Hadley Kristopher R 5279158

A MULTI-PROXY INVESTIGATION OF ECOLOGICAL CHANGES DUE TO MULTIPLE ANTHROPOGENIC STRESSORS IN MUSKOKA-HALIBURTON, ONTARIO, CANADA

Kristopher R. Hadley

A thesis submitted to the Department of Biology

in conformity with the requirements for

the degree of Doctor of Philosophy.

Queen’s University

Kingston, Ontario, Canada

(September, 2012)

Freshwater ecological issues are increasingly being recognized within the context of multiple stressors, even though relatively little is known about the limnological and biological consequences of the interactions between various environmental impacts. Moreover, long-term monitoring data are often lacking. To help address these issues, the overall goal of this thesis was to use paleolimnological approaches to document how multiple environmental stressors have altered limnological communities in south-central Ontario lakes. During the last two centuries,

Ontario lakes have been subjected to varying intensities of different environmental impacts, including increases in shoreline residential development, forest clearance and regrowth, the deposition of strong acids via the atmosphere, invasion by non-indigenous species, and climate change. I used multiple paleolimnological approaches to: 1) demonstrate how multiple natural and anthropogenic stressors have affected biological assemblages across lakes in the Muskoka-

Haliburton region of south-central Ontario, and 2) reconstruct the limnological histories of four lakes from Algonquin Park that have recorded the near complete extirpation of native crayfish species.

In the Muskoka-Haliburton lakes, I assessed the extent of limnological changes that have occurred during the past ~15 years by resampling lakes from an earlier survey, using identical paleolimnological methods. Limnological monitoring data document that, since 1992, the lakes have experienced declines in lakewater calcium and SO4 concentrations, while pH declined marginally; in contrast dissolved organic carbon, silica and Total Kjeldahl Nitrogen increased.

Marked regional increases in planktonic diatom taxa, including Cyclotella stelligera, Asterionella formosa and Fragilaria crotonensis, occurred in many lake systems, while colonial scaled chrysophyte algae have undergone a widespread decline in favour of unicellular forms (i.e.,

Mallomonas spp.), driven by interactions between resource limitation and climate change.

In the Algonquin Park study lakes, crustacean zooplankton remains revealed a marked decline in daphniid species with high Ca requirements, in favour of smaller Bosmina spp., while diatom and chrysophyte analysis suggest varying degrees of industrial acidification in the four study lakes. The paleolimnological data suggest that the crayfish decline may have began prior to the long-term monitoring record, likely as a result of declines in pH and lakewater Ca related to atmospheric acid deposition.

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Co-Authorship

Chapter 2 (Hadley et al. 2012, Aquatic Sciences) was co-authored with Andrew M.

Paterson and John P. Smol. In addition to being the primary writer of the manuscript, my responsibilities included the collection and statistical analysis of the diatom fossil data, as well as the drafting of all figures and tables.

Chapter 3 was co-authored with my co-supervisors John P. Smol and Andrew M. Paterson.

I was responsible for all of the field work, lab work, scaled chrysophyte counts, and statistical analyses. In addition, I drafted all figures, tables and was the primary writer of the manuscript.

In addition to my co-supervisors, Chapter 4 was co-authored with Keith Somers, James

Rusak and Ron Ingram from the Dorset Environmental Sciences Centre. I was the primary writer of the manuscript and was also responsible for field work, lab work, collection of diatom, scaled chrysophyte and zooplankton data and pH reconstructions. I was also responsible for drafting all figures and tables.

Acknowledgements

First and foremost, many thanks go to my supervisors, John Smol and Andrew Paterson, whose guidance and encouragement was crucial to my development as a scientist. Thanks to John for the opportunity to study at PEARL, beside so many talented scientists, and to Andrew, for his unwavering support and timely advice.

Thanks to the staff at the Dorset Environmental Science Centre, in particular Anna

DeSellas, Bev Clark, Ron Ingram, Ron Reid and Jim Rusak, without whose support this research would not have been possible. I would also like to thank my committee members, Shelley Arnott,

Brian Cumming, and Paul Treitz, for taking the time out of their schedules to be part of this process.

Thank you to all my colleagues and friends at PEARL both past and present, few are lucky enough to be surrounded by so many remarkable people. I am particularly grateful to Neal

Michelutti, Adam Jeziorski, Josh Thienpont , Jennifer Korosi and Alyson Paul for being supportive, motivating, and selfless colleagues, and friends.

Thanks to my family, my parents, uncle and brother, for their interest, encouragement and support throughout these many years. And above all, my thanks and love to my amazing wife, Stephanie. Without your patience, love, laughter and support, I surely could not have come this far.

Table of Contents

Abstract∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ii

Co-Authorship∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ iv

Acknowledgements∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ v

List of Figures∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ix

List of Tables∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ xi

Chapter 1: General Introduction and Literature Review∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 1

1.1 A paleolimnological approach∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙3

1.2 Primary thesis objective∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙6

1.3 Literature cited∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙9

Chapter 2: Effects of multiple stressors on lakes in south-central Ontario: 15 years of change in lakewater chemistry and sedimentary diatom assemblages∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙15

2.1 Abstract∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙15

2.2 Keywords∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙16

2.3 Introduction∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 16

2.4 Methods∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 20

2.4.1 Study lakes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 20

2.4.2 Water-sampling procedures∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙21

2.4.3 Surface-sediment sample collection∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 22

2.4.4 Diatoms∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙22

2.4.5 Statistical analysis∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 23

2.5 Results∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙25

2.5.1 Physical and chemical variables∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 25

2.5.2 Diatoms∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙26

2.6 Discussion∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 28

2.7 Conclusions∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙33

2.8 Acknowledgements∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 34

2.9 Literature cited∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙34

Chapter 3: Marked regional declines of colonial scaled chrysophytes in south-central Ontario lakes: resource limitation drives declines in nuisance algal blooms∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙46

3.1 Abstract∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙46

3.2 Keywords∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙46

3.3 Introduction∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙∙47

3.4 Methods∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 51

3.4.1 Study lakes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 51

3.4.2 Water-sampling procedures∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙53

3.4.3 Surface-sediment sample collection∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 53

3.4.4 Scaled chrysophytes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 54

3.4.5 Statistical analysis∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 54

3.5 Results∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙55

3.5.1 Water chemistry analysis∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 57

3.6 Discussion∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 57

3.6.1 The role of water chemistry and nutrient limitation in colonial chrysophyte declines∙ ∙ ∙58

3.6.2 The role of climate in scaled chrysophyte change∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 62

3.6.3 Other considerations∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙63

3.7 Conclusions∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙66

3.8 Acknowledgements∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 67

3.9 Literature cited∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙67

Chapter 4: Multiple environmental stressors drive extirpation of native crayfish in Algonquin

Park, Ontario, Canada∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙77

vii

4.1 Abstract∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙77

4.2 Keywords∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙78

4.3 Introduction∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 79

4.4 Methods∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 82

4.4.1 Study area∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙82

4.4.2 Water sampling procedures∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 82

4.4.3 Crayfish∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙83

4.4.4 Sediment sampling∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 83

4.4.5 Diatoms and scaled chrysophytes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 84

4.4.6 Cladocera∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 84

4.4.7 Sediment chronology∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙85

4.4.8 Statistical analysis∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 86

4.5 Results∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙86

4.5.1 Crayfish, water chemistry and climate data∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 86

4.5.2 Diatom assemblages and pH inferences ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙88

4.5.3 Scaled chrysophytes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 90

4.5.4 Cladocerans∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 91

4.6 Discussion∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 92

4.6.1 Lake acidification and its possible role in crayfish decline∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙93

4.6.2 The possible role of calcium decline in the crayfish decline∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 96

4.6.3 Direct and indirect effects of warming temperatures on crayfish populations∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 98

4.6.4 Interactive effects of multiple stressors and fish predation∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙99

4.7 Conclusions∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙100

4.8 Acknowledgements∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 101

4.9 Literature cited∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙101

Chapter 5: General Conclusions∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 112

5.1 Biological changes over decades reflect interactions of multiple anthropogenic

viii

stressors∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙115

5.2 Potential for ecological recovery and management in south-central Ontario lakes∙ ∙ ∙ ∙ ∙ ∙ ∙ 117

5.3 How does lake selection and temporal scale shape limnological interpretations?∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 119

5.4 Future directions∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 120

5.5 Literature cited∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 122

Appendix A: Water chemistry and count data∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 125

Appendix B: Water chemistry comparisons∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙173

Appendix C: Reproducability analysis ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙∙ ∙ ∙ ∙ ∙ ∙175

C.1 Triplicates∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 175

C.2 Hall and Smol recounts∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 176

Appendix D: Sediment chronologies in Algonquin Park∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙177

List of Figures

Chapter 2:

Figure 2.1. Map of the Muskoka/Haliburton region showing the location of the study lakes (circles) and towns (squares)∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 39

Figure 2.2. Yearly mean maximum temperature and total annual precipitation data from the

Muskoka Automated Weather Observation System (AWOS) climate station. ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 39

Figure 2.3. Boxplots comparing: a) Ca concentration, b) SO4 concentration, c) pH, d) DOC concentration, e) TKN concentration, and f) SiO2 concentration measured in 1990-1992 and 2007 or 2008. Variables which changed significantly between 2007 and 1992 are indicated with an asterisk∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 40

Figure 2.4. Spring ice-off data with piece-wise linear regression lines and breakpoint indicating threshold change in 1996∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 41

Figure 2.5. One to one plots of selected diatom taxa, showing changes from 1992 to 2007/8.

Numbers in brackets indicate the percent contribution of the taxa to the overall dissimilarity between 1992 and 2007 (i.e., SIMPER results). Wilcoxon Signed Rank results comparing 2007 and

1992 data are presented for all 6 taxa. The second analysis for Cyclotella stelligera is based on an identical Wilcoxon test excluding Gravenhurst Bay, Muskoka Bay and Young Lake∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 42

Figure 2.6. RDA analysis for modern water chemistry and diatom species assemblages. a)

Backwards selected RDA showing the overall relationship between diatoms and environmental variables. Variables removed with VIF> 5 and physical lake variables are plotted passively (grey).

x b) RDA with forward selection showing key environmental variables. c) Site centroids with standard error bars showing change in diatom species assemblages since pre-1850∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙43

Figure 2.7. Changes in relative abundances (relative abundance 2007 - relative abundance 1992) of selected diatom taxa between 2007 and 1992, with lakes arranged from lowest to highest [TP] ∙

∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙44

Chapter 3:

Figure 3.1. Map of the Muskoka-Haliburton region showing the location of the study lakes (circles) and towns (squares) ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙73

Figure 3.2. One to one plots of selected chrysophyte taxa, showing changes from: A) 1992 to

2007/8, B) pre-1850 to 1992, C) pre-1850 to 2007. Numbers in brackets, following the taxa name, indicate the percent contribution of the taxa to the overall dissimilarity between 1992 and 2007

(i.e., SIMPER results). Wilcoxon Signed Rank results comparing 2007, 1992 and pre-1850 data are presented for all 5 taxa, with p-values corrected for FDR. D) Boxplots summarize the overall relative abundance change of selected chrysophyte taxa across all three time periods∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 74

Figure 3.3. Boxplots comparing the decline in colonial chrysophytes across groups of key environmental variables including: a) DOC concentration, b)DOC, c) TP concentration, and d) TP. ANOVA results testing for significant differences (significant results bolded) amongst the groups are included for each chemical variable, both including and excluding (asterisked) Synura echinulata. Numbers in bold beneath the boxes indicate the number of lakes in each group∙ ∙ ∙ ∙ ∙75

xi chemical variable, both including and excluding (asterisked) Synura echinulata. Numbers in bold beneath the boxes indicate the number of lakes in each group.∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 75

Figure 3.5. Summary of the changes in the relative abundance since 1992 of the three major colonial chrysophyte taxa in the Muskoka-Haliburton lakes. Lakes are sorted left to right by percent change in colonial chrysophyte relative abundance. Lakes located on the marble intrusion are indicated by open bars∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 76

Chapter 4:

Figure 4.1. Map of central Ontario showing the approximate location of Algonquin Park and the location of the study lakes (closed circles) and towns (open circles) ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙108

Figure 4.2. Summary of monitoring data of Cradle, Delano, Pincher and Westward lakes. Top:

Catch per unit effort data of crayfish species Cambarus bartonii. Bottom: Boxplots summarizing changes in key water chemistry variables during the monitoring period∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙109

Figure 4.3. Relative abundance of key diatom taxa and downcore diatom-inferred pH in a) Cradle

Lake, b) Delano Lake, c) Pincher Lake and d) Westward Lake, plotted against core depth (cm) and

210Pb age∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙110

Figure 4.4. Relative abundance of key scaled chrysophyte taxa in a) Cradle Lake, b) Delano Lake, c)

Pincher Lake and d) Westward Lake∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 111

Figure 4.5. Summary of the relative abundance of key cladoceran taxa in a) Cradle Lake, b) Delano

Lake, c) Pincher Lake and d) Westward Lake∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙112

xii

Figure 4.6. DCA axis-1 scores plotted against core depth and 210Pb age in a) Cradle Lake, b) Delano

Lake, c) Pincher Lake and d) Westward Lake, showing species turnover∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙113

Appendix B:

Figure B.1. Comparison of the 2008 spring water [Ca], [DOC] and pH collected by Hadley et al.

(Chapter 2 and 3) with 1992 spring and 1990-1992 ice-free mean values collected by Hall and Smol

(1996)∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 173

Figure B.2. Comparison of the 2008 spring SO4, SiO3 and [TP] collected by Hadley et al. (Chapter 2 and 3) with 1992 spring and 1990-1992 ice-free mean values collected by Hall and Smol (1996) ∙ ∙ ∙

∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙174

Appendix C:

Figure C.1. Principal component analysis (PCA) of the surface sediment diatom communities the triplicate analysis of the three lakes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 176

Appendix D:

Figure D.1. Summary of the total 210Pb activity with sediment depth downcore for Cradle, Delano,

Pincher and Westward Lakes.∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙177

Figure D.2. Approximate downcore sediment age, based on 210Pb analysis of Cradle, Delano,

Pincher and Westward lake cores∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙178

xiii

List of Tables

Chapter 2:

Table 2.1. Summary of the key physical and chemical limnological variables from the Muskoka

Haliburton lakes∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 45

Chapter 4:

Table 4.1 Summary of the key physical and chemical limnological variables from the long-term

monitoring record of our four study lakes from Algonquin Park∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙114

Appendix A:

Table A.1. Relative abundances of diatom microfossils for the "top" intervals of the Muskoka

Haliburton survey lakes in Chapter 2∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 127

Table A.2. Relative abundances of scaled chrysophyte microfossils for the "top" intervals of the

Muskoka-Haliburton survey lakes in Chapter 3∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙148

Table A.3. Relative abundances of diatom microfossils for the Algonquin Park downcore analysis

in Chapter 4∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙151

Table A.4. Relative abundances of scaled chrysophyte microfossils for the Algonquin Park

downcore analysis in Chapter 4∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 163

Table A.5. Relative abundances of cladoceran remains for the Algonquin Park downcore analysis

in Chapter 4∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙167

Table A.6. Codes and full species names for diatom taxa used in Appendix tables∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 171

xiv

Table A.7. Summary of key water chemistry variables in the Muskoka-Haliburton Lakes from

Chapter 2 and 3∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ ∙ 172

Chapter 1

General Introduction and Literature Review

Long-term data are imperative to support effective science-based management of freshwater resources (Smol 2008), particularly as the magnitude and number of environmental stressors grows (e.g., intensification of human land-use activities, acid deposition, exotic species’ invasions, and climate change). In Canada, a region of particular interest is the lake-rich Muskoka-

Haliburton region of south-central Ontario where, despite its relatively remote location, lakes have been influenced by multiple environmental stressors during the past several decades (e.g.,

Hall and Smol 1996; Paterson et al. 2004; DeSellas et al. 2008; Yan et al. 2008a). The interactions between these stressors are often antagonistic or synergistic, resulting in either enhanced or muted ecological consequences (Folt et al. 1999). However, significant potential also exists for non-additive ecological effects, which may further complicate predictions regarding global change

(Christensen et al. 2006). In order to support educated policy and management decisions, a better understanding of the net effects of multiple stressors on the trajectory of biological change in these systems must be developed.

During the last two centuries, south-central Ontario has been subjected to varying intensities of several stressors, including increases in shoreline residential development, forest clearance and regrowth, the deposition of strong acids via the atmosphere, invasion by non- indigenous species, and climate change. Long-term lake monitoring projects at the Dorset

Environmental Science Centre (DESC) have tracked the effects of anthropogenic stressors on the aquatic and terrestrial ecosystems in the Muskoka area since the mid-1970s, while paleolimnological data have been used to extend the regional environmental record beyond the scope of the DESC monitoring data (e.g., Hall and Smol 1996; Paterson et al. 2001; Quinlan and

Smol 2002). Complex interactions amongst stressors have resulted in significant changes to

1 physical, chemical and biological conditions of lakes during the past several decades (Yan et al.

2008b), which are seldom equal to the sum of the individual impacts (Christensen et al. 2006).

Some of the stressors affecting Muskoka-Haliburton lakes have been relatively well studied. For example, the effects of acid deposition on lakes have been well documented in the region and elsewhere (Schindler 1986; Dillon et al. 1987; Schindler 1988). Research on the negative consequences of acid deposition on terrestrial and aquatic ecosystems eventually led to legislation which restricted industrial emissions (SO2 and NOX) and thereby reduced acid emissions to a fraction of the peak values reached in the 1970s (Keller et al. 1992; Keller et al. 2001). With reductions in acid deposition, the regional research focus shifted to documenting chemical and biological recovery (Yan et al. 1996; Keller et al. 1999; Dillon et al. 2007; Keller et al. 2007).

However, despite significant declines in sulphate deposition during the past three decades (ca. 40-

45%; Dillon et al. 2007), both chemical and biological recovery rates remain lower than expected

(Driscoll et al. 1989; Stoddard et al. 1999; Jeffries et al. 2003). This is hypothesized to be, in large part, a long-term consequence of acidification. For decades, acidification accelerated base-cation leaching from soils and led to soil acidification. This, in turn, has resulted in a measured loss of as much as ~40% of exchangeable Ca over a 16-year period (Watmough and Dillon 2001), and a marked decline in calcium concentrations in softwater lakes on the boreal shield (Stoddard et al.

1999; Watmough et al. 2003). South-central Ontario soils are generally thin and underlain by weather-resistant granite, leading to leaching rates of base cations which typically exceed the replenishment rates from weathering and atmospheric deposition (Watmough et al. 2005). As primary weathering rates of granitic bedrock are still exceeded by acid deposition in several watersheds in south-central Ontario (Jeffries et al. 2003; Vet et al. 2005), declines in lake-water calcium concentrations are expected to continue (Watmough et al. 2005; Watmough and Aherne

2008; Jeziorski et al. 2008). Additionally, calcium decline in the Muskoka-Haliburton region is

2 exacerbated by a legacy of logging, which results in a loss of Ca via forest biomass removal and timber regrowth (Huntington et al. 2000; Watmough et al. 2003). The consequences of Ca decline in both terrestrial and aquatic environments are still being explored; however, it has become increasingly apparent that several groups of Ca-dependant organisms are being adversely affected

(e.g., Jeziorski et al. 2008; Cairns and Yan 2009; Edwards et al. 2009).

Across the Northern Hemisphere, recent climate change has exerted discernible effects on lake ecosystems, inducing changes in their physical, chemical and biological conditions. For example, increases in annual mean air temperature in south-central Ontario have resulted in significant increases in the duration of the ice-free period since the mid-1970s (Futter 2003) which, in turn, results in a longer growing season and enhanced thermal stratification (Stainsby et al. 2011). Changes in thermal stability may then affect water-column mixing, light availability and nutrient distribution (Rühland et al. 2008; Smol 2010). Reduced mixing and increased thermal stability can significantly affect plankton in lakes by altering cell settling velocity and nutrient redistribution. Smaller taxa, such as some of the diatom species of the genus Cyclotella, have been shown to be favoured under such conditions (Winder et al. 2008). For example, Rühland et al. (2008, 2010) have shown significant correlations between spring ice-out date and relative abundance of Cyclotella spp. in many lakes across the Northern Hemisphere, which they have attributed mainly to a lengthening of the ice-free period and resultant changes in physical lakewater conditions. Such limnological changes can affect many other limnological conditions (e.g.,

Smol 2010).

1.1 A paleolimnological approach

A lack of long-term monitoring data is a common problem in assessing environmental change in lake ecosystems. Researchers are often limited to measurements from a few years, a

3 single year, or no data at all (Smol 2008). Paleolimnological techniques, using biological indicators, such as scaled chrysophytes, diatoms and zooplankton remains, are valuable tools in assessing and reconstructing ecological change due to a variety of natural and anthropogenic phenomena (Smol 2008).

1.1.1 Diatoms and scaled chrysophytes

The most common biological proxy used in paleolimnological studies are diatoms, whose siliceous frustules are highly resistant to chemical and physical degradation, are taxonomically distinct to the species/sub-species level, and the assemblage changes of which are sensitive to shifts in key environmental variables such as temperature, pH, salinity and nutrient concentrations

(Smol and Stoermer 2010). These species-specific responses allow diatom taxa to be calibrated to a suite of environmental variables making diatoms the most powerful proxy indicators available to paleolimnologists (Smol and Stoermer 2010). Like diatoms, scaled chrysophytes are composed of highly-resistant silica which can often be identified to species level. Chrysophytes have been shown to respond sensitively to changes in lakewater pH, making them ideal environmental indicators in paleolimnological reconstructions (Smol 1995). The usefulness of diatoms and scaled chrysophytes in pH reconstructions has been previously demonstrated in the Ontario region (e.g.

Dixit et al. 1992; Paterson et al. 2001), as well as in Nova Scotia (e.g. Ginn et al. 2007), the

Adirondacks (e.g. Cumming et al. 1992; Dixit et al. 1993), and many other regions globally

(Battarbee et al. 2010).

