Download ACIDIFICATION AND CLIMATE WARMING: UNDERSTANDING THE IMPACT OF

ACIDIFICATION AND CLIMATE WARMING: UNDERSTANDING THE IMPACT OF
MULTIPLE ANTHROPOGENIC STRESSORS ON ADIRONDACK (NY, USA) LAKES
by
Kristina Marie Allport Arseneau
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
May, 2014
Copyright © Kristina M. A. Arseneau, 2014
ABSTRACT
Lakes in the Adirondack Park (NY, USA) are undergoing chemical recovery from
acidification. There is now a pressing research need to define recovery targets for acid-impacted
sites. Researchers attempting to designate such targets are hampered by two issues: 1) a lack of
long-term monitoring data, and 2) the influence of multiple stressors on recovering lakes. This
thesis addresses both difficulties by applying paleolimnological techniques within a regional
reference lake framework. Using a set of stringent selection criteria, 31 lakes protected from
acidification, eutrophication, road salt seepage, and piscivore introductions were identified from
1,469 Adirondack lakes. Ordination techniques showed that the lakes are representative of 2436% of the chemical/morphological variation of Adirondack lakes. Qualitative and quantitative
historic analyses found that many of the lakes experienced early watershed and/or fisheries
disturbance, highlighting the danger of assuming that a lake’s condition remains static over time.
A top-bottom paleolimnological study revealed that the reference lakes have undergone a
‘shifting baseline’ in species assemblages, with increases in colonial and/or warm-water
chrysophyte taxa from pre-1900 to present, changes most likely due to regional warming and/or
oligotrophication. A subset of three reference lakes were then paired with two Adirondack lakes
that acidified and are undergoing chemical recovery from acidification. The acidified lakes
underwent a significant shift in species composition since the 1995 implementation of the US
Acid Rain Program, indicating biological recovery from acidification. However, both reference
and acidified lakes showed increases in colonial chrysophytes since ca. 1970-1980, a trend
correlated with mean annual air temperature and ice-cover measures in the two reference lakes.
Long-term species changes in acidified/reference lakes suggest that the recovering lakes will not
return to their pre-disturbance state but will instead move to a state characterized by an increased
abundance of colonial taxa/warm-water species. Overall, this thesis demonstrates the utility of
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pairing paleolimnological techniques with a regional reference site dataset for tracking shifting
baselines and defining recovery targets, a method that could be applied to examine other
stressors in other regions, thereby addressing a critical management need.
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CO-AUTHORSHIP
Chapter 3 was co-authored by Dr. Charles T. Driscoll, Graham Pope, Cassandra Cummings, and
Dr. Brian F. Cumming, and represents original work completed as part of my Ph.D. thesis. I was
responsible for project design, led all field work, performed the qualitative historical analyses
and ordination analyses, and am the principal author on the paper.
Chapter 4 was co-authored by Dr. Charles T. Driscoll and Dr. Brian F. Cumming, and
represents original work completed as part of my Ph.D. thesis. I was responsible for project
design, led all field work, collected and analyzed all chrysophyte data, performed the ordination,
univariate, and ANOSIM analyses, and am the principal author on the paper.
Chapter 5 was co-authored by Dr. Charles T. Driscoll, Cassandra M. Cummings, Ayla Fenton,
and Dr. Brian F. Cumming, and represents original work completed as part of my Ph.D. thesis. I
was responsible for project design, led all field work, collected and analyzed chrysophyte data
from Moss Lake, South Lake, and Queer Lake, performed the ANOSIM, ordination, and
correlation analyses, and am the principal author on the paper.
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ACKNOWLEDGEMENTS
If the old adage that it takes a village to raise a child is true, then it certainly must take a
lifetime of relationships to create a Ph.D. thesis. If I neglect to mention one of the seemingly
endless number of people who have supported me in this, that oversight is entirely mine and not
a reflection of the value of their contribution.
First, I must thank my supervisor and mentor Dr. Brian F. Cumming, who taught me that
above all else science is about constantly asking and refining questions. His guidance in this
project has been absolutely invaluable and his constant support has been deeply appreciated. I
must also thank my committee: Dr. John P. Smol, Dr. Neal Scott, and Dr. Andrew Paterson,
whose support and suggestions have greatly improved these manuscripts. I also want to thank
the collective members of the Paleoecological Environmental Assessment and Research
Laboratory both past and present, for their support, friendship, and their willingness to discuss
any obscure matter of paleo-science with gusto at any time of day or night.
This dissertation required a great deal of field work and would have been impossible
without the hard work and dedication of my numerous field crews. I would like to sincerely
thank: Cassandra Cummings, Ayla Fenton, Iain MacKenzie, Jerome Velasco, Colin Robinson,
Mark Kelly, and Brendan Wiltse. I also want to acknowledge the support I received from
individuals at the New York State Department of Environmental Conservation when navigating
the permitting process for this project, particularly Mr. Scott Healy.
The support of my colleagues and friends in New York State was absolutely invaluable in
this project. I would like to thank Dr. Curt Stager for his many informative discussions of
Adirondack ‘Heritage Lakes’ which were inspirational in this study design. I want to
acknowledge the researchers and staff at the Adirondack Ecological Centre who hosted my field
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crew and I at the Huntington Wildlife Forest, particularly Dr. Colin Beier who also provided icecover records for two reference lakes in this project. This project would not have been possible
without the long-term monitoring data collected by the Adirondack Lakes Survey Corporation. I
want to particularly acknowledge the help I have received from Mr. Jed Dukett, who was always
happy to answer any questions I had about their data. Lastly, I must acknowledge the collective
members of the Adirondack Research Council, whose yearly conference introduced me to the
new datasets like the Adirondack GIS portal and greatly deepened my understanding of the
intricacies of the private-public partnerships at the heart of the Adirondack Park.
A variety of funding agencies have provided support to the work undertaken in this
dissertation. The majority of the funding for this research was supported by an NSERC
Discovery Grant awarded to my supervisor, Dr. Brian Cumming. I have received funding support
from NSERC, Ontario Graduate Scholarships, and Queen’s University and my co-author Dr.
Charles T. Driscoll received funding from the New York State Energy Research and
Development Authority (NYSERDA). Funding to support water chemistry monitoring by the
Adirondack Lakes Survey Corporation has been provided by the NYSERDA and the US
Environmental Protection Agency.
Lastly, but certainly not least, I must thank my family and friends for their support: my
father Réjean Arseneau who was my first true science mentor, my mother Deborah Allport who
taught me so many of the things not covered in university, my brother Vincent Arseneau, my
dear friends Katie Griffiths, Katy Marsh, Matt Marsh, and Heather Mawby, and, of course, my
partner Laura Schaefli, who was always my first reviewer and editor-in-chief, my toughest critic
and biggest cheerleader. Danke, mein Herz.
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TABLE OF CONTENTS
Section
Page
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
ii
Co-authorship . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
iv
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
v
Table of contents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
vii
List of tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xi
List of figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xiii
List of abbreviations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xv
Chapter 1 – General introduction
General introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1
Literature cited. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6
Chapter 2 - Literature review
Impact of acid deposition on lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
9
Sources, forms, and the terrestrial-aquatic link . . . . . . . . . . . . . . . . . . . . . . .
9
Influence on the chemical & biological characteristics of lakes . . . . . . . . . .
11
Emission controls and chemical/biological recovery . . . . . . . . . . . . . . . . . .
12
Impact of climate change on lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
15
Physical/chemical effects of climate change on lakes . . . . . . . . . . . . . . . . . .
17
Biological effects of climate change on lakes . . . . . . . . . . . . . . . . . . . . . . . .
20
Understanding the effects of multiple stressors in lake environments . . . . . . . .
26
Reference sites & reference conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . .
vii
27
Chrysophytes – general ecology & use as paleo-indicators . . . . . . . . . . . . . . .
30
The Adirondack Park . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
34
Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
38
Literature cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
39
Chapter 3 – Tracking shifting baseline conditions due to climate change in lakes: a
novel use of regional reference sites and paleolimnology
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
57
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
58
Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
60
Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
79
Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
81
Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
85
Literature cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
86
Tables & table captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
95
Figures & figure captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
97
Chapter 4 – Adirondack (NY, USA) reference lakes show a pronounced shift in
chrysophyte species composition since ca. 1900
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
102
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
103
Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
106
Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
114
Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
118
Literature cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
124
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Tables & table captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
130
Figures & figure captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
131
Chapter 5 – Understanding biological recovery from acidification: applying a
reference-site framework in the Adirondacks (NY, USA)
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
135
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
136
Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
140
Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
149
Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
153
Literature cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
160
Tables & table captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
166
Figures & figure captions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
168
Chapter 6 – General discussion and conclusions
General discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
174
Future directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
180
Literature cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
183
Appendices
Appendix A. Histogram plots of Adirondack Lakes Survey (ALS) variables . . .
186
Appendix B. List of analytes measured in 2010 and 2011 water chemistry
samples from reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
187
Appendix C. Location information for 31 Adirondack reference lakes . . . . . . .
188
Appendix D. Summary diagram of species assemblages in 31 Adirondack
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
ix
189
Appendix E. Relationship between colonial chrysophyte taxa and DOC in 31
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
190
Appendix F. Stocking history, recent netting survey results, land-use
characteristics, and watershed disturbance histories for South
Lake, Queer Lake, Moss Lake, Arbutus Lake, and Wolf Lake . . . .
191
Appendix G. Seasonal Kendall τ results for long-term monthly chemistry data.
193
Appendix H. Stratigraphies of chrysophyte relative abundance . . . . . . . . . . . .
194
Appendix I. List of taxon codes used in nMDS ordination diagram . . . . . . . . .
196
Appendix J. Principal Components Analysis (PCA) results for acidified and
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
x
197
LIST OF TABLES
Table
Page
Chapter 3
Table 3.1. Summaries of lake and watershed characteristics, lake
morphometrics, and water chemistry variables for 31 Adirondack
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
95
Table 3.2. Summaries of early watershed disturbance and fisheries disturbance
in 31 Adirondack reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
96
Chapter 4
Table 4.1 SIMPER test results showing taxa that contributed to the significant
difference in species composition . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
130
Chapter 5
Table 5.1. ANOSIM and SIMPER test results assessing whether chrysophyte
assemblages have changed between two a priori defined time periods.
166
Table 5.2. Kendall τ correlation test results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
167
Appendices
Table B1. List of analytes measured in 2010 and 2011 water chemistry samples
from reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
187
Table C1. Location information for 31 Adirondack reference lakes . . . . . . . . . . .
188
Table F1. Stocking history, recent netting survey results, land-use
characteristics, and watershed disturbance histories for South Lake,
Queer Lake, Moss Lake, Arbutus Lake, and Wolf Lake . . . . . . . . . . . . .
xi
191
Table G1. Seasonal Kendall τ results for long-term monthly chemistry data . . . .
193
Table I1. List of taxon codes used in nMDS ordination diagram . . . . . . . . . . . . .
196
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LIST OF FIGURES
Figure
Page
Chapter 3
Figure 3.1. Conceptual diagram outlining the biological response of two lakes
(Lake A and Lake B) to acidification and climate change since preindustrial times . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
97
Figure 3.2. Conceptual spatial diagram of a paleolimnological reference site
study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
98
Figure 3.3. Schematic outlining the reference lake selection and characterization
process followed in this investigation . . . . . . . . . . . . . . . . . . . . . . . . .
99
Figure 3.4. Map of the Adirondack Park (NY, USA) showing the locations of a
set 31 Adirondack reference lakes minimally impacted by
anthropogenic stressors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
100
Figure 3.5. PCA ordinations of ALS survey data . . . . . . . . . . . . . . . . . . . . . . . . . .
101
Chapter 4
Figure 4.1. Schematic showing the paleolimnological study design . . . . . . . . . . .
131
Figure 4.2. PCA of environmental variables and Redundancy analysis (RDA)
using PCA-axes scores as input . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
132
Figure 4.3. PCA of chrysophyte assemblages in top (present-day) and bottom
(pre-1900) samples in 26 reference lakes . . . . . . . . . . . . . . . . . . . . . .
133
Figure 4.4. Total relative abundance (%) of colonial chrysophyte taxa in top
(present-day) and bottom (pre-1900) samples in 26 reference lake . .
xiii
134
Chapter 5
Figure 5.1. Long-term chemistry trends in acidified and reference lakes . . . . . . .
168
Figure 5.2. 210Pb and 214Bi activities in sediment cores from acidified and
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
169
Figure 5.3. Total relative abundance of colonial chrysophytes (% colonial) and
chrysophyte-inferred pH (CI-pH) values for two acidified lakes
(South Lake and Queer Lake) and three reference lakes (Arbutus
Lake, Wolf Lake, Moss Lake) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
170
Figure 5.4. Non-metric Multidimensional Scaling (nMDS) plots for acidified and
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
171
Figure 5.5. Non-metric Multidimensional Scaling (nMDS) plots relating to South
Lake species assemblages . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
172
Figure 5.6. Non-metric Multidimensional Scaling (nMDS) plots relating to
Queer Lake species assemblages . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
173
Figure A1. Histogram plots of Adirondack Lakes Survey (ALS) variables . . . . . .
186
Appendices
Figure D1. Summary diagram of species assemblages in 31 Adirondack
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
189
Figure EI. Relationship between colonial taxa and DOC in 31 reference lakes . .
190
Figure H1. Stratigraphies of chrysophyte relative in acidified lakes . . . . . . . . . . .
194
Figure H2. Stratigraphies of chrysophyte relative abundance in reference lakes .
195
Figure I1. Principal Components Analysis (PCA) results for acidified and
reference lakes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xiv
197
LIST OF ABBREVIATIONS
Symbol
Meaning
ABC-top
AEZ
ALS
ALSC
ANC
ANOSIM
APA
BC
BDL
CAAA
CFCs
CI-pH
DEM
DCA
DIC
DI-pH
DOC
ELS
EMAP
FAM
FDM
GIS
NAAQS
NAPAP
NYSDEC
PAR
PCA
PIRLA
RDA
RMSEP
NY
US
USA
US ARP
US EPA
SA
SIMPER
SUNY-ESF
SWAP
TP
TN
WS-SA
Zmax
Average Bray-Curtis similarity of replicate top samples
Adirondack Ecological Zone
Adirondack Lakes Survey
Adirondack Lakes Survey Corporation
Acid neutralizing capacity
Analysis of Similarities
Adirondack Park Agency
Bray-Curtis similarity
Below detection limit
Clean Air Act Amendments (1970 & 1990)
Chlorofluorohydrocarbons
Chrysophyte-inferred pH
Digital elevation model
Detrended Correspondence Analysis
Dissolved inorganic carbon
Diatom-inferred pH
Dissolved organic carbon
Eastern Lakes Survey
Environmental Monitoring and Assessment Program
Flow accumulation model
Flow direction model
Geographic information system
National Ambient Air Quality Standards
National Acid Precipitation Program
New York State Department of Environmental Conservation
Photosynthetically active radiation
Principal Components Analysis
Paleoecological Investigation of Recent Lake Acidification
Redundancy Analysis
Root-mean squared error of prediction
New York
United States
United States of America
United States Acid Rain Program
United States Environmental Protection Agency
Surface area
Similarities Percentages test
State University of New York School of Environmental Science and Forestry
Surface Water Acidification Programme
Total phosphorus
Total nitrogen
Watershed to surface area ratio
Maximum depth
xv
CHAPTER 1
GENERAL INTRODUCTION
In 2000, the Nobel-prize winning chemist Paul Crutzen and the renowned diatomist
Eugene Stoermer argued that the past three centuries should be renamed the Anthropocene, a
new geological epoch reflecting the profound influence of human activity on ecosystems around
the globe (Crutzen and Stoermer 2000). Though controversial, the term serves to highlight how
humans have affected all levels of ecosystem functioning in marine, terrestrial, and freshwater
environments. Indeed, even remote alpine and arctic sites have recorded the effects of
anthropogenic stressors (Catalan et al. 2013), suggesting that ‘pristine’ sites protected from all
human activities are unlikely to exist (Bennion et al. 2011).
As anthropogenic influence has increased around the world, a growing number of
environmental stressors have begun to act in tandem on freshwater ecosystems. Since the mid1990s, there has been a shift in the ecological literature away from a focus on single-stressor
impacts (e.g. acidification or eutrophication) to an emphasis on multiple stressor systems (e.g.
acidification and/or, land-use change, invasive species, climate change, etc.). For example, in a
highly cited manuscript, Schindler et al. (1996) noted that the interactive effects of acidification
and climate change increased UV-B penetration in an experimentally acidified lake more than in
an unmanipulated reference lake in the Experimental Lakes Area (ON, Canada). Similarly, in a
seminal case-study review of both terrestrial and aquatic environments, Paine et al. (1998)
argued that multiple stressors can lead to ‘ecological surprises’, or novel species communities,
which may be irreversible.
Systems impacted by multiple stressors pose a particular management challenge.
Fundamentally, most environmental stressors have legislative controls which are largely separate
from other, possibly interacting, stressors. To take the example of Schindler et al. (1996) above,
1
in the United States acid deposition is regulated under Title IV of the US Clean Air Act, ozonedepleting compounds like chlorofluorocarbons (CFCs) are regulated under Title VI of the Clean
Air Act, and the US Environmental Protection Agency is currently engaged in developing
regulatory strategies for CO2 emissions under the Clean Air Act mandate. While each stressor
may be subject to different legislative controls under the Clean Air Act, acidification, UV-B
radiation, and climate change can all influence the same waterbody. Moreover, while legislative
and management strategies dealing with single environmental stressors like acidification often
have either an explicit or implicit recovery goal of a return to pre-disturbance state (Baker et al.
