138
028

 

 

TH E308

.08 '

4_, r If”?
5401“ 3/?

This is to certify that the
thesis entitled

UNDERSTANDING ZEBRA MUSSEL (DRE/SSENA
POL YMORPHA) IMPACTS ON AUTOTROPHIC AND
HETEROTROPHIC PLANKTON OF INLAND LAKE
ECOSYSTEMS

presented by

LESLEY BETH KNOLL

has been accepted towards fulﬁllment
of the requirements for the

MS. degree in FISHERIES AND WILDLIFE

 

 

ﬂw/ﬂyc

Major Professor’s Signature
Ma? 47 >00 7

Date

 

MSU is an Afﬁrmative Action/Equal Opportunity Institution

 

LIBRARY
Michigan State
University

 

 

 

PLACE IN RETURN BOX to remove this checkout from your record.
TO AVOID FINES return on or before date due.
MAY BE RECALLED with earlier due date if requested.

 

DATE DUE DATE DUE DATE DUE

 

I {NOV 1 1.12““

 

 

 

 

 

 

 

 

 

 

 

 

 

 

6/01 c:/CIRC/DateDue.p65-p. 15

UNDERSTANDING ZEBRA MUSSEL (DREISSENA POL YMORPHA) IMPACTS ON
AUTOTROPHIC AND HETEROTROPHIC PLANKTON OF INLAND LAKE
ECOSYSTEMS
By

Lesley Beth Knoll

A THESIS
Submitted to
Michigan State University
in partial fulﬁllment of the requirements
for the degree of
MASTER OF SCIENCE
Department of Fisheries and Wildlife

2004

ABSTRACT

UNDERSTANDING ZEBRA MUSSEL IMPACTS ON AUTOTROPHIC AND
HETEROTROPHIC PLANKTON OF INLAND LAKE ECOSYSTEMS

By

Lesley Beth Knoll
Zebra mussels (Dreissena polymorpha) invaded the Great Lakes of N. America in the
mid-1980’s, and have rapidly spread into inland lakes, particularly in Michigan. Most
research on the impact of Dreissena invasion has been conducted in well-mixed aquatic
ecosystems, while little is known about such impacts in thermally-stratiﬁed inland lakes.
Dreissena invasion has recently been associated with an increase in the toxic
cyanobacterium (bluegreen), Microcystis aeruginosa, in low-nutrient lakes but here
again, there is little compelling survey data from inland lakes indicating significantly
elevated M. aeruginosa biomass in invaded habitats. In addition, no studies have
examined whether bluegreen toxins are elevated in Dreissena-invaded lakes.
To address these needs, I conducted a large-scale survey of inland lakes in Michigan that
contain or lack D. polymorpha. The surveyed lakes were otherwise similar in nutrients,
morphometry, and location. Microcystis aeruginosa biomass was 3.6 times higher and
total particulate toxin concentration was 3.3 times higher in invaded lakes. Ciliate
biomass was 45% lower and rotifer biomass was 44% lower in invaded lakes. Rotifer
richness and diversity were also lower in lakes with D. polymorpha. In addition, a shift in
the size distribution of ciliates was found between lake categories indicating that zebra
mussels may be reducing larger, algivorous ciliates in favor of smaller bacterivorous
ciliates. In general, zebra mussel impacts on microzooplankton biomass in the stratiﬁed

lakes I sampled were weaker than those reported from well-mixed systems.

ACKNOWLEDGEMENTS

I thank my advisor Dr. Orlando Samelle and my committee members Dr. Patricia
Soranno and Dr. Stephen Hamilton for their assistance during all stages of my research.
Special thanks to my graduate student collaborator, Carrie Scheele, for all of her help.
Thanks to Joan Rose and Mechelle Woodall for ELISA analysis. Thanks to Alan Tessier
and Spencer Hall for their bathymetric map of Warner Lake. Thanks to all of those of
helped in the ﬁeld and laboratory: Chelsea Stephen, Kendra Cheruvelil, Brian Elkington,
Katherine Gurin, Scott Kissman, Konrad Kulacki, Drew Kramer, Sherry Martin, Andrea
Stoddard, Alan Wilson, and Craig Wynne. Financial support was provided by: Michigan
Sea Grant, MSU Research Enhancement Award, MSU Ecology, Evolutionary Biology
and Behavior Research Award, Sigma Xi Grant-in—aid of Research, MSU Minority and

Women Graduate Assistantship Fellowship, and George H. Lauff Research Award.

iii

TABLE OF CONTENTS

LIST OF TABLES .................................................................................... v

LIST OF FIGURES ................................................................................. vi

CHAPTER 1

ZEBRA MU SSEL INVASION IS ASSOCIATED WITH INCREASED MICROCYSTIS

BIOMASS AND TOXIN CONCENTRATION IN LOW-NUTRIENT LAKES ............ 1
Abstract ....................................................................................... 1
Introduction ................................................................................... 2
Methods ....................................................................................... 4
Results ......................................................................................... 8
Discussion ................................................................................... 10
Literature Cited ............................................................................. 17

CHAPTER 2

INFLUENCE OF ZEBRA MUSSELS (DREISSENA POLYMORPHA) ON
MICROZOOPLANKTON COMMUNITIES IN STRATIFIED LOW-NUTRIENT

LAKES ................................................................................................ 31
Abstract ...................................................................................... 3 1
Introduction .................................................................................. 32
Methods ...................................................................................... 34
Results ....................................................................................... 36
Discussion ................................................................................... 38
Literature Cited ............................................................................. 44

APPENDIX

INDIVIDUAL SURVEY LAKE INFORMATION ............................................ 58

iv

LIST OF TABLES
CHAPTER 1
Table 1. Average, range and standard deviation of limnological parameters in survey
lakes. P values are for t-tests. For NHX-N, alkalinity, conductivity, DOC, and pH, P
values are for t-tests only using 2002 data ....................................................... 21

Table 2. Average, range, and standard deviation of variables related to M. aeruginosa and
microcystin toxin. P values are for t-tests ........................................................ 22

CHAPTER 2

Table 1. Average, range, and standard deviation of physical and biological parameters in
uninvaded and invaded lakes. P values are for t-tests ........................................... 49

LIST OF FIGURES

CHAPTER 1

Figure 1. Location of survey lakes in the lower peninsula of Michigan. The state is
divided into Albert ecoregion subsections. Closed circles represent invaded lakes and
open circles represent uninvaded lakes ........................................................... 23

Figure 2. Microcystis aeruginosa dry mass in lakes with and without D. polymorpha (P =
0.01). P value for a t-test. Bars are SE ............................................................ 24

Figure 3. Factor scores from the phytoplankton (exclusive of M. aeruginosa) PCA. Factor
1 (P = 0.00002) and Factor 2 (P = 0.65), which indicates a positive effect of D.
polymorpha on Factor 1. P values are for t-tests ................................................. 25

Figure 4. Relative biomass values (%) of the phytoplankton genera affected by D.
polymorpha presence (Anabaena, Ceratium, Cryptomonas, and Microcystis) and the
remaining phytoplankton grouped together as “other” ......................................... 26

Figure 5. Microcystin concentration in lakes with and without D. polymorpha (P =
0.009). P value for a t-test. Bars are SE .......................................................... 27

Figure 6. (A) Microcystin concentration per microcystin-producing biomass in lakes with
and without D. polymorpha (P = 0.0009). Microcystin producing genera are Anabaena,
Microcystis, Oscillatoria, and Nostoc. (B) Microcystin concentration per M. aeruginosa
biomass (P = 0.5). P values are for t-tests. Bars are SE ........................................ 28

Figure 7. Colony size of M. aeruginosa as average area (P = 0.47). P value for a t-test.

Bars are SE ............................................................................................ 29
Figure 8. The effect of D. polymorpha on chlorophyll a (P = 0.008). P value is for a t-test.
Bars are SE. .......................................................................................... 30
CHAPTER 2

Figure 1. Schematic representation of hypothesized direct and indirect effects of D.
polymorpha on lower trophic levels of lake ecosystems. Arrows represent a positive
effect and circles represent a negative effect .................................................... 50

Figure 2. Location of survey lakes in the lower peninsula of Michigan. The state is

divided into Albert ecoregion subsections. Closed circles represent invaded lakes and
open circles represent uninvaded lakes ........................................................... 51

vi

Figure 3. Inﬂuence of D. polymorpha on ciliate and rotifer dry biomass. Both ciliate and
rotifer biomass were signiﬁcantly lower in D. polymorpha lakes (P = 0.009, 0.01,
respectively). P values are for t-tests. Bars are SE ............................................. 52

Figure 4. The effect of D. polymorpha on (A) chlorophyll a (P = 0.02) and (B) DOC (P =
0.23) in the survey lakes. P values are for t-tests. Bars are SE ................................ 53

Figure 5. (A) Relative biomass (%) and (B) actual biomass of ciliates below 30 pm in
size (bacterivorous) and above 30 um (algivorous) in the survey lakes. Ciliates above 30
um were higher in uninvaded lakes for both relative biomass and actual biomass (P =
0.01, 0.003, respectively). The actual biomass of ciliates below 30 um were similar
between lake categories (P = 0.75). P values are for t-tests .................................... 54

Figure 6. The inﬂuence of D. polymorpha on (A) rotifer richness (P = 0.01) and (B)
diversity (P = 0.02). Both richness and diversity were lower in invaded lakes. P values are

for t-tests. Bars are SE .............................................................................. 55

Figure 7. Factor scores from the rotifer PCA. Factor 1 (P = 0.036) and Factor 2 (P = 0.2).
P values are for t-tests ............................................................................... 56

vii

CHAPTER 1

Zebra mussel invasion is associated with increased Microcystis biomass and toxin
concentrations in low-nutrient lakes

ABSTRACT
Bluegreens typically comprise a minor portion of the phytoplankton community in low-
nutrient lakes. However, previous studies have shown a positive effect of exotic zebra
mussels (Dreissena polymorpha) on the relative dominance (as percentage of total
phytoplankton biomass) of Microcystis aeruginosa, a bloom-forming, toxin-producing
bluegreen, in low-nutrient lakes (total phosphorus 10 - 25 ug L'I). I sampled thirty-nine
low-nutrient lakes (average TP ~ 10 pg L'I), 20 with D. polymorpha and 19 without, that
were otherwise similar in nutrients, morphometry, and location. I report a positive
inﬂuence of D. polymorpha on the biomass of M. aeruginosa and on the concentration of
microcystin, a toxin produced by M. aeruginosa and a few other bluegreen taxa. An
increase in M. aeruginosa and the toxins produced by it are unexpected and undesirable,
particularly in low-nutrient lakes. Microcystis aeruginosa colony sizes were similar
invaded and uninvaded lakes. Soluble reactive phosphorus concentrations were similar
between lake categories, but NHf-N was 1.8 times higher in invaded lakes. Chlorophyll
a and phytoplankton dry mass were 30 % lower in invaded lakes. Other phytoplankton
species inﬂuenced by D. polymorpha presence in the survey were: Anabaena and

C eratium (both lower in invaded lakes), and Cryptomonas (higher in invaded lakes).

