with Roditi
et al. (2001), who demonstrated the ability of zebra mussels to absorb DOC from the
water column and argued that up to 50% of zebra mussel carbon demand could be
satisfied by the uptake of DOC. An alternative possibility is that reduced phytoplankton
abundance in the presence of zebra mussels is the proximate cause of the reduction in
DOC. The reduction of DOC in lake waters by zebra mussels may pose a significant
threat to the biota since DOC attenuates UV-B radiation in the water column (Schindler
et al. 1996). If the DOC removed by zebra mussels includes those humic substances
which contribute most to UV-B attenuation, UV-B radiation will penetrate further in to
the water column. The reduction of DOC in aquatic ecosystems has been proposed as a
more important factor regulating the exposure of aquatic ecosystems to UV-B than
further destruction of the atmospheric ozone layer (Williamson et al. 1996). It is possible,
however, zebra mussels might be reducing the concentration of relatively labile forms of
DOC that contribute little to UV-B attenuation.

The reduction of DOC concentrations in lakes with zebra mussels was not
consistent with the mesocosm experiment (see chapter 4), where zebra mussels did not
reduce DOC. Resolution of this discrepancy between studies is difficult due to the
heterogeneous composition of what is collectively measured as DOC. To speculate, the

pond where the mesocosm experiment was conducted had many macrophytes and large

132

amounts of macrophytic detritus relative to total pond volume. Thus the pond could have

been dominated by DOC derived from macrophyte tissues, rather than DOC derived from
phytoplankton. Changes in phytoplankton abundance within limnocorrals would not have
affected the chief source of DOC in the pond. Lakes, in contrast, generally had much less
macrophyte biomass relative to total volume, and so reductions in phytoplankton

abundance could have resulted in reduced sources of DOC.

Nitrogen m Phosphorus - No effect of zebra mussels on the concentration of ammonium
was observed, contrary to expectations. This result contradicts experimental observations
of ammonium regeneration by zebra mussels (e.g. Heath et al. 1991, James et al. 1997).
Lavrentyev et a1. (2000) demonstrated increased nitrification rates in the presence of
zebra mussels in experimental systems with sediment cores. Loss of ammonium through
nitrification might account for the lack of an increase in ammonium in inland lakes with
zebra mussels.

No effect of zebra mussels on the concentration of SRP was observed, contrary to
expectations. This result corroborates recent observations of the inland Oneida Lake
(Idrisi et al. 2001), although increases in SRP concentration have been observed in both
the Hudson River and Lake Erie following zebra mussel invasion (Holland et al. 1995,
Caraco et a1 1997). This result illustrates that inland lakes respond to zebra mussel
invasion differently than do other ecosystems. Since Michigan inland lakes are
commonly phosphorus limited, SRP regenerated by zebra mussels might be taken up

quickly by phytoplankton.

133

Magritude of zebra mussel effect on phytoplankton community structur - A potential
explanation for the subtle effect of zebra mussels on overall phytoplankton community
composition is simply that zebra mussel populations are too small to have a large effect.
Mellina et al. (1995), for example, suspected that zebra mussel populations in European
lakes might be too small to alter the TP-chlorophyll relationship. Physical factors present
in inland lakes that are not present in the Great Lakes or large rivers may contribute to
limitation of zebra mussel population sizes (Karateyev et al. 1998). Freezing, for
example, kills zebra mussels and inland lakes commonly freeze over entirely. Thus the
shallowest areas of inland lake littoral zones may be extirpated of zebra mussels in
winter. Soft substrates limit zebra mussel attachment and are also common in inland
lakes. In addition, oxygen depletion may be common in deeper waters during summer
stratification, and under ice during winter.

The temporal dynamics of zebra mussel populations in lakes we surveyed are not
known. In addition, the date of first sighting of zebra mussels in a particular lake is a
questionable indicator of the time of colonization. If zebra mussel populations in North
American lakes behave as populations have in European lakes, an initial spike in
population size may occur after invasion with the population size subsequently stabilizing
at a much lower level (Mackie and Schloesser 1996, Karatayev et al. 1997). Thus lakes
we surveyed could have had very different population sizes even if they have similar
habitat availability. If large populations are required to exert biologically significant
effects on phytoplankton community composition, then the greatest potential for such
change may occur only shortly after invasion. This may explain why cyanobacterial

blooms were observed in inland lakes shortly after invasion, but not observed at later

134

times, including perhaps during this survey. From a strictly logical perspective,
Microcystis blooms observed in two inland lakes prior to this study could have been
simply coincidental with zebra mussel invasion and caused by other factors.

Lastly, the response of Microcystis abundance to zebra mussel invasion may be
dependent on the alga’s toxicity (microcystin production) and/or colony size. In a
mesocosm experiment, zebra mussels reduced the relative abundance of unicellular
cultured Microcystis while colonies of wild Microcystis increased in relative abundance,
but only after resuspension of biodeposited material (Bastviken et al. 1998). Zebra
mussels displayed lower feeding rates when toxic strains of Microcystis were present and
more readily rejected large colonies in microcosm experiments (Vanderploeg et al. 2001).
Thus if small colonies of Microcystis that do not produce toxins are present in inland

lakes, zebra mussels may not not actively reject them.

135

Acknowledgements

We thank Sue Raikow and Jessie Davis for assistance in the field. Gary
Mittelbach and Alan Tessier provided guidance and helpful comments about this
manuscript. Everyone who attended KBS brown bag and EEBB colloquium seminars
provided useful feedback. R. Jan Stevenson, Tara Darcy, Natalie Dubois, and Laura
Broughton kindly answered methodological questions. Madhav Machavaram, Nina
Consolatti, and John Gorentz provided invaluable logistic support. This research was
based in part upon work supported by the US. National Science Foundation (DEB-
9701714), the KBS Graduate Research Training Group (RTG) funded by the US.
National Science Foundation (DBI-9602252), and by grants fi'om the Environment Now

Fund of The Kalamazoo Community Foundation.