1990), the interactive effects of multiple stressors may prevent recovering systems from
returning to pre-disturbance conditions. Such ‘shifting baselines’ make defining recovery targets
difficult – in the face of changing climate conditions, what should a ‘recovered’ site look like?
Researchers and managers who are attempting to designate recovery targets for impacted
systems must address two issues: 1) the confounding influence of multiple stressors and, more
fundamentally, 2) a lack of long-term monitoring data. In a recent meta-analysis of both
terrestrial and aquatic recovery studies, Jones and Schmitz (2009) found that only 20% of the
studies had access to pre-disturbance data. In aquatic ecosystems, long-term monitoring records,
where they exist, are typically short (< 50 years) and are insufficient to track the effects of
stressors like climate change which work on decadal-to-centennial timescales. Fortunately,
paleolimnological techniques can be used to reconstruct the environmental histories of lakes
from yearly-to-centennial (or indeed, millennial) timescales (Smol 2008). Paleolimnological
techniques have been used to understand the effects of acidification and eutrophication in aquatic
ecosystems, as well as define biological recovery from those stressors (Battarbee 1999; Battarbee
2
et al. 2005) and study the effects of multiple co-occurring stressors in lake ecosystems (Smol
2010).
When it comes to understanding shifting baselines, paleolimnological studies have a rich
history of using remote, minimally-disturbed alpine or arctic sites to track the effects of climate
change (Battarbee et al. 2002; Catalan et al. 2013; Gregory-Eaves et al. 1999). However, such
sites are typically removed from areas heavily influenced by anthropogenic stressors like
acidification, eutrophication, and/or land-use change and may be subject to different climate
controls (i.e. high elevation/arctic vs. temperate lowlands), making them less than ideal for
defining recovery targets. Conversely, there is an expansive limnological bioassessment
literature that uses minimally-disturbed reference sites within anthropogenically-impacted
regions to define recovery targets for impacted sites, often by using modelling techniques (Bailey
et al. 2004; Downes et al. 2002; Stoddard et al. 2006). However, these studies usually lack the
long-term perspective offered by paleolimnological studies and so are unable to define predisturbance conditions for impacted sites and identify whether or not shifting baselines are
occurring in the reference/impacted systems due to climate change.
This project aims to bridge the paleolimnological and bioassessment literature by
applying paleolimnological techniques in a study of a regional set of minimally-disturbed
reference lakes in the Adirondack Park (NY, USA). The main goal of this project is to use the
regional reference sites to track shifting baselines due to regional stressors like climate change
and then use that information to define reasonable recovery targets for lakes recovering from
acidification in that region. The Adirondack region is an ideal study location for this project
because the area has been heavily influenced by acid deposition (Driscoll et al. 1991), the
acidification history of the region has been documented in previous paleolimnological studies
3
(Cumming et al. 1992; Cumming et al. 1994), and large synoptic surveys have been undertaken
in the region (Krester et al. 1989). Additionally, long-term chemical monitoring data exist for
many Adirondack lakes (Roy et al. 2011) and recent studies have found that some lakes are
undergoing chemical recovery from acidification (Driscoll et al. 2007; Waller et al. 2012).
However, while chemical recovery may be progressing, the Adirondacks are also undergoing
regional warming, a trend that has been most pronounced since the 1970s (Jenkins 2010).
This dissertation contains six chapters, including the general introduction (Chapter 1)
and literature review (Chapter 2). In Chapter 3, a conceptual framework is presented for
pairing paleolimnological techniques with a regional reference site approach. Key concepts are
discussed such as: what ‘good’ reference sites are and how they can be identified; the useful role
reference sites play in studies of multiple stressors; why reference site studies should be regional
in nature, rather than relying on a small number of sites; and how reference sites can be used to
help define recovery targets. This framework is then applied in the Adirondack Park. A set of 31
reference sites is identified out of the 1,469 lakes surveyed as part of the Adirondack Lakes
Survey using a set of stringent selection criteria. The reference sites are then evaluated critically
using both qualitative and quantitative historical methods and ordination techniques.
In Chapter 4, a paleolimnological ‘top-bottom’ study of the reference sites is carried out
using scaled chrysophytes as a proxy group (Chrysophyceae and Synurophyceae). The
relationship between the present-day distribution of scaled chrysophyte taxa and
physical/chemical variables in the reference lakes is examined using direct ordination techniques.
The change in chrysophyte assemblages from pre-1900 times to present is then examined in the
reference lakes using indirect ordination methods and quantified using both univariate and
multivariate analyses. Because of the reference lake study design, it is possible to conclude that
4
changes in chrysophyte assemblages from pre-1900 to present are not due to acidification,
eutrophication, road salt seepage, or piscivores introductions. Significant shifts in the
composition of chrysophyte species assemblages in the reference lakes are discussed in the
context of regional environmental change and shifting baselines for biological recovery.
In Chapter 5, a subset of the three reference lakes is paired with two Adirondack lakes
which acidified and are undergoing chemical recovery from acidification. Long-term changes in
the chrysophyte assemblages of the five lakes are reconstructed and compared. In the acidified
lakes, multivariate analyses are used to identify if the species assemblages have undergone a
significant shift in species composition since the implementation of the 1995 US Acid Rain
Program and are experiencing biological recovery from acidification. Then, indirect ordination
methods are used to assess if the post-1995 species assemblages of each acid-impacted lake are
returning to a species assemblage characteristic of pre-disturbance times. In a subsequent
analysis, the pre-1900 assemblages of the acid-impacted lakes are compared with those of the
entire regional reference lake set to identify which reference lakes were most similar to the acidlakes in species composition pre-1900. A subset of reference lakes are then used to define
probable recovery endpoints for the acidified lakes. Lastly, changes in chrysophyte species
composition in two reference lakes are compared with long-term temperature and ice-cover
records. Finally, Chapter 6 presents a general discussion of the results of this investigation and
suggests possible future research directions.
This dissertation makes several important contributions to the ecological literature. First,
it improves our understanding of biological recovery from acidification, a subject that is
currently under-documented (Verdonschot et al. 2013). Second, it contributes to our
understanding of how climate change, possibly interacting with the long-term effects of acid
5
deposition, influences algal assemblages, a topic that is of keen research interest as the impact of
climate change on lakes continues to grow (Winder and Sommer 2012). Lastly, this dissertation
provides a critical link between the paleolimnological and bioassessment literature and provides
a framework for examining the effects of multiple stressors on biological recovery. Though the
focus of this dissertation is on acidification, the framework outlined could be easily adapted and
applied to look at other pressing issues, including eutrophication, invasive species introductions,
or land-use change. Researchers, policy makers, and lake managers no longer have the luxury of
considering stressors in isolation. Frameworks like the one used in this dissertation that combine
the long-term perspective of paleolimnology with the rigorous study design used in
bioassessment work will become increasingly useful and necessary as lakes both within and
outside protected areas are increasingly influenced by a growing number of anthropogenic
stressors.
Literature Cited
Bailey RC, Norris RH, Reynoldson TB. (2004) Bioassessment of freshwater ecosystems: using the reference
condition approach. Kluwer Academic Publishers, New York, 170 p.
Baker JP, Bernard DP, Christensen SW, Sale MJ. (1990) Biological effects of changes in surface water acid-base
chemistry. Report 13. Acidic deposition: state of science and technology. National Acid Precipitation
Assessment Program (NAPAP), Washington, DC, pp. 13-1-13-381.
Battarbee RW. (1999) The importance of palaeolimnology to lake restoration. Hydrobiologia 395/396:149-159.
Battarbee RW, Anderson NJ, Jeppesen E, Leavitt PR. (2005) Combining palaeolimnological and limnological
approaches in assessing lake ecosystem response to nutrient reduction. Freshwater Biology 50:1772-1780.
Battarbee RW, Thompson R, Catalan J, Grytnes J-A, Birks HJB. (2002) Climate variability and ecosystem
dynamics of remote alpine and arctic lakes: the MOLAR project. Journal of Paleolimnology 28:1-6.
6
Bennion H, Battarbee RW, Sayer CD, Simpson GL, Davidson TA. (2011) Defining reference conditions and
restoration targets for lake ecosystems using palaeolimnology: a synthesis. Journal of Paleolimnology
45:533-544.
Catalan J, Pla-Rabés S, Wolfe AP, Smol JP, Rühland KM, Anderson NJ, Kopáček J, Stuchlík E, Schmidt R, Koinig
KA, Camarero L, Flower RJ, Heiri O, Kamenik C, Korhola A, Leavitt PR, Psenner R, Renberg I. (2013)
Global change revealed by palaeolimnological records from remote lakes: a review. Journal of
Paleolimnology 49:513-535.
Crutzen PJ, Stoermer EF. (2000) The "Anthropocene". Global Change Newsletter 41:17-18.
Cumming BF, Davey KA, Smol JP, Birks HJB. (1994) When did acid-sensitive Adirondack lakes (New York,
USA) begin to acidify and are they still acidifying?. Canadian Journal of Fisheries and Aquatic Sciences
51:1550-1568.
Cumming BF, Smol JP, Kingston JC, Charles DF, Birks HJB, Camburn KE, Dixit SS, Uutala AJ, Selle AR. (1992)
How much acidification has occurred in Adirondack region lakes (New York, USA) since preindustrial
times?. Canadian Journal of Fisheries and Aquatic Sciences 49:128-141.
Downes BJ, Barmuta LA, Fairweather PG, Faith DP, Keough MJ, Lake PS, Mapstone BD, Quinn GP. (2002)
Monitoring ecological impacts: concept and practice in flowing waters. Cambridge University Press,
Cambridge, 452 p.
Driscoll CT, Driscoll KM, Roy KM, Dukett J. (2007) Changes in the chemistry of lakes in the Adirondack region of
New York following declines in acidic deposition. Applied Geochemistry 22:1181-1188.
Driscoll CT, Newton RM, Gubala CP, Baker JP, Christensen S. (1991) Adirondack mountains. In: Charles DF (ed),
Acidic deposition and aquatic ecosystems: regional case studies. Springer-Verlag, New York, pp. 133-202.
Gregory-Eaves I, Smol JP, Finney BP, Edwards ME. (1999) Diatom-based transfer functions for inferring past
climatic and environmental changes in Alaska, USA. Arctic, Antarctic, and Alpine Research 31:353-365.
Jenkins J. (2010) Climate change in the Adirondacks: the path to sustainability. Cornell University Press, Ithaca,
New York, 183 p.
Jones HP, Schmitz OJ. (2009) Rapid recovery of damaged ecosystems. PLoS ONE 4:e5653-1- e5653-6.
Krester W, Gallagher J, Nicolette J. (1989) Adirondack lakes survey 1984-1987: an evaluation of fish communities
and water chemistry. Adirondack Lakes Survey Corporation, Ray Brook, NY, 437 p.
7
Paine RT, Tegner MJ, Johnson EA. (1998) Compounded perturbations yield ecological surprises. Ecosystems
1:535-545.
Roy K, Houck N, Hyde P, Cantwell M, Brown J. (2011) The Adirondack long-term monitoring lakes: a
compendium of site descriptions, recent chemistry and selected research information. New York State
Energy Research and Development Authority, Albany, 298 p.
Schindler DW, Curtis PJ, Parker BR, Stainton MP. (1996) Consequences of climate warming and lake acidification
for UV-B penetration in North American boreal lakes. Nature 379:705-708.
Smol JP. (2008) Polluion of lakes and rivers: a paleoenvironmental perspective. Blackwell Publishing Ltd, Malden.
Smol JP. (2010) The power of the past: using sediments to track the effects of multiple stressors on lake ecosystems.
Freshwater Biology 55 (Suppl. 1):43-59.
Stoddard JL, Larsen DP, Hawkins CP, Johnson RK, Norris RH. (2006) Setting expectations for the ecological
condition of streams: the concept of reference condition. Ecological Applications 16:1267-1276.
Verdonschot PFM, Spears BM, Feld CK, Brucet S, Keizer-Vlek H, Borja A, Elliott M, Kernan M, Johnson RK.
(2013) A comparative review of recovery processes in rivers, lakes, estuarine and coastal waters.
Hydrobiologia 704:453-474.
Waller K, Driscoll C, Lynch J, Newcomb D, Roy K. (2012) Long-term recovery of lakes in the Adirondack region
of New York to decreases in acidic deposition. Atmospheric Environment 46:56-64.
Winder M, Sommer U. (2012) Phytoplankton response to a changing climate. Hydrobiologia 698:5-16.
8
CHAPTER 2
LITERATUTE REVIEW
Outline
The goal of this literature review is to provide the necessary background information for
a greater appreciation of chapters 3-5. Several topics are discussed including: the effects of acid
deposition on lakes, the impact of climate change on lakes, tools for understanding the influence
of multiple stressors on lakes (with a focus on using reference sites), the ecology of scaled
chrysophytes and their use as paleo-indicators, and the influence of acid deposition and climate
change in the Adirondack Park.
The impact of acid deposition on lakes
Sources, forms, and the terrestrial-aquatic link
Acid deposition is a globally-occurring phenomenon that affects North America
(primarily eastern North America and the western US), Europe, and parts of Asia (Driscoll
2012). Acid deposition is primarily a by-product of fossil fuel combustion. The combustion of
fossil fuels from sources like electric utilities and motor vehicles leads to the production of SO2
and NOx emissions. SO2 and NOx react with water in the atmosphere to produce sulphuric and
nitric acids (H2SO4 and HNO3, respectively), which then return from the atmosphere in the form
of wet and dry deposition1. Wet deposition includes acid rain, snow, and sleet whereas dry
deposition includes acids bound to particles, gases, or a combination of both (aerosols) (Munson
and Gherini 1991).
1
In agricultural areas, NH3 emissions may also be an important source of acid deposition (Munson and Gherini
1991).
9
In the absence of acid deposition, the acid-base status of waterbodies is primarily
controlled by bicarbonate buffering and organic acids (Munson and Gherini 1991). Organic acids
can act as buffers but their presence tends to decrease lake-water pH (Kullberg et al. 1993). In
areas influenced by acid deposition, the acid-base status of waterbodies is primarily determined
by mineral acids and is intimately linked to the terrestrial environment. In general, the primary
determinant of the acid-base status of a lake is the hydrologic flow path (Goldstein et al. 1984).
In the soil matrix, H+ associated with incoming acids will undergo cation exchange. Cations with
a higher valence (3+, 2+) will be preferentially exchanged for the H+ and sulphate will adsorb to
iron and aluminum oxides (Jenkins et al. 2007). Soil pH tends to increase with soil depth, and so
the deeper precipitation percolates through the soil matrix, the higher the likelihood that H+ will
be replaced by a cation, neutralizing the acids prior to groundwater flow entering a nearby
waterbody (Munson and Gherini 1991). For this reason, lakes surrounded by thicker watershed
soils tend to have more alkaline surface waters (Newton and Driscoll 1990).
The buffering function of watershed soils can be exceeded by high levels of acid
deposition. If the leaching of base cations occurs at a rate greater than they are supplied to the
soil matrix by weathering, the soils will acidify. Sulphate adsorption is concentration dependent,
and so increasing soil sulphate concentration leads to a greater release of sulphate to nearby
waterbodies (Jenkins et al. 2007). The solubility of Al also increases as soil pH drops below 5.5,
leading to the export of toxic inorganic Al to nearby waterbodies (Munson and Gherini 1991). In
general, when soil base-cation concentration is less than 20%, the input of strong acids will result
in an incomplete neutralization of H+ and a mobilization of toxic inorganic aluminum (Cronan
and Schofield 1990).
10
While sulphate deposition may make the largest contribution to acid anion concentration
in surface waters (Baker et al. 1990b), nitrate is often the main constituent of episodic
acidification (Wigington et al. 1996). Terrestrial environments are typically N-limited and so
NO3- is actively taken up during the growing season. During the winter, NO3- builds up in the
snowpack and the spring snowmelt results in an acid pulse to receiving waterbodies (Munson
and Gherini 1991). In areas heavily influenced by acid deposition, episodic acidification can be
sufficient to result in spring fish kills (Baker et al. 1996). In general, waterbodies with an acid
neutralizing capacity (ANC) greater than 50 μeq L-1 are considered protected from episodic
acidification whereas lakes with an ANC < 0 μeq L-1 are considered chronically acidic, and any
additional input of acid will result in a decrease in pH (Driscoll et al. 2001).
Influence of acid deposition on the chemical & biological characteristics of lakes
Acid deposition causes a suite of chemical changes in freshwaters. Most importantly, acid
deposition leads to acidification (i.e. decreases in lake-water pH and ANC) when the buffering
capacity of surrounding watershed soils and in-lake processes are surpassed. High levels of acid
deposition also lead to an increased concentration of strong acid anions and toxic forms of
inorganic aluminum in surface waters (Munson and Gherini 1991). An increased flux of cations
increases the total solute concentration (and conductivity) of receiving waters and can lead to
calcium depletion in watershed soils (Likens et al. 1996). Acid deposition is also associated with
reductions in lake-water DOC concentrations, allowing for increased UV-B penetration
(Schindler et al. 1996b) and changes in lake thermostructure (Snucins and Gunn 2000). The
influence of acid deposition on TP concentrations is less well-understood. An increased
concentration of SO42- in sediments may reduce the sequestration of TP (Caraco et al. 1991). In
contrast, soil acidification may lead to a decreased export of TP from watershed soils (Eimers et
11
al. 2009). Overall, reviews of acidification research have not revealed a consistent link between
acidification and eutrophication/oligotrophication (Olsson and Pettersson 1993; Schindler 1988).