Introduction

Bloom-forming bluegreen algae (species in the genera Anabaena,
Apham'zomenon, Microcystis, and Oscillatoria in temperate lakes) have long been the
focus of intense research and concern because of their unpleasant odor and appearance,
and their ability to produce toxins harmful to animals and humans (Chorus and Bartram
1999). In addition, poor nutritional quality of bluegreens, coupled with the toxins they
produce, can reduce the growth and survivorship of herbivorous zooplankton (de
Bemardi and Giussani 1990, DeMott et a1. 1991, DeMott 1999), and so reduce food-
chain efﬁciency. It is widely accepted that the relative dominance of noxious bloom-
forming bluegreens increases with nutrient enrichment, and that these taxa usually
comprise a small percentage of the phytoplankton community in nutrient poor systems
(Trimbee and Prepas 1987, Downing et a1. 2001). As a result, lake management practices
aimed at improving water quality generally focus on reducing the nutrient loads coming
into lakes. However, recent evidence suggests that the invasive zebra mussel (Dreissena
polymorpha) may alter the well-established positive relationship between nutrients and
bluegreen dominance (Vanderploeg et a1. 2001, Raikow et a1. 2004), which may require
re-evaluation of nutrient management strategies in lakes with D. polymorpha.

Although changes in phytoplankton community structure are variable following
D. polymorpha invasion across freshwater systems, one surprising trend is emerging. In
shallow, mesotrophic areas of the Great Lakes, Microcystis aeruginosa seems to have
increased following D. polymorpha invasion, despite recent reductions in phosphorus
loading (Vanderploeg et a1. 2001). Similarly, a survey of low-nutrient (total phosphorus,

TP, 10 - 25 pg L'I) inland lakes showed higher relative dominance of M. aeruginosa in

lakes invaded by D. polymorpha. Such lakes would not be expected to harbor high levels
of bloom-forming bluegreens given their relatively low nutrient status. Thus, zebra
mussels seem to be affecting the dominance relationships of phytoplankton communities
in unexpected ways.

The aforementioned survey found higher M. aeruginosa dominance in inland
lakes invaded by D. polymorpha, but no signiﬁcant inﬂuence of invasion on the actual
biomass of this bluegreen species (Raikow et al. 2004). In addition, no measurements of
toxin concentrations were reported in that survey. Thus, the primary focus of my study
was to determine whether lakes invaded by D. polymorpha have higher M. aeruginosa
biomass and higher microcystin concentrations. The previous survey indicated that the
response of M. aeruginosa to invasion was limited to lakes within a relatively narrow
range of TP (10 — 25 ug L'I), so I focused on such lakes in my survey.

The mechanism by which D. polymorpha promotes M. aeruginosa dominance is
not fully understood, but one proposed mechanism involves reduced ﬁltering rates of D.
polymorpha on M. aerugz'nosa. This may result in a selective advantage for the bluegreen
over other phytoplankton taxa (Vanderploeg et al. 2001 ). Reduced ﬁltration rates may be
a function of the colonial habit and/or toxicity of M. aeruginosa (V anderploeg et al.
2001). Some strains of Microcystis aeruginosa and other bluegreen genera (Anabaena,
Oscillatoria, and Nostoc) produce microcystin, a hepatotoxic cyclic peptide toxin
(Chorus and Bartram 1999) that might act as an herbivore deterrent. Thus, D. polymorpha
invasion could result in an increase in per capita microcystin production by these species,

via phenotypic or evolutionary responses. Such an effect would have ecological, as well

as public health consequences, given that increased per capita toxicity might further
decrease food-chain efﬁciency.

To address the above issues, I conducted a lake survey of 39 invaded and
uninvaded, low-nutrient (total phosphorus < 20 ug/L) inland lakes in Michigan that
thermally stratify in the summer. In addition to examining the responses of M.
aeruginosa biomass and toxin concentration, I also investigated how basic limnological
parameters were inﬂuenced by D. polymorpha in these lakes, to provide some insight into
potential mechanisms leading to increased M. aeruginosa. Based on previous studies, I
expected to see lower chlorophyll a (Holland 1993, Nicholls and Hopkins 1993,

F ahnenstiel et al. 1995 Caraco et al. 1997, Yu and Culver 2000, Idrisi et al. 2001, Raikow
et al. 2004), leading to higher water clarity (Holland 1993, Caraco et al. 1997, Yu and
Culver 2000, Idrisi et al. 2001) and dissolved nutrients (Holland et al. 1995, Efﬂer et al.
1996, Caraco et al. 1997, Raikow et al. 2004) in invaded lakes. I also examined the
inﬂuence of D. polymorpha on major solutes (Ca, Mg, Na, K, Cl, 804, and Si), which has
rarely been attempted (Holland et al. 1995, J ohengen et al. 1995).

Methods

Lake Selection

For the survey, 1 selected low-nutrient lakes (TP < 20 ug L'I) in southern
Michigan (Figure 1) with a maximum depth of a least 9 m, such that thermal stratiﬁcation
was likely to be present during the summer. For the purposes of lake selection, the
presence/absence of D. polymorpha was initially assessed using the list of invaded lakes
assembled by the Michigan Sea Grant College Program (www.miseagrant.org).

Presence/absence of D. polymorpha was veriﬁed in each lake by searching for adults in

the ﬁeld and veligers in zooplankton samples. To ensure that invaded and uninvaded
lakes included in the survey were similar in depth, mean depth was determined by
digitizing bathymetric lake maps in ArcView GIS 3.2. Based on these criteria, 39 lakes
were selected, 20 with D. polymorpha and 19 without, for the survey.
Field Sampling

Each lake was visited once, in late summer of 2002 or 2003 (2 August — 4
September 2002, 3 August — 20 August 2003). I took samples from the deepest spot in
each lake, as determined by bathymetric maps and a depth ﬁnder. Temperature,
conductivity, and pH were measured at 1 at intervals with a Hydrolab Surveyor and
Datasonde. Photosynthetically available radiation (PAR) was measured at 0.5 m intervals
with a LiCor model Li-1000 quantum photometer radiometer equipped with a spherical
underwater sensor. Vertical extinction coefﬁcients were calculated as the slope of In PAR
versus depth via linear regression. A depth-integrated water sample was taken through
the entire mixed layer (epilimnion) with a ﬂexible tube (5 cm inner diameter) and placed
on ice until processing (approximately 6 hours). Subsamples of epilimnetic water were
taken for chlorophyll a, nutrients, and water chemistry. A 125 ml subsample of
epilimnion water was preserved in Lugol’s solution for phytoplankton enumeration.
Laboratory Analysis

For chlorophyll a analysis, samples were ﬁltered onto a Gelman A/E glass ﬁber
ﬁlter and kept frozen. The ﬁlter was subsequently extracted with 95% ethanol and
chlorophyll a quantiﬁed using a Turner Model 10-AU ﬂuorometer calibrated with
standards (Welschmeyer 1994). Dissolved nutrients were ﬁltered through a Gelman A/E

glass ﬁber ﬁlter. NHf-N was measured colorimetrically following an adapted version of

the phenylhypochlorite method (Aminot et al. 1997). Soluble reactive phosphorus (SRP)
and total dissolved phosphorus (TDP) were measured colorimetrically following the acid
molybdate method (Wetzel and Likens 2001); the TDP colorimetric analysis was
preceded by a persulfate digestion to decompose organically bound P (Valderrama 1981).
Total phosphorus (TP) was analyzed colorimetrically (Langner and Hendrix 1992) after
persulfate oxidation. Upon return to the laboratory, water chemistry samples were
refrigerated (alkalinity, conductivity) or ﬁltered through Gelman A/E glass ﬁber ﬁlters
and then refrigerated (anions, silica), or acidiﬁed with 8 N I'INO3 (cations) until analysis.
Conductivity was measured in the laboratory using an Orion model 135 conductivity
meter. Ca, Mg, Na, and K were measured by ﬂame atomic absorption spectrometry, and
total alkalinity, which generally represents HCO3 in lake waters, was determined by
titration with 0.3 N HCl and calculation of the Gran function (Wetzel and Likens 2001).
S04 and Cl were measured by membrane-suppressor ion chromatography. Si was
measured colorimetrically by the ammonium molybdate method (Wetzel and Likens
2001). DOC was measured with a high-temperature combustion DOC analyzer.
Phytoplankton were generally identiﬁed to species and were enumerated using the
inverted microscope method (Hasle 1978). Subsamples were settled in tubular chambers
(Hydro-Bios), the bottoms of which were divided into inner and outer zones of equal area
(Sandgren and Robinson 1984). Within each zone, at least 20 random ﬁelds were counted
at 100x, 400x, and 1000x. I determined phytoplankton biovolume by measurements of
cell dimension at 1000x of at least ten individuals of common species using a digital
camera system and image-analysis software. Biovolume was calculated by using a

geometric volume equation appropriate for the species shape and converted to dry

biomass assuming a speciﬁc gravity of l g cm'3 and a dry mass to wet mass ratio of 0.10.
To estimate M. aeruginosa colony size, I searched the entire chamber at 100x and
estimated the area of each colony (umz) by counting the number of grid squares the
colony occupied.

To collect particulate microcystin, a large quantity of integrated epilimnion water
(generally 2 IL) was ﬁltered through a Gelman A/E glass ﬁber ﬁlter and the ﬁlter was
frozen until analysis. Enzyme-linked immuno sorbent assay (ELISA) (An and
Carmichael 1994) was used to obtain particulate microcystin concentrations.

Data and statistical analyses

To examine the inﬂuence of D. polymorpha on all variables, I used t-tests.
However, for a sub-set of the physico-chemical parameters (alkalinity, conductivity,
NHf-N, DOC, and pH) only data from 2002 were available.

Since the focus of this study was to investigate the impact of D. polymorpha on
M. aeruginosa, separate analyses were performed for the rest of the phytoplankton
community. To examine the impact of D. polymorpha on phytoplankton species
composition other than M. aeruginosa, I used principal components analysis (PCA) on
phytoplankton relative biomass (exclusive of M. aeruginosa). The objective of the PCA
was to reduce the phytoplankton data set to a small number of variables to avoid data
mining for signiﬁcant responses. To reduce zero values, only common taxa were
included: Anabaena, Aphanocapsa, Ceratium, C hroococcus, Cryptomonas,
Chrysochromulina, F ragilaria, Oocystis, Peridinium, Rhodomonas, Scenedesmus,

Sphaerocystz‘s, and an unidentiﬁed small ﬂagellate.

PCA was also applied to the major solute data (Ca, Mg, Na, K, Cl, SO4,and Si) to
determine the inﬂuence of D. polymorpha presence on these variables. For this analysis,
only data from 2002 were available. Data were log-transformed when necessary so that
assumptions for parametric statistics were met. All analyses were conducted in SYSTAT
9.0.

Results

To rule out the inﬂuence of potential confounding factors, I assessed whether
invaded and uninvaded lakes were similar in morphometry and nutrients. There were no
signiﬁcant differences in either mean depth or TP between the two groups of lakes (Table
l). I assessed if day of year affected any M. aeruginosa variable since bluegreens tend to
become more dominant as the summer progresses and found no effect of day of year on
any M. aeruginosa variable (P > 0.1). Average epilimnetic temperatures were similar
between years (2002 average = 25.4, 2003 average = 25.2), and for all response variables
examined, there were no interactive effects of year sampled (P > 0.1).

Microcystis aeruginosa biomass was 3.6 times higher in invaded lakes (Table 2;
Figure 2) and the relative biomass of M. aeruginosa in lakes with D. polymorpha was 5.1
times higher (mean invaded = 14.15%, mean uninvaded = 2.79%, P = 0.001).