136

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140

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141

Table 1. Dominant phytoplankton genera encountered during surveys of Michigan Lakes.

 

Anabaena
Aphanizomenon
Aphanocapsa
Binuclearia
Ceratium
Chlamydomonas
Chrococcus
Closterium
Coelosphaerium
Dinobryon
Fragilaria
Glenodinium
Gleobotrys
Lyngbya
Melosira
Merismopedia
Microcystis
Mougeotia
Navicula
Oocystis
Oscillatoria
Pediastrum
Rhodomonas
Scenedesmus
Selenastrum
Staurastrum
Tetraedron
Ulothrix

142

Figure l. Lakes surveyed in the lower peninsula of the State of Michigan. Black dots
indicate lakes invaded by zebra mussels. Grey dots indicate lakes free of zebra mussels at

the time of surveying.

143

 

 

 

144

Figure 2. Total phosphorus (TP) concentrations in lakes surveyed in 1998 and 1999. Gray
bars are lakes without zebra mussels at the time of the survey and black bars are lakes

with zebra mussels.

145

Number of lakes

 

Nw-BMONQOOC

  

l
:1 ’1
'l

l

l

H ZM (-)
l ZM (+)

l 7’ >1
3 j ‘ 1 I I 1’1 ‘
. . 2 ’-.‘ , . - ti

1 - 1 ii

15 25 35 45 55 65 75 85 95 105115

Total phosphorus (pg/L)

146

Figure 3. Effect of zebra mussels on phytoplankton abundance and transparency in inland
lakes of Michigan. Note log scales. Open circles and dashed lines represent lakes without
zebra mussels, closed circles and solid lines represent lakes with zebra mussels. A)
Chlorophyll a concentrations (11 = 61). Lakes without zebra mussels: y = 1.148x — 0.762,
R = 0.0003; lakes with zebra mussels: y = 1.407x — 1.352, R < 0.0001. B) Total algal
biovolume (n = 40). Lakes without zebra mussels: y = 2.329x + 3.161, R = 0.0021; lakes
with zebra mussels: y = 2.554x + 2.361, R = 0.0142. C) Secchi depth (n = 61). Lakes
without zebra mussels: y = -0.925 + 1.681, R < 0.0001; lakes with zebra mussels: y = -

0.893x + 1.693, P < 0.0001.

147

Total Algal Biovolume (um3/ml)

I IIILUJ

I llIIIJL

Chlorophyll-a (pg/L)

I

 

 

 

100,000,000

10,000,000

1,000,000

100,000

1 I I I I l I I

100

 

200

 

Total Phosphorus (pg/L)

148

200

Figure 3. Continued.

149

 

 

 

as soon Eooom

Total Phosphorus (rig/L)

150

Figure 4. Effect of zebra mussels on Microcystis relative abundance and transparency in
inland lakes of Michigan. Note log scales. Open circles and dashed lines represent lakes

without zebra mussels, closed circles and solid lines represent lakes with zebra mussels.

151

 

I lj TTV

 

=—-Hd_

0
O

1

— l—udu—-4_ _ -—_.__d-
O 1

1

082965 3 3.3838: ..\c wB

200

100

10

log TP (rig/L)

152

Figure 5. Effect of zebra mussels on the relative abundance of Microcystis in lakes
grouped into low TP (< 25 rig/L) and high TP (> 25 rig/L) classes (Two-way ANOVA,

ZM, F = 5.42, R = 0.0255). Error bars are standard error.

153

Percent Microcystis

40-

30‘

20-

104

 

 

10 7

ii

 

+

Dreissena
(TP < 25)

154

- +
Dreissena
(TP > 25)

Figure 6. Simple Linear Discriminant Analysis of lakes with and without zebra mussels
using log-transformed biovolumes at the genus level (eigenvalue = 4.323, canonical

correlation = 0.901).

155

Number of lakes

10-

 

 

Canonical discriminant function 1

156

Figure 7. Effect of zebra mussels on Dissolved Organic Carbon (DOC) concentrations in
lakes surveyed in 1999. Open circles and dashed line represent lakes without zebra
mussels. Closed circles represent lakes with zebra mussels. Note log scale. Lakes without

zebra mussels: y = 0.508x + 0.228, R = 0.0008. Lakes with zebra mussels: R = 0.5086.

157

Dissolved Organic Carbon (mg/L)

20

10

 

 

Q OZM(-)
. OZM(+)
I
/
I
/
09,/’
. 08 o” o
_ , .
_ 0' ’00 o
_ ,0
- I.” 00 O
o
-0 . . .
: -' '
‘ o
10 100 200

Total Phosphorus (ug/L)

158

Chapter 6

ZEBRA MUSSEL EFFECTS ON MICROZOOPLANKTON FECUNDITY

Abstract
To assess how zebra mussels might suppress microzooplankton populations via either
predation or competition for food resources, microzooplankton fecundity was
investigated as part of a larger zebra mussel experiment. Twelve 49,000 L mesocosms
were installed in Gull Lake in the summer of 2001 and stocked with zebra mussels to
form a gradient of mussel density. If microzooplankton were predator-limited rather than
resource-limited then egg ratios would not rise if population size rises in low zebra
mussel treatments. Total phosphorus, chl a, Polyarthra abundance, Keratella abundance,
nauplii abundance, Polyarthra egg ratio, and Keratella egg ratio, and Keratella and
Polyarthra population growth rates were evaluated. No variables were related to zebra
mussel density during the experiment (using mean values of days 6 through 19). The
addition of phosphorus to the experiment most likely caused an increase in algal
abundance which may have limited the potential to evaluate interactions between
microzooplankton and zebra mussels. This experiment was not able to demonstrate an
effect of zebra mussels on microzooplankton or evaluate the relative importance of

competition and predation.