1.1.2 Cladocerans

Like the phycological indicators, zooplankton remains can provide important limnological information based on species composition (Korhola and Rautio 2001). Recently research has shown that it is possible to indirectly detect calcium decline in lakes using shifts in crustacean

4 zooplankton assemblage (Jeziorski et al. 2008; Jeziorski et al. 2012). Because they have significantly higher body calcium concentration than other common cladoceran species (Jeziorski and Yan 2006), and rely on direct absorption of Ca from the lake water as their primary source of

Ca (Cowgill 1986), daphniids are sensitive to reduced lake water [Ca] (Jeziorski et al. 2008).

Laboratory studies have shown that, below a [Ca] of 1.5 mg·L-1 the growth, reproduction and survival of the most common daphniids are inhibited (Ashforth and Yan 2008; Cairns 2010). Many of south-central Ontario’s softwater lakes are already below this 1.5 mg·L-1 threshold (Neary et al.

1990) and, with future Ca decline, concentrations below 1.0 mg·L-1 are predicted to become widespread (Watmough et al. 2003).

1.1.3 Crayfish as sentinel species

Severe biological consequences have been attributed to the legacy of anthropogenic change in Ontario. For example, a recent crayfish survey of 100 lakes in south-central Ontario suggests that crayfish populations across the region are overwhelmingly in a state of decline

(Edwards et al. 2009). Edwards et al. (2009) revisited a suite of lakes and compared modern crayfish populations (2005-2007) to a historical dataset (1989-1992, 1994, 1995). They found large-scale population declines and extirpations for seven commonly found crayfish species, along with species diversity declines from approximately 2-4 species per lake to ≤1 in the recent survey.

Research suggests that the mechanisms contributing to the imperilment of crayfish may include:

1) historical acidification and metal contamination (Davies 1989; France 1993); 2) declining calcium (Ca) concentration in lake water (Rukke 2002; Cairns and Yan 2009); 3) increased water temperature driven by climate change (Hammond et al. 2006; Hartman et al. 2010); 4) shoreline development and habitat loss (Wilcove et al. 1998); and 5) non-native species invasions (Olden et al. 2006). Reduced lake-water pH may result in a plethora of negative effects on crayfish

5 populations (e.g. DiStefano et al. 1991; France 1993). For example, long-term exposure of crayfish to low pH may result in hatchling mortality, egg loss, and thus recruitment failure (Davies 1989), and has also been linked to higher instances of microsporidian parasitism (France and Graham

1985). Furthermore, these sublethal effects may, in turn, be exacerbated by other environmental stressors such as calcium decline and climate change (Davies 1989; DiStefano et al. 1991;

Hammond et al. 2006; Hartman et al. 2010). Like daphniids, crayfish are highly dependent on ambient lakewater calcium concentration particularly during their moulting period. While species specific [Ca] optima and tolerances for organisms have as yet been rarely determined, the current lowest hypothesized requirement for freshwater crayfish is between 2 and 10 mg·L-1 of Ca in the lakewater (Greenaway 1974; France 1987; Rukke 2002; Hammond et al. 2006; Cairns and Yan

2009). However, this broad threshold is based on several larger crayfish species from outside

Canada (e.g., Astacus astacus) and therefore may not readily apply to crayfish species commonly found in Ontario (e.g., Cambarus bartonii). In addition to directly decreased survival of crayfish

(Hammond et al. 2006), increased water temperature may exacerbate the effects of other environmental stressors, such as reduce tolerance to low pH (DiStefano et al. 1991) and increased susceptibility to disease (Jiravanichpaisal et al. 2004). As with other stressors (e.g., pH), increased water temperature appears to have a significantly greater negative impact on survival of juvenile crayfish (Hartman et al. 2010).

1.2 Primary thesis objectives

The primary objective of this thesis was to explore how multiple environmental stressors have altered the physical, chemical and biological limnology of lakes in south-central Ontario. Like

Palmer (2012), I defined a stressor to be an issue that may incite a measurable environmental or biological response. To better understand the trajectory of limnological change in the region, we used multiple paleolimnological proxies (i.e., diatoms, chrysophytes and cladocerans) in an effort

6 to tease apart the impacts of a complex array of highly interactive environmental stressors. My principal study area was a suite of lakes in south-central Ontario, referred to herein as the

Muskoka-Haliburton lakes. The limnological history of these lakes has been studied previously, thereby allowing for a unique comparative approach. To document how changes in climate and other environmental stressors have altered biological communities in these lakes since 1992, I combined my modern sedimentary data with data collected previously (i.e., 1992) by Hall and

Smol (1996) and Paterson et al. (2001).

Following this short introduction, Chapter 2 of this thesis uses a comparative approach to assess the trajectory of limnological change in south-central Ontario lakes over the past ~15 years, by comparing sedimentary diatoms assemblages and key water chemistry variables collected in

2007/8 to both 1992 and pre-industrial values collected by Hall and Smol (1996). This suite of lakes provides a diverse cross-section of the physical and chemical limnological variability of the region and is representative of thousands of softwater boreal shield lakes. Since the initial sampling of these lakes in 1992, significant changes in lake stratification, chemistry and biology have been observed in several of these lakes (Paterson et al. 2008; Yan et al. 2008a, b; Palmer et al. 2011). For example, biological recovery from acidification (Yan et al. 2008a), climate-driven ecological changes (Futter 2003), the recent discovery of novel stressors such as Ca decline

(Jeziorski et al. 2008), and continued increases in catchment development and recreational activities have been reported in the region. Given these recent changes, the primary goal of this study was to acquire the updated paleolimnological data needed to better understand the responses of lakes to these dynamic environmental conditions. Specifically, we used diatom microfossils and modern water chemistry data to assess the trajectory of biological changes in this region since 1992, which has been a critical recovery period. To this end, we have revisited the

Muskoka-Haliburton lakes collecting both surface sediments and modern water chemistry and,

7 using a series of univariate and multivariate statistical techniques, directly compared the chemistry and biology (i.e., diatom assemblages) of the two datasets.

In Chapter 3 we used scaled chrysophyte microfossils and modern water chemistry data to explore changes in chrysophyte assemblages in the Muskoka-Haliburton region since 1992. To this end, we have revisited the Muskoka-Haliburton lakes (Hall and Smol 1996; Paterson et al.

2001) collecting both surface sediments and modern water chemistry. By comparing pre- industrial, 1992 and modern (ca. 2007/8) limnological conditions, we have explored if the marked increases in nuisance colonial chrysophytes continued during the past ~15 years. Widespread regional blooms of colonial chrysophytes in boreal shield lakes that occurred since pre-industrial times were often characterized by near systemic increases in the nuisance taxon, Synura petersenii (Dixit et al. 1992; Paterson et al. 2001; Vinebrooke et al. 2002; Hyatt et al. 2010), and have been linked to taste and odour issues in freshwater lakes and reservoirs (Nicholls 1995;

Watson et al. 2001). These shifts in chrysophyte communities are believed to be controlled by complex interactions between climate change, acidification, recovery and related changes in dissolved organic carbon (DOC) concentrations (Paterson et al. 2004, 2008; Hyatt et al. 2010). The marked decline in colonial chrysophyte species we recorded highlights the complex nature of multiple stressor systems, and provides useful data to inform future lake management and policy decisions in the region.

In Chapter 4, using paleolimnological analyses, our goal was to determine the limnological history of four Algonquin Park lakes where, despite their isolation, long-term monitoring data have shown a marked decline in populations of three native Ontario crayfish species (i.e.,

Cambarus bartonii, Orconectes propinquus, and Orconectes virilis) during the past several decades

(Girard et al. 2006). Crayfish feed at several trophic levels, are an important component of energy transfer through the food web, and have been lauded as sentinel species for monitoring

8 ecosystem health (e.g. Antón et al. 2000). The absence of local shoreline impacts and species invasions in Algonquin Park suggests that regional stressors (i.e., climate change, pH decline and calcium decline) are likely responsible for the fate of crayfish. Existing monitoring data provided by the Ontario Ministry of the Environment indicates that the decline in crayfish abundance in this region pre-dates the monitoring record, making the historical limnological data provided by paleolimnological proxies vital. Using multiple paleolimnological indicators (diatoms frustules, chrysophyte scales, and cladoceran remains) we reconstructed lakewater pH and examined long- term changes in biological responses to changing calcium concentrations. Our data suggest that no single stressor can account for the near extirpation of crayfish in these four lakes. While it appears that pH decline played a role in the decline of crayfish, subsequent chemical recovery has not been met with increased crayfish abundance. Furthermore, pH decline is not consistent across our study sites. Loss of Ca-sensitive daphniid species was also recorded in the Algonquin lakes and while we are not able to establish a [Ca] threshold for Ontario crayfish species, our data suggests that pH recovery, without significant increases in [Ca], will not result in crayfish recovery.

Finally, Chapter 5 provides a general overview of the findings of the thesis as a whole and the implications for future investigations into the impacts of multiple stressors in south-central

Ontario lake ecosystems. Collectively, this thesis provides a quantitative assessment of lake responses to multiple environmental stressors in south-central Ontario, while highlighting the importance of interactions between stressor in structuring aquatic ecosystems and limnological interpretations.

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Watmough, S.A., and Aherne, J. 2008. Estimating calcium weathering rates and future lake calcium concentrations in the Muskoka-Haliburton region of Ontario. Can. J. Fish. Aquat. Sci. 65: 821-833.

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Watmough, S.A., Aherne, J., Alewell, C., Arp, P., Bailey, S., Clair, T., Dillon, P., Duchesne, L., Eimers, C., Fernandez, I., Foster, N., Larssen, T., Miller, E., Mitchell, M., and Page, S. 2005. Sulphate, nitrogen and base cation budgets at 21 forested catchments in Canada, the United States and Europe. Environ. Monit. Assess. 109: 1-36.

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Chapter 2

Effects of multiple stressors on lakes in south-central Ontario: 15 years of change in lakewater chemistry and sedimentary diatom assemblages

2.1 Abstract

Freshwater ecosystems, despite their societal value and environmental importance, are influenced increasingly by multiple natural and human-affected processes. To quantify long-term changes in chemical and biological properties of lakes in the Muskoka-Haliburton region of south- central Ontario to increases in shoreline residential development and regional acid deposition,

Hall and Smol (1996) used diatom-based transfer functions to estimate changes in lakewater pH and total phosphorus (TP) concentration in a suite of 54 lakes between the preindustrial era and

1992. Since 1992, however, multiple natural and anthropogenic stressors have potentially altered the trajectory of chemical and biological conditions in these systems. Here, we assess the extent of limnological changes that have occurred during the past ~15 years in the Muskoka-Haliburton study lakes by resampling the same lakes using similar limnological and paleolimnological methods as in Hall and Smol (1996). Marked changes were observed in both water chemistry (i.e., decreased pH and increased DOC) and accumulated diatom microfossil assemblages since 1992, including significant increases in planktonic taxa (e.g., Cyclotella stelligera) commonly associated with climate warming. In fact, larger diatom changes have occurred between samples spanning the past 15 years than were observed between assemblages deposited in pre-industrial and 1992 surface sediments. Breakpoint analysis of regional ice-out data from Dorset lakes suggests that a threshold response of lake ice to climate warming did not occur until after 1992, consistent with our paleolimnological data. Significant changes observed in both chemical (e.g., pH, Ca, DOC) and biological data suggests that novel stressors, such as declines in lake calcium concentrations,

acting in conjunction with climate and land-use change, have created ecosystems for which there are no historical analogs.

2.2 Keywords

Paleolimnology, Muskoka, Haliburton, multiple stressors, diatoms

2.3 Introduction

In Canada, as elsewhere, there is a need for high quality, long-term data to support effective science-based management of freshwater resources (Smol 2008). This becomes increasingly important as the magnitude and number of environmental stressors grows (e.g., including intensification of human land-use, acid deposition, exotic species’ invasions and climate change).

A region of particular interest in Canada is the lake-rich Muskoka-Haliburton region of south- central Ontario where, despite its relatively remote location, lakes have been influenced by multiple environmental stressors during the past several decades (Hall and Smol 1996, Paterson et al. 2004, DeSellas et al. 2008, Yan et al. 2008a). Although distant from major urban and industrial centres, the Muskoka-Haliburton region is a popular destination for recreational activities, and is home to many cottages, resorts and golf courses. As a result, lakes in the region are susceptible to both regional-scale stressors, such as climate change, as well as local, catchment-scale stressors associated with increased shoreline development.

The Dorset Environmental Science Centre (DESC) has tracked the effects of anthropogenic stressors on the aquatic and terrestrial ecosystems in the Muskoka area since the mid-1970s. In addition to long-term monitoring, paleolimnological methods have been used to place recent trends into the context of longer-term post-industrial changes (e.g., Hall and Smol 1996). During the last two centuries, this region has been subjected to varying intensities of several stressors,

including forest clearance and regrowth, increases in residential development, the deposition of strong acids via the atmosphere, invasion by non-indigenous species, and climate change.

Complex interactions amongst these stressors have resulted in significant changes to physical, chemical and biological conditions of lakes during the past several decades (Yan et al. 2008b).

The effects of acidification have been well documented in several lakes in the Muskoka-

Haliburton region, as well as elsewhere in North America and Europe (Dillon et al. 1987, Ginn et al.

2007a, b, Batterbee et al. 2010, Greenaway et al. 2012). Recently, with reductions in acid deposition, the research focus has shifted to documenting chemical and biological recovery (Dillon et al. 2007, Keller et al. 2007). While chemical recovery has been observed to some degree, recovery rates are lower than expected given the significant declines in sulphate deposition that have occurred during the past three decades (ca. 40-45%; Dillon et al. 2007). Acid deposition in south-central Ontario has accelerated the leaching of base cations and led to soil acidification, which has resulted in a measured loss of 37% of exchangeable Ca over a 16-year period

(Watmough and Dillon 2001). Because primary weathering rates of granitic bedrock are still exceeded by acid deposition in several watersheds in south-central Ontario, further declines in base cations, including lake-water calcium concentrations, are expected to continue (Watmough et al. 2005, Watmough and Aherne 2008, Jeziorski et al. 2008). Recovery of biological communities are still in their earliest stages (Keller et al. 2007), and, given the complex effects of multiple environmental stressors (Schindler 2001, Smol 2010), it is becoming increasingly unlikely that they will see a return to baseline pre-disturbance conditions.

Across the Northern Hemisphere, climate changes have exerted discernible effects on lake ecosystems, inducing changes in their physical, chemical and biological conditions. For example, increases in annual-mean air temperature in south-central Ontario have resulted in significant increases in the duration of the ice-free period since the mid-1970s (Futter 2003). A longer ice-

free period results in a longer growing season and enhanced thermal stratification (Stainsby et al.

2011). Changes in thermal stability may, in turn, affect water-column mixing, light availability and nutrient distribution (Smol 2008). Reduced mixing and increased thermal stability can significantly affect plankton in lakes, by altering cell settling velocity and nutrient redistribution. Smaller taxa, such as diatom species of the genus Cyclotella, have been shown to be favoured under such conditions (Winder et al. 2008). For example, Rühland et al. (2008, 2010) have shown significant correlations between spring ice-out date and relative abundance of Cyclotella spp. in many lakes in the Northern Hemisphere, which has been attributed mainly to a lengthening of the ice-free period and resultant changes in physical lake-water conditions.

In order to anticipate future trajectories of change, and to establish realistic recovery goals, we must continue to expand our understanding of the biological changes that have occurred in these systems. Previously unrecognized stressors, such as declines in lake-water Ca concentration, acting in conjunction with climate- and land-use change, have created ecosystems for which there are no historical analogs. A recent synopsis of the DESC long-term chemistry data,

Table 1 in Palmer et al. (2011), summarizes how a combination of environmental stressors have changed lake-water quality in this region. In order to gauge future recovery, and to establish realistic recovery goals, we must continue to expand our understanding of the biological trajectories of these systems. Given the lack of long-term monitoring available for many lakes in this region, there is an obvious need for indirect proxy techniques such as those employed by paleolimnologists (Smol 2008).

In response to growing concerns regarding multiple stressors on freshwater ecosystems,

Hall and Smol (1996) sampled 54 lakes in the Muskoka-Haliburton region to document long-term changes in diatom assemblages and diatom-inferred water chemistry (i.e., total phosphorus concentrations and pH). These lakes were selected to minimize the range of pH (i.e., no lakes with

pH <5.5) and DOC (i.e., no lakes with [DOC] >6.5 mg/L), to assess the influence of total phosphorus

([TP]) on diatom assemblage composition. Nevertheless, this suite of lakes provided a diverse cross-section of the physical and chemical limnological variability of the region (Table 1). In summary, Hall and Smol (1996), using a “before and after” or so-called “top-bottom” paleolimnological approach, found relatively small increases in diatom-inferred [TP] (generally less than 4 g/L) in regional lakes since pre-industrial times, despite modest residential and cottage development. Significant increases in diatom-inferred TP occurred in only two oligotrophic lakes, while half of the mesotrophic lakes (with [TP] between 10 and 20 μg/L) showed discernible declines since pre-industrial times (Hall and Smol 1996). Inferences based on diatoms suggested that many of the acidic lakes (pH < 6.5 in 1992) had acidified since pre-industrial times, the result of the long-range transport and deposition of acids from industrial activities (Dillon et al. 1987).

The Hall and Smol (1996) top-bottom paleolimnological study provided key baseline data for the pre-disturbance limnological conditions in this region, as well as a robust diatom-inferred pH transfer function.

Since publication of the Hall and Smol study (1996), significant changes in lake stratification, chemistry and biology have been observed in several of these lakes (Yan et al.

2008a, b; Paterson et al. 2008; Palmer et al. 2011). For example, biological recovery from acidification (Yan et al. 2008a), climate-driven ecological changes (Futter 2003), the recent discovery of novel stressors such as Ca decline (Jeziorski et al. 2008), and continued increases in catchment development and recreational activities have been reported in the region. Given these recent changes, the primary goal of this study was to assess the extent of chemical and biological change that has occurred in the Hall and Smol (1996) study lakes during the past ~15 years (i.e., by comparing the surface sediments collected by Hall and Smol in 1992 to a new suite of surface-

sediments collected from the same lakes in 2007 and 2008). Many of these lakes have not been sampled since 1992.

Updated paleolimnological data are needed to better understand the responses of lakes to these dynamic environmental conditions (Quinlan et al. 2008). Paleolimnological techniques, using biological indicators such as scaled chrysophytes, diatoms and zooplankton remains, are valuable tools in assessing and reconstructing ecological change due to a variety of natural and anthropogenic phenomena (Smol 2008). Specifically, we use diatom microfossils and modern water chemistry data to assess the trajectory of biological changes in this region since 1992. To this end, we have revisited the Hall and Smol (1996) Muskoka-Haliburton lakes collecting both surface sediments and modern water chemistry. These data should shed new light on the patterns of limnological and biological change during the past ~15 years in Muskoka area lakes, and will provide useful information to inform lake management and policy decisions.

2.4 Methods

2.4.1 Study Lakes

The 54 lake dataset of Hall and Smol (1996) was originally sampled for sedimentary diatom assemblages in the fall of 1992. Water chemistry variables were determined from annual averages of volume-weighted epilimnetic or metalimnetic composites, sampled biweekly, monthly, or bimonthly during the ice-free seasons from 1990 to 1992 (Hall and Smol 1996, Table

2). We revisited and obtained new sediment cores from 53 of these sites over two field seasons in

July/August of 2007 and 2008 (no core could be obtained for the Seagull Rock site at Lake of Bays).

The study sites are located in the Muskoka-Haliburton region of south-central Ontario, Canada, situated on the Precambrian Shield (Figure 1). The Shield is characterized by thin, poorly developed soil with abundant bedrock exposures.

The Muskoka-Haliburton region is situated within the boreal ecozone with a temperate climate. The mean of the maximum daily air temperature varied from 9.2 °C in 1992 to 11.7 °C in

2007, while total precipitation has fluctuated from 1163.1 mm in 1992 to 850.6 mm and 1125.6 mm in 2007/8 respectively (Figure 2). Catchments are mainly forested, dominated by sugar maple

(Acer saccharum), beech (Fagus grandifolia), yellow birch (Betula lutea), red maple (Acer rubrum), eastern hemlock (Tsuga canadensis), and white and red pine (Pinus strobus and Pinus resinosa).

Shoreline development varies greatly among sites. Several lakes have almost no development

(e.g., Plastic Lake), while others receive inputs from urban and seasonal recreational (cottages, resorts, golf course) activities.

2.4.2 Water sampling procedures

Water samples were collected from each lake by technical staff at the Dorset Environmental

Science Centre (DESC) in the spring of 2008 (42 lakes) or 2009 (4 lakes). Water chemistry sampling protocols differed slightly from those employed by Hall and Smol (1996). Samples collected in

1990-1992 were volume-weighted ice-free means from samples collected between 1990 and

1992, while our measurements were single value spring samples. Analytical protocols were identical to those followed in the initial sampling period (1990-1992) used in the Hall and Smol

(1996) study. Due to logistical constraints, water chemistry was only obtained for 46 sites of the

53 sediment sites (no chemistry data were collected for Kelly, Little Redstone, Little Kennisis, Lake

Joseph-Frazer Island, Lake Muskoka-Gravenhurst Bay, Red Pine and Redstone). Analyses of water samples for 14 key variables were performed at Ontario Ministry of the Environment’s DESC laboratory following standard OMOE protocols (Ontario Ministry of the Environment 1983) and are summarized along with four physical parameters in Table 1. Measured variables included pH,

specific conductivity (cond), Gran alkalinity (alk), and concentrations of calcium (Ca), dissolved organic carbon (DOC), potassium (K), magnesium (Mg), ammonia/ammonium (NH3/NH4), nitrate/nitrite (NO3/NO2), total Kjeldahl nitrogen (TKN), pH, total phosphorus (TP), sulphate (SO4), sodium (Na) and silica (SiO2) (Table 1).