In general, the acidification of freshwaters leads to a simplification of aquatic foodwebs.
The toxic effects of acidification on aquatic taxa is primarily related to an interruption of normal
osmoregulatory function, as well as impairment of respiratory processes by inorganic aluminum
(Baker et al. 1990a). Acidification may also affect aquatic organisms indirectly by changing food
availability, foodweb structures, and/or predation levels (Baker and Christensen 1991). Most
acid-sensitive aquatic organisms can persist when pH is greater than 6.0 (Baker et al. 1990a).
Numerous studies have shown that decreases in pH lead to a progressive loss of taxa in
phytoplankton (Siegfried et al. 1989), zooplankton (Siegfried and Sutherland 1992), and fish
assemblages (Krester et al. 1989). Coarse metrics, like biomass and production, within
taxonomic groups often do not show consistent patterns with acidification (Baker and
Christensen 1991). The fact that coarse metrics are not typically reliable indicators in
acidification studies is likely due to functional complementarity: as one taxon is lost with
acidification, another more acid-tolerant taxon is released from competition and increases in
abundance until carrying capacity is reached (Havens and Carlson 1998). Despite these species
replacements, acidification is a key management concern, as the acidification of freshwaters
leads to the loss of desirable sportfish taxa like lake trout (Salvelinus namaycush) and increases
in nuisance filamentous algae in the littoral zone (Baker and Christensen 1991).
Emission controls and chemical/biological recovery
Growing concerns about the human health effects of SO2 and NOx emissions, as well as
the effects of acid deposition on terrestrial and aquatic environments, led to the creation of
legislative control programs for SO2 and NOx in North America and Europe. Driscoll et al.
12
(2010) provides a useful summary of the main research/policy steps taken in Europe and the
United States to address acid deposition. Briefly, in the US, legislation was implemented as part
of the 1970 and 1990 Clean Air Act Amendments (CAAA). In 1970, National Ambient Air
Quality Standards (NAAQS) were introduced that focused primarily on limiting concentrations
of individual pollutants in urban areas, rather than their long-distance transport. In 1980, the
National Acid Precipitation Assessment Program (NAPAP) was created, a multi-agency research
program tasked with examining all issues associated with acid deposition. In 1990, NAPAP
released a series of reports that provided the scientific foundation for the 1990 CAAA, which
included the creation of the US Acid Rain Program (ARP). The ARP initiated an innovative ‘cap
and trade’ program for SO2 emissions with the goal of a 50% reduction below 1980 levels by
2010, as well as intensity-based NOx emissions targets for electricity generating units (Burns et
al. 2011).
S-emissions for the continental United States peaked in 1970 and have since declined
(Husar et al. 1991). In 2009, SO2 emissions from sources governed by the ARP were 67% lower
than 1980 levels (Burns et al. 2011). In the same year, NOx emissions from sources governed by
the ARP were 67% lower than 1995 levels. Overall, emissions of SO2 from all sources (including
those not governed by the ARP) were 59% lower than 1990 levels and NOx emissions were 40%
lower than 1990 levels. Chestnut and Mills (2005) estimated that the human health and
environmental benefits of the US ARP are over $100 billion, compared to an implementation
cost of $3 billion, arguably making the 1990 CAAA one of the most successful pollution-control
measures ever implemented in the United States.
Clean air legislation in the US, Canada, and Europe has resulted in chemical recovery
from acidification in some regions. Declines in lake sulphate concentration have been almost
13
universally documented in acid-impacted areas, while some regions have also experienced
increases in ANC and decreases in H+ (Garmo et al. 2014; Skjelkvåle et al. 2005). Driscoll et al.
(2001) lists three chemical recovery targets for freshwaters: 1) pH > 6.0 (to protect acid-sensitive
organisms), 2) ANC > 50 μeqL-1 (to protect against episodic acidification), and 3) inorganic
aluminum levels < 2 μmol L-1 (to protect fish from the toxic effects of aluminum). Modelling
studies have suggested that chemical recovery will be a protracted process (Chen and Driscoll
2005; Sullivan et al. 2011). Several factors are likely to slow chemical recovery in acid-impacted
regions, including the release of stored sulphate from watershed soils (Driscoll et al. 1998), base
cation depletion in soils (Likens et al. 1996), and increases in DOC concentrations. While DOC
increases should be expected with decreased acid deposition (Monteith et al. 2007), in the shortterm increased DOC concentrations may decrease pH and ANC in recovering surface waters
(Lawrence et al. 2013; Waller et al. 2012). Increases in DOC may, however, lead to a decrease in
the concentration of toxic inorganic monomeric aluminum, improving conditions for fish
(Lawrence et al. 2013).
It is expected that chemical recovery will promote biological recovery from acidification
after a lag period that will likely differ between taxonomic groups (Driscoll et al. 2001).
Recovery is expected to be characterized by a decreased abundance of acid-tolerant taxa and an
increased abundance of acid-sensitive taxa, as well as the re-establishment of extirpated species
(Baker et al. 1990b). Biological recovery can be measured using a variety of metrics including
species relative abundances, the presence/abundance of indicator species, and metrics like
species richness. Case studies in the United States and elsewhere have identified biological
recovery from acidification (Arseneau et al. 2011; Greenaway et al. 2012a; Havas et al. 1995;
Hynynen and Meriläinen 2005; Kopáček et al. 2002), though in many cases species responses are
14
modest. Beyond insufficient chemical recovery, there are several factors that may delay
biological recovery from acidification, including the limited dispersal ability for some taxa
(Keller and Yan 1998), damage to sedimentary eggs banks (Binks et al. 2005), increased
predation (Nilssen and Wærvågen 2002), and biological resistance from established acid-tolerant
species communities (Derry and Arnott 2007). There is growing recognition that recovering
species assemblages will likely not return to their pre-disturbance state, even with a return to predisturbance chemical conditions (if possible), due to the influence of additional environmental
stressors. In particular, paleolimnological studies of biological recovery from acidification have
suggested that climate change may prevent recovering species assemblages from returning to
their pre-disturbance state (Arseneau et al. 2011; Battarbee et al. 2014). Shifting baselines caused
by regional stressors like climate change pose a particular problem to lake managers and policy
makers when it comes to designing suitable recovery targets for acid-impacted systems. The next
section explores some of the physical, chemical, and biological effects of climate change on
lakes, followed by a discussion of tools ecologists can use to understand the influence of multiple
anthropogenic stressors on aquatic ecosystems.
The impact of climate change on lakes
Globally, from 1880 to 2012, mean surface temperature increased by an estimated 0.85ºC
(Hartmann et al. 2013) and models predict an additional increase between 0.3-4.8 ºC by 20812100 relative to the 1986-2005 time period (Collins et al. 2013). Paleoclimatic studies suggest
that temperatures experienced in the Northern Hemisphere during the last thirty years were likely
the warmest over the last 1,400 years (Masson-Delmotte et al. 2013).The effects of global
warming will be regionally-specific and may be difficult to predict. In North America, the
warming trend has been most pronounced over the past 50 years (Field et al. 2007) but this
15
warming trend is not evenly distributed. By the late 21st century, temperatures in the southern,
western, and eastern parts of North America will likely have warmed by an estimated 2-3 ºC
while warming in the north may exceed 5ºC (Christensen et al. 2007). Changes in precipitation
will also vary regionally. Increases in drought frequency in the US Southwest are likely to be
accompanied by increased rainfall in the US Northeast (Field et al. 2007). Such changes in
temperature and precipitation regimes are currently having, and will continue to have, wideranging and complex effects on both terrestrial and aquatic ecosystems, influencing chemical and
nutrient cycling (George 2010), seasonal match-or-mismatch between predator and prey species
(Durant et al. 2007), parasitism and disease (Marcogliese 2001), species distributions (Walther
2010), and more. Indeed, climate warming is frequently cited as one of the main conservation
concerns of the 21st century (Baron et al. 2009; Strayer and Dudgeon 2010).
Lakes can be considered ‘sentinels’ of climate change as they integrate the effects of
regional warming from the atmosphere, the catchment, and within-lake processes (Adrian et al.
2009). To understand the effects climate change is having and will have on lakes, researchers
draw on data from a variety of sources. In some cases, high-quality long-term monitoring
records of lake physical, chemical, or biological data exist (Winder and Schindler 2004), while in
other cases inferences are drawn from unusual or extreme weather events (e.g. unusually warm
years, Forsström et al. 2005; Jankowski et al. 2006). Modelling studies can also be useful when
trying to project the impacts of regional warming (e.g. Fang and Stefan 2009). While some
processes have been well-characterized (e.g. reduction in the ice-covered period, Duguay et al.
2006; Magnuson et al. 2000), others are more poorly understood (e.g. the interactions of climate
change with multiple stressors like eutrophication, Moss 2012). Complications also arise when
comparing different lake-types (i.e. deep, dimictic lakes vs. shallow, polymitic lakes), as they
16
differ in their thermal/chemical structures, as well as their response to environmental change
(e.g. nutrient loading, Taranu et al. 2010). Certainly, an exhaustive review of the effects of
climate warming across a range of lake-types would be beyond the scope of this dissertation. As
such, this section will focus primarily on the likely physical, chemical, and biological impacts of
climate change on North-temperate dimictic lakes.
Physical/chemical effects of climate change on lakes
One of the most visible effects of regional warming on lakes is the loss of ice-cover.
Regional surveys have shown a loss of seasonal ice-cover in lakes and rivers in the northern
hemisphere due to earlier ice-off dates (break-up dates in the spring) and/or later ice-on dates
(freeze-up dates in the fall) (Duguay et al. 2006; Futter 2003; Magnuson et al. 2000)2. Using a
process-oriented, water-quality model (MINLAKE96), Fang and Stefan (2009) suggested that
warming in the United States will be associated with a reduced ice-covered period of up to 90
days at 2xCO2 concentrations and a reduced ice thickness of up to 0.44 m. Changes in ice-cover
dynamics may differ regionally. For example, Beier et al. (2012) found that a reduced icecovered period in five Adirondack lakes was due primarily to a delay in ice-on rather than earlier
ice-off dates and hypothesized that heavy snowfall in the Adirondack region may minimize
changes in ice-off dates. The study by Beier et al. (2012) represents an interesting contrast to
work in other parts of North America which showed greater changes in ice-off rather than ice-on
dates (Duguay et al. 2006).
Changes in the length of the ice-covered period have an important influence on lake
thermostructure. A reduction in the ice-covered period can lead to a longer period of thermal
stratification and warming in epilimnetic waters can lead to an enhancement of thermal gradients
2
For a review of the controls on ice phenology, see Kirillin et al. (2012) and Livingston et al. (2010).
17
(Vincent 2009). Fang and Stefan (2009) suggested that in a 2xCO2 scenario, the summer
stratification period in US lakes may be extended by up to 67 days and surface waters may warm
by up to 5.2ºC. An earlier ice-off date can result in an earlier onset of stratification (De Stasio et
al. 1996), though some lakes may instead show an extended period of spring mixing
(Weyhenmeyer et al. 1999). Changes in the onset of stratification and surface-water warming
will likely influence thermocline depth by leading to either deeper and shallower thermoclines
(De Stasio et al. 1996). For lakes less than 500 ha in surface area, water transparency is likely the
primary determinant of lake thermostructure (Fee et al. 1996), with clear lakes having deeper
thermoclines (Schindler et al. 1996a). The influence of water transparency on the thermal
structure of small lakes partially explains why increases in surface-water temperatures are more
commonly documented with regional warming than increases in hypolimnetic-water
temperatures (Adrian et al. 2009). In coloured lakes, increases in air temperatures may have little
effect on hypolimnetic waters due to reduced solar heating of profundal waters (Snucins and
Gunn 2000).
Longer and/or stronger periods of thermal stratification can have important consequences
for chemical gradients in lakes. In oxic conditions, bacterial decomposition increases nutrient
concentrations in the hypolimnion. If oxygen is depleted and anoxic conditions occur, redox
reactions lead to the release of reduced iron and stored phosphate from sediments (see summaries
in Moss 2012 and Wetzel 2001). The internal loading of phosphorus can lead to hypolimnetic
nutrient concentrations greater than those of inflowing waters and can stimulate cyanobacterial
blooms (Pettersson 1998). Thus, if longer/stronger periods of thermal stratification increase the
length of summertime anoxic periods (Fang and Stefan 2009), the internal loading of phosphorus
from sediments may increase (Jankowski et al. 2006). In contrast, a longer stratification period
18
may lead to nutrient depletion in the epilimnion due to prolonged isolation from the nutrient-rich
hypolimnion (Winder and Sommer 2012).
Changes in precipitation regimes with regional climate change will also influence lakewater chemistry. Regional droughts can release previously stored SO42- and metals from
wetland/littoral sediments and lead to reacidification events (Arnott et al. 2001). In contrast,
increased precipitation can lead to an increased leaching of dissolved organic and inorganic
compounds from the watershed, as both stream and groundwater fluxes of anions, cations, and
dissolved organic matter are closely tied to precipitation events (Deyton et al. 2009; Eimers et al.
2008; Inamdar et al. 2009). Recently, studies have noted increases in DOC concentrations in
North American and European lakes (Evans et al. 2005; Monteith et al. 2007). The potential
causes of increases in DOC are currently debated and may be due to reduced sulphate deposition,
increases in temperature, and/or increases in precipitation/runoff (Erlandsson et al. 2008; Evans
et al. 2006; Monteith et al. 2007; Weyhenmeyer and Karlsson 2009). In general, there is a need
to improve understanding of the controls on DOC dynamics, because DOC trends can be sitespecific (Adrian et al. 2009) and differ between regions (Zhang et al. 2010).
This mini-review has highlighted some of the complex effects climate change may have
on physical/chemical characteristics of lakes. These complexities deepen when considering the
biological effects that climate change can have on aquatic ecosystem, as individual taxa are
influenced by the physical/chemical characteristics of a lake, as well as by their competitive
interactions and predator-prey relationships with other individuals and species. The section
below discusses some of the biological effects of climate change on lakes, focusing on
phytoplankton, zooplankton, and fish.
19
Biological effects of climate change on lakes
While lakes may appear superficially to be relatively simple ecosystems, a single
temperate lake may be home to thousands of co-existing and competing species of
phytoplankton, zooplankton, invertebrates, and vertebrates. A review of the impacts of climate
warming across so many biological families is evidently beyond the purview of this dissertation.
This section, therefore, focuses on the groups of most interest to this project: phytoplankton
(particularly the Chrysophyceae and Synurophyceae), zooplankton (particularly cladoceran
zooplankton) and, to a lesser extent, planktivorous and piscivorous fish.
Phytoplankton responses to climate change: Climate change is already influencing the
seasonality, productivity, and composition of algal assemblages. A number of studies have
shown that lakes are experiencing an earlier spring phytoplankton bloom (Adrian et al. 1999;
Weyhenmeyer et al. 1999; Winder and Schindler 2004), the timing of which is largely
determined by light regime (Sommer et al. 2012). Modelling work by De Senerpont Domis et al.
(2013) concluded that a reduced ice-covered period and an earlier onset of stratification with
regional warming can lead to advancements in the spring phytoplankton bloom in temperate
lakes. The effects of climate change on overall phytoplankton productivity may be more difficult
to predict. In a study of 205 small, oligotrophic lakes in Sweden along a latitudinal gradient,
Weyhenmeyer et al. (2013) found that, out of 7 physical/chemical variables, the duration of the
open-water (ice-free) season best explained phytoplankton biomass, suggesting that an
increasing ice-free period with regional warming may increase overall phytoplankton biomass.
Similarly, an increased internal loading of phosphorus due to longer periods of stratification may
stimulate phytoplankton production (Jeppesen et al. 2009). In contrast, increased water
20
temperatures with regional warming may increase grazing rates on phytoplankton, reducing
phytoplankton biomass (Winder and Sommer 2012).
Reynolds (1988) provides a useful framework for understanding phytoplankton ecology
and, by extension, phytoplankton responses to climate warming. Phytoplankton maximal growth
is limited by two factors: access to essential nutrients and access to light. Nutrient limitation can
be considered a form of stress on phytoplankton, and involuntary vertical translocations in the
water column by wind-induced mixing can be considered a form of disturbance. Phytoplankton
can therefore be grouped together based on their adaptive strategies to stress and disturbance as:
competitors (C-strategists; selected during periods of low stress & low disturbance), ruderals (Rstrategists; selected during periods of low stress & high disturbance), and stress-tolerant species
(S-strategists; selected during periods of high stress & low disturbance)3. Small, fast-growing
diatoms with a low sinking-rate can be considered C-strategists; large, heavily-silicified diatoms
that depend on turbulent mixing to maintain their position in the euphotic zone can be considered
R-strategists; and, mobile taxa adapted to low-nutrient conditions can be considered S-strategists.
The C-R-S groups should not be considered mutually exclusive; rather, taxa may show a variety
of adaptations that represent a gradation between the three strategies (see, for example, Olrik
1994).
The emphasis Reynolds (1988) puts on nutrients and turbulence (as well as its corollary,
stratification) can help us understand which algal groups may benefit from regional warming.