159

Introduction

Direct effects of zebra mussels (Dreissena polymorpha) on North American
aquatic ecosystems are well known. As filter feeders, zebra mussels remove
phytoplankton, microzooplankton and other particles from the water column, decreasing
the standing stock of algae (Caraco et al. 1997, Idrisi et al. 2001) and microzooplankton
(MacIsaac et al. 1995, Pace et al. 1998) and increasing water transparency (Holland 1993,
Skubinna et al. 1995). Particulate waste as feces and pseudofeces is deposited,
transferring energy and nutrients from the water column to the benthos in an example of
pelagic-benthic coupling (Roditi et al. 1997, Ackermen et al. 2001). Epizootic
colonization of, and competition for food with, native unionid mussels contributes to
their continuing demise (Ricciardi and Rasmussen 1999, Strayer and Malcom 2001).
New benthic habitats are created by both living and dead zebra mussel shells which can
benefit some benthic macroinvertebrates (Wisenden and Bailey 1995, Botts et al. 1996).

Less well understood are indirect food web interactions involving zebra mussels.
While consumption of microzooplankton has been demonstrated (Shevtsova et al. 1986,
MacIsaac et a1. 1991, MacIsaac et al. 1995), the relative strengths of impacts via
predation of microzooplankton compared to impacts via competition for shared food
resources (phytoplankton) on microzooplankton are not known. Observations of reduced
microzooplankton population sizes following zebra mussel invasion have identified
predation as the likely cause, but have not ruled out exploitative competition for food as a
contributing factor (MacIsaac et al. 1995, Pace et al. 1998). Jack and Thorp (2000)
observed reduced rotifer abundances without reductions in egg ratios in an ip _sltp river

enclosure experiment. Daphnia fecundity has been shown to be negatively affected by

160

reduced phytoplankton abundance following zebra mussel invasion of Oneida Lake
(Horgan and Mills 1999).

This purpose of this study was to examine the effects of zebra mussels on the
reproductive rates of Polyarthra and Keratella, two rotifers which are commonly found in
North American freshwater ecosystems susceptible to zebra mussel invasion, and which
are food for larval fish (see chapter 3). This study was part of a larger experiment using a
gradient of mussel density in mesocosms arrayed in an meso-oligotrophic inland lake.
Although zebra mussels were added to mesocosms, this was a zebra mussel removal
experiment because Gull Lake, where the experiment was located, is infested with zebra
mussels at densities near the upper end of those stocked in the mesocosms. Thus
mesocosms with zebra mussels stocked at densities lower than that present in the lake
might recover from zebra mussel impact over time. We hypothesized that the abundance
of both phytoplankton and rotifers would increase with decreasing mussel density. If
competition for phytoplankton is an important interaction between zebra mussels and
microzooplankton, egg ratios of microzooplankton should also increase with decreasing
mussel density. If predation by zebra mussels is the primary interaction, then with
reduced mussel density the microzooplankton egg ratios should remain constant as
microzooplankton population size rises, at least until food resources become limiting to

the growing population.

161

Methods

An array of 12 mesocosms was deployed in Gull Lake, Michigan, USA offshore
of the Kellogg Biological Station, Michigan State University from 5 July 2001 to 5
August 2001. Gull lake is a meso-oligotrophic inland lake that is 882 ha in area with a
maximum depth of 31 m. Zebra mussels were first observed in Gull Lake in 1994, and it

was one of the first in the region to be infested (Klepinger 2000). A previous survey
determined a mean zebra mussel density in Gull Lake of 6.4 g dry mass / m2 (upper 95 %

density = 10g dry mass / m2, 0. Samelle, unpublished data).

Each mesocosm measured 2.5 m in diameter, ~10 m in depth, with a surface area

of 4.9 m2, and a volume of ~49,000 L. Mesocosms were constructed of 8 mil
polyethylene tubes manufactured by Greentek. We folded the tube back onto itself to
form a cuff at the top of the tube to reinforce the attachment of the tube to a supporting
ring and to prevent holes from forming at the water surface. The mesocosms were left
uncovered, not stocked with fish, and filled with Gull Lake water filtered through a 100-
nm net. Mesocosms were then stocked with natural densities of macrozooplankton
collected from the lake with a 100-um net and mixed in a large container before
removing aliquots for each replicate. Mesocosms were attached to a floating platform
constructed using the EZ-Dock modular dock system in a line oriented parallel to the
shore (north-south). The platform was located about 100-m offshore and anchored in
place with cement blocks.

Zebra mussels were collected from Gull Lake and allowed to attach to 27 x 36 cm

PVC sheets within tanks fed by a continuous flow-through of Gull Lake water for two

162

weeks. Feces and pseudofeces were periodically cleaned from tanks. We used mussels

20-mm in length. Each mesocosm was stocked with 12 Plexiglass sheets attached to each
other with string and suspended vertically from 50-cm below the water surface to a depth
of 7m. One mesocosm received dummy sheets that had no mussels. The other mesocosms

were stocked to form a gradient of mussel density with the following densities in g dry

mass/m2 (total number in mesocosm): 6.2 (1260), 5.7 (1152), 5.2 (1056), 4.7 (960), 4.2
(840), 3.7 (744), 3.1 (636), 2.1 (432), 1.1 (216), 0.5 (108), 0.3 (60). The health and
activity of zebra mussels were evaluated using a submersible video camera. Zebra mussel
mortality was enumerated by a technician.