2.4.3 Surface sediment sample collection

A short gravity sediment core was collected from the deepest portion of each lake using a

7.6-cm internal-diameter Glew (1989) gravity corer, and sectioned on site using a Glew (1988) vertical extruder, using identical equipment and procedures as Hall and Smol (1996). Sediment cores were sectioned at 0.25 cm resolution for the first 5 cm for diatom analysis but only the top- most 0.25 cm interval was used. This differs from the Hall and Smol protocol where the modern diatom assemblage was represented by a single 1 cm slice of sediment removed from the top of the core. All comparisons to the pre-industrial time period used the sediment samples collected and enumerated by Hall and Smol (1996). The sediment samples were stored in Whirlpak® bags and refrigerated in the dark until analysis.

2.4.4 Diatoms

For diatom analysis, 0.2 - 0.3 g of wet sediment was digested in a 50:50 (v/v) mixture of concentrated nitric and sulphuric acid, and rinsed repeatedly with deionized water until a neutral pH was achieved. Resulting cleaned diatom slurries were dried onto coverslips and permanently mounted on microscope slides using Naphrax® mounting medium. Diatom microfossils were then identified and enumerated at 1000X under oil immersion using a Leica DMR2 microscope with differential interference contrast. A minimum of 400 diatoms valves were counted for each

sample and identified using standard taxonomic sources (Camburn et al. 1984-1986, Krammer and

Lange-Bertalot 1986-1991, Round et al. 1990, Cumming et al. 1995, Camburn and Charles 2000).

In order to ensure taxonomic consistency between the 2007 and 1992 datasets, limited recounts of the Hall and Smol slides were undertaken (K.R. Hadley, Queen’s University, unpublished data).

2.4.5 Statistical Analysis

Environmental data

Shapiro-Wilk (1965) tests were performed using Systat v. 11 to test whether the assumption of normality in the chemical data was met for subsequent analyses. Shapiro-Wilk test results varied between 2007 and 1992 data. Therefore, a choice was made to utilize a non-parametric statistical technique for these comparisons (i.e., Wilcoxon Signed Rank test), rather than attempt to transform the data for individual paired t-tests.

All 14 environmental variables were tested independently in Systat v. 11 using the non- parametric Wilcoxon Signed Rank test to compare 1992 and 2007 data sets. Analysis of multiple comparisons can result in increased risk of Type-I error. To counter this effect, we calculated the false discovery rate (FDR) for the Wilcoxon Signed Rank results using the “Multtest” package in R v. 2.13 (Benjamini and Hochberg 1995).

To explore long-term changes in ice-dynamics in the study lakes (as a measure of lake responses to recent climate change), we analyzed a regional lake ice dataset available from the

DESC; the regional value is considered to be an average of ice-on and ice-off dates collected since

1976 from several small lakes near Dorset, Ontario. Breakpoint analysis (BPA) was run on spring ice-off dates in order to determine if a significant climate threshold could be identified.

Breakpoint analysis was run in R v. 2.13 using the “piecewise.linear” function of the SiZer package

(Sonderegger 2011). This function performs a piecewise-regression, joining two straight lines at a single sharp breakpoint, as outlined by Toms and Lesperance (2003).

Diatoms

Prior to statistical analysis, rare diatom taxa whose relative abundances did not exceed 1% in at least 3 lakes were removed from the dataset. The species in the reduced dataset accounted for an average of 93% of the total relative abundance. ANOSIM tests were used to compare the modern (2007) diatom assemblages with assemblages in both the surface sediment collected in

1992 and the pre-industrial “bottom” samples (pre-1850) (Hall and Smol 1996). One-way ANOSIM tests of the full regional datasets (i.e., all sites from 2007, 1992 and 1850) were performed comparing 2007 vs. 1992, 2007 vs. pre-1850, and 1992 vs. pre-1850. ANOSIM tests were performed on Bray-Curtis dissimilarity metrics using non-transformed species data with 9999 permutations. A Similarity Percentages (SIMPER) analysis was used to determine the contribution of key diatom taxa to the dissimilarity between the datasets. Because ANOSIM and SIMPER do not allow for paired sampling, we also included paired univariate analyses (i.e., Wilcoxon Signed

Rank test) of key diatom taxa over time to determine if the changes in relative abundance between 2007 and 1992 were significant. ANOSIM tests and SIMPER analyses were performed using the software Primer v. 6 (Clarke and Gorley 2006).

Redundancy analysis (RDA) was used to relate sedimentary diatom assemblage composition in the samples collected in 2007 to physical and chemical limnological variables. To summarize the species-environmental relationships, a series of initial RDAs, starting with all environmental variables were performed and variance inflation factors (VIF) were used to perform backwards elimination of highly correlated variables. Using these procedures, the variable with the highest VIF was eliminated from the RDA, and subsequent RDAs were run until all VIFs

dropped below 5. In addition to the backwards selected RDA, diatom assemblages from 2007/8 were compared against limnological variables using manual forward selection of environmental variables with 499 Monte Carlo permutations. Only significant (p ≤ 0.05) environmental variables were included in the final RDA analysis. Sedimentary diatom assemblage composition data from

1992 and pre-industrial (ca. 1850) samples were run as supplementary data in the RDA. To illustrate how diatom assemblages have changed over time, sample (i.e., lake) scores from all three time periods (2007, 1992, and pre-1850) are presented as centroids (x and y axis = average axis 1 and 2 scores), plus or minus the standard deviation (represented as error bars along axes 1 and 2).

2.5 Results

2.5.1 Physical and chemical variables

Water quality in the Muskoka-Haliburton region has undergone several notable changes during the period from 1992 to 2007. Eleven of 14 measured limnological variables have changed significantly (p ≤ 0.01) over time (Figure 3), after correcting for multiple pairwise tests. Sulphate

2- (SO4 ) concentration declined in all but one (Butterfly Lake) of the 46 sites by an average of 2.11

g/L (Figure 3, b). Despite declining sulphate deposition, decreasing pH was observed in many lakes. Lakewater pH declined in 72% of lakes, by an overall average of 0.12 pH units (Figure 3, c).

Significant decreases were also observed in concentrations of the base cations Ca2+ and Mg2+, and conductivity. Calcium declined in 83% of the sites, with a mean decrease of 0.74 mg/L (Figure 3, a), while Mg2+ declined in 76% of sites. Interestingly, [TP] showed no significant change over the study period. Dissolved organic carbon increased in 76% of the Muskoka-Haliburton lakes, by an average of 0.55 mg/L over 15 years (Figure 3, d). Total Kjeldahl Nitrogen (TKN) increased in 74% of

+ - the Muskoka Lakes (mean, 22.71 mg/L), while Na and SiO2 in increased in ~75% of lakes by an average of 1.44 mg/L and 0.80 mg/L, respectively.

As noted in the Methods, water chemistry sampling techniques differed slightly between the

DESC sampling regime in 2008/9 and Hall and Smol (1996) in 1990-1992. Samples collected by

Hall and Smol (1996) were ice-free means from samples collected between 1990 and 1992, while our measurements in 2008/09 were single value spring samples. To explore the potential implications of this difference, we plotted side-by-side comparisons, using data from intensively studied lakes near Dorset, Ontario (i.e., the Dorset A lakes). Specifically, we plotted the single measurement 2007/8 water chemistry data against both the spring 1992 and ice-free mean 1992 data (Appendix B). Results from these tests showed that most of the key water chemistry variables, including pH, [DOC], SO4, and Ca, were not substantially affected by sample strategy

(Appendix B), and showed changes that are consistent with other lakes examined in the Muskoka-

Haliburton region (Palmer et al. 2011). However, [TP] did show substantially different results when spring 2007/8 and spring 1992 values were compared. We noted a larger increase in [TP] in the Dorset A lakes when we compared values between spring 2008 and spring 1992, rather than comparing spring (2008) versus ice-free mean (1990-1992).

Breakpoint analysis indicated a significant break in the spring ice-off data for the Dorset lakes at 1996. While, the record showed a high degree of inter-annual variation (Figure 4), and bootstrapped 95% confidence intervals ranged from 1978 - 2009, no ice-out date after Julian day

115 has been recorded since 1997. Furthermore, the use of bootstrapped confidence intervals in breakpoint analysis has proved problematic when sample sizes are low (Toms and Lesperence

2003).

2.5.2 Diatoms

Along with changes recorded in several water chemistry variables, significant changes in diatom assemblages were also documented between 2007 and 1992. Preservation of siliceous

remains was excellent in all sediment samples with no signs of dissolution, including delicate lightly-silicified chrysophyte scales and spines indicating that changes in fossil assemblages were not the result of post-depositional processes (i.e., diagenesis). The greatest dissimilarity was observed between the composition of modern (2007) and pre-industrial (pre-1850) sedimentary diatom assemblages (R = 0.587, p ≤ 0.001). Surprisingly, greater change has occurred between samples spanning the past 15 years (R = 0.334, p ≤ 0.001) than was observed between those deposited in pre-industrial and 1992 surface sediments (R = 0.191, p ≤ 0.001). According to

SIMPER analysis, Cyclotella stelligera accounted for the highest percentage of the dissimilarity between both the 1992 and 2007 (20.5%) and the pre-1850 and 1992 (19.38%) datasets. Nearly

65% of lakes sampled in 2007 had a greater relative abundance of C. stelligera compared to 1992, with a mean increase of 8.02% (Figure 5). According to Wilcoxon signed ranks test comparing

2007 and 1992, the increase in the relative abundance of C. stelligera was not significant at a regional scale. However, it appears that this result was strongly influenced by three sites with large decreases in C. stelligera over time (i.e., Young Lake, Gravenhurst Bay and Muskoka Bay,

Lake Muskoka). When these sites, whose change in relative abundance was greater than 2s.d. from the mean, were removed from a subsequent analysis, the increase in C. stelligera relative abundances was found to be significant from 1992 to 2007 (Figure 5, a). Other key taxa contributing to the overall dissimilarity between the 2007 and 1992 diatom assemblages included

Asterionella formosa (increase), Fragilaria crotonensis (increase), Tabellaria flocculosa str. IIIp

(increase), Asterionella ralfsii var. americana (decrease) and Aulacoseira ambigua (decrease)

(Figure 5). Of these taxa, all but T. flocculosa str. IIIp changed significantly between 1992 and

2007 and together with C. stelligera they account for greater than half (52.37%) of the dissimilarity in the 53 lakes studied (Figure 6).

An initial RDA analysis of the 2007 water chemistry and sedimentary diatom species data with backwards selection eliminated seven environmental variables (Lake vol, Ca, Alk, ZMAX, Cond,

TN and Mg) whose VIF were >5. These variables were plotted passively along with the remaining

11 physical and chemical variables in Figure 6a to show the overall species environmental relationships. A second RDA with forward selection identified three significant (p ≤ 0.05) environmental variables (pH, TP and lake volume) (Figure 6b). The eigenvalues for RDA axes-1

(0.070) and 2 (0.044) constrained to these three environmental variables were significant (p ≤

0.001). Movement of the site centroids within the ordination space illustrates the diatom species response to the declining [TP] and increasing pH between pre-industrial times and 1992 (Figure

6c). Changes between 1992 and 2007 are focused almost entirely along axis 1 (Figure 6c), showing a response to increasing lakewater pH.

2.6 Discussion

The chemical and biological conditions of the Muskoka-Haliburton study lakes have changed during the past 15 years. Significant changes observed in water chemistry variables are consistent with those reported elsewhere in the region (e.g. Palmer et al. 2011), but with some differences that are worth noting. On average, sulphate concentrations have continued to decline in lakes. Surprisingly, however, we also recorded small decreases in pH over the same time period. Although at first counter-intuitive, this is likely the result of a regional re-acidification event that was recorded in the spring of 2007. Previous studies in the region have shown that terrestrial environments receiving acid deposition may store excess sulphur in wetlands and littoral zones in a reduced form. During extended periods of drought, as occurred during the summers of 2005 to 2007, this sulphur may become oxidized and subsequently re-mobilized with rain events in the autumn (Dillon et al. 1997, Eimers et al. 2003). This may in turn cause short-

term declines in pH and a release of toxic metals (Dillon et al. 1997, Eimers et al. 2003, Adkinson et al. 2008). For example, one of our lakes (i.e., Dickie Lake) with long-term monitoring data showed a decline of almost 1 full pH unit in spring 2007 compared to the previous spring (A.M.

Paterson, DESC, unpublished). It is likely that chemical recovery will resume with continued reduction of sulphate deposition, however it is clear that this region is still susceptible to re- acidification events that may temporarily lower lake pH (e.g. Faulkenham et al. 2003), illustrating the value of long-term monitoring data.

Diatoms appear to be responding to chemical changes occurring over longer time periods, and not to short-term events such as the 2007 re-acidification. Thus, while a short-term decline in pH is evident in our water chemistry data (Figure 3, c), biological (diatom) communities appear to be tracking long-term pH increases associated with acid recovery (Figure 6, c). For example, we documented marked decreases in the acidophilic taxon, A. ralfsii var. americana in the majority of our study lakes (Figure 5). The estimated pH optimum of A. ralfsii var. americana in the region is

5.9, based on weighted-averaging (Hall and Smol 1996). Despite the small regional decline in lake water pH, nearly 80% of our sites still record a pH over 6.0, and almost half (47.8%) are greater than 6.5, putting them outside the environmental optimum of A. ralfsii var. americana.

While other studies have shown declines of almost 20% in lakewater total phosphorus concentration in south-central Ontario (Palmer et al. 2011), we observed no significant change in

TP across our 46 study sites since 1992. In the Palmer et al. (2011) study, lakes were sampled monthly between 1981 and 1990 and then revisited in 2004 or 2005 (Palmer et al. 2011). In contrast, our study compared samples obtained in 1990-1992 and 2007 or 2008. Long-term monitoring of a suite of eight lakes near Dorset suggests that oligotrophication in the region

ceased in the early 1990’s (Eimers et al. 2009), which may account for the discrepancy between the two studies.

RDA ordination analyses of species environmental relationships indicated that pH and nutrient concentrations (i.e., [TP]) form the two dominant ecological gradients in this dataset

(Figure 6, a). This is, in general, very similar to what was observed by Hall and Smol (1996).

Oligotrophication and acid recovery trends documented previously in south-central Ontario lakes were captured on the ordination diagram, as we observed the site centroid trend towards lower

[TP] and higher pH values (Figure 6, c). An increasing pH trend in the site centroids continued between 1992 and 2007, as regional TP trends remained relatively constant. While our data suggest that chemical recovery is occurring in this suite of lakes, trends in the ordination diagram clearly show that species assemblages are not returning to the pre-disturbance space but rather towards a novel diatom assemblage. In part, this is driven by a rise in planktonic taxa, and particularly Cyclotella stelligera in nutrient-poor lakes (Figure 7).

The planktonic taxon Cyclotella stelligera explains 20.5% of the difference between the

1992 and 2007/8 diatom datasets. In central Ontario, recent climate change has resulted in shorter periods of ice cover (this paper, Figure 2; Futter 2003), a longer diatom growing season and earlier onset and later breakdown of lake stratification (Stainsby et al. 2011). Moreover, a statistical evaluation of the regional ice record suggests that these changes are very recent.

Breakpoint analysis of regional ice-out data from a suite of 8 lakes studied extensively by the DESC in Muskoka suggests that a threshold response of lake ice to climate warming did not occur until after 1996 (Figure 4), which may explain why the increases in the relative abundances of planktonic diatom taxa were not observed by Hall and Smol (1996). These changes favour small planktonic Cyclotella species, particularly in nutrient-poor waters such as those common in south- central Ontario (Raubitschek et al. 1999, Rautio et al. 2000, Pannard et al. 2008). As Arctic and

alpine environments are more susceptible to these types of changes, they were the first to exhibit similar responses (Smol et al. 2005). However, more recent work synthesizing data on changes in relative abundance of Cyclotella over broad regional scales (Northern Hemisphere) has shown that abrupt increases in Cyclotella taxa, often at the expense of more heavily silicified Aulacoseira taxa and small benthic Fragilaria taxa, are occurring in more temperate latitudes as well (Rühland et al.

2008). Previous research has shown that increases in Cyclotella taxa are significantly correlated to spring ice-out trends (Rühland et al. 2008, 2010).

While increases in percent abundance of Cyclotella taxa are apparent in our regional study lakes, they are by no means systemic. It has been noted that in lakes at temperate latitudes, multiple stressors can often mask or supersede the influence of climate signals (Smol 2008, 2010).

This is apparent in our diatom data where the largest increases in C. stelligera occur at low TP (<10

g/L), while at higher TP concentrations we more commonly record increases in mesotrophic taxa such as Asterionella formosa and Fragilaria crotonensis (Figure 7). Together these diatom taxa (A. formosa and F. crotonensis) account for 16.8% of the variation between the 1992 and 2007/8 diatom species assemblages, and often occur in sites which record a concurrent decrease in heavily silicified Aulacoseira taxa or small benthic Fragilaria taxa.

In some alpine lakes, A. formosa and F. crotonensis have been linked to enhanced nitrogen availability from atmospheric sources (Saros et al. 2005) and, while we do observe a very small overall increase in total nitrogen in the Muskoka lakes, a broad range of nutrient requirements have been reported for these taxa (Rimet et al. 2009). Furthermore, relatively high

TN/TP (range: 22-94; median: 55) and TIN/TP (range: 1-45; median: 16) mass ratios indicate that our study lakes are predominantly P-limited (Downing and McCauley 1992, Guildford and Hecky

2000, Bergström 2010), and therefore we do not expect a strong algal response to changing nitrogen inputs or concentrations in these lakes. Finally, the species changes we observed

occurred during a period when locally-measured nitrogen deposition was stable or in slight decline (Schindler et al. 2006), suggesting a minimal link to recent changes in atmospheric inputs.

Thus, we believe that an earlier spring ice-off date, an extended diatom growing season, and enhanced lake stratification are the primary drivers of the increases in Cyclotella taxa in our lakes.

These physical changes appear to be having a similar effect on other pennate planktonic taxa in mesotrophic lakes. For example, notable increases in pennate planktonic taxa were recorded in the mesotrophic Bigstone Bay in Lake of the Woods, while the nearby oligotrophic Whitefish Bay in the same lake showed increases in small Cyclotella taxa (Rühland et al. 2010).

Previously, we discussed the implications of the methodological differences in water chemistry collection between our study, and Hall and Smol (1996). Specifically, we showed that for the majority of key chemical variables examined our use of single, spring samples from 2008/9 did not change our overall interpretation of chemical changes through time. A possible exception is for total phosphorus, where our dataset may underestimate recent increases in some lakes.

However, we do not believe this finding significantly alters our interpretation of the biological data. We believe that recent increases in A. formosa and F. crotonensis, in particular, are being driven primarily by climate and not by increases in nutrient concentrations. In several lakes, including Haliburton, Lake Muskoka - Muskoka Bay, Maple and St. Nora, we see marked increases in these planktonic taxa despite stable or decreasing spring 1992 vs. spring 2008 [TP]. For example, in Maple Lake where spring vs. spring [TP] has decreased by 2.25 g∙L-1, we see slight increases in both mesotrophic planktonic taxa, while in Haliburton Lake, a 0 g∙L-1 change in [TP]

(4.4 g∙L-1 in both 1992 and 2008) was met with a 19% combined increase in the relative abundance of A. formosa and F. crotonensis. Furthermore, in Muskoka Bay - Lake Muskoka, a slight increase in [TP] (i.e., 0.65 g∙L-1 , 11.55 g∙L-1 (1992) to 12.20 g∙L-1 (2008), has resulted in a similar 21% combined increase in these two planktonic taxa. These increased relative abundances

occurring at a variety of TP concentrations, and seemingly unrelated to [TP] change, strongly suggests that nutrient concentration changes are not driving the increased relative abundance of these mesotrophic planktonic taxa.

2.7 Conclusions

By comparing chemical and biological data from 2007 to those collected by Hall and Smol

(1996), we have assessed the trajectory of limnological change over the past 15 years. Our study lakes cover a range of anthropogenic disturbance (Hall and Smol 1996), and therefore it is not surprising to see a variable response. The warming climatic trend during this 15 year period has coincided with widespread increases in Cyclotella species, particularly in the oligotrophic lakes in the region. We contend that longer duration of the ice-free season, longer diatom growing season and enhanced lake stratification are important factors driving this change. However, while we have observed a marked diatom response to climate in oligotrophic lakes, the attenuating effects of other environmental stressors often mask the impact of climate warming in acidified and eutrophied lakes. Cyclotella species are often absent from lakes that were impacted by acidification (Smol and Stoermer 2010), nutrient loading (Rühland et al. 2008) and the overdevelopment of shorelines (Little et al., 2000). Documented changes in key indicator taxa

(i.e., A. formosa, F. crotonensis) often suggest nutrient enrichment. However this is not reflected in modern water chemistry measurements. Previous studies of alpine lakes have linked these taxa to enhanced nitrogen availability from atmospheric sources (Saros et al. 2005); but again, no apparent correlation exists between these taxa and changing nitrogen inputs in our lake set. It is possible that the longer ice-free periods, longer diatom growing season and enhanced lake stratification, which are primarily driving increases in planktonic Cyclotella taxa, are also providing

a competitive advantage to these other planktonic diatoms, despite relatively small changes in regional nutrient concentrations. Our study adds to the growing body of evidence showing the impact of recent climatic changes on diatom populations in many lake regions (e.g., Smol and

Douglas 2007, Rühland et al. 2008).

2.8 Acknowledgements

The authors would like to acknowledge Adam Jeziorski, Jennifer Korosi and Josh Thienpont for their assistance in the field. We also thank Neal Michelutti for his comments on early drafts of this manuscript. We thank the researchers and staff at the Dorset Environmental Science Centre for logistical support, water chemistry data collection and analysis. This research was made possible by an NSERC Discovery grant to JPS.

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Figure 2.1. Map of the Muskoka/Haliburton region showing the location of the study lakes (circles) and towns (squares).

Figure 2.2. Yearly mean maximum temperature and total annual precipitation data from the Muskoka Automated Weather Observation System (AWOS) climate station.