For the phytoplankton community exclusive of M. aeruginosa, PCA reduced the
data to two factors that explained 34% of the total variance. PCA factor 1 was
signiﬁcantly higher in invaded lakes (Figure 3), while PCA factor 2 was not different
between invaded and uninvaded lakes (Figure 3). Factor 1 PCA scores were negatively
correlated with Anabaena (r = -0.63, P = 0.009, t-test) and Ceratium (r = -0.79, P =

0.00003, t-test) and positively correlated with Cryptomonas (r = 0.69, P = 0.002, t-test).

This suggests that D. polymorpha had a negative effect on Anabaena and Ceratium, but a
positive effect on Cryptomonas (Figure 4).

Microcystin toxin concentrations were 3.3 times higher in invaded lakes (Table 2;
Figure 5). However, the total biomass of microcystin-producing genera was not different
between lake categories (Table 2). Toxin concentration per unit of microcystin-producing
algal biomass was 2.9 times higher in lakes with D. polymorpha (Table 2; Figure 6A), but
toxin concentration per unit of Microcystis biomass was not different between lake
categories (Table 2; Figure 6B).

To determine if D. polymorpha affected colony sizes of M. aeruginosa, analyses
were conducted on average colony area and median colony area as viewed under the
microscope. Colony sizes were similar in uninvaded and invaded lakes for both average
colony area (Figure 7) and median colony area (P = 0.62).

Dreissena polymorpha also signiﬁcantly affected several limnological parameters.
Both chlorophyll a and phytoplankton dry biomass were 30 % lower in invaded lakes
(Table 2; Figure 8). Extinction coefﬁcients were signiﬁcantly lower in invaded lakes
(Table 2). No difference was found for SRP (Table 2) or TDP (mean invaded = 6.02,
uninvaded = 6.74, P = 0.46) between lake categories, while NI-If—N was 1.8 times higher
in D. polymorpha lakes (Table 2). There was no difference between lake categories for
molar ratios of NHf-N: SRP (mean invaded = 14.93, uninvaded = 19.18, P = 0.31).

Invaded and uninvaded lakes were similar in alkalinity, conductivity, DOC, and
pH (Table 1). Principal components analysis was performed on the major solutes (Ca,
Mg, Na, K, Cl, S04, and Si) to determine the relationship between D. polymorpha

presence on these variables. The PCA reduced the data to two factors that explained 81%

of the total variance. PCA factor 1 was signiﬁcantly higher in invaded lakes (P = 0.04, t-
test) and PCA factor 2 was not different between lake categories (P = 0.23, t-test). Factor
1 PCA scores were positively correlated with Na (r = 0.88, P = 0.2, t-test), K (r = 0.85, P
= 0.14, t-test), C1 (r = 0.89, P = 0.13, t-test), and 804 (r = 0.66, P = 0.0008, t-test), but
only 804 was signiﬁcant.

Discussion

There was a strong positive inﬂuence of D. polymorpha on M aeruginosa
biomass in the survey lakes. This result builds upon a previous lake survey that found the
relative dominance of M. aeruginosa to be higher in invaded lakes (Raikow et al. 2004).
The conﬂicting results between my study and Raikow et al.’s study (2004) may be
because they did not sample enough lakes in the ‘low’ nutrient range, or because they did
not control for lake mean depth or thermal stratiﬁcation. These factors are likely
important because some bloom-forming bluegreens, such as Microcystis, prefer stable
water columns (Dokulil and Teubner 2000). For example, in the well-mixed Hudson
River and in enclosures in the Ohio River, bluegreen abundance actually declined and
diatoms increased in the presence of D. polymorpha (Smith et al. 1998, Jack and Thorp
2000), which suggests mixing regime may play a role in the D. polymorpha — Microcystis
relationship.

I also documented an increase in microcystin toxin concentrations and toxin per
microcystin-producing biomass in D. polymorpha lakes, which has not been shown
previously. However, it is difﬁcult to determine if higher toxin concentrations are a result
of M. aeruginosa or Anabaena, the only other common microcystin-producing genus in

the survey lakes. It may be possible that the increased toxin concentrations in invaded

10

lakes is simply a consequence of increased M. aeruginosa biomass. However, this
explanation assumes the toxin concentrations produced by Anabaena remain the same in
invaded and uninvaded lakes. Thus, the differences in toxin concentrations cannot be
fully explained from my results.

The mechanism by which D. polymorpha promotes M. aeruginosa in low-nutrient
lakes is not fully understood, but three mechanisms have been proposed: 1) the colony
size of M. aeruginosa may be too large for D. polymorpha to feed on effectively, 2)
toxins produced by M. aeruginosa may cause D. polymorpha to preferentially feed on
non-toxic phytoplankton (Vanderploeg et al. 2001) giving M. aeruginosa selective
advantage over non-toxic phytoplankton and 3), the low N: P excretion of D. polymorpha
(Amott and Vanni 1996) may favor phytoplankton that can take advantage of low N
conditions (such as bluegreens).

Although I cannot positively determine which of these mechanisms is operating in
inland lakes, my data can provide some general insights. First, the size of M. aeruginosa
colonies were similar between lake categories and this suggests D. polymorpha did not
shift the size distribution of colonies to be dominated by larger individuals (Figure 7).
The third proposed mechanism may not explain the increased M. aeruginosa dominance
either. NHf-N: SRP ratios were similar in invaded and uninvaded lakes, suggesting that
changes in nutrient availability of N and P may not be promoting M. aeruginosa in D.
polymorpha lakes. A similar lake survey (Raikow et al. 2004) also concluded that the
ratio of available nutrients does not appear to explain bluegreen dominance in low-
nutrient lakes with zebra mussels. Based on my results, the most plausible explanation

for the dominance of M. aeruginosa in D. polymorpha lakes is explained by the second

11

mechanism. Microcystis aeruginosa toxins may cause D. polymorpha to feed on non-
toxic phytoplankton because toxin concentrations were higher in invaded lakes. Feeding
studies have shown that D. polymorpha generally prefer to feed on other, non-toxic
phytoplankton over colonial, toxic M. aeruginosa (Bastviken et al. 1998, Vanderploeg et
al. 2001).

Because microcystin is toxic to animals (Chorus and Bartram 1999), the
promotion of the toxin-producing M. aeruginosa by D. polymorpha could have serious
ecological implications. An increase in the actual biomass of M. aeruginosa may reduce
food-chain efﬁciency because bluegreens are generally of poor nutritional quality and are
not a preferred food source for herbivorous zooplankton (DeMott et a1. 1991, DeMott
1999). The impacts of toxic bluegreens on several aquatic organisms have been identiﬁed
(Fulton and Paerl 1987, Rabergh et al. 1991, Kotak et al. 1996). In particular, toxic
bluegreens have been found to inhibit feeding, reproduction and survivorship of Daphnia
(Fulton and Paerl 1987, de Bemardi and Giussani 1990, DeMott et al. 1991, DeMott
1999). In addition, Lurling (2003) found that as the microcystin concentrations of M.
aeruginosa increased, Daphnia growth rates decreased. Thus, higher toxin concentrations
in D. polymorpha lakes may seriously harm zooplankton and possibly other aquatic
organisms.

I found Dreissena polymorpha to also signiﬁcantly impact the overall
phytoplankton community (exclusive of M. aeruginosa). Dreissena polymorpha appear
to be ﬁltering large quantities of phytoplankton out of the water column because both
chlorophyll a and phytoplankton biomass were signiﬁcantly lower in invaded lakes.

F urtherrnore, extinction coefﬁcients show that D. polymorpha lakes are clearer than lakes

12

without mussels, apparently directly related to the lower phytoplankton biomass in
invaded lakes. The trend of decreased chlorophyll a concentration and increased water
clarity is commonly observed after D. polymorpha invasion in rivers (Efﬂer et al. 1996,
Caraco et al. 1997), the Great Lakes (Holland 1993, Nicholls and Hopkins 1993,
Fahnenstiel et al. 1995), and inland lakes (Yu and Culver 2000, Idrisi et al. 2001, Raikow
et al. 2004).

I also found species-level effects on the phytoplankton community. For example,
Anabaena, a colonial, ﬁlamentous bluegreen and Ceratium, a large dinoﬂagellate, were
negatively affected by D. polymorpha presence, while Cryptomonas, a cryptophyte, was
positively affected by D. polymorpha presence. Bastviken et al. (1998) found Anabaena
relative abundance to decrease in the presence of D. polymorpha although longer
Anabaena ﬁlaments were not as affected as shorter ﬁlaments and Smith et al. (1998)
found Anabaena to be sensitive to D. polymorpha presence in the Hudson River. Thus,
my Anabaena results are consistent with past studies. The effect of D. polymorpha on
Ceratium has been scarcely investigated although Smith et al. (1998) found Ceratium
(relative abundance) to be sensitive to D. polymorpha in the Hudson River, which is in
agreement with my results. However, previous studies show that Cryptomonas does not
always respond in the same manner to D. polymorpha presence. Experimental studies
have found cryptophytes (includes species other than Cryptomonas) (Bastviken et al.
1998) and Cryptomonas (Lavrentyev et al. 1995) to be highly preferred by D.
polymorpha, although Smith et al. (1998) found Cryptomonas (relative abundance) to be

indifferent to D. polymorpha in the Hudson River and Makarewicz et al. (1999) found no

13

difference in cryptophyte biomass before and after D. polymorpha invasion in Lake Erie.
Thus, it is unclear what the general trends associated with Cryptomonas are.

I also found that few water chemistry variables differed between invaded and
uninvaded lakes. For example, a PCA on all ions (Ca, Mg, Na, K, Cl, Si, and 804)
revealed that only S04 was signiﬁcantly higher in D. polymorpha lakes. The effect of D.
polymorpha on chemical parameters, such as major solutes, has been poorly investigated.
In western Lake Erie, silica and chloride increased after D. polymorpha invasion
(Holland et al. 1995), and in Saginaw Bay, particulate silica decreased while dissolved
silica increased (J onengen et al. 1995). In my study, only S04 was signiﬁcantly higher in
invaded lakes and there is no obvious explanation for this result.

The only nutrient that was inﬂuenced by invasion status was NHf-N, which was
1.8 times higher in invaded lakes. My results are only partially consistent with previous
studies. Similar to other studies, I found an increase in NHf-N in D. polymorpha lakes
(Holland et al. 1995, Efﬂer et al. 1996, Heath et al. 1995, Wilson 2003), but no difference
in SRP. It is reasonable to expect NH4+-N and SRP to increase in the presence of D.
polymorpha because mussels can excrete dissolved nutrients at high rates (Quigley et al.
1993, Arnott and Vanni 1996). However, changes to SRP concentrations following D.
polymorpha invasion have been somewhat inconsistent. For example, SRP increased in
the Hudson River (Caraco et al. 1997), Seneca River (Efﬂer et al. 1996), and western
Lake Erie (Holland et a1. 1995), while there was no change in Oneida Lake (Idrisi et al.
2001) and SRP slightly decreased in Saginaw Bay of Lake Huron (J ohengen et al. 1995)
after D- POIymorpha invasion. The lack of a D. polymorpha effect on SRP in my study

may be attributable to the degree of nutrient limitation in the lakes I surveyed. In lakes

l4

where phosphorus strongly limits phytoplankton growth, the demand for SRP may still be
high, despite reductions in phytoplankton biomass from D. polymorpha ﬁltering. SRP
concentrations in the survey lakes were low (~l .6 pg L'l) indicating that any increase in
supply induced by D. polymorpha may be quickly taken up by phytoplankton, making it
difﬁcult to detect an impact on SRP concentration. In systems where D. polymorpha
invasion positively inﬂuenced SRP, concentrations prior to invasion were higher than in
my study (Hudson River ~10 pg L'1 and western Lake Erie ~ 6 pg L'I).