Phosphorus was periodically added to mesocosms to compensate for loss of
phosphorus to uptake by periphyton and isolation from benthic sources. Chlorophyll a
(chl a) was measured by fluorometry (Welschmeyer 1994). Microzooplankton was
sampled approximately once a week during the experiment as part of the standard
sampling regime. A tube sampler was used to collect an integrated sample of the water
column to a depth of 8m. From this sample 10 L of water was passed through a 24-um
mesh sieve to collect rotifers, veligers and nauplii. Microzooplankton samples were
preserved in a 1% Glutaraldehyde solution. Samples were settled in a graduated cylinder
overnight to separate microzooplankton, which sank, from Microcystis colonies, which
floated. Microcystis was then collected and stored in a separate container. The
supernatant was discarded and the settled microzooplankton collected and placed in a
phytoplankton settling chamber overnight. Use of the phytoplankton settling chamber

allowed inspection of microzooplankton samples using an inverted microscope. The

163

entire sample was enumerated including counts of eggs attached to Keratella and
Polyarthra.

The analysis of zebra mussel mortality was used to determine when to terminate
the experiment. Analysis of the experiment was limited to those days when zebra mussel
mortality was at a minimum. Experimental effects were analyzed by averaging response
variables following initial conditions and looking for relationships with zebra mussel
density recalculated based on mortality data. The following variables were analyzed:
total phosphorus, chl a, Polyarthra abundance, Keratella abundance, nauplii abundance,
Polyarthra egg ratio, and Keratella egg ratio, and Keratella and Polyarthra population

growth rates.

Results

Zebra mussels experienced high rates of mortality by the end of experiment. After five
days, 21.5% of mussels had died on average within each enclosure, but most of the
mortality occurred after day 18 (Fig. 1). Microzooplankton population size, egg ratios,
and population growth rates were evaluated up to day 19. The sheets supporting zebra
mussels fell off in the enclosure stocked with 60 mussels, and although it was retrieved
and reinstalled soon thereafter only a few mussels remained necessitating its exclusion
from analysis.

The community of microzooplankton included the rotifers Keratella cochlearis,

Polyarthra vulgaris, Monosgla, spp., Lecane spp., Sychaeta spp., copepod nauplii and

 

zebra mussel veligers. Neither total phosphorus, chl _a, nor any of the microzooplankton

164

variables was related to zebra mussel density on day 0, or during the experiment (using
mean values of days 6 through 19). Considering the lack of a treatment effect, all
treatments were averaged together. Comparing these experiment-wide means between
initial conditions and mean of days 6-19, total phosphorus and chl-a increased (ANOVA,
P < 0.0001 and P = 0.0009, respectively, Table 1). Keratella abundance increased while
Keratella egg ratio did not change (ANOVA, P < 0.0001 and P = 0.1207, respectively).
Polyarthra abundance and egg ratio decreased (ANOVA, P < 0.0001 and P < 0.0001,
respectively). Nauplii abundance decreased (ANOVA, P < 0.0001). Zebra mussels had no
effect on the population growth rates of either Keratella or Polyarthra (0.021 ind/day and

—0.024 ind/day, respectively).

Discussion

This experiment was not able to demonstrate an effect of zebra mussels on
microzooplankton or evaluate the relative importance of competition and predation. The
addition of phosphorus to the experiment most likely caused the very modest increase in
algal abundance as measured by chlorophyll-a. The increase in algal abundance possibly
limited the potential to evaluate the effects of competition between microzooplankton and
zebra mussels. The increase in algal abundance may also have caused an increase in
Keratella abundance, again hampering attempts to evaluate hypotheses. The reduction of .
Polyarthra in all treatments was unexpected and may have reflected an experimental

effect of the enclosures. A similar drop in Polyarthra abundance was observed in a

165

smaller mesocosm experiment (see chapter 3). This particular rotifer may suffer damage
during collection and transfer (Alan Tessier, pers. com).

I attribute the failure of this experiment to the difficulty in balancing loss of
phosphorus in mesocosm experiments with intentional phosphorus additions. An
increase in algal abundance reduced competition for food resources. A lack of effect of
zebra mussel consumption of micr0200plankton at high zebra mussel densities was still
surprising, however. Such a lack of pattern may indicate that consumption of
microzooplankton by zebra mussels may not be important enough to affect population
sizes in pelagic systems. This conclusion, however, would contradict the results and
conclusions of the smaller mesocosm experiment described in Chapter 3. If the lack of
effect of zebra mussels on microzooplankton population sizes observed in this
experiment is real, and not due to methodological issues, the effects seen in the smaller
mesocosm experiment might be due to forcing pelagic zooplankton populations into
close proximity with benthic consumers. Thus the scale of mesocosm experimentation
becomes critical when attempting to evaluate the food web and ecosystem effects of
zebra mussels. This is speculation, however, due to the unexpected response of algae in

this experiment.

166

Acknowledgements

The ZMEX project would not have been possible were it not for the daily efforts of Justin
Chiotti, Elizabeth Milroy and Allison Schuerer who collected and reared mussels, built
bags, assembled the dock, collected samples and made sure the whole thing didn't float
away. Thanks also to Alan Wilson and Carrie Scheele. This project was funded by
Michigan Sea Grant (project number R/NIS-4) and the Environment Now Fund of the

Kalamazoo Community Foundation.

167

Literature cited

Ackerrnan, J. D., M. R. Loewen, and P. F. Hamblin, 2001, Benthic-pelagic coupling over
a zebra mussel reef in western Lake Erie, Limnology and Oceanography 46:892-
904.

Botts, P. S., B. J. Patterson, and D. W. Schlosser, 1996, Zebra mussel effects on benthic
invertebrates: physical or biotic?, Journal of the North American Benthological
Society 152179-184.