Figure 2.4. Spring ice-off data with piece-wise linear regression lines and breakpoint indicating threshold change in 1996.

Figure 2.5. One-to-one plots of selected diatom taxa, showing changes from 1992 to 2007/8. Numbers in brackets indicate the percent contribution of the taxa to the overall dissimilarity between 1992 and 2007 (i.e., SIMPER results). Wilcoxon Signed Rank results comparing 2007 and 1992 data are presented for all 6 taxa. The second analysis for Cyclotella stelligera is based on an identical Wilcoxon test excluding Gravenhurst Bay, Muskoka Bay and Young Lake.

Figure 2.6. RDA analysis for modern water chemistry and diatom species assemblages. a) Backwards selected RDA showing the overall relationship between diatoms and environmental variables. Variables removed with VIF> 5 and physical lake variables are plotted passively (grey). b) RDA with forward selection showing key environmental variables. c) Site centroids of all lakes with standard error bars showing change in diatom species assemblages since pre-1850.

Figure 2.7. Changes in relative abundances (relative abundance 2007 - relative abundance 1992) of selected diatom taxa between 2007 and 1992, with lakes arranged from lowest to highest [TP] (see top of figure for values).

Table 2.1. Summary of the key physical and chemical limnological variables from the Muskoka-Haliburton lakes.

Surface Depth Lake Area (max) Elevation Volume Ca Conductivity DOC K Mg ha m m asl *106·m3 mg·L–1 S·cm–1 mg·L–1 mg·L–1 mg·L–1 Min 11.0 4.5 223.50 0.71 1.28 16.40 2.00 0.18 0.25 Max 1675.0 80.5 403.90 442.91 10.20 127.00 7.90 5.99 10.00 Mean 260.1 29.0 320.02 41.22 3.74 42.19 4.18 0.82 1.41 Std 347.3 18.7 46.67 85.79 2.14 23.87 1.32 1.11 2.17

NH3 NO3 TKN pH TP Alkalinity SO4 Na SiO2 –1 –1 –1 –1 -1 –1 –1 –1 g·L g·L g·L μg·L mg·L as CaCO3 mg·L mg·L mg·L Min 2.00 2.00 165.00 5.50 3.70 0.20 3.10 0.49 0.32 Max 154.00 160.00 479.00 7.88 22.80 23.30 7.05 21.00 1.92 Mean 21.15 95.63 268.93 6.47 7.61 7.32 5.24 3.34 1.11 Std 24.98 34.23 68.22 0.54 4.08 6.50 0.90 4.57 0.43

Chapter 3

Recent reversals in competitiveness of colonial scaled chrysophytes in south-central Ontario lakes: paleoecological evidence of declines in nuisance algae

3.1 Abstract

Anthropogenic stressors, operating both regionally and locally, have drastically changed the limnological properties of lakes in Ontario over the past several decades. To quantify these long-term changes in chemical and biological properties in south-central Ontario, previous work from our lab used paleolimnological techniques to estimate changes in scaled chrysophyte taxa in a suite of 54 lakes between the preindustrial era and 1992. One of the observed consequences of these stressors has been wide-spread blooms of nuisance colonial chrysophyte algae. However, since 1992, the trajectory of chemical and biological conditions in these systems has been altered by multiple natural and anthropogenic stressors. Here, we assess the extent of limnological changes that have occurred during the past ~15 years in the Muskoka-Haliburton study lakes by resampling the same lakes using identical limnological and paleolimnological methods to those used in the 1992 comparison. We observed changes in both water chemistry (i.e., decreased pH and increased dissolved organic carbon) and scaled chrysophyte microfossil assemblages since

1992. Despite continued warming, the regional-scale increases in nuisance species Synura petersenii that had previously been identified has not continued and in many cases have begun to reverse. We believe these changes are being driven by complex interactions between resource limitation (specifically DOC and total phosphorus), acidification and recovery, and climate change.

3.2 Keywords

Paleolimnology, Muskoka, Haliburton, reacidification, DOC, scaled chrysophytes

3.3 Introduction

Despite their relatively remote location, lakes in Muskoka-Haliburton have been influenced by multiple environmental stressors during the past several decades (e.g., Hall and

Smol 1996; Paterson et al. 2004; DeSellas et al. 2008; Yan et al. 2008). Although distant from major urban and industrial centres, these lakes are a popular destination for recreational activities, and are home to many cottages, resorts and golf courses. As a result, lakes in the region are susceptible to both regional scale stressors, such as climate change, as well as local, catchment-scale stressors associated with increased shoreline development. During the past several decades, the Dorset Environmental Science Centre (DESC) has tracked the effects of anthropogenic stressors on the aquatic and terrestrial ecosystems in the Muskoka area. In addition to long-term monitoring, paleolimnological methods have been used to place recent trends into the context of longer-term post-industrial changes (e.g., Paterson et al. 2001).

Complex interactions amongst multiple stressors, including forest clearance and regrowth, increases in shoreline residential development, the deposition of strong acids via the atmosphere, invasion by non-indigenous species, and climate change, have resulted in significant changes to physical, chemical and biological conditions of lakes during the past several decades (Yan et al.

2008). One consequence of these changes has been a regional increase in the relative abundance of colonial scaled-chrysophytes in boreal shield lakes, often characterized by near systemic increases in relative abundance of the nuisance taxon Synura petersenii (Dixit and Smol 1992;

Paterson et al. 2001; Vinebrooke et al. 2002; Hyatt et al. 2010). Blooms of Synura petersenii have been linked to taste and odour events in freshwater lakes and reservoirs (Nicholls and Gerrath

1985; Nicholls 1995; Watson et al. 2001), however, despite extensive research, the mechanisms responsible for these blooms remain unknown.

Several hypotheses have been generated to explain the recent increases in colonial chrysophytes in lakes across North America. Mechanisms believed to contribute to taxonomic shifts in chrysophyte species include: 1) variation in species- or group-level tolerances to pH

(Paterson et al. 2001), 2) increased transparency associated with declining [DOC] (Leavitt et al.

1999), 3) changes in deposition of metals (Davis et al. 2006), 4) climate-driven changes in lake thermal stability (Paterson et al. 2004; Ginn et al. 2010), and 5) resource limitation driven by changes in nutrients and light regime (Findlay et al. 2001). Previous research has shown scaled chrysophytes to be clearly sensitive to changes in lakewater pH, with taxonomically distinct pH optima and tolerances which has made them an ideal indicator in paleolimnological reconstructions (Cumming et al. 1992; Dixit et al. 1999; Paterson et al. 2001). Meanwhile, with reduced transparency, deepwater chrysophyte populations are at a selective advantage due to reduced exposure to harmful ultraviolet radiation, reduced grazing pressure, and increased proximity to nutrient-rich hypolimnetic waters (Leavitt et al. 1999; Vinebrooke et al. 2002;

Paterson et al. 2004). Based on their study of 2 lakes in New Hampshire, Davis et al. (2006) have suggested that chrysophyte species shifts were the result of increased metal deposition caused by both catchment disturbance and atmospheric inputs. Finally, it has been suggested that, because scaled chrysophytes are a highly motile, planktonic algal group (Sandgren 1988), they are favoured under conditions of increased thermal stability (Paterson et al. 2004, 2008; Ginn et al. 2010). For example, several lakes surveys have shown that, in general, lakes which lack substantial chrysophyte populations tend to be shallow lakes, which would be unlikely to stratify during the ice-free season and thus provide advantageous habitat for chrysophyte species (Ginn et al. 2010).

Recent climate warming has led to earlier ice-off (Futter 2003) which may, in turn, lead to enhanced thermal stratification and thereby change important properties of lake water columns.

Similar changes in thermal stability have been shown to have drastic, widespread biological

consequences on other planktonic algal groups that find themselves at a competitive advantage

(e.g., Rühland et al. 2008, 2010).

In addition to many other stressors, nutrient concentrations are in a state of flux in lakes in south-central Ontario. For example, lake [DOC] has shown an inverse relationship with ionic flux (i.e., Ca, Mg and SO4; Montieth et al. 2007), suggesting that recovery from acidification may be playing an important role in the recent increases in [DOC]. Furthermore, reduced atmospheric deposition of acidifying compounds results in increased soil pH and reduced Al mobilization which, in turn, increases soil organic matter mobility and decreases the binding of Al with organic molecules, thereby contributing to increased DOC concentration in recovering lakes (Montieth et al. 2007). [DOC] in lakes is also partially climate dependant, with supply being influenced by both temperature (affecting the production and decomposition of organic matter) and precipitation

(exporting DOC to lakes via precipitation) (Dillon and Molot 2005; Keller et al. 2008; Eimers et al.

2008a). DOC is a major determinant in the transparency of lakes and therefore is vital in shaping lake thermal structure. DOC is not the only key water quality variable that has changed in the

Muskoka-Haliburton Lakes since 1992. Regional lake surveys suggest that prior to 1992 total phosphorus concentration was, in general, declining in many lakes despite increased shoreline development and recreational use (Hall and Smol 1996). However, since 1992 evidence suggests that [TP] has stabilized or begun to increase in many of the Muskoka-Haliburton lakes (Eimers et al. 2009, Fig. 2; A.M. Paterson, DESC, unpublished data).

In order to anticipate future trajectories of change in lakes, we must continue to expand our understanding of the biological changes that are occurring in these systems. Given the lack of long-term monitoring available for many lakes in this region, there is an obvious need for indirect proxy techniques such as those employed by paleolimnologists (Smol 2008). To evaluate changes

in water quality since preindustrial times, Paterson et al. (2001) examined 54 lakes in the

Muskoka-Haliburton region to document long-term changes in chrysophyte assemblages and chrysophyte-inferred water chemistry (i.e., pH). These lakes were originally selected by Hall and

Smol (1996) to minimize the range of pH (i.e., no lakes with pH <5.5) and DOC (i.e., no lakes with

[DOC] >6.5 mg∙L-1), so as to assess the influence of total phosphorus ([TP]) on diatom community composition. This suite of lakes provided a diverse cross-section of the physical and chemical limnological variability of the region (Hadley et al. 2012, Table 1). Paterson et al. (2001) used a

“before and after” or so-called “top-bottom” paleolimnological approach (Smol 2008) to assess changes in chrysophyte-inferred pH since pre-industrial times. Although absolute changes were found to be relatively small, chrysophyte-based pH inferences suggested that many of the acidic lakes (pH < 6 in 1992) had acidified since pre-industrial times (Paterson et al. 2001), the result of the long-range transport and deposition of acids from industrial activities (Dillon et al. 1987).

Since the initial Paterson et al. (2001) study, a synopsis of the DESC long-term chemistry data

(Table 1, Palmer et al. 2011) documented marked changes in water quality in this region, while

Hadley et al. (2012) have demonstrated significant changes in diatom species assemblages associated with climate warming during the past 15 years. In general, Palmer et al. (2011) found that regional stressors, including acidic deposition, climate change, and shoreline development, resulted in decreasing acidity, calcium and metals, and increased concentrations of dissolved organic carbon (DOC) and ammonium. Hadley et al. (2012) recorded a marked biological response in diatoms since 1992, driven primarily by an increased relative abundance of Cyclotella stelligera in oligotrophic lakes, previously shown to be indicative of recent climatic warming and associated changes in the duration of ice cover and thermal stability (Rühland et al. 2008).

Given the many changes to water chemistry and physics that have been observed in regional lake surveys, and recent changes reported for diatoms, we explore how these recent

limnological changes have altered chrysophyte communities during the past 15 years. In many oligotrophic, softwater lakes, chrysophytes may comprise 25-75% of the total phytoplankton biomass (Siver 1995; Paterson et al 2008) and are of particular interest to lake managers because of undesirable taste and odour events produced by some taxa (Nicholls and Gerrath 1985; Nicholls

1995). Paleolimnological techniques, using biological indicators such as scaled chrysophytes, diatoms and zooplankton remains, are valuable tools in assessing and reconstructing ecological change due to a variety of natural and anthropogenic phenomena (Smol 2008). Here, we used scaled-chrysophyte microfossils and modern water-chemistry data to evaluate the importance of changes in DOC and TP concentrations to biological changes in this region since 1992. To this end, we have revisited the Muskoka-Haliburton lakes (Paterson et al. 2001), collecting both surface sediments and modern water chemistry. By comparing pre-industrial, 1992 and modern (ca.

2007/8) limnological conditions, these data shed new light on the patterns of limnological and biological change during the past ~15 years in Muskoka area lakes, highlight the complex nature of multiple stressor systems, and provide useful knowledge to inform future lake management and policy decisions.

3.4 Methods

3.4.1 Study Lakes

The 54 lake dataset of Hall and Smol (1996) and Paterson et al. (2001) was originally sampled for sedimentary diatom and scaled-chrysophyte assemblages in the fall of 1992. Water chemistry variables were determined from annual averages of volume-weighted epilimnetic or metalimnetic composites, sampled biweekly, monthly, or bimonthly during the ice-free seasons from 1990 to 1992 (Hall and Smol 1996, Table 2). Due to logistical and other issues, the total number of sampling sites was reduced in our study, as outlined below. We revisited and obtained

new sediment cores from 53 of these sites over two field seasons in July/August of 2007 and 2008

(no core could be obtained for the Seagull Rock site at Lake of Bays). Additionally, chrysophyte scales were extremely rare or showed signs of dissolution in the sediments of six lakes (Three Mile

Lake-Main Basin, Bruce, Butterfly, Plastic, Heney, and Frazer Island) and therefore they were eliminated from further analysis. However, the remaining lakes showed excellent preservation of siliceous material including subtle features such as spines on chrysophyte remains. As a result of logistical constraints within the DESC, water chemistry was only obtained for 46 sites of the 53 sediment sites (no chemistry data were collected for Kelly, Little Redstone, Little Kennisis, Lake

Joseph-Frazer Island, Lake Muskoka-Gravenhurst Bay, Red Pine and Redstone lakes). The combined effect of these issues has reduced the overall dataset from the original 54 to 41 lakes, for which we have full chrysophyte and water-chemistry data.

The Muskoka-Haliburton region is situated within the boreal ecozone with a temperate climate. Mean maximum annual air temperature has increased from 9.2 °C during the 1992 field season to 11.7 °C during the 2007 field season, while total precipitation has fluctuated from

1163.1 mm in 1992 to 850.6 mm and 1125.6 mm in 2007/8, respectively (Hadley et al. 2012, Fig.

2). Catchments are mainly forested, with mixed hardwood (i.e., birch, oak and maple) and coniferous (i.e., hemlock, spruce, white and red pine and tamarack) species. Currently, anthropogenic impacts in the region vary from lake to lake, and are largely associated with urban and recreational (cottages, resorts, golf course) activities. The study lakes are located in the

Muskoka-Haliburton region of south-central Ontario, Canada, situated on the Precambrian Shield

(Figure 1). Twelve of the study lakes are underlain by a marble intrusion found in the eastern portion of the study area. Soils on the Canadian Shield are generally thin, contain little or no carbonate-bearing material, and are underlain by resistant granitic bedrock. These thin soils,

combined with a rough topography and a cool climate, have resulted in little agricultural development in the region.

3.4.2 Water sampling procedures

Water samples were collected from each lake by technical staff at the Ontario Ministry of the Environment’s Dorset Environmental Science Centre (DESC) in the spring of 2008 (42 lakes) or

2009 (4 lakes). Water-chemistry sampling protocols differed slightly from those described by

Paterson et al. (2001). Samples collected by Hall and Smol (1996) were volume-weighted, ice-free means from samples collected between 1990 and 1992, while our measurements were single value spring samples. However, analytical protocols were identical to those summarized in

Paterson et al. (2001) and detailed in Hall and Smol (1996). Analyses of water samples for 14 key variables were performed at the DESC following standard OMOE protocols (Ontario Ministry of the Environment 1983), and are summarized along with four physical parameters in Hadley et al.

(2012, Table 1).

3.4.3 Surface sediment sample collection

A short gravity sediment core was collected from the deepest portion of each lake using a

3-inch (7.6-cm) internal-diameter, high-resolution Glew (1989) gravity corer, and sectioned on site using a Glew (1988) vertical extruder, using identical equipment and procedures as Paterson et al.

(2001). The top-most 0.25 cm interval of our sediment cores were extruded and used for chrysophyte analysis. This differs from the Paterson et al. (2001) protocol where the modern assemblage was represented by a single 1 cm slice of sediment removed from the top of the core.

All comparisons to the pre-industrial time period used the sediment samples collected by Hall and

Smol (1996) and later enumerated and analyzed by Paterson et al. (2001). The sediment samples were stored in Whirlpak® bags and refrigerated in the dark until analysed. 53

3.4.4 Scaled chrysophytes

Following standard protocols for siliceous remains (Battarbee et al. 2001), we digested 0.2

- 0.3 g of wet sediment in a 50:50 (v/v) mixture of concentrated nitric and sulphuric acid, and rinsed repeatedly with deionized water until a neutral pH was achieved. The resulting cleaned scaled chrysophyte slurries were dried onto coverslips and mounted on microscope slides using

Naphrax® mounting medium. Siliceous microfossils were then identified and enumerated at

1000X (100X objective, 10X ocular) under oil immersion using a Leica DMR2 microscope with differential interference contrast optics. A minimum of 300 chrysophyte scales were counted for each sample and identified to the species level wherever possible. Due to the difficulty in identifying some scales to species level using light microscopy, a few taxa were grouped into broader taxonomic categories (e.g., Mallomonas ‘small’ and ‘medium’), using the identical criteria described in Paterson et al. (2001), thus allowing for direct comparisons. However, these grouped categories never accounted for more than 5% relative abundance.

3.4.5 Statistical analysis

Prior to statistical analysis, rare taxa whose relative abundances did not exceed 1% in at least three lakes were removed from the dataset. The 25 scaled-chrysophyte species that remained in the reduced dataset accounted for a minimum of 96.96% of the total relative abundance. Analysis of similarities (ANOSIM; Clarke 1993) tests were used to compare the modern (2007/8) assemblages with assemblages in both the surface sediment collected in 1992 and the pre-industrial “bottom” samples (pre-1850) (Paterson et al. 2001). ANOSIM is a non- parametric multivariate test which uses the rank order of dissimilarity values from a dissimilarity matrix (e.g., Bray-Curtis) to test the significance of dissimilarity between two or more groups that have been defined a priori. If the groups are different in their species composition, then the

dissimilarities between the groups will be greater than those within the groups. Here we used a one-way ANOSIM test of the regional chrysophyte datasets (i.e., 45 lakes from 2007/8, 1992 and

1850) to compare 2007 vs. 1992, 2007 vs. pre-1850, and 1992 vs. pre-1850. This analysis was performed on Bray-Curtis dissimilarity metrics using non-transformed species data with 9999 permutations. In addition to ANOSIM, a Similarity Percentages (SIMPER) analysis was used to determine the contribution of individual chrysophyte taxa to the dissimilarity between the datasets. ANOSIM and SIMPER do not allow for paired sampling (i.e., direct comparison of species composition lake by lake), therefore we also conducted several paired univariate analyses (i.e.,

Wilcoxon Signed Rank test) of key chrysophyte taxa, to determine if the changes in relative abundance between the three time periods were significant. Analysis of multiple comparisons can result in increased risk of Type-I error. To counter this effect, we calculated the false discovery rate (FDR) for the Wilcoxon Signed Rank results using the “Multtest” package in R v. 2.13

(Benjamini and Hochberg 1995). Colonial chrysophyte changes since 1992 plotted relative to category groupings of key environmental variables (i.e., pH, DOC, and TP; Sandgren 1988) were summarized in a series of boxplots, and ANOVA analyses with Tukey post-hoc tests were run to determine whether the changes in relative abundances among groupings were significant.

ANOSIM and SIMPER analyses were performed using the software Primer v. 6 (Clarke and Gorley

2006), while ANOVAs and Wilcoxon Signed Rank tests were performed using Systat v. 11.

3.5 Results

Surface-sediment chrysophyte assemblages of the 45 Muskoka-Haliburton study lakes were composed of a total of 25 taxa and groups. In general, the chrysophyte assemblages of all three time periods were dominated by Synura petersenii Korshikov, S. sphagnicola (Korshikov)

Korshikov, S. echinulata Korshikov, S. uvella Ehrenberg, Mallomonas crassisquama (Asmund) Fott

and M. duerrschmidtiae P.A.Siver, J.S.Hamer & H.J.Kling. We observed regional declines in relative abundances of colonial chrysophyte taxa since 1992 (e.g., S. petersenii and S. sphagnicola), with subsequent increases in small, unicellular species (i.e., M. duerrschmidtiae and M. crassisquama)

(Figure 2). Significant regional increases in S. echinulata were also observed in many of the study lakes, with only four of 45 lakes showing declines (Figure 2). In contrast, marked declines in the relative abundance of S. sphagnicola and other colonial chrysophytes occurred particularly in lakes whose pH had fallen below 6.0 (Figure 4), but according to Wilcoxon tests these declines were not regionally significant. In 75% of the 12 lakes located within the marble intrusion, S. petersenii either increased in relative abundance (by ≥ 5%) or did not change (± ≤5%) (Figure 5). Similarly, S. sphagnicola remained relatively stable, increasing (by ≥ 5%) in two lakes and remaining stable in the remaining 10 (Figure 5). The single exception amongst the marble intrusion lakes was

Boshkung Lake, which had a >10% decline in colonial chrysophyte relative abundance between

1992 and 2007.

According to SIMPER analysis, S. petersenii, S. sphagnicola, S. echinulata, S. uvella, M. crassisquama and M. duerrschmidtiae accounted for 71.43% of the cumulative dissimilarity between the 2007 and 1992 datasets (Figure 2). Synura petersenii accounted for 19.13% of the dissimilarity between 2007 and 1992, with an average relative abundance decrease from 27.91%

(1992) to 22.95% (2007). Between 1992 and pre-1850, S. petersenii accounted for 17.38% of the dissimilarity with an average relative abundance increase from 8.25% (1850) to 27.91% (1992).