There are many health-related and management consequences of low-nutrient
lakes with D. polymorpha having more toxins than uninvaded lakes. This is especially a
concern since the impacted lakes are low-nutrient and therefore considered to be highly
desirable for recreational purposes. Microcystins are potentially harmful to humans
through consumption and skin exposure and can cause a variety of ailments from rashes
to liver damage as well as the promotion of tumors (Chorus and Bartram 1999). The
World Health Organization has set drinking water guidelines so that microcystin
concentrations should not exceed 1 pg L'1 (WHO 1996). Although none of the survey
lakes exceed this guideline (highest concentration = 0.097 pg L’I), the values reported in
my study may be underestimated. Samples were collected from the deepest spot of the
lake, which was generally in the center of the lake. Bluegreen blooms are often buoyant
and tend to blow towards the shoreline and accumulate densely there. For example,
Johnston and J acoby (2003) found microcystin to range from 1.5 — 3.1 pg L'1 throughout
most of Lake Sammamish, but near the boat launch the concentration was much higher
(43 pg L'l), which they attribute to wind causing accumulations of M. aeruginosa. Thus,

the values I report are likely at the low end of the range and there is the possibility for

15

microcystin values in low-nutrient lakes to exceed guidelines set by the World Health
Organization. In particular, the toxin concentrations may be higher in locations of high
human contact and use, such as along the shoreline and in swimming areas.

Because of the negative consequences that can be associated with microcystin
toxins, it should be a high priority to determine how D. polymorpha are able to promote
M. aeruginosa and microcystin in lakes they invade. Although previous feeding studies
have attempted to answer if D. polymorpha choose other phytoplankton over M.
aeruginosa, more studies are needed that mimic natural conditions. In particular, studies
investigating how a range of colony sizes, rather than unicellular Microcystis versus
colonial Microcystis (Bastviken et al. 1998, Dionisio Pires and Van Donk 2002, Dionisio
Pires et al. 2004) or colony sizes split only into two categories (Vanderploeg et a1. 2001),
are needed. In addition, D. polymorpha may be enhancing strains of M. aeruginosa that
produce high levels of toxins or may be stimulating a phenotypic or evolutionary
response by inducing M. aeruginosa to increase the amount of toxin produced. Thus,
studies aimed at determining if D. polymorpha induce greater toxin production by M.
aeruginosa are also necessary. Many inland North American lakes, particularly those in
Michigan and in bordering states, are relatively low in TP and capable of supporting D.
polymorpha populations (Raikow et al. 2004), and therefore determining how D.

polymorpha increase M. aeruginosa and microcystin is crucial.

16

LITERATURE CITED

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 56: 59-75.

An, J., and W. W. Carmichael. 1994. Use of a colorimetric protein phosphatase inhibition

assay and enzyme linked immunosorbent assay for the study of microcystins and
nodularins. Toxiocon 32: 1495-1507.

Amott, D. L. and M. J. Vanni. 1996. Nitrogen and phosphorus recycling the zebra mussel
(Dreissena polymorpha) in the western basin of Lake Erie. Canadian Journal of
Fisheries and Aquatic Sciences 53: 646-659.

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

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 78: 588-602.

Chorus, I., and J. Bartram, editors. 1999. Toxic cyanobacteria in water: a guide to their
public health consequences, monitoring and management. E & FN Spon, London,
UK.

de Bemardi, R., and G. Giussani. 1990. Are bluegreen algae a suitable food for
zooplankton? An overview. Hydrobiologia 200/201: 29-41.

DeMott. W. R. 1999. Foraging strategies and growth inhibition in ﬁve daphnids feeding
on mixtures of a toxic cyanobacterium and a green alga. Freshwater Biology 42:
263-274.

DeMott, W. R., Q. Zhang, and W. W. Carmichael. 1991. Effects of toxic cyanobacteria
and puriﬁed toxins on the survival and feeding of a copepod and three species of
Daphnia. Limnology and Oceanography 36: 1346-1357.

Dionisio Pires, L. M., R. R. Jonker, E. Van Donk, and H. J. Laanbroek. 2004. Selective
grazing by adults and larvae of the zebra mussel (Dreissena polymorpha):
application of ﬂow cytometry to natural seston. Freshwater Biology 49: 116-126.

Dionisio Pires, L. M., and E. Van Donk. 2002. Comparing grazing by Dreissena

polymorpha on phytoplankton in the presence of toxic and non-toxic
cyanobacteria. Freshwater Biology 47: 1855-1865.

17

Dokulil, M. T., and K. Teubner. 2000. Cyanobacterial dominance in lakes. Hydrobiologia
438: 1-12.

Downing, J. A., S. B. Watson, and E. McCauley. 2001. Predicting cyanobacteria
dominance in lakes. Canadian Journal of Fisheries and Aquatic Sciences 58:
1905-1908.

Efﬂer, S. W., C. M. Brooks, K. Whitehead, B. Wagner, S. M. Doerr, M. Perkins, C. A.
Siegfried, L. Walrath, and R. P. Canale. 1996. Impact of zebra mussel invasion on
river water quality. Water Environment Research 68: 205-214.

Fahnenstiel, G. L., T. B. Bridgeman, G. A. Lang, M. J. McCormick, and T. F. Nalepa.
1995. Phytoplankton productivity in Saginaw Bay, Lake Huron: effects of zebra
mussel (Dreissena polymorpha) colonization. Journal of Great Lakes Research
21: 465-475.

Fulton, R. S. and H. W. Paerl. 1987. Effects of colonial morphology on zooplankton
utilization of algal resources during blue-green algal (Microcystis aeurginosa)
blooms. Limnology and Oceanography 32: 634-644.

Hasle, G. R. 1978. The inverted-microscope method, p. 88-96 in A. Sournia, editor.
Phytoplankton manual. UNESCO, Paris, France.

Heath, R. T., G. L. Fahnenstiel, W. S. Gardner, J. F. Cavaletto, and S. J. Hwang. 1995.
Ecosystem-level effects of zebra mussels (Dreissena polymorpha): an enclosure
experiment in Saginaw Bay, Lake Huron. Journal of Great Lakes Research 21:
501-516.

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.

Holland, R. E., T. H. Johengen, and A. M. Beeton. 1995. Trends in nutrient
concentrations in Hatchery Bay, western Lake Erie, before and after Dreissena
polymorpha. Canadian Journal of Fisheries and Aquatic Sciences 52: 1202-1209.

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

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.

Johengen, T. H., T. F . Nalepa, G. L. Fahnenstiel, and G. Goudy. 1995. Nutrient changes

in Saginaw Bay, Lake Huron, after the establishment of the zebra mussel
(Dreissena polymorpha). Journal of Great Lakes Research 21: 449-464.

18

Johnston, B. R. and J. M. J acoby. 2003. Cyanobacterial toxicity and migration in a
mesotrophic lake in western Washington, USA. Hydrobiologia 495: 79-91.

Kotak, B. G., R. W. Zurawell, E. E. Prepas, and C. F. B. Holmes. 1996. Microcystin-LR
concentration in aquatic food web compartments from lakes of varying trophic
status. Canadian Journal of Fisheries and Aquatic Sciences 53: 1974-1985.

Langner, C. L., and Hendrix, P. F. 1982. Evaluation of a persulfate digestion method for
particulate nitrogen and phosphorus. Water Research 16: 1451-1454.

Lavrentyev, P. J ., W. S. Gardner, J. F. Cavaletto, and J. R. Beaver. 1995. Effects of the
zebra mussel (Dreissena polymorpha) on protozoa and the phytoplankton from
Saginaw Bay, Lake Huron. Journal of Great Lakes Research 21: 545-557.

Lurling, M. 2003. Effects of microcystin-free and microcystin-containing strains of the
cyanobacterium Microcystis aeurginosa on growth of the grazer Daphnia magna.
202-210

Makarewicz, J. C., T. W. Lewis, and P. Bertram, 1999. Phytoplankton composition and

biomass in offshore waters of Lake Erie: pre- and post-Dreissena introduction
(1983-1993). Journal of Great Lakes Research 25: 135-148.

Michigan Sea Grant Inland Lakes Zebra Mussel Infestation Monitoring Program Record.
2001. www.miseagrant.org

Nicholls, K. H., and G. J. HOpkins. 1993. Recent changes in Lake Erie (North Shore)
phytoplankton: cumulative impacts of phosphorus loading reductions and the
zebra mussel introduction. Journal of Great Lakes Research 19: 637-647.

Quigley, M. A., W. S. Gardner, and W. M. Gordon. 1993. Metabolism of the zebra
mussel (Dreissena polymorpha) in Lake St. Clair of the Great Lakes in Zebra

mussels: impacts and control. T. F. Nalepa and D. W. Schloesser, editors. Lewis
Publishers, Ann Arbor, Michigan.

Rabergh, C. M. I., G. Bylund, and J. E. Eriksson. 1991. Histopathological effects of MC-
LR, a cyclic peptide toxin from the cyanobacterium (blue-green alga) Microcystis
aeurginosa, on common carp (Cyprinus carpio L.). Aquatic Toxicology 20: 131-
146.

Raikow, D. F ., O. Samelle, A. E. Wilson, and S. K. Hamilton. 2004. Dominance of the
noxious cyanobacterium Microcystis aeruginosa in low-nutrient lakes is
associated with exotic zebra mussels. Limnology and Oceanography 49: 482-487.

Sandgren, C. D., and J. V. Robinson. 1984. A stratiﬁed sampling approach to

compensating for non-random sedimentation of phytoplankton cells in inverted
microscope settling chambers. British Phycological Journal 19: 67-72.

19

Smith, T. E., R. J. Stevenson, N. F. Caraco, and J. J. Cole. 1998. Changes in
phytoplankton community structure during the zebra mussel (Dreissena

polymorpha) invasion of the Hudson River (New York). Journal of Plankton
Research 20: 1567-1579.

Sprung, M. and U. Rose. 1988. Inﬂuence of food size and food quality on the feeding of
the mussel Dreissena polymorpha. Oecologia 77:526-532.

Ten Winkel, E. H., and C. Davids. 1982. Food selection by Dreissena polymorpha Pallas
(mollusca: bivalvia). Freshwater Biology 12: 553-558.

Trimbee, A. M., and E. E. Prepas. 1987. Evaluation of total phosphorus as a predictor of
the relative biomass of blue-green algae. Canadian Journal of Fisheries and
Aquatic Sciences 44: 1337-1342.

Valderrama, J. C. 1981. The simultaneous analysis of total nitrogen and total phosphorus
in natural waters. Marine Chemistry 10: 109-122.

Vanderploeg, H. A., J. R. Liebig, W. W. Carmichael, M. A. Agy, T. H. Johengen, G. L.
Fahnenstiel, and T. F. Nalepa. 2001. Zebra mussel (Dreissena polymorpha)
selective ﬁltration promoted toxic Microcystis blooms in Saginaw Bay (Lake
Huron) and Lake Erie. Canadian Journal of Fisheries and Aquatic Sciences

58:1208-1221.