Idrisi, N., E. L. Mills, L. G. Rudstam, and D. J. Stewart, 2001, Impact of zebra mussels
(Dreissena polyrpogpha) on the pelagic lower trophic levels of Oneida Lake, New
York, Canadian Journal of Fisheries and Aquatic Sciences 58:1430-1441.

Caraco, N. F., J. J. Cole, P. A. Raymond, D. L. Strayer, M. L. Pace, S. E. G. Findlay, and
D. T. Fischer, 1997, Zebra mussel invasion in a large, turbid river: phytoplankton
response to increased grazing, Ecology 782588-602.

Holland, R. E., 1993, Changes in planktonic diatoms and water transparency in Hatchery
Bay, Bass Island Area, Western Lake Erie since the establishment of the zebra
mussel, Journal of Great Lakes Research 19:617-624.

Horgan, M.J., and E. L. Mills, 1999, Zebra mussel filter feeding and food-limited
production of Daphnia: recent changes in lower level trophic dynamics of Oneida
Lake, New York, U.S.A., Hydrobiologia 79-88.

Jack, J. D., and Thorp, J. H., 2000, Effects of the benthic suspension feeder Dreissena
polymorpha on zooplankton in a large river, Freshwater Biology 44: 569-579.

Klepinger, M., 2000, Michigan Sea Grant Inland Lakes Zebra Mussel Infestation
Monitoring Program Record December 2000, http://www.msue.msu.edu/
seagrant/zmfiles/ lake01 1501.htrnl.

MacIsaac, H. J., and W. G. Sprules, 1991, Ingestion of small-bodied zooplankton by
zebra mussels (Dreissena polymorpha): can cannibalism on larvae influence
population dynamics, Canadian Journal of Aquatic Sciences 48:2051-2060.

MacIsaac, H. J ., C. J. Lonnee, and J. H. Leach, 1995, Suppression of microzooplankton
by zebra mussels: importance of mussel size, Freshwater Biology 34:379-387.

Pace, M. L., S. E. G. Findlay, and D. Fischer, 1998, Effects of an invasive bivalve on the
zooplankton community of the Hudson River, Freshwater Biology 39:103-116.

Ricciardi, A., and J. B. Rasmussen. 1999. Extinction rates of North American fi'eshwater
fauna. Conserv. Biol. 13:1220-1222.

168

Roditi, H. A., D. L. Strayer, and S. E. G. Findlay, 1997, Characteristics of zebra mussel

(Dreissena polymorpha) biodeposits in a tidal freshwater estuary, Archiv fur
Hydrobiologie 140:207-219.

Shevtsova, L. V., G. A. Zhdanova, V. A. Movchan, and A. B. Primak, 1986,
Experimental interrelationship between Dreissena and planktic invertebrates,
Hydrobiological Journal 22236-39.

Skubinna, J. P., T. G. Coon, and T. R. Batterson, 1995, Increased abundance and depth of
submersed macrophytes in response to decreased turbidity in Saginaw Bay, Lake
Huron, Journal of Great Lakes Research 21 :47 6-488.

Strayer, D. L., and H. M. Malcom, 2001, Declines of unionids and other benthic
planktivores following the zebra mussel invasion of the Hudson River, Abstract
#199, Bulletin of the North American Benthological Society 18:187.

Welschmeyer, N. A. 1994. Fluorometric analysis of chlorophyll a in the presence of
chlorophyll b and pheopigments. Limnology and Oceanography 3921985-1992.

Wisenden, P. A., and R. C. Bailey, Development of macroinvertebrate community
structure associated with zebra mussel (Dreissena polymorpha) colonization or
artificial substrates, Canadian Journal of Fisheries and Aquatic Sciences 7321438-
1443.

169

Table l. Experiment-wide response of variables. All zebra-mussel treatments have been

combined because no effect of mussel density on these variables could be detected.

 

 

 

 

 

 

 

 

 

Variable Unit Initial (SD) Mean of days 6-19 (SD) ANOVA (P)
Total Phosphorus rig/L 5.0 (0.7) 9.8 (1.7) <0.0001
Chlorophyll-a ug/L 3.5 (0.4) 4.8 (1.5) < 0.001
Keratella Abundance ind./L 13.8 (6.1) 90.3 (73.6) < 0.0001
Keratella egg ratio #/ind. 0.26 (0.1) 0.19 (0.1) 0.1207
Polyarthra Abundance ind./L 59.5 (22.1) 13.9 (15.0) < 0.0001
Polyarthra egg ratio #/ind. 0.08 (0.04) 0.05 (0.07) < 0.0001
Nauplii #/L 2.8 (1.0) 0.7 (0.6) < 0.0001

 

 

 

 

 

170

 

Figure 1. Zebra mussel mortality during the experiment. Points are means for 11
enclosures. Bars are standard errors.

171

0.
1

 

 

 

- u A d u
9. 8. 7. 6. 5. A
0 0 0 0 0 0

20332 wfizgm mo :oEomoE

20 30 40

Day

10

172

APPENDICES

173

APPENDIX A

METHODS FOR THE DESIGN AND CONSTRUCTION OF LIMNOCORRALS

I used a design commonly used around KBS as described to me by Gary Mittelbach and
Chris Steiner (Figure A1). Materials needed to build frames that fit a 140" circumference

bag are listed in Table A1. The notes that follow were provided to me by Jessica Rettig:

Building Bags:

- Sweep floor so it is very clean (helps prevent punctures of plastic) and put roll of plastic
on wooden stake between two chairs.

- Put a layer of clean plastic on floor (helps protect plastic for the bag).

- Roll out plastic and iron 3-6 inches of the bottom to seal the bottom. Place cardboard
under the plastic and notebook paper on top of the plastic so the heat from the iron
melts the plastic but doesn’t goop up on the iron or the plastic floor cover. set the
iron at around 6 power.