For example, as detailed above, a reduced period of ice-cover may increase the length or strength
3
Reynolds (1988) argues that there is no adaptive strategy for plankton in high-stress, high-disturbance conditions
(e.g. oligotrophic, polymictic sites). While this is likely an over generalization, such sites would most likely be
dominated by benthic algal taxa. In another classification scheme, Reynolds et al. (2002) lists 31 trait-differentiated
functional groups of algae. While useful for understanding why certain algal groups may be present in certain
habitats seasonally, the schematic is arguably less intuitive to use for hypothesis testing of algal responses to climate
warming than Reynolds (1988).
21
of thermal stratification (Vincent 2009). As such, we may expect that an increased period of
thermal stratification (i.e. a reduction in disturbance from mixing) may favour C- and/or Sstrategists over R-strategists. Indeed, a recent paleolimnological meta-analysis demonstrated that
recent pan-continental increases in small Cyclotella taxa (C-strategists) have been accompanied
by declines in large, heavily-silicified diatoms (R-strategists) (Rühland et al. 2008). Similarly,
increases in the abundance of colonial chrysophyte taxa (highly mobile taxa adapted to low
nutrient conditions, S-strategists4) documented in paleolimnological studies across North
America (Arseneau et al. 2011; Flear 2011; Ginn et al. 2010; Paterson et al. 2004) are likely
related, at least in part, to warming. In general, recent reviews have suggested that longer periods
of thermal stratification may result in an increased abundance of small taxa with a high surfaceto-volume ratio (and, therefore, lower sinking rates), buoyant taxa (such as cyanobacteria) and
motile taxa such as flagellated species (Winder and Sommer 2012).
Frameworks such as the one proposed by Reynolds (1988) are appealing because they are
intuitive and provide researchers with the ability to make predictions about how phytoplankton
species composition may change with climate warming. However, the difficulty remains that
climate warming influences physical, chemical, and biological factors in tandem and so
unexpected species responses may occur. For example, Shatwell et al. (2008) found that a shift in
the timing of spring succession events in Lake Mügglesse associated with regional warming led
to a an unexpected increase in cold-adapted diatom species. Typically, it is suggested that warmwater plankton species will increase with increases in air temperatures (Winder and Sommer
4
Olrik (1994) suggests that the scaled chrysophytes can be considered a transitional group between C- and Rstrategists and between C- and S-strategists. Similarly, Kristiansen (2005) suggests that most chrysophyte taxa
cannot be exclusively grouped into any of the 3 strategies. Presumably, some of this ambiguity arises from the
different seasonal distributions of chrysophyte taxa (Siver & Hamer 1992). However, as scaled chrysophytes are
motile taxa most commonly occurring in low nutrient conditions, the ‘S-strategy’ described by Reynolds (1988) may
be a suitable as a general description for the group. For additional information, see ‘Chrysophytes (the golden-brown
algae): general ecology and use as paleo-indicators’ in this introduction.
22
2012), though the study by Shatwell et al. (2008) emphasizes the point that exceptions to general
ecological assumptions occur. As an additional complicating factor, changes in DOC
concentration will also likely play an important role in controlling phytoplankton seasonality and
vertical structure, as DOC influences lake thermostructure (Fee et al. 1996), irradiance regimes
(Snucins and Gunn 2000), and the concentration of certain micronutrients (i.e. chelated iron,
Fuss et al. 2010). Part of the difficulty in predicting algal species responses to regional warming
arises from the fact that phytoplankton are subject to both bottom-up and top-down control. For
this reason, some of the effects of climate change on zooplankton and fish taxa are discussed
below.
Zooplankton responses to climate change: In North-temperate lakes, the spring phytoplankton
bloom is typically followed by a peak in zooplankton abundance, sometimes resulting in a clearwater phase (Sommer et al. 2012). The timing and magnitude of the spring phytoplankton bloom
determines the timing and size of the subsequent zooplankton bloom. For this reason, there has
been a considerable amount of research interest in a possible ‘mismatch’ between phytoplankton
and zooplankton with the advancement of the spring phytoplankton bloom (Durant et al. 2007;
Winder and Schindler 2004). Such a decoupling of phytoplankton and zooplankton peak
abundances has not been seen universally (Adrian et al. 2006; Preston and Rusak 2010). De
Senerpont Domis et al. (2007) argued that daphniid overwintering strategy would likely
determine whether or not a mismatch in spring phenology would occur, as taxa that overwinter
are less likely to suffer a mismatch than taxa that must recruit from the sedimentary eggbank.
Beyond the potential decoupling of phytoplankton and zooplankton phenology, changes
in water temperature will likely have an influence on zooplankton species assemblages. Different
zooplankton taxa have different physiological tolerances to temperature (Moore et al. 1996).
23
Taxa with a low thermal optimum and taxa with a thermal optimum just below their lethal
thermal limit may be more likely to undergo local extinctions with increases in surface
temperatures (De Senerpont Domis et al. 2007). Changes in temperature may also influence
zooplankton size structure. In a recent meta-analysis, Daufresne et al. (2009) summarized a set of
three biological rules that suggest that small-bodied taxa may benefit from regional warming5. In
a recent top-bottom study of 44 south-central Ontario lakes, Korosi et al. (2010) found that
Daphnia decreased in size from pre-industrial times to present and attributed that change to
either an effect of regional warming and/or an acid deposition. However, studies of cladoceran
zooplankton have not universally demonstrated the group’s sensitivity to climate change. For
example, in a transect study of arctic and subarctic lakes, Sweetman et al. (2008) found no
consistent response in cladoceran assemblages to regional warming from pre-industrial times to
present. Because zooplankton are subject to top-down control, the effects of climate change on
fish populations will likely influence the overall abundance of zooplankton, as well as the
composition of the group. The following section examines some of the possible effects of
climate change on populations of planktivorous/piscivorous fish.
Fish responses to climate change: Different planktivorous and piscivorous fish taxa have
different thermal preferences and are often described as cold-water, cool-water, or warm-water
taxa. There is a concern that surface-water warming will result in a reduction in cold-water
habitat for desirable taxa like lake trout (Salvelinus namaycush) and an expansion in warm-water
habitat for taxa like yellow perch (Perca flavescens) (Ficke et al. 2007; Stager and Thill 2010).
An increased hypoxic zone in lakes with longer periods of stratification would place cold-water
5
As cited in Daufresne et al. (2009), the three rules are:
Bergmann’s rule – small-bodied species are more common in warmer environments
James’ rule – within a species, small-sized organisms are generally found in warmer environments
Temperature-size rule – for ectotherms, individual body size decreases with increasing temperature
24
taxa under a “temperature-oxygen squeeze”, forcing cold-water taxa to abandon cold-water
refugia in the hypolimnion and move into warmer waters where they may experience an
increased incidence of stress and disease (Ficke et al. 2007).
While the loss of cold-water habitat is a key management concern, observational studies
and modelling studies have not consistently shown a reduction in cold-water habitat with
regional warming. For example, Snucins and Gunn (2000) showed an increase in cold-water
habitat during a warm year, likely because an early establishment of strong thermal gradients
inhibited mixing between surface and bottom waters. Similarly, a modelling study by De Stasio
et al. (1996) showed an increase in habitat available to warm, cool, and cold-water fish taxa with
a 2xCO2 scenario, though some modelling simulations did predict a reduction in habitat for coldwater taxa. Modelling studies by Fang and Stefan (2009) did show an increased potential for
fishkills during summer time anoxic events with a 2xCO2, but also showed a decreased potential
for winterkill due to a shortened ice-cover period. Overall, changes in the
composition/abundance of freshwater fishes will likely have an important top-down influence on
zooplankton/phytoplankton assemblages through trophic cascades that may be difficult to predict
(Sommer et al. 2012).
This mini-review has highlighted how climate change can affect the
physical/chemical/biological processes of lakes. These interactions can be complex, difficult to
predict, and are likely to increase in intensity over time. As lakes are increasingly influenced by
multiple anthropogenic stressors, these stressors may interact with climate change to produce
unexpected ‘ecological surprises’ (Paine et al. 1998). The following section discusses some of
the strategies ecologists can use to understand the effects of multiple stressors on lakes.
25
Understanding the effects of multiple stressors in lake environments
The introduction sections above discuss the effects of acid deposition and climate change
on lakes in isolation from other stressors. This is something of an oversimplification as a
growing number of lakes are increasingly influenced by multiple anthropogenic stressors (Keller
2009). Acidification, eutrophication, land-use change, invasive species, and climate change can
all have complex and interactive effects on aquatic ecosystems. Regional droughts can lead to
reacidification events and alter species assemblages (Arnott et al. 2001; Faulkenham et al. 2003).
Changes in land-use can change the quantity of DOC entering aquatic environments (Mattsson et
al. 2009), influencing lake thermal structure and chemical gradients, a trend that may be
exacerbated by regional warming. The fact that all of these stressors and more may co-occur in
lakes makes both understanding the environmental effects of multiple stressors and the
remediation of impacted sites difficult.
Ecologists have employed a variety of approaches to understand the influence of multiple
stressors on individual taxa and whole species assemblages. Microcosms can be used to test for
the influence of toxicants in isolation and in combination (e.g. Locke 1991), though such
experiments may be criticized for being an oversimplification of aquatic ecosystems. Mesocosm
experiments have also been used to document the synergistic and antagonistic effects of multiple
stressors (Christensen et al. 2006; Folt et al. 1999), though again, mesocosms are unable to
mimic the full complexity of freshwater ecosystems. More rarely, whole lake manipulations have
been used to elucidate the effects of multiple stressors on waterbodies (e.g. acidification and
climate change, Schindler et al. 1996b). Such large-scale manipulations are uncommon but
provide valuable insights into changes in the structure and function of aquatic ecosystems.
26
In general, a lack of long-term monitoring data for aquatic ecosystems can make
understanding the effect of multiple stressors on species assemblages difficult. Here,
paleolimnological studies can play an important role by providing the long-term perspective
needed for management and environmental stewardship (Wolfe et al. 2012). Often,
paleolimnologists employ a multi-proxy approach when examining the influence of multiple
stressors on aquatic ecosystems (Michelutti and Smol 2013), an effective but time-consuming
approach. In such studies, multiple paleolimnological indicator groups are reconstructed and the
different ecological sensitivities of the organism groups are used to reconstruct the influence of
different stressors in the waterbody (Quinlan et al. 2008). Alternatively, mathematical models
can be applied in an attempt to help distinguish between the effects of individual stressors on
species assemblages (Simpson and Anderson 2009). However, such approaches are complex,
require a deep understanding of statistics, and are constrained by the available data, the
correlational structure of underlying data, and many assumptions.
Arguably, the most intuitive approach to understand the effects of multiple stressors on
aquatic ecosystems is the reference site (or reference condition) approach. When impacted sites
are paired with reference sites, changes in species assemblages in the reference sites can be used
to understand regional environmental changes (for example, in climate; Yan et al. 1996). The
reference sites can be used to determine recovery targets for the impacted sites that account for
regional stressors (Yan et al. 2004, USEPA 2012). The reference lake concept has been applied
in a variety of ways in ecological studies, and so is examined in greater detail below.
Reference sites & reference conditions
The concept of a ‘reference site’ or ‘reference condition’ is frequently applied in
limnological (and paleolimnological) work but different studies apply the reference term in
27
different ways, resulting in confusion about its meaning. First, a reference site is not normally a
control site in the traditional use of the word. The word ‘control’ is normally used to designate
sites (or, more accurately, sampling units; Bailey et al. 2004) that are identical to impacted sites
in physical, chemical and/or biological characteristics, except for the presence of a stressor (or
stressors) under study (Downes et al. 2002). True controls are rare in most non-manipulative
limnological studies however (though, for an example see Keatley et al. 2008), and so
limnologists often rely on ‘reference’ frameworks for comparison between impacted and nonimpacted sites.
In its simplest application, a reference site is a site defined by the absence of a stressor of
interest (e.g. logging, Paterson et al. 1998; acidification, Jeziorski et al. 2013; eutrophication,
Hawryshyn et al. 2012; metal contamination, Ilyashuk et al. 2003). The main appeal of this
approach is that it is relatively easy to implement (Chessman et al. 2008) and can be effective at
detecting the influence of stressors on species assemblages, especially if executed in a balanced
study design (Persson 2008). The main weakness of this approach, however, is that these
reference sites may be influenced by environmental stressors other than the main stressor of
interest. The uncontrolled-for stressors may have a significant influence on the species
assemblages independent of the stressor of interest, affecting the study results.
An alternative, more rigorous reference approach is to designate sites as being in
‘reference’ condition based on the absence of multiple anthropogenic stressors (e.g.,
acidification, eutrophication, introduced species, road salt seepage, etc.). In this framework, the
reference sites represent the best condition that could be expected for impacted sites in the region
(Reynoldson et al. 1997). Minimally-disturbed reference sites possess a high degree of biotic
integrity (or ‘naturalness’) because they have been protected from multiple anthropogenic
28
stressors (Hamilton et al. 2010; Stoddard et al. 2006). A variety of methods can be used to
identify minimally-disturbed reference sites, including selection criteria and best professional
judgement, multivariate or multimetric approaches, and/or modelling or ordination techniques
(Hawkins et al. 2010; Soranno et al. 2011). The minimally-disturbed reference site approach is
commonly used in bioassessment studies. In such studies, the present-day ecological condition of
the reference sites is used to define recovery targets for impacted systems, often by using
modelling techniques (e.g., which taxa would be present at this impacted site if it were in
reference condition?, sensu Bailey et al. 2004). Whatever approach is used, the criteria used to
designate reference sites must be clearly defined; otherwise, the sites have little interpretative or
comparative value (Brucet et al. 2013).
Arguably, the ideal reference site would be a hypothetical ‘pristine’ site that has never
experienced any anthropogenic impact (Bennion et al. 2011). Unfortunately, the pervasive
influence of atmospheric deposition and climate change make it highly unlikely that such a site
exists. Alternatively, in paleolimnological studies a pre-disturbance time period, typically one
prior to the industrial revolution (e.g. ca. 1850, Battarbee et al. 2011), can be designated to serve
as a recovery target for lakes impacted by anthropogenic stressors. While this historical
perspective is critically important, especially for identifying if the present-day species
composition of impacted sites is significantly different from their historical norm, a historical
condition may not always be an appropriate recovery target. There is growing recognition that
the influence of multiple anthropogenic stressors on lakes can prevent impacted sites from
returning to their pre-disturbance, or historical, condition, even with management intervention
(Bennion et al. 2011; Hobbs et al. 2009). This phenomenon is often referred to as a “shifting
baseline”; specifically, the influence of an additional, non-remediated stressor (such as invasive
29
species, land-use change, or climate warming) can prevent recovering sites from returning to
their pre-disturbance state (or baseline condition). The best-attainable condition for that impacted
site may then become a state that is still ‘degraded’ when compared to its historical (or baseline)
condition. A growing number of recovery case studies have suggested that shifting baselines due
to climate change may already be occurring (Arseneau et al. 2011; Battarbee et al. 2012;
Greenaway et al. 2012b; Helliwell and Simpson 2010). Indeed, in a recent report the USEPA
called for long-term monitoring at regional reference sites to specifically monitor the effect of
climate warming on the sites and track shifting baseline conditions (USEPA 2012). In a shifting
baseline scenario, reference sites can be used to define recovery targets for impacted sites in the
same geographic area (Yan et al. 1996; Yan et al. 2004).
Regardless of the approach ecologists use to understand the influence of multiple
stressors on aquatic ecosystems and designate recovery targets, they must employ taxa in their
studies that are sensitive to the stressors being investigated. This dissertation focuses on the
scaled chrysophytes, an algal group that has frequently been employed in paleolimnological
studies of lake acidification. The following section discusses the ecology of chrysophytes and
their use as paleolimnological indicators.
Chrysophytes (the golden-brown algae) - general ecology & use as paleo-indicators
In general terms, chrysophyte algae have two flagella unequal in length, and possess
chlorophyll a and c, as well as accessory pigments (β-carotene and phytoxanthylls) which give
the algae their characteristic golden-brown colour (Kristiansen 2005). Morphologically,
chrysophytes are a diverse group, including motile and non-motile taxa, unicellular and colonial
taxa, and taxa with branched structures (Kristiansen 2005). The cells of chrysophytes can be
naked (unadorned), protected in loricas, or covered in siliceous scales and bristles (Sandgren
30
1995). This last group of chrysophytes, the scaled chrysophytes, are members of the classes
Synurophyceae and Chrysophyceae. The scaled chrysophytes are of particular interest to
paleolimnological studies because their siliceous scales/bristles preserve well in lake sediments
and possess morphologically distinct features that can be used for identification (Smol 1995).
Scaled chrysophytes are used as an indicator group in this investigation and so when the term
‘chrysophytes’ is used in this dissertation, it is primarily in reference to the scaled chrysophytes.
Members of the Synurophyceae and Chrysophyceae differ in morphology, size, and life
strategies. Scaled chrysophytes are euplanktonic and both unicellular (Mallomonas,
Paraphysomonas, Spiniferomonas) and colonial taxa (Synura, Chrysosphaerella,
Chrysodidymus) exist. For colonial taxa, the colonies range in size from 2 cells (Chrysodidymus
synuroides, Wujek and Wee 1983) to colonies > 200 µm in diameter (Synura uvella, Siver 2003).