Welschmeyer, N. A. 1994. F luorometric analysis of chlorophyll a in the presence of
chlorophyll b and pheopigrnents. Limnology and Oceanography 39: 1985-1992.

Wetzel, R. G. and G. E. Likens. 2001. Limnological analyses, 3rd ed. Saunders.

Wilson, A. E. 2003. Effects of zebra mussels on phytoplankton and ciliates: a ﬁeld
mesocosm experiment. Journal of Plankton Research 25: 905-915.

World Health Organization. Guidelines for drinking water quality, 2nd Edition, WHO,
Geneva, 1996, Volume 2.

Yu, N. and D. A. Culver. 2000. Can zebra mussels change stratiﬁcation patterns in a
small reservoir? Hydrobiologia 431: 175-184.

20

Table 1. Average, range and standard deviation of limnological parameters in survey
lakes. P values are for t-tests. For NHX-N, alkalinity, conductivity, DOC, and pH, P
values are for t-tests only using 2002 data.

 

 

 

Variable Uninvaded Invaded P values
Average (range) SD Average (range) SD

Mean depth 4.93 2.11 6.16 2.8 0.16

(m) (2.31 — 8.8) (2.13 — 12.41)

Alkalinity 2670.43 525.36 2804.83 342.36 0.2

(peq L") (1700 - 3579 (2163 - 3402)

pH 8.06 0.32 8.08 0.12 0.76
(7.66 — 8.89) (7.82 - 8.29)

Conductivity 329.36 92.29 400.67 114.86 0.073

(mg L") (188-483) (270—681)

DOC 8.38 2.95 7.18 1.53 0.23

(mg L'l) (5.35 — 15.82) (4.76 — 9.3)

TP 10.4 3.83 10.62 3.43 0.73

(118 L") (4.9 — 19.32) (5.32 - 19.25)

SRP 1.63 0.87 1.61 1.57 0.94

(pg L") (0.08 — 3.72) (0 — 6.08)

NHf-N 6.81 4.98 12.4 7.04 0.015

(pg N L") (O — 15.6) (6.5 - 28)

Extinction -0.275 0.07 -0.212 0.05 0.003

coefﬁcient (- 0.428 —- - 0.16) (- 0.308 — - 0.06)

Chlorophyll a 4.59 1.77 3.23 1.1 0.008

(Mg L’I) (2.36 -- 9.29) (0.89 - 5.25)

Phytoplankton 50.77 19.44 35.33 22.16 0.006

biomass (18.71 — 95.08) (12.88 — 115.8)

SHE L")

 

21

Table 2. Average, range, and standard deviation of variables related to M. aeruginosa and
microcystin toxin. P values are for t-tests.

 

 

 

Variable Uninvaded Invaded P values
Average SD Average SD
(range) (rage)
M. aeruginosa 1.33 1.34 4.8 5.44 0.01
biomass (0 — 4.17) (0.11 — 20.5)
(118 L")
Microcystin- 5.18 4.49 5 .69 5.52 0.83
producing (0 — 13.2) (0.16 — 21.67)
biomass (pg L'I)
Microcystin 9.87 8.25 33.06 30.55 0.009
concentration (2.14 — 31.38) (1.6 — 96.64)
(n8 L")
Microcystin per 2.76 2.62 7.88 5.39 0.00009
producing (0 — 11.49) (0.83 — 20.6)
biomass (ng pg“)
Microcystin per 0.019 0.035 0.013 0.011 0.5
M. aeruginosa (0 — 0.158) (0.004 - 0.05)

biomass (11g pg")

22

 

 

 

 

Figure 1. Location of survey lakes in the lower peninsula of Michigan. The state is
divided into Albert ecoregion subsections. Closed circles represent invaded lakes and
open circles represent uninvaded lakes.

23

 

oo

N

 

 

 

 

 

 

 

O

 

Uninvaded Invaded

Microcystis dry mass (pg L'1)
4:. c»

Figure 2. Microcystis aeruginosa dry mass in lakes with and without D. polymorpha (P =
0.01). P value for a t-test. Bars are SE.

24

 

1 ‘ o o o o: 00;). .g o
N0 0 008082......
a -1 - '

“8 -2- - ’
U- . Invaded

'3 I 0 Uninvaded

-4 - o

-5 .

-3 -2 -1 0 1
Factor 1

 

 

 

 

 

Figure 3. Factor scores from the phytoplankton (exclusive of M. aeruginosa) PCA. Factor
1 (P = 0.00002) and Factor 2 (P = 0.65), which indicates a positive effect of D.
polymorpha on Factor 1. P values are for t-tests.

25

100

§
780
U)
m
E 60-
.2
m.
a, 40-
.2
E 20-
8

o

 

 

 

 

_\

N

   

 

 

 

Uninvaded Invaded

 

% Others

’ %Anabaena

lﬂ-—l

% Ceratium
% Cryptomonas
% Microcystis

Figure 4. Relative biomass values (%) of the phytoplankton genera affected by D.
polymorpha presence (Anabaena, Ceratium, Cryptomonas, and Microcystis) and the
remaining phytoplankton grouped together as “other”.

26

 

 

 

 

‘7

_l

m 30

C

v 20

.E

2‘, 10 T
I-

 

 

 

 

 

 

 

Uninvaded Invaded

Figure 5. Microcystin concentration in lakes with and without D. polymorpha (P =
0.009). P value for a t-test. Bars are SE.

27

.3
O

 

moo

 

Toxin (ng pg'1)

 

 

 

ON-b

 

 

 

 

Uninvaded Invaded

0.04

 

0.03 .

0.02 : I

 

0.01

Toxin (ng pg'I)

 

 

 

 

 

 

 

0.00
Uninvaded Invaded

Figure 6. (A) Microcystin concentration per microcystin-producing biomass in lakes with
and without D. polymorpha (P = 0.0009). Microcystin producing genera are Anabaena,
Microcystis, Oscillatoria, and Nostoc. (B) Microcystin concentration per M. aeruginosa
biomass (P = 0.5). P values are for t-tests. Bars are SE.

28

 

 

 

 

 

 

 

 

 

6?
E 10000
:3.
V
m 8000 -
3 l
< 6000 l
8) 4000 I
co
5
g 2000 ~
< o

Uninvaded Invaded

Figure 7. Colony size of M. aeruginosa as average area (P = 0.47). P value for a t-test.
Bars are SE.

29

 

 

 

CANOO-hU'IO}

 

 

 

 

 

 

Chlorophyll 3 (pg L-1)

Uninvaded Invaded

Figure 8. The effect of D. polymorpha on chlorophyll a (P = 0.008). P value is for a t-test.
Bars are SE.

30

CHAPTER 2

Inﬂuence of zebra mussels (Dreissena polymorpha) on microzooplankton communities in
stratiﬁed low-nutrient lakes

ABSTRACT
Although planktonic rotifers and ciliates are an important component of secondary
production in lakes, few studies have investigated how zebra mussels (Dreissena
polymorpha) affect these organisms. Existing studies have largely been conducted in
well-mixed systems, yet D. polymorpha is invading stratiﬁed inland lakes at a rapid rate.
It is very likely that the effects of D. polymorpha may vary with mixing regime. Thus, I
conducted a lake survey to examine how D. polymorpha inﬂuences the biomass and
community structure of rotifers and ciliates in stratiﬁed lakes. Forty-six low nutrient
lakes were sampled, 24 with D. polymorpha and 22 without. The total biomass of rotifers
and ciliates were lower in invaded lakes (44% and 45%, respectively) and in both cases,
the difference between invaded and uninvaded lakes was highly signiﬁcant. Thus, I found
no evidence that D. polymorpha invasion has stronger negative effects on ciliates than
rotifers in general. Lakes with zebra mussels also had lower rotifer species richness and
diversity. In addition, a shift in the size distribution of ciliates was found between lake
categories indicating that zebra mussels may be reducing larger, algivorous ciliates in
favor of smaller bacterivorous ciliates. In general, zebra mussel impacts on
microzooplankton biomass in the stratiﬁed lakes I sampled were weaker than those

reported from well-mixed systems.

31

Introduction

Microzooplankton (heterotrophic ﬂagellates, ciliates and, in freshwaters, rotifers)
comprise, at times, a major fraction of total zooplankton biomass in freshwater lakes
(Pace and Orcutt 1981, Gates and Lewg 1984). Microzooplankton function as consumers
of bacteria and small phytoplankton and as a food resource for larger organisms (e. g.
macrozooplankton, larval ﬁsh), and thus can link microbes to higher trophic levels (Sherr
et al. 1986, Sherr et al. 1987, Sanders et al. 1989, Amdt 1993). Microzooplankton
biomass is strongly regulated by predation and resource competition with
macrozooplankton (Gilbert 1988, Pace and Funke 1991, Wickham and Gilbert 1991, Jack
and Gilbert 1994, Marchessault and Mazumder 1997). Zebra mussels directly consume
microzooplankton (MacIsaac et al. 1991, MacIsaac et al. 1995) and act as a resource
competitor by reducing phytoplankton abundance (Fahnenstiel et al. 1995, Heath et a1.
1995, Caraco et al. 1997, Idrisi et al. 2001). Thus, it is not surprising that invasion of
freshwater habitats by the exotic zebra mussel (Dreissena polymorpha) has been
associated with large declines in microzooplankton abundance (MacIsaac et a1. 1995,
Pace et al. 1998).

Negative effects of D. polymorpha on abundance of ciliates and rotifers have been
observed largely in well-mixed systems as before and after studies (MacIsaac et al. 1995,
Pace et al. 1998) or experimentally (MacIsaac et al. 1995, Lavrentyev et al. 1995, Jack
and Thorp 2000, Thorp and Casper 2002, Wilson 2003). The few studies conducted in
stratiﬁed systems have only examined macrozooplankton (crustaceans, copepods) (Idrisi
et a1. 2001) or combined zooplankton (crustacean + rotifer) biomass (Yu and Culver

1999). In thermally stratiﬁed systems, contact between the pelagic biota and benthic ﬁlter

32

feeders like D. polymorpha should be reduced relative to well-mixed systems (MacIsaac
and Sprules 1991, MacIsaac 1996, Noonburg et al. 2003). Thus, we might expect D.
polymorpha impacts on the plankton to be weaker in stratiﬁed systems. The ongoing
Dreissena invasion of inland lakes in the upper Midwest US, especially in Michigan,
provided an opportunity to examine this question.

Dreissena polymorpha ﬁltering also has the potential to affect ciliate and rotifer
biomass differently. The preferred food size range of zebra mussels (5-45 pm, Ten
Winkle and Davids 1982, Sprung and Rose 1988) indicates that they should inﬂict greater
mortality on ciliates than rotifers, since ciliate cell sizes are commonly within the
preferred range, while rotifers are generally larger (often > 100 pm). However, the
abundance of bacteria, a potential food source for microzooplankton, is typically not
affected by D. polymorpha presence (Cotner et al. 1995, Findlay et al. 1998) because of
the small size of bacteria. Ciliates generally consume bacteria more effectively than
rotifers. Thus, the relative magnitude of Dreissena's predatory effect may be larger for
ciliates, while Dreissena's competitive effect may be larger for rotifers (Figure 1). It is
not obvious whether mussel invasion will have a greater overall impact on ciliates or
rotifers. Previous studies have considered the effects of D. polymorpha on ciliates and
rotifers separately and have not compared differences in magnitude.