- Loosely pleat bottom, once plastic has cooled, and tie a knot. This step takes 2 people!
Insert an anchor rope into the knot.

- Measure 2.2m above the knot and cut the plastic from the roll. This length will give you
a bag of about 1.5 m. Adjust your cut to get the bag length you need.

- Fold the top into a cuff about 4-6 inches wide and tape the cuff with poly-tape (used 3.5
rolls to tape 24 cuffs). This step works best with two people.

- Survey finished bag for any small nicks or holes and cover with poly-tape.

174

- Put finished bags into a vat lined with plastic. Stack no more than 8 bags high, with a
plastic liner between every 3 or 4 bags (helps prevent punctures to bags).

- I think it can take around 30 minutes to build one bag, but I can't remember for sure.

Pond Prep:

- Cut opening into cattails so you have access to the pond. Opening should be around 3m
wide and all cattails should be trimmed to the level of the pond substrate (cattails
can puncture the bags if the stems are sticking up under the water).

- Also snorkel the transect upon which you will deploy the frames and remove any stiff
vegetation that might puncture the bags.

- set up ropes across the pond, 2 ropes per row of bags. Ropes are tied to metal stakes set
beyond the lip or hump of the pond. In the past I've set up 6 ropes for 3 rows of

bags using 4 stakes.

Attaching Bags:

- Attaching bags to the frames works best with three people, two of which need to be in
the water (the third may be in the water or in a boat leaning out to hold the frame
or bag).

- Before using the frames check for rough edges, nails, staples, etc. and remove. Any of
these could puncture the bag.

- Pass one frame into the water and have the person in the boat unfold a bag so the
bottom can be sent through the center of the frame. Try to keep the edges of the

bag from snagging on the frame.

175

- Stretch the cuff opening along the top of the flame and staple into place. This step
works best if two people stretch and staple the cuff at the same time (2 staple
guns) while the third person holds the flame steady. Staples should be parallel to
the edges of the flame so they are less likely to tear off.

- Find the bottom of the bag (knot) and fold it over the side of the flame and into the bag,
so no parts of the bag are dragging as you push the flame and bag through the
water.

- Once you have enough flames to fill a row, attach the flames to the ropes using cable-
ties. Adjusting the distance between flames as needed.

-Note that working with 3 or 4 people, it can take 1-1 .5 hours to deploy 8 bags.

Filling Bags:

- Attach anchors (half brick with rope and clip) to the anchor rope on each bag and set the
end of the bag into the water. Try to center the knot under the flame so the bag
hangs vertically. Once the bag is filled, it is difficult to adjust how it hangs.

- Put big red pump into boat (fill tank with gas prior to setting into boat and check oil in
Pump)-

- Attach the flexible blue hose to the pump outflow and cable-tie a zooplankton net over
the end to filter out zooplankton (if needed). The zooplankton net should be tied
shut.

-Attach the stiff gray hose to the intake.

-Prime the pump before starting it by flooding the open end of the gray hose and raising it

upwards toward the sky to force water into the pump.

176

- Start the pump, keeping the gray hose under the water. I tie it to the boat to prevent it
flom sinking to the pond bottom and sucking up muck.

- Filling bags often works best with two people, one holding the blue hose to fill the bags
and the other monitoring the gray intake and holding the boat in position.

- F 111 each bag for a similar period of time and check the volume by pressing on the sides
of the bag under the flame (does it feel full?). Every 3-5 bags, point the blue hose
away flom the bags, untie the knot in the zoop net, and rinse the net clean. I filled
22.5 bags 91 .5cm deep) with one tank of gas. Pump gets very hot and you may

need to let it cool down before adding more gas.

177

Table A1: Materials needed to build limnocorrals.

 

Materials for 18 corrals:

72

18

15

576

864

72

8’ (6x1) boards

8’ (3x1) boards

4’x8’ styrofoam panels
screws

nails

eyehooks

Materials for 1 corral:

32

48

8’ (6x1) board

8’ (3x1) board
4’x8’ styrof. panel
screws

nails

eyehooks

 

178

Figure A1. Design of Limnocorral Frames.

179

 

 

*—-——34 in ———>

 

 

 

 

 

 

=' '= ""' 4 screws

+
eyehook

seven layers
dense styrofoam

 

 

 

180

APPENDIX B

METHODS FOR THE COLLECTION AND REARING OF ZEBRA MUSSELS

Collection: I collected zebra mussels from Gull Lake. There are plentiful populations
along the KBS shoreline. For larger specimens you need to snorkel deeper water. Take a
razor blade and cut the byssal threads to remove mussels flom rocks (don't just pull them
off). This prevents injury to the mussels, and should encourage them to secrete new
threads. You can use a sieve to sort for a specific size class. Watch for empty shells that

are still closed.

Holding: 1 held mussels in a tank with continuous flow-through of lake water. Place
a submersible pump in Gull Lake just under the dock. Run the power cord and a hose into
the boat house by fishing them through the fan duct next to the hood. This prevents any
windows or doors flom being unlocked during operation. Nina allowed the bending of a
few bars of the fan guard to accommodate the hose. Put the hose in a bucket into which
you have placed coarse grade mesh. Cut a hole in the bucket so that it drains into the
oversize tank in the boat house. Position the top hole of the tank over the cistern so that it
may drain directly into the pit. This set up will allow fine sediment to collect in the tank,
depending on how much wave action is present in the lake. Check the set-up daily and do
the following: clear the pump intakes of debris, rinse the mesh, siphon sediment flom the

mussels.