Due to their size, large chrysophytes such as M. caudata and large synuran colonies are only
susceptible to predation by large herbivores (i.e. large Daphnia taxa) whereas smaller taxa can be
predated upon by small zooplankton (i.e. Bosmina/Eubosmina, copepods, and small daphniids)
(Sandgren and William 1995). Some scaled chrysophyte taxa such as Chrysosphaerella have
been reported to be phagotrophic (Holen and Boraas 1995). Scaled chrysophytes can reproduce
both asexually and sexually and all chrysophytes form siliceous resting stages known as
stomatocysts or statospores (or more simply as cysts) (Kristiansen 2005). Like their siliceous
scales, chrysophyte cysts preserve well in sediments and can be used in paleolimnological
studies, though often it is not known which cyst is derived from which species (Duff et al. 1995;
Wilkinson et al. 2002).
Scaled chrysophytes are known to occur on all of the continents except for Antarctica and
have been identified in both temperate and tropical locales (Kristiansen 2008). In general,
31
chrysophytes are considered characteristic of cool, oligotrophic, soft-water lakes, with low
conductivity and neutral-to-slightly acidic pH (Sandgren 1988). However, numerous
investigations have demonstrated that this is an oversimplification. For example, chrysophyte
taxa have been identified in warm, tropical locales (e.g. Northeast India, Saha and Wujek 1990)
and eutrophic waters (Kristiansen 1988). The fact that chrysophytes tend to be a dominant flora
at low-TP conditions does not appear to be because they are indifferent to TP concentrations. In
a survey of 91 lakes, Watson et al. (1997) demonstrated that the total abundance of chrysophytes
increased with TP concentrations when values were < 10 µg L-1, after which their summertime
biomass levelled out. The authors argued that chrysophytes may have been excluded from
higher-TP lakes due to increased herbivory by large Daphnia or because the high-TP lakes in the
dataset were primarily shallow, polymictic ponds that were unsuitable for large colonial
chrysophytes.
The seasonal occurrence of chrysophyte taxa is controlled by several factors including
water column stability, light conditions, temperature, nutrient availability, and predation
(Kristiansen 2005). Typically, chrysophytes are considered a major component of the spring
flora (Siver 1995). While true in many lakes, different unicellular and colonial chrysophyte taxa
have different seasonal occurrences in the water column; some taxa are known to be tightly tied
to the summer months (e.g. Mallomonas punctifera, Siver and Hamer 1992) or to form blooms
under-ice (e.g. Synura petersenii, Watson et al. 2001). During periods of thermal stratification,
large stable metalimnetic populations of both unicellular (M. acaroides, M. caudata, M.
crassisquama) and colonial chrysophyte taxa (Synura, Chrysosphaerella) can occur near or
below the thermocline (Nicholls 1995; Sandgren 1988; Siver 2003). Such blooms have been
associated with taste and odour issues in lakes and reservoirs (e.g. S. petersenii blooms, Nicholls
32
and Gerrath 1985; Watson et al. 2001) and are disrupted with the deterioration of thermal
stratification at the fall overturn (Fee 1976). The causes of metalimnetic chrysophyte blooms are
not well understood. While chrysophytes can persist in low-light, low-temperature conditions,
such conditions may not be optimal for their growth (Healey 1983). Nutrient conditions may be
more favourable at or below the thermocline due to increased nutrient concentration from
decomposition (Sandgren 1988). Additionally, the metalimnetic position of the blooms may
provide a refuge from predation by large Daphnia (Nicholls 1995) or protection from harmful
UV-B radiation (Vinebrooke et al. 2002). As such, during extended periods of thermal
stratification, flagellated algae like chrysophytes may have a competitive advantage over nonflagellated taxa like diatoms because of their ability to seek out optimal conditions in the water
column (Winder and Sommer 2012). Furthermore, at an assemblage level, large colonial
chrysophytes may have an advantage over unicellular taxa because large flagellates are often
superior swimmers (i.e. have faster swimming speeds and show larger diurnal migrations,
Sommer 1988).
As proxies of past environmental change, the occurrence of scaled-chrysophytes along
environmental gradients has been well-characterized. pH is found consistently to an important
chemical variable structuring chrysophyte assemblages in different regions (e.g. New York:
Cumming et al. 1992a, Connecticut: Siver and Hamer 1989, the US Northeast: Hyatt et al. 2010,
Ontario: Dixit et al. 1988; Paterson et al. 2001, maritime Canada: Ginn et al. 2010).
Chrysophytes are also sensitive to specific conductance (Siver and Hamer 1989), temperature
(Siver and Hamer 1989; Siver and Lott 2012), and lake trophic status (Siver and Marsicano
1996). As indicators, they have primarily been used in paleolimnological studies of acidification
(Cumming et al. 1992b; Dixit et al. 1988), though other applications include salinization (Zeeb
33
and Smol 1991), eutrophication (Siver and Marsicano 1996), and seasonality reconstructions
(Pla-Rabes and Catalan 2011). Scaled-chrysophytes may respond earlier than diatoms to
acidification (Cumming et al. 1992b; Cumming et al. 1994; Dixit et al. 2002) and chemical
recovery (Arseneau et al. 2011; Greenaway et al. 2012b) because they are euplanktonic and
because some taxa are common in the spring and so may be more sensitive to episodic
acidification (e.g. Mallomonas hindonii, Cumming et al. 1992a).
Scaled-chrysophytes were chosen as a paleo-indicator in this investigation primarily
because of their sensitivity to pH and because they have previously been shown to be abundant
in the surface waters of the Adirondack Park, the study location of this dissertation. The
following section provides information on the Adirondack Park, particularly the influence of acid
deposition and climate change on waterbodies in the region.
The Adirondack Park
The Adirondack Park is the largest protected area in the contiguous United States at ~6
million acres6. Unlike many protected areas in North America, the Adirondack Park follows a
mixed-use model, including both publically and privately-owned lands. Approximately 51% of
the land in the Adirondack Park is privately owned, 43% is state land, and 6% is water (APA
2011a). Based on 2010 census data, the Adirondack Park Agency (APA) estimated that
approximately 130,000 people live within the Adirondack Park (Barge 2011), though millions of
seasonal tourists visit the region each year. Areas within the Adirondack Park have different
land-use classifications and are subject to different levels of protection. The classifications range
from intensive-use areas (e.g. the Whiteface ski area near Lake Placid) to wild forest and
wilderness areas (the most restrictive land-use designation). While recreational activities like
6
The Adirondack Ecological Zone (AEZ) extends beyond the boundaries of the Adirondack Park and is defined as
the areas greater than 305 m (1000 ft) in elevation in the Adirondack Mountains (Krester et al. 1989).
34
hiking and canoeing are permitted in wild forest and wilderness areas, logging and development
are not (APA 2011b).
The climate of the Adirondacks can be described as cool continental. Average monthly
temperatures range from approximately -7ºC in the winter to 21-27ºC in the summer (Jenkins
and Keal 2004). Precipitation is evenly distributed throughout the year, with precipitation
amounts of ~3-4 inches (~76-100 mm) per month (Jenkins and Keal 2004). Precipitation falls as
snow or as snow mixed with rain from November until April and the snowpack accumulates
from December until spring (Driscoll et al. 1991). The growing season is short, approximately
100-120 days per year (Jenkins and Keal 2004). Northern hardwoods comprise approximately
50% of Adirondack forests and are dominated by yellow birch (Betula alieghaniensis), beech
(Fagus grandifolia), and sugar maple (Acer saccharum). Another 25% of Adirondack forests are
mixed forests of hardwoods and softwoods, including red spruce (Picea rubens), balsam fir
(Abies balsamea), eastern hemlock (Tsuga canadensis), and white pine (Pinus strobus). Ten
percent of Adirondack forests are coniferous stands, and approximately 5% of all Adirondack
vegetation are wetlands (Driscoll et al. 1991).
The Adirondack region receives high levels of acid deposition. Major wind patterns in the
United States move from west to east, and so SO2 and NOx emissions from the industrial
Midwest are transported to the Adirondacks (Driscoll et al. 1991). A precipitation gradient exists
in the Adirondack Mountains, with the highest amounts of the most acidic precipitation falling in
the Southwest (Ito et al. 2002). The sensitivity of Adirondack lakes to acid precipitation is
caused by the fact that the region is primarily underlain by Ca-poor granitic bedrock, that soils in
the region are primarily shallow acidic Spodosols, and that the steep slopes in the region
contribute to waterbodies having a relatively short water residence time (< 1 year), limiting the
35
potential for in-lake processes to neutralize incoming acids (Driscoll et al. 1991). A 1980s survey
of 1,469 Adirondack lakes found that 27% were chronically acidic and 24% had pH values less
than 5 (Krester et al. 1989). The majority of lakes with a pH < 5 were located in the
Southwestern Adirondacks whereas lakes with a pH > 6 were more evenly dispersed throughout
the region (Gallagher and Baker 1990). Lakes insensitive to acid deposition (> 50 μeq L-1) were
either associated with thick glacial till deposits or were carbonate-influenced (i.e. associated with
CaCO3 bearing rock in till or exposed bedrock) (Jenkins et al. 2007).
Since the implementation of the 1990 CAAA, Adirondack lakes have begun to undergo
chemical recovery from acidification. Long-term monitoring data have documented a decrease in
SO42-, NO3-, and inorganic monomeric aluminum in surface waters since the 1990s, as well as
increases in pH and ANC (Driscoll et al. 2007; Waller et al. 2012). A recent analysis of longterm monitoring data also found that many Adirondack lakes are experiencing increases in DOC,
but that the trend is not universally documented (Lawrence et al. 2013). Modelling studies
suggest that chemical recovery will be a protracted process in the Adirondacks, as many
Adirondack lakes will still have ANC values less than 50 μeq L-1 by 2050 (Chen and Driscoll
2005). Without further reductions in sulphate deposition, many Adirondack watersheds will
experience deposition levels greater than target loads for maintaining lake-water ANC > 50 μeq
L-1 by 2100 (Sullivan et al. 2012).
While lakes in the Adirondacks may be undergoing chemical recovery from acidification,
the region is also experiencing the effects of climate change. Annual air temperatures in the
Adirondacks have increased at a rate of 1.3ºC per 100 years since 1900 (Jenkins 2010). The rate
of warming in the Adirondacks is somewhat higher than the rate documented for the Northeast as
a whole (0.8ºC per 100 years) (Hayhoe et al. 2007). In both the Adirondacks and the Northeast as
36
a whole, the rate of temperature change has increased since approximately 1970, with winter
temperatures increasing at a greater rate than summer ones (Hayhoe et al. 2007; Jenkins 2010).
However, in an analysis of monthly temperature data from 1975-2005, Stager et al. (2009) found
the June and September temperatures have increased significantly over the time period, though
many other months showed non-significant increases. Stager et al. (2009) also found no
statistically significant trends in total monthly precipitation trends in Adirondack records from
1975-2005, though precipitation levels in the Northeast are expected to increase by 7-14% by
2070-2099 (Hayhoe et al. 2007). Total snowfall (as snow-water-equivalent) in the Northeast has
decreased at a rate of ~3.5 mm per decade from 1970-2000 (Hayhoe et al. 2007) and monitoring
records in the Adirondacks either showed a significant decrease in snowfall from 1948-2005, or
no significant change (Jenkins 2010).
Changes in temperature/precipitation regimes in the Northeast have already been
associated with an earlier timing of peak spring stream flows, earlier ice-out dates, and an earlier
timing of first leafing and flower blooming (Hayhoe et al. 2007; Hodgkins et al. 2002) and in the
Adirondacks, regional warming has been associated with a decreased ice-covered period on lakes
(Beier et al. 2012; Stager et al. 2009). The effects of climate change are likely to have a profound
influence on Adirondack flora and fauna. For example, Momen et al. (2006) noted an increase in
productivity in seven Adirondack lakes (measured as chlorophyll a) from 1994 to 2003. Though
the authors attributed this increase in productivity to chemical recovery from acidification,
paleolimnological studies have attributed recent increases in chlorophyll a concentration to
regional warming (Michelutti et al. 2005), a phenomenon not examined by Momen et al. (2006).
Importantly, the Adirondack region represents the southernmost range limit of many boreal
species, many of which may be extirpated from the region with significant warming (Jenkins
37
2010). Regional warming may result in the loss of cold-water fish taxa, rare boreal habitats
(high-elevation Krummholtz and alpine-tundra areas), and winter recreation opportunities in the
region (skiing, snowmobiling) (Frumhoff et al. 2007). Climate change will also affect how lake
ecosystems are undergoing biological recovery from acidification and may lead to novel (and
potentially undesirable) aquatic species communities in recovered systems (Arseneau et al.
2011).
Summary
This literature review has highlighted the complex effects acid deposition and climate
change have on aquatic ecosystems. Ecologists have a variety of methods available to them to
understand the influence of multiple stressors on species assemblages, one of the most intuitive
being the use of reference systems. In a recent review, Gray and Arnott (2009) argued that the
ideal recovery study would include the use of reference sites, historical data, and a temporal
analysis. This is precisely the method followed in this investigation. First, a regional set of
minimally-disturbed reference sites were identified for the Adirondack Park and critically
evaluated (Chapter 3). Second, the reference sites are used in a ‘top-bottom’ paleolimnological
study of scaled chrysophyte assemblages to identify if regional stressors have led to a significant
shift in species assemblages since pre-1900, creating a ‘shifting baseline’ (Chapter 4). Lastly,
chrysophyte assemblages are compared between a subset of reference lakes and two Adirondack
lakes recovering from acidification. Paleolimnological techniques are used to identify if
biological recovery from acidification is occurring and reference sites are used to define
reasonable recovery endpoints for the acid-impacted systems (Chapter 5). Overall, this project
will improve our understanding of biological recovery from acidification, the influence of
multiple stressors on aquatic ecosystems and, perhaps most importantly, highlight the utility of
38
pairing a rigorous regional reference site approach commonly used in bioassessment studies with
the valuable long-term perspective of paleolimnology. In Chapter 6, the general implications of
this project are discussed, along with some possible directions for future research.
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56
CHAPTER 3
TRACKING SHIFTING BASELINE CONDITIONS DUE TO CLIMATE CHANGE IN
LAKES: A NOVEL USE OF REGIONAL REFERENCE SITES AND
PALEOLIMNOLOGY
Abstract
Climate warming may prevent lakes recovering from acidification or eutrophication from
returning to their pre-disturbance state, creating the ‘shifting baseline’ phenomenon. To define
recovery targets in a multiple stressor environment, researchers must: 1) have knowledge of predisturbance conditions; and 2) understand how stressors/recovery processes influence species
assemblages, a difficult task as long-term monitoring data are rare. Paleolimnological studies of
minimally-disturbed reference sites can address both problems. Paleolimnological studies use the
physical, chemical, and biological properties of sediments to infer how aquatic environments
have changed over time. By reconstructing past biotic assemblages of reference lakes, we can
infer if/how climate warming is influencing species assemblages, which can help characterize the
shifting baseline and define recovery targets. In this manuscript, we present a framework for
using paleolimnological techniques in a regional reference site approach and apply it in the
Adirondack region (NY, USA). Simple selection criteria were used to identify reference lakes
that were relatively unimpacted by acidification, eutrophication, road salt seepage, and nonnative piscivore introductions (31 out of 1469 lakes). Ordination methods found that the
reference lakes were representative of 24-36% of the chemical/physical characteristics of
Adirondack lakes. GIS methods, historic maps, and historic fisheries information were used to
quantify historic disturbance and separate the reference lakes into classes. Only 9 lakes
experienced little or no historic watershed disturbance, highlighting the danger of assuming a
site’s condition has been static over time. As the effects of climate change become more
57
prevalent, the historical perspective offered by paleolimnological studies will become
increasingly important for management.
Introduction
Limits on sulphur dioxide emissions implemented via ‘clean air’ legislation in North
America and Europe have led to significant declines in acid deposition. In some regions,
waterbodies have begun to undergo chemical recovery from acidification, with long-term
increases acid neutralizing capacity (ANC) and decreases in H+ (Garmo et al. 2014; Skjelkvåle et
al. 2005). Given this chemical recovery, there is now an acute research need to characterize
biological recovery from acidification. However, researchers tasked with identifying biological
recovery are faced with two major problems: 1) a lack of long-term monitoring data; and 2) the
difficulty of assessing recovery in systems influenced by multiple stressors.
Paleolimnological methods can be used to identify biological recovery when there is a
lack of long-term monitoring data. Paleolimnological studies use the chemical, physical, and
biological properties of lake sediments to infer the environmental histories of lakes (Smol 2008).
Commonly, researchers will identify the remains of aquatic organisms in radiometrically-dated
sediments and reconstruct how the species assemblages have responded to environmental change
over time. This approach has been used widely to identify the onset of acidification (Battarbee
1990; Cumming et al. 1994) and, more recently, biological recovery from acidification
(Arseneau et al. 2011; Hesthagen et al. 2011).
Interpretations of both paleolimnological studies and studies that rely solely on modern
limnological sampling are difficult when systems are impacted by multiple anthropogenic
stressors (e.g. acidification, eutrophication, landscape alteration, climate warming), an
increasingly common problem (Keller 2009). Researchers working on lakes that are influenced
58
by multiple stressors may face challenges in attributing the response of species or assemblages to
a single stressor. Climate change is of particular concern in this regard, given its regional nature.