The dissimilarity between the composition of 1992 and pre-industrial (pre-1850) sedimentary chrysophyte assemblages (ANOSIM R = 0.449, p ≤ 0.001), and between the 2007 and pre-1850 assemblages were similar (ANOSIM R = 0.436, p ≤ 0.001). Dissimilarity between the 2007 and

1992 assemblages was the lowest (ANOSIM R = 0.164, p ≤ 0.001), suggesting that scaled

chrysophyte assemblage changes since 1992 were small relative to changes since pre-industrial times.

3.5.1 Water chemistry analyses

As noted in the Methods, Hall and Smol (1996) water chemistry sampling technique differed slightly from the DESC sampling regime in 2008/9. Samples collected in 1992 were ice- free means from samples collected between 1900 and 1992, while our measurements were single value spring samples. To explore the potential implications of this difference, using eight lakes that are intensively studied by the Ontario Ministry of the Environment (i.e., Dorset A lakes), we ran several comparisons. First, we checked the 1992 spring data from the Dorset A lakes against the ice-free mean data from 1992 to determine if the single point measurements were similar to the yearly ice-free means (Appendix B). Second, we plotted the single measurement 2007/8 water chemistry data against both the spring 1992 and ice-free mean 1992 data (Appendix B).

Results from these tests showed that pH, [DOC], SO4, and Ca were not substantially affected

(Appendix B), and showed long-term trends that were similar to those presented for 36 lakes in south-central Ontario from the 1980s to 2004-05 (Palmer et al. 2011). However, [TP] did show substantially different results when comparing spring 2007/8 to spring 1992 values. In general, we see greater overall increase in [TP] in the “A” lakes when comparing spring vs. spring rather than spring (2007/8) vs. ice-free mean (1990-1992). This is likely because of the high spring values which would be expected following snow and ice melt.

3.6 Discussion

Biological changes since pre-industrial times have been characterized by marked regional increases in the relative abundance of colonial chrysophytes, particularly S. petersenii, in many boreal shield lakes (Dixit and Smol 1992; Paterson et al. 2001; Vinebrooke et al. 2002). While the 57

specific causes for these increases have not yet been determined, data suggest the timing is consistent with recent climate warming and industrial activities (Paterson et al. 2004; Hyatt et al.

2010). The initial survey of the Muskoka-Haliburton lakes by Paterson et al. (2001) showed widespread increases in S. petersenii since pre-industrial times to 1992. However, despite continued warming in the region (Hadley et al. 2012, Figure 2), we have documented regional declines of several colonial chrysophytes (e.g., S. petersenii and S. sphagnicola) over the past 15 years. Despite the increased abundance of S. echinulata, decreases in S. petersenii and S. sphagnicola have resulted in an overall decrease in colonial chrysophytes. Documented declines in colonial chrysophytes, despite continued warming and the associated increases in thermal stratification, suggests that thermal stability hypotheses described elsewhere (e.g., Ginn et al.

2010) is not driving chrysophyte assemblages shifts in these lakes. One of the limitations of this type of paleolimnological study (i.e., top-bottom) is that we only have data from distinct points in time; therefore it is difficult to be certain about the timing of biological change. Therefore, future research to further elucidate the timing of changes in the Muskoka-Haliburton lakes, involving detailed downcore analyses to more precisely estimate the timing of biological change, would be advantageous.

3.6.1 The role of water chemistry and nutrient limitation in colonial chrysophyte declines

We believe the recent declines in colonial scaled chrysophytes are linked to changes in nutrient availability, and specifically the complex interactions between climate warming, acidification and recovery, and their influence on lakewater pH, DOC and TP. DOC concentrations in south-central Ontario lakes have been linked to reduced ionic flux (i.e., Ca, Mg and SO4;

Montieth et al. 2007), and despite reduced atmospheric sulphate deposition, [SO4] remained high in the Muskoka-Haliburton lakes until after the initial sampling by Hall and Smol (1996) in the early

1990s (Eimers et al. 2004). Since that time, a steady decrease in [SO4] and subsequent increase in

[DOC] throughout the Muskoka-Haliburton region has been documented (this paper; Eimers et al.

2004; Palmer et al. 2011). Declining atmospheric deposition of SO4 documented since 1992 would lead to increased soil organic matter mobility and decreased Al binding with organic molecules, and thereby contribute to increased [DOC] in recovering lakes (Montieth et al. 2007). DOC is the most important factor controlling light penetration in small (< 500 ha) Canadian Shield lakes, and will therefore strongly control the thermocline depth and light attenuation of lakes (Perez-

Fuentetaja et al. 1999). For example, in Northeastern Ontario lakes, DOC was considered the best individual predictor of late summer epilimnion thickness (i.e., high DOC levels equate to a reduced epilimnion) (Keller et al. 2006). Furthermore, small changes in DOC can significantly alter the penetration of harmful UV-B radiation, particularly in lakes with historically lower DOC concentrations (Scully & Lean 1994).

Increases in [DOC] appear to correlate well with biological change in colonial chrysophytes in the Muskoka-Haliburton lakes. Lakes with increasing DOC concentrations tend to record the largest declines in colonial chrysophyte taxa, whereas less chrysophyte change was observed in lakes whose DOC concentration has remained stable or in rare cases increased (Figure 3a, b).

Colonial chrysophyte species form deepwater populations and may therefore have had a selective advantage over unicellular epilimnetic forms in lakes with lower DOC concentrations (i.e.,

Mallomonas spp.), allowing them to avoid harmful UV radiation. This ability to form deepwater populations might also provide the added benefits of reduced grazing intensity, and an increased proximity to nutrient-rich hypolimnetic waters (Fee et al. 1977; Nakamoto et al. 1983; Sandgren

1988; Leavitt et al. 1999; Vinebrooke et al. 2002). Since 1992, the nearly systemic increases documented in [DOC] in the Muskoka-Haliburton study lakes may have negated or reduced this

competitive advantage by reducing transparency and thereby reducing light availability in deeper water.

Further evidence for suggesting that DOC is driving chrysophyte change comes from a closer examination of the 12 well-buffered (Alkalinity: 5.97 to 23.30 mg∙L-1; mean 15.31 mg∙L-1) study lakes that are located on a marble intrusion. These lakes possess unique chemistry and, as a result, demonstrate the potential for a highly variable response in the region. All of the marble

-1 intrusion lakes have higher [SO4] (5.8 - 7.05 mg∙L ) compared to the average of all 41 of our study lakes (5.24 mg∙L-1). Eight of these 12 lakes also have atypically high [Ca] and [Mg] (>5.0 mg∙L-1and

>1.5 mg∙L-1, respectively) relative to the other softwater lakes typical of the Canadian Shield.

Interestingly, these eight marble lakes are also the only lakes whose [DOC] remained unchanged

(2 lakes) or experienced a decrease (6 lakes) since 1992. This is likely because the unique geology of these lakes has resulted in increased buffering of both lakewater and soils, which would have tempered lakewater pH and DOC change. Consequently, in contrast to the majority of our study lakes, colonial chrysophytes have remained stable or increased slightly in these lakes (Figure 5), as they continue to realize a selective advantage.

ANOVA analysis comparing changes in relative abundance across key environmental categories were not significant, this is likely in part due to variability between lakes (e.g., the marble intrusion lakes). The lakes within this dataset vary considerably in physical parameters including both size and depth. Hanson et al. (2011) have shown that differences in size can affect the extent to which lakes process DOC. Therefore, large lakes, which tend to have longer residence times, would retain DOC longer and thereby allow more time for DOC to be mineralized.

Furthermore, evidence also suggests that lakes, whose DOC concentrations are highest, will have a restricted depth of mineralization of DOC due to lower light penetration (Hanson et al. 2011). It is

also important to note that not all colonial chrysophyte taxa are declining in relative abundance.

For example, Synura echinulata increased from 1992 to 2007 in the majority of the study lakes regardless of measured lakewater pH or [DOC] (Figure 2). S. echinulata has been previously identified as a cold-water taxon, having a bimodal pH distribution (Cumming et al. 1992) which blooms in late summer to early autumn, persists under the ice in the winter and then declines in early spring (Siver 1995). Therefore, it is surprising that S. echinulata is increasing in all but 4 of the Muskoka-Haliburton lakes despite warmer waters and increasingly shorter periods of ice cover

(Futter 2003). Some evidence suggests that S. echinulata prefers humic waters (Siver 1987), which may explain why it has thrived with increasing [DOC] since 1992; however the mechanism responsible for increases in this taxon remains largely unknown. The overall increased relative abundance of the colonial taxon, Synura echinulata (Figure 2c), is also unusual and likely significantly influencing the ANOVA analysis. It may be worth noting that Paterson et al. (2001) previously suggested this taxon may potentially be Synura leptorrhabda (Asmund) Nicholls comb. nov., a taxon that may only be reliably separated from S. echinulata using electron microscopy and whose ecological preferences are largely unknown. When excluding S. echinulata from the

ANOVA analysis, colonial chrysophytes changes across the four %DOC categories were significant

(F=3.818, p=0.017, df=3) when %DOC increased by more than 20% (pairwise Tukey post-hoc p=0.034 between groups 3 and 4, Figure 3b). However, further research is needed to establish what is driving increases in colonial taxon, S. echinulata.

In addition to [DOC], changes to other key variables may be contributing to the loss of colonial chrysophytes since 1992. For example, we see relatively smaller declines in colonial chrysophytes in lakes with decreasing TP. In contrast, lakes in which TP has increased showed the most pronounced declines in the relative abundance of colonial chrysophytes (Figure 3c, d). One of the potential mechanisms suggested for the increases observed between the pre-industrial

period and the 1990s was oligotrophication, or the decrease in TP observed in south central

Ontario lakes (Paterson et al. 2004; Hyatt et al. 2010). Decreasing TP across the Muskoka-

Haliburton region has been well documented between pre-industrial times and the early 1990s

(Hall and Smol 1996; Eimers et al. 2009), and may lead to resource shortages in stratified, oligotrophic lakes that are common to south-central Ontario. Reduced TP, particularly when coupled with decreased [DOC], may have favoured highly motile colonial scaled chrysophytes that form deepwater populations, and can therefore take advantage of nutrient-rich hypolimnetic waters (Findlay et al. 2001), as well as the absence of epilimnetic predators and reduced levels of harmful UV-B radiation. However, since the early 1990s, [TP] in south-central Ontario has begun to stabilize or increase in ~50% of our study lakes (Eimers et al. 2009, Fig. 2; Hadley et al. 2012).

Increased [TP] may reduce the competitive advantage of colonial chrysophytes, whose high motility and proximity to nutrient-rich hypolimnetic water, allow them to outcompete small unicellular forms which are restricted to the epilimnion. Moreover, as mentioned in the results, comparing the spring [TP] from 2007/8 to the ice-free means calculated by Hall and Smol (1996) appears to have slightly underestimated the increases in TP since 1992 (suppl. data). Thus, the importance of increased [TP] to the decline in colonial chrysophytes may be greater than our data suggests.

3.6.2 The role of climate in scaled chrysophyte change

Climate change in the Muskoka-Haliburton region presents a constantly shifting set of variables with complex interactions operating at multiple scales (Smol 2010), making it difficult to forecast its effect on our study lakes. Changes in air temperature and precipitation affect allochthonous inputs of DOC, which subsequently may influence the absorption of solar irradiance and thermal conditions in lakes (Schindler et al. 1996b; Snucins and Gunn 2000). DOC supply from

the terrestrial environment is also partly temperature dependent (Schiff et al. 1997; Davidson and

Jannsens 2006), while the export of DOC to lakes, via runoff (Dillon and Molot 2005; Eimers et al.

2008b), is also impacted by climatic factors such as precipitation, evaporation and evapotranspiration (Keller et al. 2008). Additionally, the export of sulphate in streams has been shown to be higher in years with warm, dry summers (Eimers et al. 2004). The impact of climate on [DOC] is further confounded by the presence and extent of wetlands within the lakes catchment. For example, Keller et al. (2008) found that long-term average temperature was the most important predictor of [DOC], except in areas with the highest percent cover of wetlands, where precipitation outstripped temperature as a predictor of DOC concentrations in lakes.

Furthermore, wetlands are also an important factor in drought-induced re-acidification events.

Wetland-mediated re-acidification tends to result in increased loads of sulfate and DOC, in addition to decreases in pH (Yan et al. 1996; Evans et al. 1997; Dillon and Evans 2001; Faulkenham et al. 2003). For example, the inflow of Blue Chalk Lake, which drains no wetland, had a mean (10 yr) [DOC] of 2.2 mg∙ L-1, while inflows to Chub Lake, which drain 4% wetlands, measured >8.1 mg∙

L-1 (Faulkenham et al. 2003). Because of the variability in wetlands coverage between these two lakes, water chemistry (i.e., pH) in Chub Lake has been more variable during the past 20 years, whereas Blue Chalk Lake showed signs of chemical recovery from acidification following reductions in sulfate emissions (Dillon and LaZerte 1992; Dillon et al. 1997). Wetland coverage around lakes in the Muskoka-Haliburton region varies between 0 and 19.5% (Dillon 1986; A.M.

Paterson, DESC, unpublished data) and may therefore contribute to the variability recorded across the Muskoka Lakes. While wetland data are not available for all 45 of our study lakes, making it difficult to fully elucidate its impact, climate-driven increases in wetland-derived SO4 may temper increases in [DOC].

3.6.3 Other considerations

While we observed no significant correlation between lakewater pH change and colonial chrysophytes, our data suggest that pH still plays an indirect role in scaled chrysophyte species assemblages, particularly in the most acidic lakes (Figure 4a). While similar changes to the relative abundance of colonial chrysophytes occurred in lakes regardless of the direction or magnitude of pH change (Figure 4b), lakes with lakewater pH below 6.0 appeared to have undergone more marked declines in colonial chrysophytes since 1992 (Figure 4a). Surprisingly, the most marked declines in S. sphagnicola occurred in our most acidic lakes (pH<6), despite evidence suggesting a relatively acidic pH optimum (Siver 1995; Kristiansen 2000; Paterson 2001, Fig. 3b). This suggests that in these lakes, rather than directly influencing chrysophyte scales, pH has had a more indirect effect. As a result of acidification, DOC may be lost by: 1) photochemical processes such as photobleaching (Findlay et al. 1999; Graham et al. 2007), 2) bacterial conversion to CO2 and subsequent evasion to the atmosphere or, 3) precipitation with metals and subsequent incorporation into sediments (Clair et al. 1999; Gennings et al. 2001). Furthermore, the degree of

DOC loss as a result of acidification is often greater at lower pH (Schindler et al. 1996a).

Therefore, those lakes which acidified the most between pre-industrial times and 1992 may have likewise seen the greatest declines [DOC] and increases in colonial chrysophytes. With reduced acid deposition, these lakes would have the most potential for both chemical (i.e., increased pH and DOC) and biological (i.e., decreased colonial chrysophytes) recovery.

A final consideration that warrants discussion is the possibility that chrysophytes are changing in response to loss of CO2. In circumneutral lakes recovering from acidification, increased pH may result in shifting inorganic carbon speciation (i.e., an overall shift from CO2 to bicarbonate). The inability of many chrysophyte taxa to secrete alkaline phosphatase and therefore use bicarbonate as an inorganic carbon source makes them dependent on CO2 (Saxby-

Rouen et al. 1998; Bhatti and Colman 2005, 2008; Maberly et al. 2009). It has therefore been

suggested that shifting DIC speciation at high pH may contribute to the decline in Synura petersenii and other Synurid chrysophyte taxa in more circum-neutral lakes (Maberly et al. 2009).

The optimal pH range for two Synura species has been demonstrated at between 5.0 and 6.5, after which photosynthetic rate drops off significantly (Bhatti et al. 2008). However, greater than half of our study lakes fall within this optimal pH and, moreover, we see greater declines in colonial chrysophyte species within the optimal pH range than outside it. This suggests that pH- induced declines in CO2 concentration in the water are unlikely to be causing the assemblages shift which we have documented since 1992. Additionally, if CO2 limitation was driving chrysophyte communities in our study lakes, we would expect to see declining chrysophyte populations in favour of other algal groups (e.g., diatoms). Instead, Paterson et al. (2008) has shown that over the past decade, during a time of increasing lakewater pH (Palmer et al. 2011), the biovolume of chrysophytes relative to other algal groups has significantly increased in south- central Ontario lakes. Again, this suggests that the shifts we have documented in chrysophyte species assemblages are not being driven by changes in dissolved inorganic carbon (DIC) speciation, at least at an annual scale.

While any future increases in [DOC] should continue to favour small unicellular chrysophyte taxa, top-down controls may partially mitigate algal response. The anticipated loss of large, high Ca-demand taxa (i.e., Daphniids) from these lakes should favour colonial chrysophytes species that are too large to be consumed by smaller grazers (Sandgren and Walton 1995). For example, Lehman and Sandgren (1985) have shown that large Daphnia pulex can feed on

Uroglena colonies, which are amongst the largest chrysophyte colonies. Over half of our study lakes have lakewater [Ca] of 2.0 to 2.5 mg∙L-1 and, with continued calcium decline, will quickly approach critical biological thresholds for some algal grazers (Jeziorski et al. 2008; Cairns and Yan

2009). How these two opposing stressors (i.e., increased [DOC] vs. decreased [Ca]) will ultimately

affect chrysophyte assemblages remains to be seen, and presents an interesting avenue for future research.

3.7 Conclusions

Our data suggest that complex interactions between climate, pH, DOC and TP may be driving recent changes in scaled chrysophyte species assemblages in the Muskoka-Haliburton lakes. Since sampling in 1992, the Muskoka-Haliburton lakes have been exposed to varying degrees of anthropogenic disturbance. Recovery from long-term regional acidification has resulted in gradual chemical and biological change, but water chemistry data indicate that lakes in this region are still susceptible to short-term re-acidification events (Hadley et al. 2012).

Importantly, despite continued warming (Hadley et al. 2012), the regional-scale increase in the nuisance species Synura petersenii (Paterson et al. 2001) has not continued. We argue that the net result of a myriad of factors influencing lakewater [DOC], both positively and negatively since

1992, has been a regional increase in [DOC] and an ensuing decline in colonial chrysophyte algae.

Regional declines in ionic flux (i.e., SO4 and Ca), a key factor influencing [DOC] are expected to continue (Watmough and Aherne 2008), and may contribute to future increases in lakewater DOC concentration leading to further declines in light penetration. How climate will affect DOC in the future is much more difficult to predict, particularly given that hydrological and temperature- based processes are operating at different scales (Keller et al. 2008).

Although our data illustrate the importance of increased DOC concentration to biological recovery of chrysophyte algae following acidification, our ability to forecast future changes remains hampered. The complexity of the interactions between multiple environmental stressors in these lakes makes them a highly desirable target for continued long-term monitoring, as

changes observed here may be mirrored in other regions recovering from more extreme acid deposition.

3.8 Acknowledgments

The authors would like to thank Jennifer Korosi, Josh Thienpont and Adam Jeziorski for their assistance in the field. This research was made possible by the NSERC discovery grant of JPS.

3.9 Literature cited

Bhatti, S., and Colman, B. 2005. Inorganic carbon acquisition by the chrysophyte alga Mallomonas papillosa. Can. J. Bot. 83: 891-897.

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Figure 3.1. Map of the Muskoka-Haliburton region showing the location of the study lakes (circles) and towns (squares).

Figure 3.2. A) One-to-one plots of selected chrysophyte taxa, showing changes from 1992 to 2007/8, B)pre-industrial to 1992, and C) pre-industrial to 2007. Numbers in brackets indicate the percent contribution of the taxa to the overall dissimilarity between 1992 and 2007 (i.e., SIMPER results). Wilcoxon Signed Rank results comparing 2007, 1992 and pre-1850 data are presented for all 5 taxa. D) Boxplots summarize the overall relative abundance change of selected chrysophyte taxa across all three time periods.

Figure 3.4. Boxplots comparing the decline in colonial chrysophytes across groups of: a) 2007 lakewater pH, and b) % change in pH between 1992 and 2007. ANOVA results testing for significant differences (significant results bolded) amongst the groups are included for each chemical variable, both including and excluding (asterisked) Synura echinulata. Numbers in bold beneath the boxes indicate the number of lakes in each group.

Figure 3.5. Summary of the change in the relative abundance between 1992 and 2007 of the three major colonial chrysophyte taxa in the Muskoka-Haliburton lakes. Lakes are sorted left to right by percent change in colonial chrysophyte relative abundance. Lakes located on the marble intrusion are indicated by open bars.

Chapter 4

Multiple environmental stressors drive extirpation of native crayfish in Algonquin Park, Ontario, Canada

4.1 Abstract

Exposure of freshwater environments to multiple natural and anthropogenic stressors poses a serious threat to the health and diversity of their biological communities. A growing body of research suggests that one of the consequences of these stressors has been a regional-scale loss of crayfish populations across south-central Ontario. Research suggests that mechanisms contributing to the imperilment of crayfish likely include historical acidification and metal contamination, shoreline development and habitat loss, increased water temperature driven by climate change, declining calcium (Ca) concentration in lake water, and non-native species invasion. Long-term monitoring data of crayfish populations in Algonquin Provincial Park (Ontario) has captured marked declines in crayfish abundance despite the relatively isolated nature of the study lakes. Furthermore, monitoring data indicate that the onset of this decline may pre-date the monitoring record creating a need for the use of paleolimnological techniques to infer past conditions. Here, we have used a multiple proxy approach (i.e., diatom, chrysophytes and cladocerans) to reconstruct key environmental variables (i.e., pH) and long-term changes in lake chemistry (i.e., [Ca]) as they pertain to crayfish population loss. We have found that no single stressor can account for the near extirpation of crayfish in these four lakes. The complex interactions between pH, [Ca] and climate warming, together with top-down pressure exerted by predatory fish, led to the decline in crayfish abundance and to the continued lack of recovery of these populations. Our data suggest that given the expected trajectory of stressors such as climate change and calcium decline, recovery of crayfish in Algonquin Park is unlikely.