181

Substrate: I allowed zebra mussels to attach to Plexiglass sheets for two weeks.
Sprinkle zebra mussels onto Plexiglass (or PVC pipe). They will creep around very
slowly. The next day some mussels will have begun to attach while others will have
wandered off your substrate. Reposition the stray mussels in the vicinity of attached
mussels. Allow 7-14 days for the mussels to become affixed. I separated my plexiglass
sheets by wedging foam insulation into the tank, creating "walls" between the plexiglass.
This allowed me to add a set number of mussels to each plexiglass sheet without fear of

them wandering over to another sheet.

Size and Density: 1 used a size structured population of 200 mussels per treatment.
Natural densities impose an upper natural limit, but the actual density used should be less
because the corrals are closed systems. In addition to density, size structure of the
population must be considered. Large mussels (>30 mm) will exert a large influence on
filtering activity compared to smaller individuals. If one large mussel filters at a different
rate than another large mussel in another treatment, overall filtering activity will differ

between treatments. Very large mussels should thus be excluded.

182

APPENDIX C

METHODS FOR THE COLLECTION AND REARING OF LARVAL BLUEGILL.

Thanks again to Jessica Rettig for providing valuable advice.

Collection: I collected hatchling bluegill directly flom nests in Warner lake (Fig. A2).
Bluegill nests are bowl-shaped depressions in soft sediment close to the shore. Up to 50-
100 cm wide and 20 cm deep, nests are clustered into colonies. The nests collect organic
debris but are cleaned out when active. When hatchlings are present, the nests appear to
have a gray or pink "fuzzy" substance within them, sometimes glittering with gold (the
gold-eye stage). Adult bluegill guard active nests and so try not to scare fish away flom a
colony as you approach (guarding fish act as flags to active nests). To collect hatchlings,
swim slowly up to a nest and use a turkey baster with its tip cut off. Take two baster's
worth and squirt the fish into a jar and cap the jar, all while underwater. Transport jars to
the lab in a cooler filled with lake water to prevent the jars flom overheating. Don't get
greedy during collection. Call Lyle Champion for permission to work in Warner Lake.
Lawrence Lake had many colonies that were not active. Middle Three Lakes had one
colony that yielded some active nests. Warner Lake is the best. I made collections in late

June to practice, fine tune the tanks, and be sure I'd have larvae swimming at all times.

Holding: 1 reared hatchling bluegill in the lab for 4-7 days. Acclimate jars by placing

them into prepared tanks without removing the lids for 30 minutes. A ten-gallon tank can

accommodate up to three jars. Add methylene blue to the tank water just until there is a

183

slight blue tint to the water. Have an air stone running. Use an out-o-tank filter, and wrap
the intakes with filter floss. After 30 minutes, open jars and lay them on their side. Place
the airstone at the mouth of the jar, but not inside. This creates a gentle current within the
jar without disturbing the fish. After a few days, the hatchlings will be swimming in the
tank (swimming stage). This method results in several hundred fish in a tank. If

something goes wrong and the fish die, their bodies become white.

Feeding: 1 fed larval bluegill brine shrimp nauplii in the lab. Have a brine shrimp
culture prepared. Brine shrimp require a rearing chamber to ensure that nauplii can be
collected without their egg shells. The culture also needs to be warm (80°F) and saline.
Sea monkeys hatch overnight and are orange. You can tell if the bluegill are eating the
shrimp because the bluegill's bodies are translucent and their stomachs turn orange. Also
add copepods collected flom Duck Lake so the larvae can get some experience feeding

on naturally occurring plankton.

Size and Density: I used 20 fish (7.5mm-long) per treatment. The fish are usable when
they begin eating. Since they grow quickly, they must be small to qualify as larval and
not juvenile. 20 fish per bag represents a density at the high end of, or just higher than

natural density.

Sea Monkey (brine shrimp) rearing tank: Use a 10 gallon tank. Affix two opaque
Plexiglass dividers inside the tank, one with space above and one with space below. This

forces the nauplii to swim and shed their egg shells. Affixing the light to the side of the

184

tank heats the tank nicely. Put the eggs in the far right area, and they will swim to the
light and congregate. Add eggs every day near the time you need them so you will have

nauplii ready for your fish larvae.

185

 

Figure A2: Location of some bluegill nest colonies in local lakes.

186

 

Lawrence lake

 

Warner lake

“is
39 amp
w put-in
‘\

Tree with
nest box

Middle 3 lakes

 

187

 

Figure A3. Design of brine shrimp rearing tank.

188

 

 

   
  

Light source ,W. 4— Water level

Dividers

 

 

l

O
O
MAirstone

 

189

APPENDIX D

SAS program codes with dummy data

AN COVA and Linear Regression

title ' title';

data logchla;

input ZM LOGTP LOGChla;
cards;

0 1.401 1.275

0 1.323 0.782

0 1.584 0.908

1 1.509 0.653

1 1.821 1.606

1 1.331 0.079

proc print data=logchla;

proc glrrr;

class zm;

model logchla = ZM LOGTP/SS3;
output out=logchla2 1=resid;

proc glm;

class zm;

model logchla = ZM LOGTP ZM*LOGTP/SS3;
proc glm;

model logchla = LOGTP;

by ZM;

nm;

quit;

190

 

title ' K-S test';

data lakeTP;

input ZM TP;
cards;

0 18.8

0 28.4

0 16.9

1 30.8

1 17.8

1 64.5

proc print data=lakeTP;

proc nparlway data=LakeTP;
class zm;

run;

quit;

Kolmogorov-Smirnov

191

 

Correspondence Analysis

Title 'Correspondence analysis';
data 'genera';

input Binuclearia Oocystis Scenedesmus Pediastrum Mougeotia
Chrococcus Rhodomonas Aphanocapsa Dinobryon Staurastrum
Gleobotrys Selenastrum Merismopedia Fragilaria Diatom
Chlamydomonas Cerratium Anabeana Crysophyte Lyngbya
Ossilatoria Microcystis Coelosphaerium Ulothrix
Aphanizomenon;

cards;