Climate warming may result in particular “syndromes” of species change in lakes, such as an
increase in the abundance of warm-water taxa, small planktonic algae, or flagellated/buoyant
algae as water temperature increases and/or longer stratification periods occur with regional
warming (Winder and Sommer 2012). In such a scenario, how are researchers to distinguish
between a species’ response to a novel stressor such as climate change and, for example, the
recovery of algal or invertebrate assemblages from acid deposition? Furthermore, climate change
acting either alone or in conjunction with other stressors may prevent recovering systems from
returning to their pre-disturbance state, even with remediative action, creating a phenomenon
known as the “shifting baseline” (Bennion et al. 2011). When the baseline of an undisturbed lake
changes, it becomes increasingly difficult to define a reasonable recovery target for an impacted
site.
The goal of this manuscript is to provide a framework for using paleolimnological
techniques in regional studies of reference lakes (sites) to track shifting baseline conditions and
define recovery targets for impacted sites. In a recent report on environmental assessment
practices in the United States, the USEPA highlighted the problem that climate change poses to
bioassessment programs and called for long-term monitoring of regional reference sites (i.e. sites
minimally disturbed by anthropogenic stressors) so that they could serve as ‘sentinels’ of climate
warming and track shifting baseline conditions (USEPA 2012). However, long-term monitoring
data of aquatic ecosystems, where they exist, typically cover short time periods (< 50 years) and
may not be suited for studies of climate change. Here, paleolimnological studies can play an
important role. Using minimally-disturbed reference sites in paleolimnological studies can allow
59
researchers to identify a species’ response to climate change independent of other stressors and
identify if a ‘shifting baseline’ effect has occurred. An understanding of the shifting baseline can,
in turn, provide the necessary information to define recovery targets for impacted systems.
Unfortunately, most environmental assessment studies lack the historical perspective offered by
paleolimnological work (Stoddard et al. 2006). Similarly, while paleolimnological studies of
environmental stressors often include a small number of reference sites, the regionally-based
reference site approach that is often used in bioassessment studies (e.g. Bailey et al. 2004) has
been less commonly applied in paleolimnological investigations (though for an early example,
see Battarbee et al. 2002). As such, in this manuscript we will:
1) provide a framework for pairing paleolimnological techniques with a regional
reference site approach and;
2) describe a case study in the Adirondack Park (NY, USA) where a set of 31 minimallydisturbed reference sites were identified and assessed for use in a regional
paleolimnological study of recent climate warming and biological recovery from
acidification.
The methods used in this investigation could be easily modified and applied in other protected
areas to fulfill a critical research and management objective (Baron et al. 2009). To that end, we
will elaborate on the rationale behind the proposed framework and the methods used, while
future manuscripts will focus on the results of our paleolimnological investigations of changes in
algal and invertebrate assemblages in these reference systems.
Methods
A conceptual framework for using paleolimnological techniques in regional reference site
studies
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In the following section we outline some key concepts that should be considered when
designing a paleolimnological study of biological recovery that utilizes regional reference sites.
These concepts include: what are ‘good’ reference sites and how are they identified; the useful
role of reference sites in studies of multiple stressors and in identifying reasonable recovery
targets for impacted systems; and why paleo-studies should be regional in nature, rather than
relying on a small number of reference sites. Subsequently, we present a case study that
identified and assessed a group of reference lakes in the Adirondack region (NY, USA).
Reference sites in limnological vs. paleolimnological studies – a distinction: Before
elaborating on our framework, it is useful to distinguish between the use of reference sites in
studies that rely solely on modern limnological sampling and those that use paleolimnological
techniques. In limnological studies, the regional reference site approach (or reference condition
approach, sensu Bailey et al. 2004) involves the identification of sites minimally-disturbed by
anthropogenic stressors (reference sites) whose present-day ecological condition is used to define
recovery targets for impacted sites, often by using modelling techniques (i.e., what species would
be present at this site if it were in reference condition?). The modelling approach may be
necessary because the pre-disturbance condition of an impacted site is not usually known. In
contrast, paleolimnological studies can reconstruct the pre-disturbance or historic species
assemblages (e.g., of diatoms, chrysophytes, invertebrates) of an impacted site. The purpose of
including reference sites in paleolimnological studies is to track regional changes in species
composition over time which are occurring independent of a stressor of interest (e.g.,
acidification or eutrophication); information which, in turn, can be used to define recovery
targets for impacted sites.
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What makes a ‘good’ reference site in paleolimnological studies?: Given the pervasive
influence of environmental stressors such as acid deposition or climate change, it is unlikely that
researchers will be able to identify ‘pristine’ reference sites (i.e., sites that have never
experienced anthropogenic influence, Bennion et al. 2011). Nevertheless, the reference sites
should possess a high degree of biotic integrity (or, ‘naturalness’), in that they have been largely
protected from human impact over the time period being investigated (Hamilton et al. 2010;
Stoddard et al. 2006). To be useful in research on ecosystem recovery, the reference sites must
have been protected from the main stressor of interest (e.g., acidification or eutrophication), and,
ideally, other stressors common in the study region (e.g., watershed development, introduced
species, road salt application, Bailey et al. 2004).
How can reference sites be identified?: The methods used to identify reference sites will
depend in part on the research question being addressed. Regardless of the question, however,
the selection criteria used to designate reference sites must be clearly defined. These criteria may
differ among regions, depending on levels of human impact or natural variability in variables of
interest (e.g. total phosphorus, Cl- concentration; Herlihy et al. 2008). Methods used to identify
reference sites include the use of selection criteria or ‘best professional judgement’, multivariate
or multimetric techniques, or a hybrid approach using both (see recent reviews by Hawkins et al.
2010; Soranno et al. 2011). Ultimately, the selection process used will depend on the availability
of information on the candidate sites. Paleolimnological studies that use data unsuited to
quantification (e.g., descriptive regional histories) may need to use methods that include a
reliance on selection criteria and/or best professional judgement.
How can reference sites be used to understand the effects of multiple stressors?: The
implicit null hypothesis of paleolimnological studies is one of ‘no change over time’ or, more
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specifically, no change exceeding the natural variability of the system at a decadal/sub-decadal
temporal resolution (Smol 2008). When paleolimnologists identify a species change that appears
to be outside natural variability, they must attribute that change to a mechanism based on their
knowledge of the study site/region and the ecology of the taxa they have identified in the paleorecord. Attributing species response to a particular mechanism can be difficult in ecosystems
affected by multiple anthropogenic stressors (Quinlan et al. 2008). The growing influence of
climate change on the physical, chemical, and biological properties of lakes adds complexities to
a system’s response to other environmental stressors (Smol 2010). Studies of climate change
typically involve a two-step process: a change must be first detected in a particular system and
then that change must be attributed to a causal mechanism (Hegerl et al. 2007). By using
reference sites, the paleolimnologist is following one of the basic principles of good scientific
design; namely that researchers should attempt to simplify their systems so that they can examine
the effect(s) of stressors in isolation (Jager and Loonman 1995). The careful use of reference
systems may allow researchers to identify shifts in species composition due to regional climate
change independent of other stressors, an inference that can be strengthened by comparing
paleolimnological data with long-term climate monitoring data (Battarbee et al. 2012). Thus, in
a multiple stressor environment, using reference sites to track long-term species response to
regional climate change can help identify whether or not a ‘shifting baseline’ effect is likely
occurring in a region.
How can reference sites be used to define recovery targets?: In paleolimnological studies
of environmental stressors, the recovery target for a lake is usually defined as the pre-disturbance
condition of that lake (e.g. ca. 1850, Battarbee et al. 2011). However, a lake responding to
multiple stressors over the recovery period is unlikely to return to its pre-disturbance state, even
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with management intervention (Bennion et al. 2011; Hobbs et al. 2009). In this case, a
reasonable recovery target may be defined by comparing impacted species assemblages to those
of suitable reference sites (i.e. lakes with higher pH/ANC values that have not experienced
chronic acidification, See Figure 3.1). For example, if the reference sites document a recent
increase in warm-water taxa or small planktonic algae, changes predicted to occur with regional
climate warming (Winder and Sommer 2012), a recovery target for an impacted site should
reflect the regional shifts in species composition. Effectively, the recovery target would be
recovery to a “novel” state that is similar to suitable reference systems (Hobbs et al. 2006). The
‘recovered’ species assemblages of the impacted site would contain elements of the predisturbance assemblage such as previously abundant acid-sensitive taxa and ‘novel’ additions
such as an increased abundance of warm-water taxa. Note that in this framework, the reference
sites do not serve as ‘controls’ for the impacted sites in the traditional sense (i.e. sites with
identical species composition, morphology; Downes et al. 2002) but should, nevertheless, be
similar to the impacted site(s) in general characteristics (i.e. surface area, depth, watershed
characterstics) so that meaningful comparisons among the sites can be made.
Why is it important that paleolimnological studies of reference sites be ‘regional’ in
nature?: Though paleolimnological studies of environmental stressors may use reference sites
(e.g. Alahuhta et al. 2009; Charles et al. 1990; Jeziorski et al. 2013), such studies often have
fewer reference sites than impacted sites, sometimes relying on only a single reference site. Such
a study design is understandable given the time-intensive nature of paleolimnological sampling.
However, a limitation in this approach is the possibility that the chosen reference site(s) is/are
not characteristic of the region (Downes 2010). Thus, rather than identifying a common regional
response to climate warming, the reference site may be responding to a local phenomenon such
64
as an unknown introduced species or watershed disturbance. Similarly, paleolimnological studies
have a robust history of using remote, minimally-disturbed sites in arctic or alpine areas to track
the influence of environmental stressors like climate change (Catalan et al. 2013). While
important, these sites are by design remote from populated areas and so are not ideal for defining
recovery targets (e.g. alpine areas are under different climate-controls than more populated,
temperate lowland areas).
An alternative approach to using a small number of reference sites in paleolimnological
studies or to using a number of remote sites would be to combine two different types of
paleolimnological approaches: i) a ‘top-bottom’ approach (Cumming et al. 1992) and ii) a
traditional ‘down-core’ paleolimnological study embedded within an anthropogenicallyimpacted area (See Figure 3.2). In ‘top-bottom’ studies, researchers only analyze a present-day
sediment sample (the ‘top’ of a sediment core, normally between 0-1 cm) and a sample from a
set depth in the core that is taken to represent pre-industrial conditions (e.g., 20 cm to represent
~1850). The top-bottom approach is similar to a BACI design, in that it compares a time period
minimally influenced by regional stressors like climate change (the ‘bottom’ sample) with a
recent time where such stressors may be more important (the ‘top’ sample). However, preindustrial conditions may not represent a “pristine” time-period and so researchers should
attempt to characterize the early disturbance history of the sites (e.g., land-clearance/logging
history), if possible. The top-bottom approach gives a ‘snap-shot’ of change throughout the
region. Because the temporal resolution at each study site is low (i.e., 2 samples), researchers can
analyze data from a much larger number of sites than is possible in a traditional ‘down-core’
paleolimnological study. With top-bottom data, researchers can test for a significant difference in
species composition from pre-industrial times to present and, depending on the study design,
65
attribute changes to stressors such as regional warming (Enache et al. 2011). A variety of
statistical frameworks can be used to analyze top-bottom paleo-data, including univariate
methods (Wilcoxon Sign-Rank test, e.g., Korosi and Smol 2012) or multivariate methods (e.g.
ordination techniques, e.g., Hyatt et al. 2010; or Analysis of Similarities, e.g., DeSellas et al.
2011). Analyses can be performed on the entire paleo-species assemblage, on key individual
taxa, or, if transfer functions are applied to the paleo-species data, on reconstructed inferred-pH
(or TP, etc.) values (Juggins and Birks 2012). However, interpretations of such mathematical
reconstructions must be rooted in an understanding of the biological assemblage as a whole, as
inferred-water chemistry values may ‘mask’ important changes in individual taxa of
management concern (e.g., increases in taxa that cause taste and odour issues, Paterson et al.
2004).
While the top-bottom approach is effective at building a regional picture of species
change from pre-industrial times to present, the method provides limited information for each
individual site. If one of the goals of the reference site approach is to use the reference sites to
define recovery targets for impacted sites, more detailed temporal information may be required.
Thus, a complementary approach to the top-bottom study is to compare a subset of the reference
lakes with a set of impacted lakes in a down-core study. In a down-core study, many sediment
sub-samples (e.g., n > 20) are analyzed per core, providing a more detailed temporal
understanding of changes in species assemblages over time at a site. The down-core study would
allow researchers to compare the timing of species changes between the reference and impacted
sites, as well as compare paleolimnological data to long-term monitoring data, if available.
Again, this comparison can be done using univariate methods (e.g., correlations between key
taxa and long-term chemical/climatic monitoring data, Battarbee et al. 2012), multivariate
66
methods (e.g., plotting both impacted and reference sites in ordination space, Yan et al. 2004) or
a combination of both. Ultimately, researchers must be conscious of scale in their study design
(Allen and Hoekstra 1992). The effects of climate change are regional in nature but those effects
are filtered at the local scale by a given watershed and its waterbody and are reflected in the
physical, chemical, and/or biological characteristics of the system (Adrian et al. 2009). A
reference lake study design should therefore allow for an understanding of both broad regional
trends and lake-specific responses, and combining the top-bottom and down-core approaches can
provide both. While it is difficult to determine the minimum number of sites required in a
regional study, given the variation in lake physical/chemical characteristics, a minimum of at
least 25 reference sites is likely needed (Bailey et al. 2004).
Implementing a regional paleolimnological study of reference sites in the Adirondacks (NY,
USA)
The study region: The Adirondack Park presents an excellent opportunity to demonstrate how
paleolimnological techniques can be used to investigate biological recovery from acidification
within a regional reference-site framework. The Adirondack region is a mountainous area that is
a largely forested mosaic of public and private lands (Liroff 1981). Adirondack lakes have been
characterized extensively in regional surveys (e.g., Krester et al. 1989) and long-term chemical
monitoring data are available for many sites (Roy et al. 2011). The region has been heavily
impacted by acid deposition (Driscoll et al. 1991) and the acidification history of the region has
been well-documented in previous paleolimnological studies (Cumming et al. 1992b; Cumming
et al. 1994). Following declines in acid deposition, many lakes in the region are now undergoing
chemical recovery from acidification (Driscoll et al. 2007; Waller et al. 2012). However, while
chemical recovery may be progressing, the Adirondacks are also experiencing the influence of
67
regional warming. Mean annual temperature has increased by a rate of 1.3°C per 100 years, a
warming trend that has been most pronounced since the 1970s (Jenkins 2010).
Project goals: Following the principles outlined above, the goal of this investigation was to
identify a set of reference lakes to be used in a two-part paleolimnological analysis that included
a top-bottom study and a down-core study. The purpose of the top-bottom study was to identify
if changes in species assemblages were occurring in Adirondack lakes independent of several
common stressors. Subsequently, a subset of the reference lakes were compared with a set of
Adirondack lakes that were known to have acidified by acid deposition and are now experiencing
chemical recovery from acidification (based on long-term monitoring data) in a down-core
paleolimnological study. By comparing long-term trends in reference sites and impacted sites,
we sought to identify if a shifting baseline in background conditions was occurring and, if so, use
the reference sites to define recovery targets for the impacted sites.
Reference site selection criteria: This study builds on the work of Stager and Sanger (2003) who
introduced the concept of “Heritage Lakes”, or minimally-disturbed lakes in the Adirondacks.
Our primary interest was to identify a set of Adirondack lakes that had been largely protected
from several stressors common in the Adirondack Park, namely: acidification, eutrophication
from watershed/shoreline development, road salt seepage, and the introduction of non-native
piscivores. The reference lakes were identified from an online database of 1,469 Adirondack
lakes sampled from 1984-1987 as part of the Adirondack Lakes Survey (ALS) (Krester et al.
1989) (database available at: www.adirondacklakessurvey.org). The ALS used a modified
random sampling design. In 1984-1986, the survey focused on three of the five large watersheds
in the Adirondack Ecological Zone (AEZ), a region that extends beyond the boundary of the
Adirondack Park and covers the area that falls within the 305 m (1000 ft.) contour line of the
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Adirondack Uplands. The three watersheds contained 1,862 lakes greater than 0.2 ha in surface
area, of which 1,200 lakes were sampled. Sixty percent of the lakes were chosen randomly and
then additional waterbodies within 2 km of the randomly selected sites were sampled. In 1987, a
pure randomized design was used to sample a minimum of 250 waters in the remaining two
watersheds. In total, the 1,469 lakes sampled in the ALS survey are considered a representative
sample of the 2,759 lakes greater than 0.2 ha that fall within the AEZ.
The six selection criteria used to identify reference lakes in the ALS are listed below. The
criteria are based on the findings of previous limnological or paleolimnological studies of the
Adirondack Park and elsewhere.
1. Public access: The search was restricted to lakes with public access. Lakes that were not
accessible via hiking trails and lakes with excessively long or difficult hikes were also
excluded.
2. 1980s pH > 6.5; 1980s ANC > 50 μeq L-1: These paired criteria were used to exclude
lakes that were likely to have acidified. Cumming et al. (1992) demonstrated that few
Adirondack lakes with a pH > 6.5 in the late 1970s and early 1980s underwent significant
chronic acidification. Additionally, an acid neutralizing capacity (ANC) > 50 μeq L-1
largely protects acid-sensitive lakes from episodic acidification (Driscoll et al. 2001).7
3. Watershed & shoreline development < 5%: These paired criteria were used to exclude
lakes that were likely to have experienced eutrophication due to development along their
shorelines or in their watersheds. A 5% development cut-off was used so that Adirondack
lakes with small camps would not be excluded.