4.2 Keywords

Crayfish, Algonquin Park, multiple stressors, calcium decline, acidification

4.3 Introduction

Freshwater ecological issues are now, more than ever, being recognized as multiple stressor issues, where a myriad of natural and anthropogenic factors are driving significant declines in water quality, and the loss of species biodiversity. Crayfish represent a vital component of energy flow in aquatic food webs, as they can act as detritivores, herbivores or keystone benthic predators (Momot 1995), and are also a key element of fish diets (Saiki and

Ziebell 1976, Stein 1977). Despite their importance, research on crayfish populations in Canada is scarce.

A recent crayfish survey of 100 lakes in south-central Ontario suggests that crayfish populations across the region are overwhelmingly in a state of decline (Edwards et al. 2009).

Edwards et al. (2009) revisited a suite of lakes and compared modern crayfish populations (2005-

2007) to a historical dataset (1989-1992, 1994, 1995). They found large-scale population declines and extirpations for seven commonly found species of crayfish since 1989 – 1995, with species diversity declines from approximately 2-4 species per lake to ≤1 in the recent survey. The three crayfish species commonly found in Algonquin Park lakes, the focus of our study (i.e., Orconectes virilis, O. propinquus and C. bartonii), declined by 72%, 91% and 91%, respectively, in the Edwards et al. (2009) study.

Recent research suggests that multiple stressors have contributed to the declines in native crayfish populations in Ontario. Mechanisms contributing to the imperilment of crayfish include:

1) historical acidification and metal contamination (Davies, 1989; France 1993), 2) declining

78 calcium (Ca) concentration in lake water (Rukke 2002, Cairns and Yan 2009), 3) increased water temperature driven by climate change (Hammond et al. 2006; Hartman et al. 2010), 4) shoreline development and habitat loss (Wilcove et al. 1998), and 5) non-native species invasions (Olden et al. 2006). The most widely researched of these stressors is lake-water pH, which has been extensively studied over the past several decades. The effects of acidification and recovery in south-central Ontario lakes have been well documented (e.g., Dillon et al. 1987), with reductions in acidifying sulfur compounds since 1970s, leading to chemical and biological improvements

(Dillon et al. 2007; Keller et al. 2007). However, given the rate of decline in sulphate deposition in the region over the past three decades (ca. 40-45%), chemical recovery has been only moderate

(Dillon et al. 2007), and the recovery of biological indicators are still in their earliest stages (Keller et al. 2007).

Reduced lake-water pH may result in a plethora of negative effects on crayfish populations (e.g. DiStefano et al. 1991; France 1993). While field and laboratory studies suggest that short-term acid exposure may not be sufficient to cause crayfish mortality, population declines may result from the sublethal effects of long-term acid exposure, and the loss of buffering capacity (France 1993; Rukke 2002). For example, long-term exposure of crayfish to low pH may result in hatchling mortality, egg loss, and thus recruitment failure (Davies 1989), and has also been linked to higher instances of microsporidian parasitism in Orconectes virilis in some cases (France and Graham 1985). Furthermore, these sublethal effects may, in turn, be exacerbated by other environmental stressors such as calcium decline and climate change (Davies

1989; DiStefano et al. 1991; Hammond et al. 2006; Hartman et al. 2010).

As a long-term consequence of acidification, calcium concentrations in softwater lakes on the boreal shield have declined (Stoddard et al. 1999; Watmough et al. 2003). South-central

Ontario soils are thin and underlain by weather resistant granite, leading to leaching rates of base cations which typically exceed the replenishment rate from weathering and atmospheric deposition (Watmough et al. 2005). Within Algonquin Park this problem is compounded by a legacy of logging, which results in a loss of Ca via forest biomass removal and timber regrowth

(Huntington et al. 2000, Watmough et al. 2003). Additionally, at low pH (5.75), Ca uptake by crayfish may be less efficient, resulting in a reduced Ca content and rigidity of crayfish carapaces

(France 1987; 1993). Crayfish, like other crustaceans, are highly dependent on ambient lakewater calcium concentration particularly during their moulting period. Species specific [Ca] optima and tolerances for organisms are rare within the literature. However, the lowest hypothesized requirement for freshwater crayfish is between 2 and 10 mg∙L-1 (Greenaway 1974, France 1987,

Rukke 2002, Hammond et al. 2006; Cairns and Yan 2009). At ambient Ca concentrations of less than 5 mg∙L-1, Rukke (2002) showed decreased survival and retarded growth in the European crayfish (Astacus astacus). However, these results may not translate readily to our study region as

A. astacus are much larger (up to 16 cm) than the crayfish species commonly found in Ontario

(e.g., Cambarus bartonii; maximum size 3.9 cm).

Recent research has shown that it is possible to use crustacean zooplankton assemblages to indirectly detect calcium decline in lakes (Jeziorski et al. 2008; Jeziorski et al. 2012). The outstanding preservation of cladoceran exoskeletal remains in lake sediments, and their wide dispersal, makes crustacean zooplankton useful indicators for use in paleolimnological investigations (Frey 1960). Specifically, daphniid species have been shown to have significantly higher body calcium concentration than other common cladoceran species (Jeziorski and Yan

2006) and, as a result, are sensitive to reduced lake water [Ca] (Jeziorski et al. 2008). Previous laboratory and field research has identified 1.5 mg∙L-1 as a critical value below which the growth, reproduction and survival of the most common daphniids in south-central Ontario lakes are

80 inhibited (Ashforth and Yan 2008; Cairns 2010). Using paleolimnological techniques to track changes in the relative abundance of these high-Ca demanding taxa downcore can provide an indication of the history of Ca concentration in softwater lakes.

Due to the isolated nature of lakes in Algonquin Park, Ontario (i.e., the focus of our study), stressors such as shoreline development and recreational use are virtually non-existent.

Furthermore, there have been no documented occurrences of the invasive crayfish Orconectes rusticus in the park. Despite their isolation, provincial monitoring data from our four study lakes within Algonquin Park indicate that crayfish populations have undergone significant declines over the past several decades (Girard et al. 2006). Specifically, populations of three native Ontario crayfish species (i.e., Cambarus bartonii, Orconectes propinquus, and Orconectes virilis) show marked population declines beginning ca. 1988. In the absence of local shoreline impacts and species invasions, these declines suggest that regional stressors (i.e., climate change, pH decline and calcium decline) are likely responsible for the fate of crayfish within Algonquin Park. Due to logistical constraints, long-term data from these lakes are relatively sparse, with little to no data available on the pre-industrial limnological properties of these lakes. Furthermore, existing data provided by Ontario Ministry of the Environment (OME) suggest that the decline in crayfish abundance in this region may pre-date the monitoring record, making the historical limnological data provided by paleolimnological proxies vital.

Our goal was to determine the limnological history of four Algonquin Park lakes, as it pertains to declining crayfish populations. Using paleolimnological analysis we reconstructed lakewater pH from sedimentary diatom and scaled-chrysophyte assemblages, and examined long- term changes in biological responses to changing calcium concentrations. Downcore analysis of diatom microfossils was used in conjunction with a regional pH training sets (Hall and Smol 1996)

81 to establish pre-industrial pH, while preserved cladoceran remains were used to attempt to estimate the extent and timing of historical calcium decline. These data, combined with long-term monitoring data provided by the Dorset Environmental Science Centre (DESC), shed new light on the complex nature of the collapse of native crayfish populations in Algonquin Park, and provide valuable information towards future management and recovery efforts.

4.4 Methods

4.4.1 Study area

Dating to 1893, Algonquin Park was the first provincial park established in Ontario. The park is located in central Ontario between the Georgian Bay and the Ottawa River and covers approximately 7653 km2. Industrial logging within the park dates to the early 1800s, and continues in the park’s interior via clear cutting, selection cutting and shelterwood cutting. We focused our research on four small (18 ha to 63 ha) headwater lakes within the

Algonquin/Haliburton Highlands, located on the Precambrian Shield of south-central Ontario

(Figure 1). Historically, these lakes have remained relatively undisturbed compared to similar lakes outside Algonquin Park. However, while these lakes are highly isolated from direct human- impacts via cottages, roads and agriculture within their catchments, atmospheric sulphate deposition in the region has historically been high (0.75 - 1.25 g.m-2 .yr-1; Neary and Dillon 1988).

4.4.2 Water sampling procedures

Water samples were collected in late-autumn (end-Oct to early-Nov) from each lake by technical staff at the Dorset Environmental Science Centre (DESC). Due to logistical constraints, sampling times vary slightly from lake to lake. Cradle and Pincher lakes were both sampled between 1986 and 2009 but are missing data from 1994 to 1999 and 2003 to 2004. Westward

Lake was sampled from 1981 to 2008 and also had missing data between 1993 and 1999, as well as 2005. Finally, Delano Lake was sampled from 1982 to 2008 with the same missing data as

Westward. Analyses of water samples for 14 key variables were performed at Ontario Ministry of the Environment’s DESC laboratory following standard OMOE protocols (Ontario Ministry of the

Environment 1983). Measured variables included pH, specific conductivity (cond), Gran alkalinity, and concentrations of calcium (Ca), dissolved organic carbon (DOC), potassium (K), magnesium

(Mg), ammonia/ammonium (NH3/NH4), nitrate/nitrite (NO3/NO2), total Kjeldahl nitrogen (TKN), pH, total phosphorus (TP), sulphate (SO4), sodium (Na) and silica (SiO2) (Table 1). Physical properties (e.g., depth, surface area) are summarized in Table 1 as are the range of key chemical variables (i.e., pH, Ca and DOC) since monitoring began in the early to mid-1980s.

4.4.3 Crayfish

Crayfish sampling began in 1988 by researchers at the Ontario Ministry of the

Environment’s DESC laboratory. Crayfish were caught in baited traps following procedures outlined in detail by Girard et al. (2006). A total of seven crayfish species have been captured and identified by DESC researchers in south-central Ontario (i.e., Cambarus bartonii, C. robustus,

Orconectes immunis, O. obscurus, O. propinquus, O. rusticus and O. virilis). However, only three of these species have historical populations in the four Algonquin Park lakes considered here (i.e., C. bartonii, O. propinquus and O. virilis; Girard et al. 2006). Of these three species, Cambarus bartonii has the largest populations in our study lakes (based on CUE), but these population have been declining for the duration of the monitoring record (Figure 4.2).

4.4.4 Sediment sampling

Lakes were accessed by float plane and sampled from canoe on October 3rd and 4th, 2008.

A short gravity sediment core was collected from the deepest basin of each lake using a 3″ (7.6 cm) diameter Glew (1989) corer. Sediment cores were 30 cm to 50 cm in length and were sectioned on site using a Glew vertical extruder (Glew 1988) at 0.5 cm resolution. Sediment samples were stored in Whirlpak® bags and kept cool and dark until return to the laboratory.

4.4.5 Diatoms and scaled chrysophytes

For diatom and chrysophyte analysis, 0.2 - 0.3 g of wet sediment was digested in a 50:50

(v/v) mixture of concentrated nitric and sulphuric acid, and rinsed repeatedly with deionized water until a neutral pH was achieved. Diatom slurries were then dried onto coverslips and permanently mounted on microscope slides using Naphrax® mounting medium. Diatom and chrysophyte fossils were enumerated at 1000X under oil immersion using a Leica® DMR2 microscope with differential interference contrast. Diatom and chrysophyte data were collected at each 0.5 cm interval between 0 and 10 cm core depth and at every other interval from 10 - 20 cm. A minimum of between 350 and 400 diatoms valves were counted for each sample and identified using standard taxonomic sources (e.g. Camburn et al. 1984-1986; Krammer and Lange-

Bertalot, 1986-1991; Round et al. 1990; Cumming et al. 1995). Chrysophyte scales were counted to a minimum of 300 scales in each interval in both the Westward and Pincher lake cores.

Chrysophyte scales became sparse in Delano and Cradle lakes below core depths of 14 cm and 15 cm, respectively. For these sparse intervals, a minimum of 100 scales were enumerated in each sample. Scales were identified to the lowest possible taxonomic level using standard sources

(Asmund and Kristiansen 1986; Kling and Kristiansen 1983; Nicholls 1982; Nicholls and Gerrath

1985; Siver 1991) and the image archives of Andrew Paterson (A.M. Paterson, DESC, unpublished data).

4.4.6 Cladocera

Standard sample preparation techniques of Korhola and Ratio (2001) were followed with only minor deviations. Sediment was deflocculated using a 10% KOH solution at approximately 80

°C for 40 minutes. Sediment was then rinsed with deionized water through a 38 mm sieve to remove all undesired particulate matter. Several drops of ethanol and safranin-glycerin were added to prevent fungal growth and stain the cladoceran remains, after which 50 μL aliquots were pipetted onto slides and mounted using glycerin jelly.

Cladocerans were enumerated at 150X using a Leica® DMR2 microscope using bright field optics. Individual cladoceran remains (carapaces, headshields, postabdominal claws) were counted separately and the most abundant segment was used to represent the number of individuals for a given interval (Frey 1986). Counts were performed to a minimum of 75 individuals and were identified using Szeroczyńska and Sarmaja-Korjonen (2007) and Korosi and

Smol (2012a,b)(. Daphnia taxa were split into two species complexes based on the presence/absence of stout middle pecten on the postabdominal claw. D. pulex, D. pulicaria, D. catawba and D. minehaha form the D. pulex complex (stout pecten present), while the D. longispina complex (stout pecten absent) is comprised of D. ambigua, D. mendotae, D. longiremis,

D. dubia, D. parvula and D. retrocurva (Witty 2004; Yan et al. 2008).

4.4.7 Sediment chronology

All four sediment cores were dated using gamma spectrometry at continuous 0.5 cm intervals from 0 - 20 cm (Appleby 2001). 210Pb dates were determined using the constant rate of supply model (CRS), with calculations being performed using the “Binford” package developed by

Adam Jeziorski and Josh Thienpont for the statistical programming language R (Binford 1990;

Jeziorski 2011, Appendix C).

4.4.8 Statistical analysis

Diatom, chrysophyte and cladoceran species assemblage changes through time were summarized using detrended correspondence analysis (DCA) by plotting axis -1 species scores against depth and age. DCA analysis was performed on non-transformed diatom species relative abundances using CANOCO ver. 4.5 (ter Braak and Šmilauer 2002).

The regional diatom-based transfer function developed by Hall and Smol (1996) from 54 nearby lakes in Muskoka-Haliburton was used to infer pH downcore in all four of our study lakes.

The Muskoka-Haliburton lakes pH inference model used weighted-averaging calibration and regression and yielded a strong relationship between measured and diatom-inferred pH values in

2 all four Algonquin Park lakes (r boot= 0.81-0.85, RMSEboot = 0.20-0.23). In addition, to assess the quality of the pH reconstructions, we performed analogue matching on Bray-Curtis dissimilarity metrics, using the “analog” function of the “analogue” package in R v. 2.13.2 (Flower et al. 1997;

Simpson 2007).

4.5 Results

4.5.1 Crayfish, water chemistry and climate data

Water chemistry changes observed in the Algonquin Park lakes were similar to those documented regionally. We recorded gradual declines in sulphate concentrations (SO4) and subsequent increases in pH in all four lakes. All four lakes showed evidence of a re-acidification event, marked by a pH decline of between 0.75 and 1 in 2006, before returning to their previous

86 recovery trajectory in 2008/9. The earliest recorded calcium (Ca) concentrations in the study lakes ranged between 1.5 and 2.5 mg∙L-1, and since 1986 all sites have declined in lakewater [Ca] (Figure

2). Dissolved organic carbon (DOC) concentrations varied across lakes. However, we observed an increase in DOC concentrations in all lakes over the course of the monitoring record (Figure 2).

Three species of crayfish have been identified by the DESC crayfish capture surveys in the

Algonquin Park study lakes. Cambarus bartonii is present historically in all four lakes, while only

Westward and Delano lakes have documented populations of Orconectes propinquus. Orconectes virilis has only been captured in Delano Lake in 1988 (2 individuals in 180 traps) and 1989 (1 individual in 60 traps). Because of the rarity of Orconectes species only C. bartonii is shown

(Figure 2). The decline in abundance of Cambarus bartonii observed in Cradle, Pincher and

Westward lakes may have occurred prior to the monitoring record (ca. 1986). As the catch per unit effort (CUE) of C. bartonii in these lakes was quite low at the beginning of the monitoring record (i.e., 15, 3 and 6, respectively) (Figure 2). In Cradle Lake, C. bartonii declined gradually until

1993. However, when monitoring resumed in 2000, and in all years since, CUE in Cradle Lake has been virtually nil (i.e., 0 - 2 individuals in 54 traps, 2000 - 2009). In Pincher Lake, the decline of C. bartonii occurred by 1990 (CUE <0.5). However, when sampling resumed a small increase was observed between 2001 and 2005, but was subsequently followed by a decline in 2006 (CUE < 0.3 since 2005; Figure 2). As with Pincher Lake, the Westward Lake C. bartonii population declined rapidly from 1988 to 1990, before rebounding slightly between 1996 and 1997 (CUE ~ 2.0), and subsequently crashing again post 1999. C. bartonii populations in Westward Lake remained low

(CUE <0.5) from 1999 until 2009, when there was a CUE over one for the first time since 1998

(Figure 2). Westward Lake, unlike Cradle and Pincher, has a documented historical population of

Orconectes propinquus. O. propinquus populations in Westward have fluctuated historically from a maximum of 2.56 individuals per trap in 1993 to a minimum of 0.20 individuals per trap in 2007.

Unlike the other study lakes, no significant population of C. bartonii has been recorded in Delano

Lake during the 20-year monitoring period (Figure 2). While O. propinquus was captured by DESC researchers, numbers were very low (CUE ~0.5 to 1) for the duration of the Delano Lake monitoring record.

4.5.2 Diatom assemblages and pH inferences

We observed notable changes in diatom species relative abundances in all of our four

Algonquin Park lakes, however the magnitude of these changes was highly variable. In Cradle

Lake, prior to a significant ecological shift ca. 1950, the diatom species assemblage was dominated by the oligotrophic, planktonic taxon Cyclotella stelligera, which accounted for approximately 40-

55% of the diatom relative abundance (Figure 3a). A sharp increase in acidophilic taxon,

Asterionella ralfsii var. americana, began ca. 1950 and increased steadily to a maximum relative abundance of 60% ca. 1990 (Figure 3a). An analogous increase occurred in the early acidification indicator Tabellaria flocculosa str. IIIp, which reached its maximum relative abundance in ca. 1995.

A. ralfsii var. americana underwent a marked decline in relative abundance immediately following the 1990 peak, but increased in the youngest sediment intervals (Figure 4a). Subsequent to reaching a maximum ca. 1995, T. flocculosa str. IIIp declined gradually in relative abundance to

20% in the modern (ca. 2006) sediments, where we also observed a small increase in the relative abundance of C. stelligera (Figure 3a). Increased relative abundance of A. ralfsii v. americana and

T. flocculosa str. IIIp closely coincides with decreased diatom-inferred pH values in Cradle Lake

(Figure 3a). Based on analogue matching, 32 of 35 intervals in Cradle Lake had close modern analogues within the modern training set.

Historical diatom assemblages in Delano Lake prior to ca. 1940 were dominated (15-35%) by small, benthic Fragilaria pinnata and large, heavily silicified tychoplanktonic Aulacoseira taxa

(Figure 3b). Between 1940 and 1985, we observed the first increase in circumneutral, planktonic taxon Cyclotella stelligera and mesotrophic, planktonic taxon Asterionella formosa, with gradual declines in the relative abundances of F. pinnata and Aulacoseira spp. After 1985, Cyclotella stelligera abundance continued to increase to a maximum of ~25%, while F. pinnata abundance declined to a historical low of <10%. In the most recent sediments, we noted slight declines in the relative abundance of C. stelligera and increases in Aulacoseira spp. (Figure 3b). In Delano Lake, diatom-inferred pH remained relatively stable at ~6.75 prior to ~1970 after which we recorded a gradual decline in pH to ~6.5 (Figure 3b). However it is worth noting that only 1 of 35 sediment intervals from Delano Lake had a close modern analog in the training set.

The pre-industrial diatom assemblage in Pincher Lake was comprised of a variety of tychoplanktonic Aulacoseira spp. (i.e., A. ambigua, A. alpegina, A. distans and A. perglabera), as well as acidophilic taxa Asterionella ralfsii var. americana and Tabellaria flocculosa (Figure 4c).

Relative abundances of A. ralfsii var. americana and T. flocculosa increased slightly between ~1915 and ~1953 at the expense of Aulacoseira species. Following this increase, A. ralfsii var. americana decreased gradually to a relative abundance of less than 2 % in modern sediments, while T. flocculosa decreased to ~2% before an abrupt increase to 10% relative abundance in the mid-

2000s (Figure 3c). Diatom-inferred pH suggests that pH in Pincher Lake has increased since the mid-1990s from ~5.9 to ~6.5 (Figure 3c). Prior to the mid-1990s the pH in Pincher Lake, while stable, was relatively lower than what was observed in the other Algonquin Park Lakes (Figure 3c).

However, Pincher Lake had no close analogues with the modern training set and therefore the pH inferences should be interpreted with caution.

Diatom species assemblages in Westward Lake have been dominated by the Cyclotella stelligera complex (60 - 85% relative abundance) for the duration of the sediment core (Figure 3d).

An increase in Tabellaria flocculosa str. IV from historically negligible numbers to ~10% relative abundance was observed beginning in the late-1960s and has continued into the most recent sediments (Figure 3d). Diatom-inferred pH in Westward Lake remained relatively stable (ph ~6.5) throughout the duration of the core (Figure 3d). Based on analogue matching, all (35 of 35) of the sediment intervals in Westward Lake had close modern analogues within the training set.

4.5.3 Scaled chrysophytes

Pre-disturbance chrysophyte scale assemblages in Cradle Lake were dominated by

Mallomonas duerrschmidtae and Mallomonas lychenensis. A major shift occurred ca. 1933, with a marked increase in the colonial taxon, Synura petersenii (Figure 4a), from <10 to 20% relative abundance by the mid 1980s, before declining sharply to <10%. Following this shift, S. petersenii rose sharply again to a maximum of 40% in the most recent sediment interval (Figure 4a).