0 9418 0 0 0 0 14127 0 0 0 0 0
1056 0 0 0 1420000 8852803 0 1421405
0 46172 0 0 7539625

0 0 0 0 0 15249 4709 0 0 0 0 0
0 496353 0 0 94667 63016 0 0 0 0
0 0 164801

6669 3973 0 0 0 26303 58027 0 14236 0 0 0
0 0 492505 0 0 134660 0 0 0
0 327762 0 0

0 0 0 0 0 0 16023 770 27703 0 0 0
0 0 0 0 189333 143377 0 0 0
4329 35634 27703 89267

0 0 0 0 0 0 O 6541 21372 0 0 0
0 0 0 0 189333 0 0 0 0 0
399479 0 0

proc corresp;

var Binuclearia Oocystis Scenedesmus Pediastrum Mougeotia

Chrococcus Rhodomonas Aphanocapsa Dinobryon Staurastrum
Gleobotrys Selenastrum Merismopedia Fragilaria Diatom
Chlamydomonas Cerratium Anabeana Crysophyte Lyngbya
Ossilatoria Microcystis Coelosphaerium Ulothrix
Aphanizomenon;

quit;

192

data survey;
input zm

cards;

t—u—t—sn—rr—rHmHH—rflw—‘OOOOOOOOOOOOOO

21.12
68.65
22.00
30.26
1.00

1.00

21.77
31.28
1.00

84.54
32.52
71.57
62.41
36.64
34.19
12.44
19.32
1.00

48.52
25.87
24.67
48.22
1.00

37.56
15.68
15.77
26.32

proc glm;
class zm;

MANOVA

Chloro Chryso Crypto Cyano Pyrro;

39.91
79.18
88.52
40.28
89.96

24.17
16.76
38.72
25.21
1.00

175.60 29.17

63.80
57.81
74.93
25.33
1.00

1.00

78.30
73.68
43.64
19.96
26.61
32.07
54.55
28.81
85.29
1.00

1.00

49.56
42.26
46.35
57.62

40.08
34.59
24.17
1.00

36.33
1.00

36.04
25.45
26.40
1.00

21.12
39.00
26.92
1.00

25.15
24.17
22.87
37.75
1.00

32.31
34.40

261.40 112.40
62.41 45.58

78.77
64.90
74.05

1.00
57.42
57.42

248.41 101.36
107.58 72.35

14.83
65.64
6.93

60.42
41.93
64.09

1.00
1.00
1.00
56.81
1.00
56.81

208.54 45.58

33.68
31.88
59.99
43.86

1.00
1.00
1.00
57.42

436.89 1.00

65.53

1.00

147.65 72.35
307.78 123.71
39.40 45.58
60.34 1.00

95.68

1.00

47.94 3.08
189.95 58.00

model Chloro Chryso Crypto Cyano Pyrro= zm/ss3;
output out=phyto r=resid;
manova H = zm;

run;
quit;

193

Repeated Measures AN OVA

data trans;

input rep zm bg dayl day2 day3 day4;

cards;

1 l 1 0.8 0.6 0.6 0.6
2 1 1 0.6 0.3 0.9 0.5
3 1 0 1.0 0.3 0.4 0.3
4 l 0 1.0 0.4 0.5 0.4
5 0 0 2.4 1.6 1.0 1.1
6 0 0 2.1 1.7 2.1 1.3
7 1 0 0.6 0.4 0.5 0.6
8 0 l 2.3 2.0 1.0 1.4
9 0 1 1.6 2.0 1.7 1.7
10 0 0 2.2 1.9 1.4 1.2
11 l 1 0.7 0.4 0.5 1.4
12 0 1 2.2 1.1 1.2 1.9
proc print data=trans;

proc glm;

class zm bg;

model dayl-day4 = zm bg zrn*bg/nouni;

repeated time 4 (1 4 7 11) profile/summary printrn printe;
run;

quit;

194

APPENDIX E

Complete References with Additions

Ackerrnan, J. D., M. R. Loewen, and P. F. Hamblin, 2001, Benthic-pelagic coupling over
a zebra mussel reef in western Lake Erie, Limnology and Oceanography 46:892-
904.

Adams, S. M., R. B. McClean, and M. M. Huffman, 1982, Structuring of a predator
population through temperature-mediated effects on prey availability, Canadian
Journal of Fisheries and Aquatic Sciences 39:1175-1184.

Aminot, A., D. S. Kirkwood, and R. Kerouel, 1997, Determination of ammonia in
seawater by the indophenol-blue method: evaluation of the ICES NUTS I/C 5
questionnaire. Marine Chemistry: 56259-75.

Baker, S. M., and D. J. Hornbach, 1997, Acute physiological effects of zebra mussel
(Dreissena polmorpha) infestation on two unionid mussels, Actinonaias
ligmentina and Amblema plicata, Canadian Journal of Fisheries and Aquatic
Sciences 54:512-519.

Baker, S. M., J. S. Levinton, J. P. Kurdziel, and S. E. Shumway, 1998, Selective feeding
and biodeposition by zebra mussels and their relation to changes in phytoplankton
composition and seston load, Journal of Shellfish Research 172 1207-1213.

Barkoh, A., and T. Modde, 1987, Feeding behavior of intensively cultured bluegill fly,
Progressive Fish-Culturlist 49:204-207.

Bastviken, D. T. E., N. F. Caraco, and J. J. Cole, 1998, Experimental measurements of
zebra mussel (Dreissena polmorpha) impacts on phytoplankton community
composition, Freshwater Biology 39:375-3 86.

Berg, D. J ., S. W. Fisher, and P. F. Landrum, 1996, Clearance and processing of algal
particles by zebra mussels (Dreissena polymorpha), Journal of Great Lakes
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