Dreissena polymorpha may also have species-speciﬁc effects on rotifers and
ciliates. Because D. polymorpha are size-selective (Ten Winkle and Davids 1982, Sprung
and Rose 1988), they may shift the size distribution of ciliates toward species larger than
45pm, for example. Additionally, few studies have examined the effects of D.

polymorpha on species diversity in general and those that have were focused mainly on

33

benthic invertebrates (Stewart and Haynes 1994) and native unionids (Herbert et al.
1991). No studies have assessed whether D. polymorpha invasion affects the diversity of
pelagic assemblages.

In this paper, I examine the inﬂuence of D. polymorpha invasion in stratiﬁed
lakes on two groups of microzooplankton, ciliates and rotifers, via a large-scale lake
survey. The main questions I address are: 1) do zebra mussels reduce microzooplankton
biomass in stratiﬁed inland lakes as much as they do in well-mixed systems?, 2) within
microzooplankton, are ciliates or rotifers more negatively affected by zebra mussel
presence?, and 3) how do zebra mussels impact rotifer diversity?

Methods
Lake Selection

The survey was restricted to low-nutrient lakes (TP < 21 pg L") in southern
Michigan (Figure 2) with a maximum depth of a least 9 m, such that thermal stratiﬁcation
was likely to be present during the summer. For the purposes of lake selection, the
presence/absence of D. polymorpha was initially assessed using the list of invaded lakes
assembled by the Michigan Sea Grant College Program (www.miseagrant.org).
Presence/absence of D. polymorpha was later veriﬁed in each lake by searching for adults
in the ﬁeld and for veligers in zooplankton samples. To ensure that invaded and
uninvaded lakes included in the survey were similar in depth, mean depth was determined
by digitizing bathymetric lake maps in ArcView GIS 3.2. Based on these criteria, 46
lakes (24 with D. polymorpha and 22 without) were selected and rotifer biomass was
assessed in all lakes, while ciliate biomass was only assessed in 38 randomly chosen

lakes within each lake group (20 with D. polymorpha and 18 without).

34

Field Sampling

Each lake was visited once, in late summer of 2002 or 2003 (2 August — 4
September 2002, 29 July — 20 August 2003). Samples were taken from the deepest point
in each lake, as determined by bathymetric maps and a depth ﬁnder. Temperature was
measured at 1 m intervals with a Hydrolab Surveyor and Datasonde. A depth-integrated
water sample was taken through the entire mixed layer (epilimnion) with a ﬂexible tube
(5 cm inner diameter) and placed on ice until processing (approximately 6 hours).
Subsamples of epilimnetic water were taken for TP, chlorophyll a, and ciliates (125 mL
preserved in Lugol’s solution). Rotifers were collected by passing 10 L subsample
through a 35-pm mesh screen and rinsing organisms on the screen into sample bottles
containing glutaraldehyde (ﬁnal concentration: 2%).
Laboratory Analysis

TP was analyzed colorimetrically (Langner and Hendrix 1992) after persulfate
oxidation. For chlorophyll a analysis, samples were ﬁltered onto a Gelman A/E glass
ﬁber ﬁlter and kept frozen. The ﬁlter was subsequently extracted with 95% ethanol and
chlorophyll a was quantiﬁed using a Turner Model 10-AU ﬂuorometer calibrated with
standards (Welschmeyer 1994). Dissolved organic carbon (DOC) was measured with a
high-temperature combustion DOC analyzer. Ciliates were identiﬁed and enumerated
with the inverted microscope technique using a Nikon model TE2000-S inverted
microscope. Subsamples (subsample volume 30 — 100 mL) were settled in tubular
chambers (Hydro-Bios), the bottoms of which were divided into inner and outer zones of
equal area (Sandgren and Robinson 1984). Within each zone, at least 20 random ﬁelds

were counted at 100x. Ciliate cell volume was determined by measuring at least ﬁve

35

individuals per taxon at 400x using a digital camera and image-analysis software.
Biovolume was calculated by using a geometric equation appropriate for the species
shape. Biovolume was converted to dry biomass assuming a speciﬁc gravity of 1 g cm'3
and a dry mass to wet mass ratio of 0.10. Rotifers were identiﬁed to species using a
Nikon model E600 compound microscope at 100x and a Sedgwick-Rafter counting
chamber. For each lake, approximately 400 individuals (average 394, range 341 — 475)
were counted in a minimum of 2 subsarnples. Dry biomass for each species was
estimated from established literature values (Pauli 1989). Rotifer richness was quantiﬁed
as the total number of species in each lake and diversity was assessed using the Shannon
index, (Shannon and Weaver 1949), H' = - 2(pi ln p), where pi is the proportion of
individuals found in the ith species.
Statistical analyses

There was not an effect of year sampled on any variable (P > 0.10), so t-tests were
used to determine the inﬂuence of D. polymorpha on all variables. Rotifer assemblage
structure was reduced via principal components analysis (PCA) on the relative biomass of
each genus. To reduce zeros in the data set, the genus level was used for the PCA. I
restricted the PCA to the following common taxa: Ascomorpha, Colletheca, Conochilus,
Keratella, Polyarthra, Synchaeta, and T richocerca. Data were log-transformed as
necessary so that assumptions for parametric statistics were met. All analyses were
conducted in SYSTAT 9.0.
Results

To rule out the inﬂuence of potential confounding factors, I assessed whether

invaded and uninvaded lakes were similar in morphometry and nutrients. There were no

36

signiﬁcant differences in either mean depth or TP between the two groups of lakes (Table
1). Average epilimnetic temperatures were similar between years (2002 average = 25.4,
2003 average = 25.2), and for all response variables examined, there were no interactive
effects of year (P > 0.1).

Total ciliate biomass was 45% lower in invaded lakes (Figure 3). Similarly, total
rotifer biomass was 44% lower in lakes with D. polymorpha (Figure 3). Chlorophyll a
was 26% lower in invaded lakes (Table 1, Figure 4A) and there was no difference in
DOC between lake categories (Figure 4B).

In order to examine if D. polymorpha presence affected the size distribution of
ciliates, I divided ciliates into two size classes (< 30 pm and > 30 pm). The biomass of
ciliates in the < 30 pm range was similar between lake categories, but the biomass of >
30 pm ciliates was signiﬁcantly lower in invaded lakes (Figure 5B). As a consequence,
zebra mussel invasion signiﬁcantly shifted the ciliate assemblage toward dominance by
smaller size category.

Uninvaded lakes had 1.2 times higher rotifer richness (Figure 6A) and 1.2 times
higher diversity as measured by the Shannon index (Figure 6B). By using PCA, rotifer
relative biomass was reduced to two factors that explained 46% of the total variance.
Zebra mussel presence was signiﬁcantly related to PCA factor 1 (Figure 7), but not to
factor 2 (Figure 7). Factor 1 PCA scores were negatively correlated with Keratella
relative biomass although is was not signiﬁcant (r = -0.82, P = 0.15, t-test). Whereas, D.
polymorpha was positively correlated to Polyarthra relative biomass (r = 0.71 , P = 0.028,

t-test) suggesting that D. polymorpha had a positive inﬂuence on Polyarthra.

37

Discussion

My results suggest that D. polymorpha has strong negative effects on
microzooplankton in stratiﬁed inland lakes (Figure 3). However, the magnitude of effect
of zebra mussels on microzooplankton appears to be weaker in stratiﬁed lakes. In my
study, lakes with D. polymorpha had 44% lower rotifer biomass and 45% lower ciliate
biomass. Experimental studies have shown D. polymorpha presence to reduce ciliate
biovolume by 77% (Wilson 2003) and protozoan abundance by 70-80% (Lavrentyev et
al. 1995). In the Hudson River, total zooplankton biomass declined by 70% (Pace et al.
1998) and mean total zooplankton density (excluding ciliates) was 55-71% lower in Lake
Erie (MacIsaac et al. 1995) following D. polymorpha invasion, and in both cases,
reductions were mainly attributed to negative effects on rotifers. The lesser reduction in
my survey compared to Lake Erie and Hudson River may be attributable to mixing
regime. The plankton in shallow, well-mixed systems may experience greater D.
polymorpha impacts than deep, stratiﬁed systems (MacIsaac and Sprules 1991, MacIsaac
1996, Noonburg et al. 2003) because pelagic organisms are more likely to come into
contact with benthic populations of D. polymorpha in systems lacking distinct vertical
stratiﬁcation.

Although it was unclear whether D. polymorpha would have a greater impact on
rotifers or ciliates, I expected D. polymorpha to inﬂict greater predatory effects on
ciliates than rotifers and greater competitive effects on rotifers. However, my results
show that the magnitude of effect was remarkably similar (45% versus 44%). Although it
seems likely that D. polymorpha should ﬁlter ciliates more effectively than rotifers

because ciliates are generally smaller, my results indicate that D. polymorpha may ﬁlter

38

rotifers and ciliates similarly. This is surprising given that feeding studies have shown D.
polymorpha to ingest some size ranges (5-45 pm) more effectively (Ten Winkle and
Davids 1982, Sprung and Rose 1988) and the majority of ciliates in my study were in this
range (85% in uninvaded and 98% in invaded, based on the biomass in this size range).

The unexpected similar effect of D. polymorpha on ciliates and rotifers may also
be explained by bacteria abundance and the feeding behavior of microzooplankton.
Typically, D. polymorpha does not reduce the total abundance of bacteria in systems they
invade (Cotner et al. 1995, Findlay et al. 1998) despite their ability to consume bacteria
(Roditi et al. 1996, Frischer et a1. 1998). Planktonic ciliates feed primarily on other
protozoa and phytoplankton (Fenchel 1987), but they also consume bacteria and
sometimes rely on bacteria as a key resource (Christophersen et al. 1990). While rotifers
can feed on bacteria with limited ability (Amdt 1993), they generally consume
phytoplankton between 4-17 pm because particles larger and smaller are difﬁcult for
them to ingest (Bogdan et al. 1980, Bogdan and Gilbert 1984). Thus, in invaded lakes
ciliates may be able to take advantage of bacteria more easily than rotifers, allowing
ciliates to compensate for predation losses and phytoplankton reductions.

It cannot be determined from my survey results whether indirect (competition) or
direct (predation) mechanisms account for the negative impact of D. polymorpha on
microzooplankton. As mentioned above, it is likely that predation is an important factor.
However, phytoplankton biomass (measured by chlorophyll a) was signiﬁcantly lower in
invaded lakes, as found in previous studies (Fahnenstiel et al. 1995, Caraco et al. 1997,
Idrisi et al. 2001). By consuming phytoplankton, D. polymorpha are competing with

microzooplankton for resources. Thus, it is reasonable to assume that reductions in

39

phytoplankton, mediated through D. polymorpha, could indirectly affect
microzooplankton abundance. However, previous studies conclude that D. polymorpha
predation may be more important than resource competition in reducing
microzooplankton abundance (MacIsaac and Sprules 1991, MacIsaac et al. 1995, Thorp
and Casper 2002). In these studies, small-bodied zooplankton were primarily reduced
while large-bodied were not, even though both compete for resources with D.
polymorpha. However, these experiments were either conducted in small containers
(MacIsaac et al. 1991, MacIsaac et al. 1995), or short-tenn (Thorp and Casper 2002). In
small-scale experiments, predators and prey may experience greater spatial overlap than
in thermally stratiﬁed lakes, which may exaggerate the importance of predation over
resource competition, particularly on smaller organisms (Samelle 1997). Short-term
experiments may also overstate the importance of predation on microzooplankton
because the effects of resource competition often take longer to observe than those of
predation (Samelle 1997). F urtherrnore, it is possible that microzooplankton communities
could be regulated by macrozooplankton in the survey lakes. Macrozooplankton are
known to prey on ciliates (Pace and Funke 1991, Burns and Gilbert 1993, Jack and
Gilbert 1993, Wiackowski et al. 1994) and compete with rotifers for resources (Gilbert
1985). Consequently, macrozooplankton may be an extremely important covariate and
the absence of data on their biomass makes it difﬁcult to conclude that D. polymorpha is
the sole factor controlling microzooplankton in the survey lakes.