In Delano Lake, Mallomonas crassisquama and Mallomonas duerrschmidtiae dominated the pre-industrial chrysophyte species assemblages accounting for ~70% of the assemblage until

~1925. Post-1925, there was a gradual rise in the relative abundances of Synura echinulata, S. petersenii and S. sphagnicola from <2% to ~20% (Figure 4b).

The scaled chrysophyte assemblage in Pincher Lake was dominated by M. crassisquama and M. duerrschmidtiae, accounting for 70-80% of the species assemblage prior to the mid-1980s

(Figure 4c). After ~1985, there was a gradual decline in the relative abundances of M. crassisquama and M. duerrschmidtae, and subsequent increases in M. galeiformis, M. tonsurata/alpina, M. hamata and S. echinulata.

Prior to ~1940, scaled-chrysophyte assemblages in Westward Lake were dominated by

Mallomonas duerrschmidtiae, Mallomonas crassisquama and Synura uvella (Figure 4d). Between

~1940 and ~1983, Synura uvella disappeared from the sediment record, while we documented an increase in M. duerrschmidtiae, and increases followed by equivalent decreases in M. caudata and

M. hamata (Figure 4d). Between 1983 and 1996, there was a marked decrease in M. duerrschmidtiae and short-term increases in M. caudata and M. lychenensis. Post-1996 we see an increase in Synura curtispina, to 40% relative abundance ca. 2001, gradually declining to 20% in the most recent sediment interval.

4.5.4 Cladocerans

Prior to ~1980 several cycles in cladoceran assemblage composition occurred in Cradle

Lake, where increases in the relative abundance of Bosmina spp. came at the expense of D. longispina and vice versa (Figure 5a). The D. longispina complex increased from ~15% to 50% relative abundance between ~1885 and ~1920, while Bosmina spp. declined in relative abundance. This trend was reversed from ~1920 to ~2007 where Bosmina relative abundance increased to 60%, while D. longispina declined to <5%. An abrupt increase in the relative abundance of the D. pulex complex from 5% to >20% was observed post-1980.

In Delano Lake, the relative abundance of Bosmina spp. increased gradually from historical relative abundances of 40% to ~75% beginning in ~1960 (Figure 5b). Daphnia spp. declined during this period from ~25 - 30% to <10% (D. pulex complex), and >20% to ~10% (D. longispina complex) relative abundance, respectively.

Bosmina spp. dominated the Pincher Lake cladoceran assemblage throughout the duration of the core (Figure 5c). After ~1987, both the Daphnia longispina and D. pulex complexes were present. However, since ~1987 the relative abundances of daphniid species in Pincher Lake have been substantially lower (i.e., less than ~5%, Figure 5c), while Bosmina spp. increased from historical relative abundances of 60% to >90%.

In the pre-industrial period of Westward Lake, relative abundances of the dominant cladoceran species were relatively stable (i.e., 60% Bosmina spp., 20% Daphnia spp. and 20% Sida spp.; Figure 5d). Beginning ca. 1940, a gradual increase in Bosmina spp. occurred from ~60 to

90%, with a subsequent decline of daphniids and Sida spp., both of which recovered beginning ca.

1984 (Figure 5d).

Species turnover of diatoms, scaled chrysophytes and cladocerans were summarized using

DCA axis-1 scores for all four lakes (Figure 6). Species turnover in Cradle Lake occurred in all three biological indicators beginning in the early to mid-1950s, while in the other three lakes the timing of species turnover varied (Figure 6a). Biological change in all three indicators occurred much more gradually in Delano Lake than what was observed in Cradle Lake (Figure 6b). In Pincher Lake diatom and chrysophyte species change occurred later than in Cradle (ca. 2000), while cladocerans changed much more gradually beginning in the mid-1980s (Figure 6c). In contrast to Cradle,

Pincher and Delano lakes, very little species turnover occurred in the diatoms of Westward Lake, whereas the cladocerans in Westward Lake present the only clear return to pre-disturbance species composition (Figure 6d).

4.6 Discussion

Crayfish capture data collected by the OME documented historical crayfish populations in three of the four study lakes, and, while their historical populations differ, all have undergone a marked decline in catch per unit effort in recent decades. Both Cradle and Pincher lakes have supported populations of C. bartonii in the past, while Westward Lake capture data suggest historical populations of both C. bartonii and O. propinquus. Negligible crayfish populations documented in the monitoring data for Delano Lake may suggest that crayfish decline in this lake occurred prior to the monitoring record which began in 1988. This is supported by the

92 paleolimnological data which suggests multiple environmental stressors have been acting on these lakes over the past several decades (Figure 3, 4, 5). As discussed below, Cradle and Pincher lakes appear to be the most acid-sensitive of the study lakes (Table 1). Thus, the presence of only

C. bartonii is not surprising, as previous research has shown this species to be more tolerant of low pH water than Orconectes species (DiStefano et al. 1991).

4.6.1 Lake acidification and its possible role in crayfish decline

Diatom data collected from Cradle Lake shows a marked period of acidification beginning ca. 1954, which is captured in the DCA axis 1 scores (Figure 6) and punctuated by increases in acidophilic taxa A. ralfsii var. americana and T. flocculosa (Figure 4a). By the mid-1970s the diatom-inferred pH, for which there are strong modern analogues, dropped to below 6.0 for the first time in the lakes history, reaching a minimum of 5.75 in the late 1970s (Figure 4a).

Monitoring and paleolimnological data indicate that declining sulphate input (e.g., Dillon et al.

1997) has resulted in chemical recovery (pH increased to ~6.25) in Cradle Lake. A small increase in

DCA axis 1 scores in the most recent sediments (Figure 6) driven by an increased relative abundance of circumneutral taxon, C. stelligera, may be an indication of the beginning of biological recovery in the diatom communities (Figure 3a). However, given the magnitude of the reduction in sulphate concentration, the biological response remains relatively small.

While no strong acidification signal is obvious in the proxy reconstruction, both chrysophyte and diatom species assemblages suggest Pincher Lake may have been historically more acidic than the other three Algonquin Park lakes. Pincher Lake had no close analogues throughout the duration of this core, likely as a result of high relative abundances of Aulacoseira taxa (e.g., Aulacoseira alpegina) that were either absent in the training set or present in much lower abundance, and therefore the pH inferences should be interpreted with caution. The

93 presence of acidophilic taxon, Asterionella ralfsii var. americana, in pre-industrial sediment intervals suggests that pH was near or below 6 throughout the lakes history (Figure 3c).

Furthermore, while gradual increases in A. ralfsii var. americana and Tabellaria flocculosa between 1915 and ca. 1955 suggest a period of modest acidification in Pincher Lake, these increases do not result in a significant change in diatom-inferred pH data, nor are there any concurrent changes in our chrysophyte data. It is therefore not surprising that, as with Cradle

Lake, only C. bartonii are found throughout the monitoring record as Cambarus species been shown to be more acid tolerant than Orconectes species (Berrill et al. 1985).

Similar to Pincher Lake, the biological responses to acid deposition in Westward and

Delano lakes were more tempered than those observed in Cradle Lake. Increased relative abundance of the diatom taxon Tabellaria flocculosa str. IV (Figure3) and decreased DCA axis-1 scores (Figure 6) in Westward Lake beginning in the late-1940s may be indicative of mild acidification (e.g., Camburn and Charles 2000; Ginn et al. 2007). Furthermore, the disappearance of the circumneutral/alkaline scaled-chrysophyte taxon Synura uvella (Siver 1995; Paterson et al.

2001) at 11cm (ca. 1940) supports the assertion of mild acidification in Westward Lake (Figure 4d).

However, due to the overwhelming dominance (60 - 85%) of a single, circumneutral taxon (C. stelligera, pH optimum = 6.6), our diatom-inferred reconstruction does not capture any notable decrease in pH.

Chrysophyte proxy data in Delano Lake may also suggest that a period of mild acidification may have occurred. Increases in the relative abundance of acidophilic Synura taxa

(i.e., Synura echinulata, Synura sphagnicola), such as those we observe beginning ca. 1925, have been previously associated with early acidification (Charles et al. 1990; Siver 1995). However, a bimodal pH distribution of S. echinulata has been identified in previous studies (Cumming et al.

1992; Dixit et al. 1999; Paterson et al. 2001). It is also important to note that increases in the abundance of Synura scales in Ontario have also been associated with altered thermal regimes, as a result of climate change (Paterson et al. 2004) and changes in dissolved organic carbon (Hadley et al. 2012). Therefore, it is possible that increases in Synura species in Delano Lake are the result of both an acidification/recovery and a climate signal. While some acidification may have occurred, diatom-inferred pH in Delano Lake is consistently above 6.5, and previous research

(Smol 1980; Smol and Dickman 1981) also suggests the extent of acidification in Delano Lake was relatively minor.

Our paleolimnological data suggest that while varying degrees of acidification have occurred in the study lakes, Cradle Lake has undergone the most dramatic changes in pH and crayfish abundance. The reduction in lakewater pH in Cradle Lake from 6.5 to 5.75 by the late-

1990s coincided with the extirpation (i.e., CUE of 15 to 0) of the crayfish C. bartonii. Laboratory and field studies have previously identified a pH of 5.75 as a potentially critical threshold for some crayfish species, below which Ca uptake is limited thereby reducing Ca content and rigidity (i.e.,

25-35% reduction in calcification and hardness) of the carapace (France 1987, 1993). While no specific data could be found for C. bartonii, exposure to pH between 5.4 and 6.1 was toxic to juvenile stages of O. propinquus (Berrill et al. 1985), and other detrimental effects are expected with reduced pH, including hindered reproductive success, increased disease susceptibility, decreased feeding activity, and impaired chemoreception of food (Tierney and Atema 1988;

DiStefano et al. 1991; Uiska et al. 1994). The sensitivity of crayfish to acidification has been further documented in Lake 223 of the Experimental Lakes Area (North-western Ontario), where

O. virilis population declined from over 100,000 animals to extinction when pH was reduced from

6.8 to 5.1 (Davies 1989).

4.6.2 The possible role of calcium decline in the crayfish decline

Proxy-based pH reconstructions suggest that acid deposition alone cannot account for the decline in crayfish in Algonquin Park, as neither Westward nor Delano Lake show clear signs of pH decline below 6 throughout their history, yet a similar decline in C. bartonii abundance occurred in

Westward Lake in the mid-1990s (Figure 2). Furthermore, since 2000, the increases in pH we have documented in the Algonquin Park lakes monitoring record have not been met with any sustained increase in crayfish abundance (Figure 2). This is in contrast to observed recovery in other regions. For example, following the extirpation of crayfish in Lake 223 at the Experimental Lakes

Area in northern Ontario, crayfish populations showed signs of recovery when pH increased to approximately 5.8 (K. Mills, Freshwater Institute, unpublished data; France, 1993). The apparent lack of recovery with increased pH in our lakes suggests that additional stressors, such as calcium decline, are inhibiting the reestablishment of crayfish populations.

Although, it is not possible to directly reconstruct [Ca] using paleoecological indicators, both proxy evidence and monitoring data suggest substantial Ca decline has occurred in the

Algonquin Park lakes over the past several decades. Despite relatively stable pH, Ca-rich daphniids in Delano and Pincher Lake began to disappear from our cores as early as ~1960, replaced by smaller, less Ca-demanding Bosmina species (Figure 5b, 5c). Similarly, the relative abundance of

Daphnia longispina, a taxon requiring relatively high lakewater Ca levels, falls from 20% to <5% in

Cradle Lake, beginning in the early 2000s (Figure 5a), coinciding with the first instance of [Ca] <1.5 mg∙L-1 in the monitoring record. A conspicuous increase in relative abundance of the Daphnia pulex complex in the recent sediments of Cradle Lake (Figure 5a) may at first seem counterintuitive to the Ca decline scenario. However, zooplankton monitoring data shows that D. catawba, a member of the D. pulex complex which has been shown to be tolerant of low pH

(Malley and Chang 1986; Locke 1991) and low [Ca] (Jeziorski et al. 2012), accounts for this increase

(M. Palmer, Ontario Ministry of the Environment, unpublished data).

Lakewater [Ca] in all four Algonquin Park Lakes has now fallen below 2.0 mg∙L-1, the lowest hypothesized requirement for freshwater crayfish (2-10 mg∙L-1; Greenaway 1974; France 1987;

Hammond et al. 2006). Crayfish, like other crustaceans, are highly dependent on ambient lakewater calcium concentration. In their first year, young crayfish will moult as many as ten times, during which they are vulnerable to both injury and predation, making rapid calcification of the exoskeleton essential. Crayfish store only 10-20% of the required calcium required to harden a new exoskeleton in the gastrolith (Malley 1980; Lowery 1988; Taugbøl et al. 1996) but lose as much as 50% of the total body calcium in the shed exuviae (Wheatly and Ayers 1995). These exuviae may be consumed as a source of Ca, but evidence suggests they are often eaten by postmoult or intermoult individuals with fully-calcified mandibles, increasing the importance of lakewater [Ca] to newly moulted individuals (Greenaway 1985). While specific calcium optima and tolerances are not available for native Ontario crayfish, Hammond et al. (2006) have documented significantly decreased survival rate of crayfish Paranephrops zealandicus in low calcium (<10 mg∙L-1) environments. Similarly, experimental studies of Astacus astacus revealed both retarded growth and decreased survival at Ca concentrations below 5 mg∙L-1 (Rukke, 2002).

Furthermore, the impacts of low lakewater calcium concentration are likely exacerbated when pH decreases (France 1987, 1993). While pH has been recovering in the Algonquin Park lakes, calcium decline is expected to continue (Watmough et al. 2005; Watmough and Aherne 2008).

Although we cannot estimate a calcium threshold for crayfish survival in Algonquin Park, the growing body of evidence suggests that pH recovery, without significant increases in [Ca], will not result in crayfish recovery.

4.6.3 Direct and indirect effects of warming temperatures on crayfish populations

Regional long-term climate data show a steady increase in mean annual maximum temperature over the past several decades in south-central Ontario (Hadley et al. 2012), and an overall decrease in wind (H. Yao, Ontario Ministry of the Environment, unpublished data).

Warmer air temperatures as a result of climate warming lead to longer ice-free periods in south- central Ontario lakes (e.g. Futter 2003), longer growing seasons, and enhanced thermal stratification (Stainsby et al. 2011). In oligotrophic lakes, such as Delano, increased thermal stability in the water column often favours small, planktonic taxa such as C. stelligera, typically at the expense of more heavily silicified Aulacoseira taxa and small benthic Fragilaria taxa (Rühland et al. 2008). Diatom-based climate signals are typically masked in lakes which have acidified (e.g.,

Cradle Lake), as diatoms respond more sensitively to chemical changes (e.g., pH) (Smol and

Stoermer 2010). Therefore, it is not surprising we have not seen similar taxonomic changes in our other study lakes in response to regional climate change.

Data on the impact of climate warming on crayfish species are sparse, but suggest that both direct and indirect effects may result. For example, Cambarus bartonii have been shown to be particularly sensitive to increased temperature, which influences their reproductive cycle, growth, and behavior (Somers and Green 1993). Increased water temperature may exacerbate the effects of other environmental stressors, such as reduce tolerance to low pH (DiStefano et al.

1991), and increase susceptibility to disease (Jiravanichpaisal et al. 2004). Hammond et al. (2006) also found significantly decreased survival of the freshwater crayfish Paranephrops zealandicus as water temperatures exceeded 16 °C. As with other stressors (e.g., pH), increased water temperature appears to have a significantly greater negative impact on survival of juvenile crayfish (Hartman et al. 2010).

4.6.4 Interactive effects of multiple stressors and fish predation

While only limited (presence/absence) fish data are available for our study lakes, the importance of top-down pressures on crayfish populations cannot be ignored, particularly given the complex interactions between crayfish health and lake chemistry in multiple stressor systems.

Crayfish are most susceptible to fish predation immediately following ecdysis when their carapace is soft (e.g., Scott and Duncan 1967). As discussed above, evidence suggests the interactive effects of low pH and low [Ca] weaken the exoskeleton or prevent complete calcification, potentially exacerbating this vulnerability. Normally, to counter predation, crayfish try to hide during their postmoult calcification (Stein 1977). However, it has been noted that this behaviour is absent when carapace rigidity is reduced by acidification (France 1987; Davies 1989). Quantifying the effect of top-down pressures on crayfish populations in Algonquin Park is difficult as all of our lakes have historical populations of predatory fish; however data from the Ministry of Natural

Resources (MNR) shows no evidence of introduction of fish species in these lakes which could explain the decline in crayfish populations. For example, there have been no introductions of smallmouth bass in Algonquin Park lakes over the past 30 years (Brad Steinberg, Ministry of

Natural Resources, pers. comm.), and historical data indicate that the only sanctioned fish introductions in our lakes were Arctic char, alpine char and cherry salmon in Westward Lake, all in the 1960’s. Cherry salmon and alpine char stocking were both unsuccessful, however a naturally reproducing Arctic charr population was established (Mandrak and Crossman 2003; Kerr 2006).

Because the fish communities found in northern waters are relatively simple, few interactions have been noted between Arctic charr and other fish species, or crayfish, making it difficult to assess the potential impact of this introduction. While we do document a marked decline in Sida spp. in Westward Lake, a cladoceran which has been previously observed to favour conditions with very low planktivory (N. Yan, DESC, pers. comm.), the available data are not sufficient to

99 equate predation changes with crayfish extirpation. Therefore, while it is likely that fish predation has played some role in the decline in crayfish in our lakes, particularly where multiple stressors have left crayfish vulnerable, the available data do not suggest that changes in fish communities alone can account for the extirpation of crayfish in Algonquin Park.

4.7 Conclusions

Significant regional declines of crayfish have been documented throughout south-central

Ontario and are believed to be the result of anthropogenic induced stress (Edwards et al. 2009).

Long-term monitoring data collected by the DESC recorded similar crayfish declines in Algonquin

Park, but significant gaps exist in the monitoring record and it failed to capture the onset of crayfish decline, creating an opportunity for paleolimnological analysis. Here, we have used detailed downcore analysis of multiple paleolimnological proxies to establish the limnological history of four remote lakes in Algonquin Park. The remoteness of these lakes allowed us to ignore direct human impacts (e.g., land-use change and crayfish habitat destruction) and focus on indirect disturbances. Our data suggest that no single stressor can account for the near extirpation of crayfish in these four lakes. For example, while it appears that pH decline resulting from acid deposition has played a key role in the decline of crayfish in Cradle Lake, subsequent chemical recovery has not been met with increased crayfish abundance. Furthermore pH decline is not consistent across our study sites. Both Delano and Westward Lake show little evidence of marked pH change, suggesting one or more other key stressors are at play.

Using cladoceran zooplankton remains, we have documented biological response to calcium decline as early as the 1960s. Both long-term monitoring and paleolimnological data indicate lakewater [Ca] in all four Algonquin Park lakes has fallen below 2.0 mg∙L-1, the lowest hypothesized requirement for freshwater crayfish. The detrimental effects of low lakewater

100 calcium concentration on crayfish are likely to be exacerbated at low pH and the sensitivity of crayfish to both low pH and [Ca] is heightened with increasing temperature. While the complex interactive effects of multiple stressors make it impossible to quantify specific management goals, our data suggest that pH recovery alone cannot and will not result in biological recovery of native crayfish populations. As calcium decline and climate warming are expected to continue, unassisted recovery of crayfish populations in these lakes seems unlikely. Our data emphasizes the dire need for continued monitoring of crayfish populations throughout Ontario, particularly in calcium depleted shield lakes.

4.8 Acknowledgments

The authors wish to thank the staff at Dorset Environmental Sciences Centre for logistical support.

Specifically, we thank, Crystal Hyatt and Anna DeSellas for assistance in the field. This research was made possible by an NSERC Discovery Grant to JPS.

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Somers, K.M., and Green, R.H. 1993. Seasonal patterns in trap catches of the crayfish Cambarus bartonii and Orconectes virilis in six south-central Ontario lakes. Can. J. Zool. 71: 1136-1145.

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Figure 4.1. Map of central Ontario showing the approximate location of Algonquin Park and the location of the study lakes (closed circles) and towns (open circles).

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Figure 4.2. Summary of monitoring data of Cradle, Delano, Pincher and Westward lakes. Top: Catch per unit effort data of crayfish species Cambarus bartonii. Bottom: Boxplots summarizing changes in key water chemistry variables during the monitoring period.

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Figure 4.3. Relative abundance of key diatom taxa and downcore diatom-inferred pH in a) Cradle Lake, b) Delano Lake, c) Pincher Lake and d) Westward Lake, plotted against core depth (cm) and 210Pb age.

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Figure 4.4. Relative abundance of key scaled chrysophyte taxa in a) Cradle Lake, b) Delano Lake, c) Pincher Lake and d) Westward Lake.

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Figure 4.5. . Summary of the relative abundance of key cladoceran taxa in a) Cradle Lake, b) Delano Lake, c) Pincher Lake and d) Westward Lake.

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Figure 4.6. DCA axis-1 scores plotted against core depth and 210Pb age in a) Cradle Lake, b) Delano Lake, c) Pincher Lake and d) Westward Lake, showing species turnover.

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Table 4.1. Summary of the key physical and chemical limnological variables from the long-term monitoring record of our four Algonquin Park study lakes.