4. Cl- < 2 mg L-1: In the ALS, lakes were considered ‘salt impacted’ based on a cut-off
criterion of 20 µmol L-1 (~0.71 mg L-1) (Newton and Driscoll 1990). We applied a less
restrictive selection criterion, allowing lakes with a modest concentration of salt to be
included in the survey. In general, it was found that lakes excluded from the reference set
7
Middle Branch Lake was included in the survey, despite having an ANC < 50 μeq L-1 in the ALS survey (pH 6.9;
ANC 18.4 μeq L-1). Inclusion was based on the fact that the lake had a pH > 6.5 at the time of sampling and longterm monitoring data from Middle Branch Lake suggests that the lake has generally maintained a pH > 6.0 since
1993, with summertime pH values frequently reaching > 6.5 (Roy et al. 2011).
69
based on high Cl- concentration were also excluded based on other criteria, primarily
shoreline or watershed development.
5. No introduced piscivores: One aim of this paleolimnological study is to reconstruct
cladoceran zooplankton assemblages, a group that includes Daphnia species that are
sensitive to changes in fish predation (Korhola and Rautio 2001). As such, lakes that had
non-indigenous piscivores (e.g., largemouth bass, Micropterus salmoides) in the ALS
survey were excluded from the study. Lakes with rainbow smelt (Osmerus mordax), a
species introduced in the Adirondack uplands that can compete with lake trout fry, were
also excluded. While it may have been desirable to exclude lakes with any nonindigenous fish species, some taxa such as yellow perch (Perca flavescens) and the
golden shiner (Notemigonus crysoleucas) are so widespread in the Adirondacks that
excluding lakes based on the presence of these taxa would have reduced the number of
sites to an insufficient number for study. Non-native piscivores were identified using
George (1981), Kraft et al. (2006), and USGS (2013).
6. Maximum depth > 5 m; surface area > 4 ha: These selection criteria were included
because some of the paleo-indicators of interest (scaled-chrysophytes, the Chrysophyceae
and Synurophyceae; and Daphnia taxa) are planktonic taxa that are rare or absent in
small, shallow waterbodies (Dixit et al. 1999; Ginn et al. 2010; Jeziorski et al. 2012).
Recognizing that many Adirondack lakes are shallow and < 4 ha in size (Krester et al.
1989), the size selection criteria used in this investigation were chosen to maximize the
total number of possible reference sites while minimizing the inclusion of sites unlikely
to contain the paleo-indicators of interest.
Identifying historic landscape disturbance & fisheries disturbance: Once the reference sites were
identified in the ALS, we sought to distinguish between sites that had experienced historic
watershed and fisheries disturbance from those that had not. In paleolimnological studies, the
process of attributing a species response to a particular stressor can be facilitated with knowledge
of the disturbance history of a particular lake. Such historical information is often incomplete or
70
qualitative. However, even incomplete historical data can help paleolimnologists distinguish
between species responses to a local stressor versus a regional stressor such as climate change.
This investigation used both quantitative and qualitative methods to define the landscape
disturbance history of the reference sites, identify alterations to historic fish assemblages, and
group the lakes into ‘classes’ based on their disturbance history (see Figure 3.3). Two historic
maps were used to characterize early watershed disturbance in the reference lakes. The first
historic map is a digital version of an 1890 map made available online by the New York State
Archives and the Adirondack Museum Library. The map shows the condition of Adirondack
forests prior to the creation of the Adirondack Park in 1892 (Original map scale-bar shown in
digital version, geographic coordinates not provided; NYS Forest Commission 1890). The
second data source is a 1916 “fire protection map” (hereafter referred to as the 1916 fire map)
digitized by the Adirondack Park Agency (APA) for use as a tool to determine historic
disturbance in the Adirondack Park (Scale 1:126720; APA 2000).
1890 and 1916 are two important time periods in the development history of the
Adirondack Park. Prior to 1890, logging practices in what is now the Adirondack Park were
fairly benign, involving the selective logging of merchantable timber (primarily large
softwoods), a process that did not alter the fundamental composition of the forest (McMartin
1994). The original 1892 boundary of the Adirondack Park was set to encompass lands that were
primarily virgin forest or areas logged for softwoods pre-1890, while areas outside the boundary
included regions that had been burned for charcoal production (denuded) or cleared for
agriculture/grazing. Thus, 1890 is a critical time-window that represents a period where most of
the lakes within the original 1892 boundary can be considered minimally disturbed. With the
creation of the 1894 “forever wild” provision in the New York State constitution, logging was
71
prohibited on State lands in the park. However, logging increased on private lands from the
1890s until ~1910, after which harvest generally declined (McMartin 1994). Additionally, parts
of the Adirondack region (primarily private land) experienced major fires in 1903, 1908, and
1913 caused by coal sparks along now largely abandoned railways, fires that were exacerbated
by logging and land-clearance practices (McMartin 1994). From 1910 onwards, there was an
expansion of the area covered by the Adirondack Park and an enlargement of State-held lands
protected from logging (Jenkins and Keal 2004). Thus, 1916 represents a time-window that
captures much of the historic watershed disturbance experienced by Adirondack lakes on private
or public lands.
Focusing on two early time periods (1890 & 1916) is a sufficient level of historical
investigation for a top-bottom study because of the method’s coarse temporal resolution (i.e., a
two-sample comparison between the present-day ‘top’ sample and a pre-industrial ‘bottom’
sample). Using these two early maps provides us with information on historic local disturbance,
a disturbance that may be reflected in the species assemblages of the ‘bottom’ sample. It is
difficult to assign an exact date to the ‘bottom’ sample because of differences in sedimentation
rate (the rate at which sediment accumulates in a lake basin) among lakes (Cumming et al.
1992b). For example, a sediment slice from 20 cm in one lake may represent sediment from
1850, while in another it may be ca. 1900 or later. Thus, while we may have been justified to
restrict our analysis to the pre-1900s, to be conservative in our approach we chose to examine
disturbance in both 1890 and 1916. In the subsequent down-core paleolimnological study,
additional historical resources were used to reconstruct the disturbance histories of a subset of
the reference lakes over a longer time period (not discussed here).
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The two historic maps provided us information on the spatial extent of logging, fire
disturbance, and land clearance in 1890 and 1916 in what is now the Adirondack Park. From a
paleolimnological perspective, the importance of a watershed disturbance is related to its
influence on the chemical composition of lake-water (e.g., increased export of cations, anions,
nutrients, dissolved and suspended organic matter, trace metals) which, in turn, can influence the
composition of aquatic species assemblages. Selective logging can influence runoff
concentrations of cations and anions but has a short-lived impact on stream chemistry (Scott et
al. 2001; Wang et al. 2006). Fire can increase runoff and erosion rates, increase nutrient export
and, in cases where large portions of the watershed are denuded, cause vegetative succession
(Binkley 1999; Committee on Hydrologic Impacts of Forest Management 2008). Lastly,
persistent land clearance has many potentially long-lasting influences on aquatic ecosystems,
including increased organic matter export to nearby lakes, changes in stream hydrology, and
eutrophication from increased phosphorus loading (Foster et al. 2003). Thus, when the reference
lakes in this investigation were ranked based on historic disturbance, logging, fire, and land
clearance, in that order, were viewed as increasingly detrimental to water quality and
increasingly influential on aquatic species.
The 1890 map was analyzed qualitatively. The map lacks geographic co-ordinates and
contains discrepancies in the placement of some smaller waterbodies and so is not suitable for
quantitative analysis. The reference lakes were subdivided into “A” lakes and “B” lakes based on
their land-cover classification in the 1890 map. “A” lakes were classed as virgin forest or logged
for softwoods in 1890 and are considered minimally disturbed whereas “B” lakes were in
denuded or cleared areas. In cases where a reference lake was not clearly labelled in the 1890
73
map (typically smaller or more remote sites), the placement of the lake was inferred based on the
proximity to other geographic features (e.g., large waterbodies, mountains).
The 1916 fire map was analyzed quantitatively using GIS methods (see Jenson and
Domingue 1988). First, watershed maps were generated based on elevational gradients around
the lake using ARCGIS 10.0. The 10 m US National Elevation Dataset was used as the base
Digital Elevation Model (DEM) (USGS 1999). The DEM was corrected to fill pits or sinks that,
while either real-world phenomena or digitizing errors, would corrupt the hydrological models.
After this process, a flow direction model (FDM) was created. The FDM examined the corrected
DEM on a cell-by-cell basis and, for each cell, evaluated which direction water would flow
based on the elevation of the surrounding 8 cells. The FDM was then used to create a flow
accumulation model (FAM) which mirrored the drainage pattern of the landscape. Using the
FAM, the point of highest accumulation for each lake was manually digitized as a point. The
point was then used to calculate the watershed for that lake using the FDM, which stepped
backwards up slopes from the starting point until a watershed divide was reached (i.e., flow
direction = null and/or accumulation = 0).
We defined watershed disturbance by calculating the percentage of each watershed that
had been affected by logging, fire, or land clearance on the 1916 map. Many GIS studies
quantify not only the overall extent of watershed disturbance (i.e. % disturbed area) but also the
spatial structure of the disturbance (e.g., amount of riparian vegetation cleared, slope of cleared
areas, Rogan and Miller 2007; Uuemaa et al. 2009). However, we used a non-spatial metric
because the 1916 digitized fire map showed discrepancies between its ‘water’ classification and
the placement of lakes in Adirondack watersheds. These discrepancies are not uniform or
systematic and likely originate from the digitization process of the original 1916 fire map (a
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hard-copy map including warped areas, digitized into 12 coverages edge-matched and appended
together, APA 2000). Given these discrepancies, interpreting fine-scale differences in watershed
disturbance was not appropriate. Regardless, the majority of the disturbed areas in the 1916 map
cover large areas and so this approach is sufficient to provide a general understanding of historic
watershed disturbance in the reference lakes. GIS studies of digitized historic maps often
quantify disturbance using non-spatial metrics (e.g., change in ‘% forest area’ or ‘% arable land’,
Bender et al. 2005; Skaloš et al. 2012), and % forest cover is often found to be the most
important determinant of water quality (Hunsaker and Hughes 2002).
To calculate disturbed areas in the watersheds of the reference lakes, the watershed maps
were converted to polygon (from raster) datasets. The area of the watersheds’ polygons were
calculated, and intersected with the 1916 fire map. Intersecting areas retained the original size
field and so a new size calculation of the intersected polygon revealed the percentage of a land
cover layer that fell within a watershed. The 6 terrestrial land-use/land-cover classifications in
the 1916 fire map were simplified into 4 groups and scored as progressively detrimental: green
timber (0); selective logging, including areas logged for softwoods, and areas logged for
softwoods/hardwoods (-1); fire damage, including burned, with slash, and burned, denuded (-2);
and open land (-3). Each lake was assigned a score based on the percentage of the watershed that
had been impacted by a particular disturbance type. For example, a lake that had experienced
selective logging in 50% of its terrestrial watershed, with the remaining area classed as green
timber would receive a score of -0.5, and a lake that had been logged in 50% of its watershed,
burned in 25% of its watershed, and had green timber in the remaining areas would receive a
score of -1. The reference lakes were then separated into 3 disturbance classes based on these
scores: Class 1 (least disturbed), 100% of the watershed listed as green timber (score = 0); Class
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2, at least 1 form of disturbance in less than 50% of the terrestrial watershed (0 > score > -1); and
Class 3, at least 1 form of disturbance in more than 50% of the watershed (score < -1).
Lastly, reference lakes that experienced reclamation and/or had non-native piscivores in a
recent (post-ALS) survey were distinguished from those without known changes. Reclamation is
a fisheries management strategy employed by the New York State Department of Environmental
Conservation (NYSDEC) in the Adirondack Park. Reclamation involves the addition of rotenone
to a lake to eliminate all fish taxa and allow for subsequent restocking of desirable taxa, a
process that can influence the size and structure of zooplankton species assemblages (Harig and
Bain 1998). As cladoceran zooplankton taxa are sensitive to alterations in fish assemblages and
are a paleo-indicator of interest, reclamation is considered a form of disturbance in this
investigation. A list of lakes reclaimed prior to 2011 was provided by R. Preall (NYSDEC).
Recent (1990-2011) fisheries surveys (NYSDEC 2012; Roy et al. 2011; R. Preall pers. comm.)
were available for 24 of the reference lakes and were examined to identify the presence of nonnative piscivores in the most recent fisheries survey. For the 7 remaining reference sites, the
most recent fisheries survey is the ALS data. When identifying fisheries disturbance, the most
recent survey was used (rather than all surveys combined), because not all of the reference lakes
included in the NYSDEC data had been surveyed multiple times (minimum 1 survey; maximum
10). In the disturbance classes defined using the historic watershed maps, lakes that have
experienced reclamation and/or the introduction of non-native piscivores are demarked by a
negative (-) sign (i.e., a lake that was virgin forest in 1890 and 100% green timber in 1916 but
reclaimed in the 1950s is listed as Class 1A-, See Figure 3.3).
Comparing the reference lakes to the ALS population: The reference lake selection criteria were
chosen to identify lakes that are minimally influenced by anthropogenic stressors but the criteria
76
also excluded certain kinds of Adirondack lakes (e.g., naturally acidic ponds). If reference lakes
are chosen based on a priori defined selection criteria, it is important to identify how
‘representative’ the reference lakes are of the different lake-types in a region (e.g., high-DOC
sites vs. low-DOC sites, deep lakes vs. shallow ponds). Thus, we plotted the 1,469 ALS lakes in
a Principal Components Analysis (PCA) of environmental variables (Lepš and Šmilauer 2003) to
compare the range of chemical/morphological characteristics of the reference lakes with that of
the ALS population; a population that is itself representative of 2,759 lakes in the AEZ.
Prior to inclusion in PCA, missing samples in the environmental matrix were replaced
with the variable’s mean value (Lepš and Šmilauer 2003). Variables with more than 4% of their
values missing and strongly correlated environmental variables (R > 0.9) were excluded from
analysis. A Shapiro-Wilks normality test was applied to test the normality of all environmental
variables (and sqrt/log transformed datasets). However, none of the environmental variables
were found to have a normal distribution, although many were minimally skewed. Rather than
arbitrarily exclude outliers (which would affect the representativeness of the ALS population),
we opted to visually inspect histogram plots of the environmental variables. As PCA is fairly
robust to minor violations of the assumption of normality, we selected 16 environmental
variables (or their sqrt/log transformed equivalents) that were roughly normally distributed for
inclusion (See Appendix A). Three ordinations were performed: one on chemical/morphological
data, one on chemical data only, and one on morphological data only. PCA ordinations were
performed using CANOCO 5.0 (ter Braak and Šmilauer 2012).
In a PCA of environmental variables, samples that are similar in chemical/morphological
characteristics plot close together in the ordination diagram. This aspect of PCA allows us to
identify two interesting lake populations in the ALS. First, using the placement of the reference
77
sites in the PCA to define the outer bounds of a polygon in PCA space, we can identify a
population of ALS lakes that, although they failed to meet one or more of the selection criteria,
are similar in physical/chemical characteristics to the reference lakes (hereafter referred to as
polygon-sites). Secondly, assuming that the primary gradient in the PCA is a pH/ANC gradient,
ALS sites that fall within the same range as the reference-lake polygon on the secondary axis,
but are outside the polygon on the low end of the pH/ANC gradient, form a population of lowpH/ANC lakes that the reference sites can be used to define recovery targets for. This
comparison is possible because the low pH/ANC sites have physical/chemical characteristics
(size, colour, nutrient levels, etc.) to the reference lakes.
Because the ALS is considered representative of 2,759 lakes in the AEZ, we can
extrapolate the percentage of sites in the ALS that meet our selection criteria to lakes in the AEZ.
The extrapolated population is the (relatively small) number of Adirondack lakes that the
reference sites are most representative of (or, most similar to). However, the polygon-sites are
also similar in physical/chemical characteristics to the reference sites, though the polygon-sites
failed to meet one or more selection criteria. The polygon-sites form a larger ‘representative’
population that is of interest as a study population for additional research questions (e.g.
comparing species responses in reference sites to an equal number of polygon sites that have
experienced watershed development).
Field sampling & additional GIS methods: To characterize the present-day water chemistry of
the reference sites, water samples were collected by hand in sterile glass and polypropylene jars
from ~0.5 m beneath the surface. All chemical analyses were performed by the Center for
Environmental Systems Engineering (Syracuse University, NY) using standard US EPA methods
(See Appendix B). Additionally, current land-use in the watersheds of the reference was defined
78
using the 2006 National Land Cover Database (Homer et al. 2012). Percent area of land-use
categories was identified using the methods outlined above.
Results
Site descriptions: Using the selection criteria described earlier, we identified a total of 30
reference sites in the ALS but one lake had a Cl- concentration exceeding 2 mg L-1 in 2010 and
so was excluded post-hoc from the dataset, leaving 29 lakes (~2.0% of ALS lakes) (See Figure
3.4 & Appendix C). If we extrapolate this percentage (~2.0%) to the 2759 AEZ lakes, we find
that there may be up to 54 AEZ lakes that meet the selection criteria. Two additional lakes,
Arbutus Lake and Wolf Lake, which were not part of the ALS, were added post-hoc to the
reference lake dataset. Both lakes are located in the Huntington Wildlife Forest, a private
research property owned and protected by the State University of New York, School of
Environmental Science and Forestry (SUNY-ESF) since the 1930s. The lakes have been
protected from acidification, watershed and shoreline development, lack non-native piscivores,
meet the depth and surface area criteria (Roy et al. 2011; Stager and Sanger 2003), and currently
have Cl- concentrations < 2 mg L-1.