Given that zebra mussels exert more grazing pressure on phytoplankton than
bacteria, invasion might lead to a shift in ciliate assemblages toward greater dominance

by bacterivorous ciliates. One way to examine this is to divide ciliates into size classes,

40

ciliates < 30 pm can be roughly considered bacterivorous and those > 30 pm algivorous
(Fenchel 1987). By dividing ciliates into these classes, I found a shift toward smaller
(<30 pm) ciliates in invaded lakes (Figure 5), indicating bacterivorous ciliates may be
more prominent in D. polymorpha lakes. The size-selective predation hypothesis would
expect < 30 pm ciliates to be more negatively affected by D. polymorpha than > 30 pm
ciliates. It is possible that smaller ciliates are able to compensate for predatory losses
better than larger ciliates because smaller ciliates may not be experiencing such severe
resource competition with D. polymorpha. Because I did not measure bacteria, 1 cannot
determine if bacteria abundance was similar in invaded and uninvaded lakes. However,
DOC was similar between lake categories (Figure 4) and since bacteria use DOC (Figure
1), this indicates that bacteria resources may not differ between uninvaded and invaded
lakes.

Although evidence suggests that D. polymorpha increases benthic invertebrate
diversity (Stewart and Haynes 1994) but decreases native unionid diversity (Herbert et al.
1991), no studies have investigated possible diversity changes in pelagic organisms. Both
rotifer richness and diversity were lower in invaded lakes. Lower richness might be
attributed to D. polymorpha’s ability to ﬁlter large quantities of material that span a wide
size range (Ten Winkel and Davids 1982). A reduction in richness may indicate that D.
polymorpha strongly preys on and competes with rotifers. Only rotifer species with
effective defense mechanisms (e. g. spines, large size, escape mechanisms) may be able to
avoid ingestion, so rotifers without such defenses should be severely affected in zebra
mussel lakes. For example, faster swimming species may be able to escape D.

polymorpha ﬁltering currents (MacIsaac and Sprules 1991). The only taxa that differed

41

between uninvaded and invaded lakes was Polyarthra and its relative biomass was
signiﬁcantly higher in invaded lakes. Unlike most rotifers, Polyarthra is able to avoid
predators (e. g. Asplanchna, Chaoborus, Daphnia) by a jump mechanism (Gilbert and
Williamson 1978, Gilbert 1985, Gilbert 1987, Moore and Gilbert 1987). Polyarthra may
be able to escape D. polymorpha through this jump mechanism. However, Polyarthra
abundance was dramatically reduced by D. polymorpha presence in experiments
(MacIsaac and Sprules 1991, MacIsaac et al. 1995, Thorp and Casper 2002), the Hudson
River (Pace et al. 1998), and Lake Erie (MacIsaac et a1. 1995). In these studies,
particularly the experiments, Polyarthra may come into such close proximity with D.
polymorpha that its jump mechanism cannot facilitate escape. However, Polyarthra is a
selective feeder on cryptophyte phytoplankton (Gilbert and Bogdan 1984) and the
biomass of the cryptophyte, Cryptomonas was found to be higher in invaded lakes
(Chapter 1), thus possibly allowing Polyarthra to withstand relatively high predation
rates. It must also be noted that because I sampled each lake only once per summer, my
rotifer diversity results might represent natural community ﬂuctuations since rotifer
communities can change rapidly. However, I restricted my study so that lakes were
sampled within a few weeks of each other during a stable time of late summer
stratiﬁcation. Controlling for time should help alleviate problems associated with natural
ﬂuctuations.

My results show that D. polymorpha has the potential to signiﬁcantly alter the
ciliate and rotifer community assemblages. The data also suggest that similar to well-
mixed systems, microzooplankton are negatively impacted by D. polymorpha in stratiﬁed

inland lakes. However, the magnitude of effect may be weaker in stratiﬁed than in well-

42

mixed systems. Despite this potential dampening, the considerable reduction of
microzooplankton (25 %) in D. polymorpha lakes may extend to higher trophic levels.
For example, reduced zooplankton biomass may decrease larval or planktivorous ﬁsh
growth rates. However, if no negative effects of D. polymorpha on macrozooplankton are
seen, ﬁsh may not be affected because an adequate food supply may still be available.
Thus, it is important to determine how D. polymorpha affect macrozooplankton in

stratiﬁed systems.

43

LITERATURE CITED

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 56: 59-75.

Amdt, H. 1993. Rotifers as predators on components of the microbial web (bacteria,
heterotrophic ﬂagellates, ciliates) — a review. Hydrobiologia 255/256: 231-246.

Bogdan, K. G. and J. J. Gilbert. 1984. Body size and food size in freshwater zooplankton.
Proceedings of the National Academy of Sciences (USA) 81: 6427-6431.

Bogdan, K. G., J. J. Gilbert, and P. L. Starkweather. 1980. In situ clearance rates of
planktonic rotifers. Hydrobiologia 73: 73-77.

Burns, C. W. and J. J. Gilbert. 1993. Predation on ciliates by freshwater calanoid
copepods: rates of predation and relative vulnerabilities of prey. Freshwater
Biology 30: 377-393.

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 78: 588-602.

Christoffersen, K., B. Riemann, L. R. Hansen, A. Klysner, and H. B. Sorensen. 1990.
Qualitative importance of the microbial loop and plankton community structure in

a eutrophic lake during a bloom of cyanobacteria. Microbial Ecology 20: 253-
272.

Cotner, J. B., W. S. Gardner, J. R. Johnson, R. H. Sada, J. F. Cavaletto, and R. T. Heath.
1995. Effects of zebra mussels (Dreissena polymorpha) on bacterioplankton:
evidence for both size-selective consumption and growth stimulation. Journal of

Great Lakes Research 21: 517-528.

Fahnenstiel, G. L., T. B. Bridgeman, G. A. Lang, M. J. McCormick, and T. F. Nalepa.
1995. Phytoplankton productivity in Saginaw Bay, Lake Huron: effects of zebra
mussel (Dreissena polymorpha) colonization. Journal of Great Lakes Research
21: 465-475.

F enchel. T. 1987. Ecology of protozoa: the biology of free-living phagotrophic protests.
Springer-Verlag, Berlin.

Findlay, S., M. L. Pace, and D. T. Fischer. 1998. Response of heterotrophic planktonic
bacteria to the zebra mussel invasion of the tidal freshwater Hudson River.
Microbial Ecology 36: 131-140.

44

 

F rischer, M. E., S. A. Nierzwicki-Bauer, R. H. Parsons, K. Vathanodom, and K. R.
Waitkus. 2000. Interactions between zebra mussels (Dreissena polymorpha) and

microbial communities. Canadian Journal of Fisheries and Aquatic Sciences 57:
591-599.

Gates, M. A. and U. T. Lewg. 1984. Contribution of ciliated protozoa to the planktonic
biomass in a series of Ontario lakes: quantitative estimates and dynamical
relationships. Journal of Plankton Research 6: 443-456.

Gilbert, J. J. 1985. Competition between rotifers and Daphnia. Ecology 66: 1943-1950.

Gilbert, J. J. 1987. The Polyarthra escape response: defense against interference from
Daphnia. Hydrobiologia 147: 235-238.

Gilbert, J. J. 1988. Suppression of rotifer populations by Daphnia: a review of the
evidence, the mechanisms, and the effects on zooplankton community structure.
Limnology and Oceanography 33: 1286-1303.

Gilbert, J. J. and K. G. Bogdan. 1984. Rotifer grazing: in situ studies on selectivity and
rates, pp. 97-133. In D. G. Meyers and J. R. Strickler [eds], Trophic interactions
within aquatic ecosystems. American Association for the Advancement of
Science Symposium 85. Westview Press, Boulder, CO, USA.

Gilbert, J. J. and C. E. Williamson. 1978. Predator-prey behavior and its effect on rotifer
survival in associations of Mesocylops edax, Asplanchna girodi, Polyarthra
vulgaris, and Keratella cochlearis. Oecologia 37: 13-22.

Heath, R. T., G. L. Fahnenstiel, W. S. Gardner, J. F . Cavaletto, and S. J. Hwang. 1995.
Ecosystem-level effects of zebra mussels (Dreissena polymorpha): an enclosure
experiment in Saginaw Bay, Lake Huron. Journal of Great Lakes Research 21:
501-516.

Herbert, P. D., C. C. Wilson, M. H. Murdoch, R. Lazar. 1991. Demography and
ecological impacts of the invading mollusk Dreissena polymorpha. Canadian
Journal of Zoology 69: 405-409.

Idrisi, N., E. L. Mills, L. G. Rudstam, and D. J. Stewart. 2001. Impact of zebra mussels
(Dreissena polymorpha) on the pelagic lower trophic levels of Oneida Lake, New

York. Canadian Journal of Fisheries and Aquatic Sciences 58: 1430-1441.

Jack, J. D. and J. J. Gilbert. 1993. Susceptibilities of different-sized ciliates to direct
suppression by small and large cladocerans. Freshwater Biology 29: 19-29.

Jack, J. D. and J. J. Gilbert. 1994. Effects of Daphnia on microzooplankton communities.
Journal of Plankton Research 16: 1499-1512.

45

 

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: 569-5 79.

Langner, C. L., and P. F. Hendrix. 1982. Evaluation of a persulfate digestion method for
particulate nitrogen and phosphorus. Water Research 16: 1451-1454.

Lavrentyev, P. J ., W. S. Gardner, J. F. Cavaletto, and J. R. Beaver. 1995. Effects of the
zebra mussel (Dreissena polymorpha) on protozoa and the phytoplankton from
Saginaw Bay, Lake Huron. Journal of Great Lakes Research 21: 545-557.

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 Fisheries and Aquatic Sciences 48:
2051-2060.

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

MacIsaac, H. J. 1996. Potential abiotic and biotic impacts of zebra mussels on the inland
waters of North America. American Zoologist 36: 287-299.

Marchessault, P., and A. Mazumder. 1997. Grazer and nutrient impacts on epilimnetic
ciliate communities. Limnology and Oceanography 42: 893-900.

Michigan Sea Grant Inland Lakes Zebra Mussel Infestation Monitoring Program Record.
2002. www.miseagrant.org

Moore, M. V. and J. J. Gilbert. 1987. Age-speciﬁc Chaoborus predation on rotifer prey.
Freshwater Biology 17: 223-236.

Noonburg, E. G., B. J. Shuter, and P. A. Abrams. 2003. Indirect effects of zebra mussels
(Dreissena polymorpha) on the planktonic food web. Canadian Journal of
Fisheries and Aquatic Sciences 60: 1353-1368.

Pace, M. L. and J. D. Orcutt, Jr. 1981. The relative importance of protozoans, rotifers,
and crustaceans in a ﬁeshwater zooplankton community. Limnology and

Oceanography 26: 822-830.