Cradle Lake Delano Lake Pincher Lake Westward Lake Lake Area (ha) 17.89 23.9 42.06 63 Lake Volume (m3 x 105) 22.25 17.04 25.48 129.5 Mean Depth (m) 12.44 7.13 6.06 20.54 Max Depth (m) 33.3 18.6 15.5 44 Shoreline Length (km) 2.44 1.99 5.52 3.52 Acid Sensitivity 2 3 2 3 Brook Trout, Lake Trout, Lake Trout, White Sucker, Potential crayfish predators Splake, Smelt White Sucker White Sucker Creek Chub pH range 5.3 - 6.2 5.37 - 6.46 5.11 - 5.86 5.53 - 6.77 Ca range (mg/L) 1.12 - 2.00 1.78 - 3.08 0.90 - 1.62 1.58 - 2.40 DOC range (mg/L) 1.0 - 2.8 3.8 - 7.2 2.0 - 3.8 1.4 - 2.5

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Chapter 5

General Conclusions

Environmental issues, now more than ever, are being recognized within the context of multiple stressors, often involving complex interactions that have resulted in unanticipated ecological consequences (Christensen et al. 2006). Long-term monitoring projects, such as those employed by the Dorset Environmental Science Centre (DESC), are vital to our understanding of changes occurring in terrestrial and aquatic ecosystems in response to environmental pressures.

However, financial and logistical restrictions make relying strictly on monitoring data unfeasible.

By using paleolimnological techniques we can extend the sampling window and footprint spatially and temporally, and in concert with long-term monitoring data, modeling and experimental studies, provide the best possible estimate of ecosystem change and the most likely causes for such change. The overarching goal of this thesis was to use both paleolimnology and monitoring data to assess how the impacts and interactions of multiple environmental stressors have shaped the limnology of south-central Ontario lakes over the past several decades. What we found was a myriad of human-induced chemical and biological changes, which, in some cases, has created environments for which we have no historical analogs. Although the biological responses to recent change have varied, a few fundamental generalizations about regional limnology can be made.

5.1 Biological changes over decades reflect interactions of multiple anthropogenic stressors

Each chapter of this thesis, in one way or another, illustrates the importance of interactions between multiple environmental stressors to lake management in south-central

Ontario. In Chapter 2, the warming climatic trend during the past 15 years (Futter 2003; Chapter

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2, Figure 2.2) has coincided with widespread increases in Cyclotella species, particularly in the oligotrophic lakes in the region. A longer duration of the ice-free season, longer diatom growing season and enhanced lake stratification are important factors driving this change; however, while we have observed a marked diatom response to climate in oligotrophic lakes, Cyclotella species were less pronounced or absent from lakes that were more nutrient-rich. Instead, in our more mesotrophic study lakes, we saw increases in the relative abundance of Asterionella formosa and

Fragilaria crotonensis. We believe these planktonic species are also tracking recent climate warming, but are more common than small Cyclotella in mesotrophic to nutrient-rich waters.

As with diatoms, marked regional changes were also observed in the assemblages of scaled chrysophyte algae in south-central Ontario lakes. Our data suggest that, despite continued warming, the regional-scale increases in nuisance colonial chrysophyte species have slowed and in many cases begun to reverse. When DOC concentrations were lowered because of acid deposition, highly motile colonial chrysophyte species (i.e.,Synura spp.),which form deepwater populations, had a selective advantage over unicellular epilimnetic forms (i.e., Mallomonas spp.).

Since their original sampling in 1992, in addition to warming climate, the Muskoka-Haliburton lakes have experienced marked chemical change due to reduced acid deposition and associated loss of base cations (i.e., calcium). The net result of changes in climate, acid deposition and hydrology have been a near systemic regional increase in [DOC] which we believe, coupled with increases in TP, is responsible for the widespread declines in colonial chrysophytes documented here. The complexity of the interactions between natural (i.e., wetland coverage, local geology) and anthropogenically-induced changes (i.e., warming temperature and decreased precipitation), operating at varying scales, have created variable responses amongst lakes. Thus, these lakes are highly desirable targets for continued long-term monitoring, particularly as changes observed here may be mirrored in other regions recovering from acid deposition.

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Our investigation into the decline of crayfish populations in Algonquin Park (Chapter 4) suggests that no single limnological change is consistent in the study lakes that can account for the near extirpation of crayfish, and that the significant regional declines of crayfish, documented throughout south-central Ontario, are likely the result of interactions between multiple natural and anthropogenic stressors. The remoteness of these lakes allowed us to largely ignore many direct human impacts (e.g., land-use change and crayfish habitat destruction) and focus on indirect disturbances. Our data suggest that while lake acidification resulting from atmospheric acid deposition played a key role in the decline of crayfish, the degree of disturbance was not consistent across our study sites, and subsequent chemical recovery of pH has not been met with increased crayfish abundance. This implies that other regional stressors such as climate change and Ca decline also played an important role in crayfish decline, with significant implications for recovery. We documented zooplankton response (loss of Ca sensitive cladoceran species) to calcium decline as early as the 1960s. Both long-term monitoring and paleolimnological data indicate lakewater [Ca] in all four Algonquin Park lakes has fallen below 2.0 mg·L-1, the lowest hypothesized requirement for freshwater crayfish (Cairns and Yan 2009), the detrimental effects of which are likely to be exacerbated at low pH. Furthermore, the sensitivity of crayfish to both low pH and [Ca] is heightened with increasing temperature (Hammond et al. 2006). In addition to highlighting the complexity of the interactions between stressors in these Algonquin Park lakes, our data emphasize both the need for a better understanding of the biological thresholds of

Ontario crayfish and continued monitoring of crayfish populations throughout Ontario, particularly in calcium depleted shield lakes.

5.2 Potential for ecological recovery and management in south-central Ontario lakes

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In addition to highlighting the interactions between multiple stressors, this research provides some insight as to the potential for ecological recovery in lakes in south-central Ontario.

Widespread increases in planktonic algal communities, documented in Chapter 2, suggest a continuing biological impact from climate warming in south-central Ontario lakes. Given that the warming trend we have seen over the past several decades is expected to continue and intensify, the likelihood that biological change will continue in temperate lakes is high. Without addressing large-scale regional stressors such as climate change, future biological communities will remain a moving target, making lake management difficult and biological recovery, impossible.

In Chapter 3 we documented marked decreases in colonial chrysophyte algae, which we attributed, in part, to systemic increases in [DOC] in the majority of our study lakes. As regional declines in ionic flux (i.e., SO4 and Ca), a key factor influencing [DOC], are expected to continue

(Watmough and Aherne 2008), future increases in lakewater DOC concentration leading to further declines light penetration are likely. However, while increased [DOC] is thought to be a consequence of acid recovery, the potential exists that other mechanisms (e.g., climate, hydrology) may also be contributing to these increases (e.g., Eimers et al. 2008). Additionally, how climate will affect DOC in the future is difficult to predict, particularly given that hydrological and temperature-based processes are operating at different temporal scales (Keller et al. 2008).

Therefore, although our data illustrate the importance of increased DOC concentration to biological recovery of chrysophyte algae following acidification, our ability to forecast future changes remains hampered. Given the number of lakes still recovering from varying degrees of acid deposition, in south-central Ontario and beyond, improving our understanding of the interactions of climate, pH and DOC is crucial to future management efforts.

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In Chapter 4, we documented limnological change resulting from multiple environmental stressors which may have led to the near extirpation of native crayfish in four Algonquin Park lakes. Although the complex interactive effects of multiple stressors made it difficult to quantify specific management goals (i.e., a critical lakewater [Ca] threshold), our data suggest that pH recovery alone cannot and will not result in biological recovery of native crayfish populations. As calcium decline and climate warming are expected to continue, unassisted recovery of crayfish populations in these lakes seems unlikely. This raises several interesting questions for lake and resource management within the park. For example: How do we weight the importance of crayfish against the potential negatives of a major management intervention? Assuming pH and

Ca concentrations recover to pre-disturbance levels, can native crayfish survive in warming

Ontario lakes?

5.3 How does lake selection and temporal scale shape limnological interpretations?

Because we were revisiting lakes previously sampled and studied by other researchers, site selection was not a major consideration when beginning this research. However, our results illustrate the importance of lake selection and temporal scale in shaping data interpretation in this region. For example, in Chapter 2 we discussed our water chemistry observations in the context of a recent survey done by Palmer et al. (2011) and found that, because of differences in the timing of sample collection, important water chemistry trends varied somewhat. Specifically,

Palmer et al. (2011) found increased pH and decreased TP overall between 1981-1990 and

2004/5, while our study, based on recent spring chemistry found stable or increased TP in many cases, and a slight decrease in pH. Furthermore in Chapter 3, we discussed how catchment wetland cover, which varied greatly across the Muskoka-Haliburton region, can have marked effects on pH, [DOC] and climate, thereby drastically altering algal communities. The Muskoka-

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Haliburton marble intrusion lakes also demonstrate the importance of lake selection. As detailed in Chapter 3, this small subset of 12 lakes differed chemically (e.g., high Ca) from our other study lakes and therefore colonial chrysophyte algae remained relatively stable or, in many cases, increased in these sites. This type of variability in our study sites, while confounding at times, is valuable when attempting to tease apart the impacts of various stressors, and can allow us a better understanding of our data.

Despite their obvious usefulness, long-term monitoring data collection began after the initial decline in crayfish populations. By extending the temporal scale of the data back through time using paleolimnology, we were able to better establish potential mechanisms that may be partly responsible for these extirpations (Chapter 4). For example, based on the timing of acidification in Cradle Lake, it seems likely that acid stress on crayfish may have begun as early as the 1970s and therefore the CUE of Cambarus bartonii may have been considerably higher during pre-disturbance times than the monitoring data suggest. Several key questions in lake management and recovery arise from these observations, such as: What were crayfish CUEs in

Algonquin Park before acid deposition and Ca decline? What should be our realistic recovery goal?

Can we establish one that is evidence-based? Is unassisted recovery even possible?

While the use of paleolimnology allowed us to extend the monitoring record back through time, the top-bottom approach does have limitations. The most obvious limitation of the top- bottom approach is the lack of temporal resolution. Where a 210Pb-dated downcore analysis would provide us with a more comprehensive view of change through time, the top-bottom approach provides only a few brief snapshots. However, the top-bottom approach allows a greatly expanded spatial scale (i.e., dozens or hundreds of lakes), and thereby allows for a better regional understanding of limnological change then can be provided by a few downcore analyses.

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5.4 Future directions

As with most scientific research, this thesis raises several questions and also highlights several potential avenues for future research. In Chapter 2, we documented marked increases in several planktonic taxa including A. formosa and F. crotonensis. However given the limitations of the top-bottom approach, putting these changes into a precise temporal context was difficult.

Recent data suggest that blooms of these diatoms have continued and may present several opportunities for future research. First, downcore analysis of select Muskoka-Haliburton lakes could be undertaken to elucidate the timing of these changes, and second, lakes such as

Haliburton and Walker, whose nutrients decreased in our study while A. formosa and F. crotonensis increased, could be investigated more closely to better establish the mechanism responsible for biological change (i.e., climate vs. nutrients).

The analysis of scaled chrysophytes in Chapter 3 also raised several questions which should be explored in future research. First, although our data show that DOC concentration is increasing in the vast majority of lakes in Muskoka-Haliburton, the mechanisms responsible for these changes warrant further exploration. It has been suggested that this mechanism may vary between regions, for example in more acidified areas such as Sudbury, DOC recovery may be related to acid recovery while in Muskoka it has been suggested that climate play a more important role (Eimers et al. 2008). A better understanding of these mechanisms and how they will continue to affect [DOC] in the future is vital for forecasting and managing future limnological change. In addition, as with Chapter 2, our top-bottom approach allowed us to expand our study regionally but limited our ability to resolve detailed temporal changes in colonial chrysophytes.

Detailed downcore paleolimnological analysis, coupled with newly developed spectral techniques

121 for DOC (Rouillard et al. 2011), may lead to new insights regarding the interactions between DOC and chrysophyte algae.

Our research on native crayfish extirpations in Chapter 4 highlights several significant gaps in the scientific literature, which should be addressed in the future. First and foremost, it is clear that a better understanding of native crayfish calcium tolerance is necessary in order for informed lake management decisions to be made. Currently, available data places the approximate hypothesized Ca tolerance range at between 2 and 10 mg/L, a range which spans almost the entire variability of lakes in the Muskoka-Haliburton region. These numbers, while useful, are based upon research from a variety of regions and on crayfish species that are not present in south-central Ontario. Any real management solution should ideally be based on a much more precise Ca tolerance making establishing such values for native Ontario crayfish a research priority. Likewise, because all four of our lakes contained predatory fish species, it was difficult to incorporate top-down pressures into our research. There is little doubt that fish predation is an important driver of crayfish populations, therefore future research combining robust fish and crayfish data over wide Ca and pH gradients should be undertaken.

Finally, we have highlighted an apparent increase in the importance of climate change in driving limnological change in south-central Ontario lakes. As a consequence, we have begun to see a potential disconnect between proxy-based inference models and modern biological assemblages. For example, Cyclotella species are widely recognized as circumneutral, oligotrophic taxa whose pH and TP optima in regional calibration sets reflect these preferences. However, considerable evidence exists suggesting recent increases in Cyclotella spp. are being driven by climate and the associated changes in lake thermal structure (Rühland et al. 2008, 2010).

Therefore, attempting to reconstruct pH in lakes with high relative abundance of Cyclotella

122 species could potentially underestimate or overestimate lake-water pH. Similarly, increases in A. formosa and F. crotonensis, which may be partly driven by climate change, could affect TP reconstructions under certain circumstances. However, there is potential to work around these concerns by combining multiple regional models, and analogue matching is a valuable tool for diagnosing such reconstructions.

In summary, my Ph.D. research demonstrates that multiple stressors are responsible for significant chemical and biological change across south-central Ontario lakes. By combining long- term monitoring and paleolimnological data, I have provided a relatively comprehensive assessment of limnological change to date, which should aid the development of lake management guidelines and future policy decisions aimed to prevent further degradation of vital freshwater resources.

5.5 Literature cited

Cairns, A., and Yan, N.D. 2009.A review of laboratory and field evidence of the influence of low ambient calcium concentrations on daphniids, gammarids and crayfish. Environ. Reviews 17: 67- 79.

Christensen, M.R., Graham, M.D., Vinebrooke, R.D., Findlay, D.L., Paterson, M.J., and Turner, M.A. 2006. Multipleanthropogenic stressors cause ecological surprises in boreal lakes. Glob. Change Biol. 12: 2316-2322.

Eimers, M.C., Watmough, S.A., Buttle, J.M., and Dillon, P.J. 2008.Examination of the potential relationship between droughts, sulphate and dissolved organic carbon at a wetland-draining stream. Glob. Change Biol. 14: 938-948.

Enache, M.D., Paterson, A.M., and Cumming, B.F. 2011.Changes in diatom assemblages since pre- industrial times in 40 reference lakes from the Experimental Lakes Area (northwestern Ontario, Canada). J. Paleolimnol. 46: 1-15.

Futter, M.N. 2003. Patterns and trends in southern Ontario lake ice phenology. Environ. Monit. Assess. 88: 431-444.

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Hall, R.I., and Smol, J.P. 1996.Paleolimnological assessment of long-term water-quality changes in south-central Ontario lakes affected by cottage development and acidification. Can. J. Fish. Aquat. Sci. 53: 1-17.

Rouillard, A., Rosén, P., Douglas, M.S.V., Pienitz, R., and Smol, J.P. 2011. A model for inferring lakewater dissolved organic carbon (DOC) in lakewater from visible-near-infrared spectroscopy (VNIRS) measures in lake sediment. J. Paleolimnol. 46: 187-202.

Rühland, K., Paterson, A.M., and Smol, J.P. 2008.Hemispheric-scale patterns of climate-related shifts in planktonic diatoms from North American and European lakes. Glob. Change Biol. 14: 2740-2754.

Smol, J.P., and Douglas, M.S.V. 2007. From controversy to consensus: making the case for recent climatic change in the Arctic using lake sediments. Front. Ecol. Environ. 5: 466-474.

Watmough, S.A., and Aherne, J. 2008.Estimating calcium weathering rates and future lake calcium concentrations in the Muskoka-Haliburton region of Ontario. Can. J. Fish. Aquat. Sci. 65: 821-833.

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Appendix A

Water chemistry and count data

This appendix contains all the relative abundance data from diatom, chrysophyte and zooplankton counts collected for this thesis. All tables consist of species (rows) and either lakes or core depth

(columns), and are presented as percent relative abundance. Additionally, all the water chemistry data collected in 2008/9 are included with the water chemistry parameters in columns and the lake names in rows.

The data has been organized into six primary sections, as follows:

 A.1 Relative abundances of diatom microfossils for the "top" intervals of the Muskoka-

Haliburton survey lakes in Chapter 2

 A.2 Relative abundances of scaled chrysophyte microfossils for the "top" intervals of the

Muskoka-Haliburton survey lakes in Chapter 3

 A.3 Relative abundances of diatom microfossils for the Algonquin Park downcore analysis

in Chapter 4

o A.3.1 Relative abundances of diatom microfossils for Cradle Lake

o A.3.2 Relative abundances of diatom microfossils for Delano Lake

o A.3.3 Relative abundances of diatom microfossils for Pincher Lake

o A.3.4 Relative abundances of diatom microfossils for Westward Lake

 A.4 Relative abundances of scaled chrysophyte microfossils for the Algonquin Park

downcore analysis in Chapter 4

o A.4.1 Relative abundances of scaled chrysophyte microfossils for Cradle Lake

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o A.4.2 Relative abundances of scaled chrysophyte microfossils for Delano Lake

o A.4.3 Relative abundances of scaled chrysophyte microfossils for Pincher Lake

o A.4.4 Relative abundances of scaled chrysophyte microfossils for Westward Lake

 A.5 Relative abundances of cladoceran remains for the Algonquin Park downcore analysis

in Chapter 4

o A.5.1 Relative abundances of cladoceran zooplankton remains for Cradle Lake

o A.5.2 Relative abundances of cladoceran zooplankton remains for Delano Lake

o A.5.3 Relative abundances of cladoceran zooplankton remains for Pincher Lake

o A.5.4 Relative abundances of cladoceran zooplankton remains for Westward Lake

 A.6 Codes and full species names for diatom taxa used in Appendix tables

 A.7 Summary of key water chemistry variables in the Muskoka-Haliburton Lakes from

Chapter 2 and 3

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Appendix B

Water chemistry comparisons

This appendix contains all the figures comparing spring 1992, 1990-1992 ice-free mean and spring

2007/8 water chemistry data discussed in detail in Chapters 2 and 3.

Figure B.1. Comparison of the 2008 spring water [Ca], [DOC] and pH collected by Hadley et al. (Chapter 2 and 3) with 1992 spring and 1990-1992 ice-free mean values collected by Hall and Smol (1996).

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Figure B.2. Comparison of the 2008 spring SO4, SiO3 and [TP] collected by Hadley et al. (Chapter 2 and 3) with 1992 spring and 1990-1992 ice-free mean values collected by Hall and Smol (1996).

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Appendix C

Reproducibility analysis

The reproducibility of the diatom microfossil data collected for Chapters 2 was examined by counting triplicate cores for three of the Muskoka-Haliburton survey lakes (Basshaunt, Bruce and Leech Lake). At each lake, three sediment cores (labeled A, B and C) were collected, sectioned and prepared identically. Surface sediments (i.e. core depth = 0.0-0.25cm) from each of the 9 cores were counted according to standard paleolimnological practices, which are described in detail in the Chapter 2 methods. In addition, in order to harmonize taxonomy between counts performed by Roland Hall (1992) and Kristopher Hadley (2007/8), several double-blind recounts of slides previously counted by R. Hall were undertaken. I recounted three samples from the original

Hall and Smol (1996) counts, without any knowledge of the lake being counted, or the expected taxonomy.

C.1 Triplicates

The statistical analysis of the diatom assemblages from the replicate surface sediments was restricted to the 44 species that were present in an abundance >1% in at least one of the triplicate cores. To examine the similarities and differences in these cores, Bray-Curtis similarities were calculated within and between the replicate cores. Principal components analysis (PCA) of the triplicate core data showed a relatively strong separation between lakes within the ordination space, while replicate cores formed relatively tight clusters (Figure C.1). Furthermore, Bray-Curtis similarities within replicate cores (range = 69 - 85%), was much greater than between the different lakes (range = 41 - 54%). Strong separation between different lakes coupled with high similarity suggests a high degree of reproducibility between samples.

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Figure C.1. Principal component analysis (PCA) of the surface sediment diatom communities the triplicate analysis of the three lakes.

C.2 Hall and Smol recounts

Statistical analysis was also performed on the three double-blind recounts of the Hall and

Smol (1996) lakes. As with the above, diatoms assemblages were restricted to the 33 species that were present in >1% abundance in at least one interval. After counts were completed, the three lakes sampled in this manner were revealed to be Harp, Blue Chalk and Little Redstone. When comparing 1992 and 2007 counts, similarities were in a similar range to what was seen between replicate cores (range 60, 79 and 84%). The lowest of these values (i.e., 60%) was for Little

Redstone Lake and was the result of a single taxonomic inconsistency, which as a result of this analysis was rectified. Because the variability between my counts and the counts of R. Hall were, similar to the inter-core variability I determined the two datasets could be compared directly with minimal impact of the individual counter.

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Appendix D

210Pb dating analysis for Algonquin Park cores

The following figures summarize the 210Pb chronologies developed for Cradle, Delano, Pincher and Westward Lakes.

300 Cradle Delano 250 Pincher Westward

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Pb activity (DPM/g) Pb activity

210 100

Total 50

0 0 5 10 15 20 25 30

2000 Cradle Figure D.1. Summary of the total 210Pb activity with sediment depth downcore forDelano Cradle, Delano, Pincher Pincher1980 and Westward Lakes. Westward

1960

1940

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1900

Pb age (Years A.D.) Pb (Years age

210 1880 177

1860

1840 0 5 10 15 20 25 30 Sediment depth (cm) 300 Cradle Delano 250 Pincher Westward

200

150

Pb activity (DPM/g) Pb activity

210 100

Total 50

0 0 5 10 15 20 25 30

2000 Cradle Delano Pincher 1980 Westward

1960

1940

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1900

Pb age (Years A.D.) Pb (Years age

210 1880

1860

1840 0 5 10 15 20 25 30 Sediment depth (cm)

Figure D.2. Approximate downcore sediment age, based on 210Pb analysis of Cradle, Delano, Pincher and Westward lake cores.

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