Pace, M. L. and E. Funke. 1991. Regulation of planktonic microbial communities by
nutrients and herbivores. Ecology 72: 904-914.

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.

Pauli, H. R. 1989. A new method to estimate individual dry weights of rotifers.
Hydrobiologia 186: 355-361.

46

Roditi, H. A., N. F. Caraco, J. J. Cole, and D. L. Strayer. 1996. Filtration of Hudson
River water by the zebra mussel (Dreissena polymorpha). Estuaries. 19: 824-832.

Sanders, R. W., K. G. Porter, S. J. Bennett, and A. E. DeBiase. 1989. Seasonal patterns of
bacterivory by ﬂagellates, ciliates, rotifers, and cladocerans in a freshwater
planktonic community. Limnology and Oceanography 34: 673-687.

Samelle O. 1997 . Daphnia effects on microzooplankton: comparisons of enclosure and
whole-lake responses. Ecology 78: 913-928.

Shannon, C. E. and W. Weaver. 1949. The mathematical theory of communication. y
University of Illinois Press, Urbana. i

Sherr, E. B. and B. F. Sherr. 1987. High rates of consumption of bacteria by pelagic
ciliates. Nature 325: 710-711. i

 

Sherr, E. B., B. F. Sherr, and G. A. Paffenhofer. 1986. Phagotrophic protozoa as food for
metazoans: a “missing” trophic link in marine pelagic food webs? Marine
Microbial Food Webs 1: 61-80.

Sprung, M. and U. Rose. 1988. Inﬂuence of food size and food quality on the feeding of
the mussel Dreissena polymorpha. Oecologia 77:526-532.

Stewart, T. W., and J. M. Haynes. 1994. Benthic macroinvertebrate communities of
southwestern Lake Ontario following invasion of Dreissena. Journal of Great
Lakes Research 20: 479-493.

Ten Winkel, E. H., and C. Davids. 1982. Food selection by Dreissena polymorpha Pallas
(mollusca: bivalvia). Freshwater Biology 12: 553-558.

Thorp, J. H. and A. F. Casper. 2002. Potential effects on zooplankton from species shifts
in planktivorous mussels: a ﬁeld experiment in the St. Lawrence River.
Freshwater Biology 47: 107-119.

Welschmeyer, N. A. 1994. Fluorometric analysis of chlorophyll a in the presence of
chlorophyll b and pheopigrnents. Limnology and Oceanography 39: 1985-1992.

Wickham, S. A. and J. J. Gilbert. 1991. Relative vulnerabilities of natural rotifer and
ciliate communities to cladocerans: laboratory and ﬁeld experiments. Freshwater
Biology 26: 77-86.

Wiackowski, K., A. M. Ventela, M. Moilanen, V. Saarikari, K. Vuorio, and J. Sarvala.

2001. What factors control planktonic ciliates during a summer in a highly
eutrophic lake? Hydrobiologia 443: 43-57.

47

Wilson, A. E. 2003. Effects of zebra mussels on phytoplankton and ciliates: a ﬁeld
mesocosm experiment. J oumal of Plankton Research 25: 905-915.

Wong, W. H., J. S. Levinton, B. S. Twining, and N. Fisher. 2003. Assimilation of micro-
and mesoplankton by zebra mussels: a demonstration of the food web link

between zooplankton and benthic suspension feeders. Limnology and
Oceanography 48: 308-312.

Yu, N. and D. A. Culver. 2000. Can zebra mussels change stratiﬁcation patterns in a
small reservoir? Hydrobiologia 431: 175-184.

48

 

Table 1. Average, range, and standard deviation of physical and biological parameters in
uninvaded and invaded lakes. P values are for t-tests.

 

 

 

Variable Uninvaded Invaded P values
Average SD Average SD
(range) Jrange)
Mean depth 5.04 2.17 6.17 2.76 0.17
(m) (2.31-8.8) (2.13-12.41)
Total
10.75 10.53
phosphorus 4.28 3.8 0.94
- 4.9-20.76 526-1925
(“g L.) ( ) ( )
4.48 3.33
Chlirgphlﬁll a (1.91929) 1.82 (0.89-6.2) 1.19 0.04

 

49

 

 

 

wllv

Zebra mussels
A
l» :

 

Al

 

 

 

 

 

 

Crustaceans

\ l
Roti

I V
ll 11 .

 

 

 

 

 

 

 

 

fers

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Bacteria Phytoplankton
l — DOC Nutrients

 

 

 

 

 

 

Figure 1. Schematic representation of hypothesized direct and indirect effects of D.
polymorpha on lower trophic levels of lake ecosystems. Arrows represent a positive

effect and circles represent a negative effect.

50

 

 

 

 

 

 

Figure 2. Location of survey lakes in the lower peninsula of Michigan. The state is
divided into Albert ecoregion subsections. Closed circles represent invaded lakes and
open circles represent uninvaded lakes.

51

01
O

 

A O Rotifers
‘-'_l 40 _ 0 Ciliates
m

5':

m 30-

m

E l

.2 20 '

m

E: 10 -

0 § 3

 

 

 

o . .
Uninvaded Invaded

Figure 3. Inﬂuence of D. polymorpha on ciliate and rotifer dry biomass. Both ciliate and
rotifer biomass were signiﬁcantly lower in D. polymorpha lakes (P = 0.009, 0.01,
respectively). P values are for t-tests. Bars are SE.

52

 

 

 

 

 

 

 

 

 

Chlorophyll a (pg L")
Q _\ N 00 -h 01 O5

 

 

DOC(mg L")
o N h m on c:

 

 

 

 

 

 

 

Uni nvaded Invaded

Figure 4. The effect of D. polymorpha on (A) chlorophyll a (P = 0.02) and (B) DOC (P =
0.23) in the survey lakes. P values are for t-tests. Bars are SE.

53

 

 

 

 

 

 

 

 

 

 

 

,g 100 -
at A
m 80 -
(D
N
g 60 - :2 %<30pm
if, —- %>3o pm
a, 40 -
.2
E 20
6'2
0 . .
Uninvaded Invaded
B :1 <30

_>30

 

 

 

 

 

 

Dry Biomass (pg L'I)
o —x N co 4:. 01 a: \l

Uninvaded Invaded

Figure 5. (A) Relative biomass (%) and (B) actual biomass of ciliates below 30 pm in
size (bacterivorous) and above 30 pm (algivorous) in the survey lakes. Ciliates above 30
pm were higher in uninvaded lakes for both relative biomass and actual biomass (P =
0.01, 0.003, respectively). The actual biomass of ciliates below 30 pm were similar
between lake categories (P = 0.75). P values are for t-tests.

12

 

 

Rotifer richness
o:

 

 

 

 

 

 

 

 

 

Diversity (H')
O O
1:. 'c»

P
N

 

 

 

 

 

 

 

P
o

Uninvaded Invaded

Figure 6. The inﬂuence of D. polymorpha on (A) rotifer richness (P = 0.01) and (B)
diversity (P = 0.02). Both richness and diversity were lower in invaded lakes. P values are
for t-tests. Bars are SE.

55

 

 

o o

2 - . O .
(\l o
L 1 9 O O
49 00. O
8 0 “’1'
LL 0

4: “1w -
'1 ‘ O... . . Invaded
0 Uninvaded

 

 

 

 

 

 

-2 0 2 4 6
Factor 1

Figure 7. Factor scores ﬁom the rotifer PCA. Factor 1 (P = 0.036) and Factor 2 (P = 0.2).
P values are for t-tests.

56

APPENDIX

57

 

INDIVIDUAL SURVEY LAKE INFORMATION

Appendix A. Lake name, county, year sampled, mean depth, total phosphorus (TP), and
chlorophyll a of survey lakes.

 

 

Lake County Status Year Mean TP Chl a
sampled depth pg L'1 ”g L'1
Banksons Van Buren Uninvaded 2003 3 .60 14.16 4.29
Big Fish Lapeer Uninvaded 2002 5 .66 5.52 7.36
Big Portage Jackson Invaded 2002 3.16 15.57 3.48
Big Seven Oakland Uninvaded 2002 3 .07 10.41 5.50
Bird Hillsdale Invaded 2003 9.37 9.63 2.98
Bishop Livingston Uninvaded 2002 2.34 6.33 3 .93
Bristol Barry Uninvaded 2003 6.47 l 1.65 3.06
Cass Oakland Invaded 2002 9.17 9.83 2.46
Cedar Van Buren Invaded 2003 7.37 9.29 0.89
Cedar Island Oakland Invaded 2003 8.68 6.27 2.21
Clark Jackson Invaded 2002 3.56 19.25 2.37
Corey St. Joseph Uninvaded 2003 7.79 8.62 2.67
Deep Lenawee Uninvaded 2003 7.20 10.96 2.82
Devils Lenawee Invaded 2003 4.19 1 1 .84 4.29
Diamond Cass Invaded 2002 5 .08 9.32 5.25
Donnell Cass Invaded 2003 8.41 6.94 2.55
Fine Barry Uninvaded 2002 3.38 10.68 4.28
Fish Barry Uninvaded 2002 8.80 7.78 2.36
Fish St. Joseph Uninvaded 2002 8.19 7.00 3 .24
Gilkey Barry Uninvaded 2002 4.99 4.90 4.20
Gravel Van Buren Invaded 2002 5.61 12.29 3.84
Gull Kalamazoo Invaded 2002 12.41 5 .32 1.95
Gun Barry Invaded 2002 2.98 1 1.38 4.84
Halfrnoon Washtenaw Invaded 2003 8.50 10.25 2.67
Hemlock Hillsdale Invaded 2003 9.91 1 1.31 4.24
Heron Oakland Uninvaded 2002 3.33 1 1.23 4.23
Klinger St. Joseph Invaded 2003 6.43 8.12 3.84
Lake of the Woods Van Buren Invaded 2002 4.36 8.30 2.79
Lake Orion Oakland Invaded 2002 5.32 15.67 2.49
Lee Calhoun Uninvaded 2003 8.18 6.61 1.91
Lobdell Livingston Uninvaded 2002 3 .87 10.36 6.03
Magician Cass Invaded 2002 2.13 12.69 4.02
North Washtenaw Uninvaded 2003 3 .65 14.50 6.12
Orchard Oakland Invaded 2002 7.1 1 7.3 8 2.20
Payne Barry Invaded 2003 4.91 9.46 3 .63

58

 

Table l (cont’d)

 

 

Lake County Status Year Mean TP Chl a
sampled depth pg L‘1 llg L‘1
Pine Barry Uninvaded 2002 3.24 1 1.01 5 .48
Saddle Van Buren Uninvaded 2003 2.62 20.76 6.48
Sand Lenawee Invaded 2003 3.52 1 1 .67 4.14
Stone Cass Uninvaded 2003 5.73 13.16 2.40
Union Oakland Invaded 2003 8.92 5 .26 2.83
Upper Crooked Barry Uninvaded 2002 2.87 19.32 5. 1 3
Vandercook Jackson Uninvaded 2002 6.86 15.82 9.29
Vineyard Jackson Invaded 2002 4.30 6.94 3.66
Wamplers Jackson Invaded 2003 2.66 18.74 6.20
Warner Barry Uninvaded 2002 6.75 5.64 3.81
Woodland Livingston Uninvaded 2002 2.3 1 10.12 3 .99

 

59