Most of the reference lakes are protected in either ‘wild forest’ or ‘wilderness’ preserves
in the Adirondack Park. According to 2006 land-cover data, the two most dominant land-cover
types in the watersheds of the reference lakes, excluding open water, are deciduous forest and
coniferous forest. The least common land-cover type is developed land (See Table 3.1). In
general, the reference lakes are small (mean SA: 20.5 ha), moderately deep (mean Zmax: 10.9 m),
and are located at elevations greater than 300 m. The mean pH of the reference lakes was 6.4
when sampled in 2010 and 2011 and 28 lakes (90%) had ANC values greater than 50 μeq L-1. All
79
of the lakes had Cl- concentrations less than 0.5 mg L-1. The lakes are oligotrophic (mean TP =
1.3 µg L-1), calcium-poor (mean Ca2+ = 2.3 mg L-1) , with a DOC range of ~2 mg L-1 to
8.8 mg L-1.
Disturbance Rankings: A summary of the watershed disturbance experienced by each reference
lake in 1890 and 1916 is provided in Table 3.2, as well as a summary of fisheries alterations. In
1890, 24 (77%) of the reference were located in minimally-disturbed forests (virgin forest or
selectively logged for softwoods; “A” lakes). The lakes located in denuded or cleared regions
(“B” lakes) are primarily located in what is now the eastern Adirondack Park (See Figure 3.4). In
1916, 13 lakes (42%) had the terrestrial areas of their watershed listed as 100% green timber
(Class 1), of which 9 were considered minimally disturbed in 1890 (Class 1A). An additional 6
lakes (19%) had experienced at least 1 form of disturbance in less than 50% of their watershed
(Class 2), and 12 lakes (39%) had experienced a least 1 form of disturbance in more than 50% of
their watershed (Class 3). Forest fires were the most common form of anthropogenic disturbance
experienced by the reference lakes in 1916, followed by selective logging, and agricultural
clearance. In terms of fisheries disturbance, as of 2011, 20 of the reference lakes (65%) had been
reclaimed and 6 lakes (19%) had non-native piscivores identified in their most recent post-ALS
fisheries survey (See Table 3.2). In total, 5 Class-1 reference lakes (16%) (and 4 Class 1A lakes),
have no record of reclamation and lacked non-native piscivores in their most recent fisheries
survey.
Ordination Results: The first two axes of the PCA plots based on chemical/morphological data,
chemical data only, and morphological data only captured approximately 50%, 57%, and 73% of
the variation in environmental data of samples in the ALS dataset respectively (See Figure 3.5).
80
In general, when compared to the total ALS population, the reference lakes are comparatively
low in DOC/TP, are generally deeper and have larger surface areas/watershed areas, and are
intermediate in pH, conductivity, and concentration of major ions. Polygons created to bound the
reference lake dataset in the PCA contained 356 (24.2% of the population), 534 (36.4%), 451
(30.7%) sites in the chemical/morphological, chemical only, and morphological only PCAs
respectively. There was significant overlap in the polygon-sites between the 3 PCAs (48.8%91.9%). In the chemical/morphological PCA (the most restrictive test), the most common reason
why the polygon-sites were excluded from the reference lake set was the presence of non-native
piscivores.
Discussion
Understanding reference sites: The myth of the pristine & the dangers of presentism
The ideal reference site would be a hypothetical “pristine” site that has never experienced
any anthropogenic disturbance. Given the pervasive influence of stressors such as atmospheric
deposition and climate change, such a site is unlikely to exist (Stager and Sanger 2003).
However, this scarcity does not imply that the presence of present-day or historical
anthropogenic influence on a lake makes it unsuitable as a reference site. The process of
identifying reference sites is inherently subjective, even if ordination methods are used to
‘organically’ identify sites (Yates and Bailey 2010), because it depends on the research goals of
the study and the input data used. A site deemed to be in “reference” condition in one study
could be inadmissible in another. Some studies employ simple criteria (e.g., absence of stressor
of interest, Chessman et al. 2008) while others use more holistic measures of anthropogenic
disturbance (Wang et al. 2008). Researchers must therefore be able to answer the question
81
“reference for what?”. The selection criteria and methods used to identify the reference sites
must be clearly defined and correspond to the goals of the study.
While the selection criteria used in this investigation identified lakes that have been
largely protected from the detrimental effects of acid deposition, watershed/shoreline
development, and road salt seepage, historical analysis shows that many of the sites have
experienced early watershed and/or fisheries disturbance. Like many protected areas, the
Adirondack Park has a long history of human impact, including land clearance, logging, forest
fires, and non-native piscivores introductions (George 1981; McMartin 1994) and the 31
reference lakes in this investigation reflect that history. This study thus serves as a reminder of
the danger of presentism in ecological research; namely, the assumption that the present-day
physical, chemical, or biological characteristics of a system are representative of the long-term
historical condition of the site (Leira et al. 2006). Landscape disturbance may have long-lasting
effects on waterbodies in certain circumstances (Martin et al. 2011; Renberg et al. 2009), and so
limnological studies should ideally include a thorough historical analysis.
The historical perspective offered by paleolimnology can play an essential role in
developing a more complete understanding of the present-day ecology of a region and the
response of species assemblages to environmental stressors. Unfortunately, most studies that rely
solely on modern limnological sampling lack this historical perspective. For example, in a recent
meta-analysis of 240 recovery studies that included both terrestrial and aquatic studies, Jones and
Schmitz (2009) found that only 20% used pre-disturbance data (and, interestingly, only 58%
used undisturbed reference sites). This lack of pre-disturbance data is problematic because, while
many recovery studies have an implied recovery goal of a return to pre-disturbance state, without
the historical perspective it is difficult to define such a recovery target, let alone assess whether
82
or not the recovering sites are experiencing a shifting baseline. Indeed, the fact that the 2012 US
National Lakes Assessment includes a paleo-component (USEPA 2011), and that the EU Water
Framework Directive supports the use of paleolimnological data to determine ‘reference
conditions’ (Bennion et al. 2011), highlights the growing recognition of the value the paleoapproach provides in environmental assessment. Using paleolimnological techniques within a
reference site framework allows us to answer the question “what condition should impacted sites
recover to?”. While the ideal situation would be a return to pre-disturbance condition (e.g.
conditions ca. 1850), this may not be a feasible recovery goal in a period of global environmental
change.
Defining recovery targets: making meaningful comparisons between reference & impacted sites
Ideally, reference sites in paleolimnological studies should be minimally-disturbed by
multiple environmental stressors. However, such minimally-disturbed sites may not necessarily
be representative of all lake-types in a region. If reference sites are to be used to define recovery
targets for an impacted site, the site must be similar in general characteristics to the reference
systems to allow for meaningful comparisons. For example, our ordination of the ALS survey
found that the reference lakes are, in general, mid-sized, mid-elevation, low DOC/TP lakes with
circumneutral pH. As such, none of our reference sites would be appropriate to define a
recovery target for a high DOC acidic Adirondack pond. To address this data gap, we could
increase the pool of reference sites by relaxing some of the selection criteria (in this example, the
pH / depth criteria) but this could result in the inclusion of a number of acidified sites. The tradeoff between maximizing the number of reference sites and minimizing the possibility of
classifying ‘disturbed’ sites as reference sites should be carefully considered when implementing
regional paleolimnological studies in a reference site framework.
83
Testing hypotheses about climate change
Using reference lakes provides us with the opportunity to test hypotheses about climate
change on species assemblages. Given the rigor of the selection criteria used, if similar species
changes are documented in the reference lakes, it will be possible to conclude that: 1) the species
changes are not the result of acidification, eutrophication, salinization, or piscivores
introductions; 2) that a regional stressor is responsible for the species changes; and 3) that
climate change could be the cause of the species changes noted. This last inference can be made
through the use of statistical techniques such as ordination methods (Hyatt et al. 2010), a strong
understanding on the ecology of the paleo-indicators under study, and by comparing ‘down-core’
paleorecords with long-term climate monitoring data (Battarbee et al. 2012).
The historical analysis done in this study also allows us to conduct ‘natural experiments’
to test the influence of local versus regional stressors on species assemblages. For example, it
will be possible to identify the influence of fisheries disturbance on the assemblages by
comparing the species responses in an equal number of reclaimed lakes and non-reclaimed lakes
using either univariate or multivariate frameworks. Similarly, it will be possible to test the
influence of early watershed disturbance by comparing lakes in different disturbance classes.
Presumably, if local stressors are influencing species assemblages, there should be a significant
difference in species composition between disturbed and undisturbed sites at a centennial scale.
If not, a regional stressor is likely the cause of the change in species assemblage. Taken together,
these analyses can help develop a holistic view of how aquatic species assemblages are likely to
respond to climate change both independent of and synergistically with other stressors.
84
Applications in other regions
The successful implementation of this project in the Adirondack Park relied on three
factors: the availability of data from a large synoptic survey; previous paleolimnological studies
that provided the basis of appropriate selection criteria; and historical GIS data and fisheries
information for the study sites. During the acid-rain crisis, many regions in North America and
Europe were surveyed as part of large, synoptic surveys (e.g., ELS, Landers et al. 1988; EMAP,
Hughes et al. 2000; PIRLA-I and PIRLA-II, Charles and Smol 1990; Charles and Whitehead
1986; SWAP, Battarbee and Renberg 1990), many of which included paleolimnological studies.
There is also a growing availability of digital historic maps and online archival data for use in the
reconstruction of land-use change over time. Thus, the approach followed in this investigation
could be modified and applied in other regions where similar data exist. Of particular interest
would be the development of reference lakes in protected areas (Baron et al. 2009), which could
serve two roles: first, identifying the influence of climate warming on lakes in the region, and
second, serving as reference sites for lakes in developed areas. Landscape development threatens
to degrade the condition of reference sites across the United States (USEPA 2012). As such, the
importance of reference sites in protected areas will continue to increase as the number of
reference sites in or near urban, suburban, and agricultural areas decreases.
Conclusions
Lakes in the Adirondack Park and elsewhere in North America are increasingly
influenced by multiple stressors, with anthropogenic climate change posing a growing threat to
management and conservation efforts. Paleolimnological studies of minimally-disturbed
reference lakes can be used to identify the influence of climate change on aquatic species
assemblages independent of other stressors and define recovery targets for impacted sites. We
85
described and applied a conceptual framework for pairing paleolimnological techniques with a
regional reference site framework in a study of the Adirondack Park. Though not “pristine”, the
31 reference lakes have been largely protected from common environmental stressors and can be
used to define reasonable recovery targets for lakes recovering from acidification and test
hypotheses about climate change. The methods used in this investigation could be easily
modified and applied in other regions. As land-use change continues to threaten reference sites
outside of protected areas and the impact of climate warming on freshwater systems continues to
grow, reference sites in protected regions will become an increasingly important tool for
biodiversity conservation and policy development.
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Tables & Table Captions
Table 3.1. Summaries of lake and watershed characteristics, lake morphometrics, and water chemistry
variables for 31 Adirondack reference lakes. Minimum (min.), maximum (max.), mean, and median of
variables presented. Lake watershed land-cover & land-use characteristics are calculated as percent area
(%) of the terrestrial watershed. Land cover & land use types are: deciduous forest (% deciduous),
evergreen (coniferous) forest (% evergreen), woody wetland (% woody wetland), mixed forest (%
mixed), shrubs/scrub (% shrub), emergent herbaceous wetland (% emergent), and developed land (open
space) (% developed). Lake morphometric parameters are: elevation (m), surface area (ha), and
maximum depth (Zmax, m). Chemistry variables are: pH, acid neutralizing capacity (ANC, µeq L-1),
chloride (Cl-, mg L-1), total phosphorus (TP, µg L-1), dissolved organic carbon (DOC, mg C L-1), and
calcium (Ca, mg L-1). BDL = below detection limit. Note TP values are from sites sampled in 2011 only
(n = 20).
Category
Watershed
Land Cover &
Land Use
Lake
Morphometrics
Water
Chemistry
Variable
% deciduous
% evergreen
% woody wetland
% mixed
% shrub
% emergent wetland
% developed
Elevation
Surface Area
Zmax
pH
ANC
ClTP
DOC
Ca
Min
6.3
0
0
0
0
0
0
349
4.7
5.2
5.9
35.3
0.11
0 (BDL)
2.1
0.5
95
Max
88.8
73.2
48.8
44.3
5.4
7.5
1.9
731
57
23.8
6.8
376.8
0.37
4
8.8
5.2
Mean
50.2
25.7
13.3
9.5
0.5
0.7
0.1
490
20.5
10.9
6.4
120.7
0.20
1.3
4.3
2.3
Median
51.5
15.9
9.2
4.8
0
0
0
486
13.9
10.1
6.3
85.8
0.17
1.0
3.9
2.0
Table 3.2. Summaries of early watershed disturbance and fisheries disturbance in 31 Adirondack reference lakes.
Lakes are listed by their lake code given in Appendix C. A digital version of an 1890 forest commission map was
consulted to identify the watersheds as virgin forest (VF), logged for softwoods (LS), denuded (DE), or cleared
for agriculture (CL). GIS data of a digitized 1916 fire protection map were used to calculate the percent of each
lake’s terrestrial watershed characterized as green timber, logged, burned, or cleared land (for agriculture or
grazing) and assign the lake a score based on the type and amount of disturbance. The classes were sub-divided
into ‘A’ lakes (lakes that were either virgin forest or logged for softwoods as of 1890) and ‘B’ lakes (lakes that
were cleared or denuded as of 1890) and grouped into a disturbance class based on the 1916 GIS data, with Class
1 lakes being the least disturbed. Fisheries disturbance data documented whether or not the lake is known to have
been reclaimed (Recl.; Y if reclaimed), and whether non-native piscivores (Non-nat.) were identified in the most
recent NYSDEC fisheries surveys (Y if non-natives identified). Lakes that were reclaimed or had non-native
piscivores in their more recent fisheries survey are denoted by a negative sign (-) in the lake class.
Lake
Code
WolfL
SampP
StreL
CoppP
CascL
MossL
ArbL
ENestL
LongP
IslaP
USpecP
GullP
ChallP
WilcL
RockP
EPineP
FishP
LFishP
BassL
CascP
BootP
RounP
LSargP
DeerP
ClamP
BessP
NellP
ClearP
LydiP
MidBL
GrizO
1890
Forest
Map
VF
VF
VF
LS
VF
VF
VF
VF
VF
CL
CL
CL
DE
VF
VF
LS
LS
LS
DE
LS
VF
VF
VF
VF
LS
LS
LS
LS
LS
LS & DE
CL
Green
Timber
100
100
100
100
100
100
100
100
100
100
100
100
100
92.5
91.4
58.8
52.4
51.1
67.7
0
0
0
0
0
28.1
13.8
11.8
0.2
0
0
18.2
1916 Fire Map
Logged Burned Cleared
Land
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
7.5
0
0
8.6
0
41.2
0
0.3
46.9
0.3
0.3
48.2
0.3
32.3
0
0
100
0
0
3.5
84.0
12.4
0
100
0
99.8
0.2
0
94.1
5.9
0
0
71.9
0
0
86.2
0
0
88.2
0
0
99.8
0
0
81.4
18.6
4.2
95.8
0
0
1.3
80.5
96
Score
0
0
0
0
0
0
0
0
0
0
0
0
0
-0.22
-0.26
-0.82
-0.95
-0.98
-0.32
-1.00
-2.09
-2.00
-1.00
-1.06
-1.44
-1.72
-1.76
-2.00
-2.19
-1.96
-2.44
Fisheries
NonRecl.
Nat.
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
Y
-
Lake Class
1A
1A
1A
1A
1A1A1A1A1A1B
1B1B1B2A
2A2A2A2A2B3A
3A
3A3A3A3A3A3A3A3A3B
3B-
Figures & Figure Captions
Figure 3.1. Conceptual diagram outlining the biological response of two lakes (Lake A and Lake B) to
acidification and climate change since pre-industrial times. The lake response is shown across two
gradients: one of increasing acidity and one of increasing air temperature. Lake A is a higher-pH/ANC
‘reference lake’ that is resistant to acidification. Lake B is a lake susceptible to acidification. The purpose
of including a reference lake (Lake A) is to identify trends in species composition caused by regional
warming (i.e., to characterize the ‘shifting baseline’). Historically, the influences of regional warming and
acid deposition on the lakes are minimal. The lakes are similar in species composition (A hist, Bhist), though
the taxa of Lake A are characteristic of higher pH/ANC conditions. With the onset of acid deposition, Lake
B acidifies (solid arrow). Acid-sensitive taxa are lost and replaced with acid-tolerant ones (Bacid). Lake A
does not acidify; its species assemblages do respond, however, to regional warming (solid arrow). There is
an increase in warmer-water taxa, taxa that benefit from shorter-ice covered seasons, etc. (Apres). When the
stress of acid deposition is reduced, there are four possible recovery trajectories for Lake B (dashed
arrows): 1) Return to pre-disturbance state: this is only possible if Lake B does not respond to warming,
which is unlikely; 2) No Recovery: Lake B responds to warming but shows no decline in acid-tolerant
taxa; 3) Partial recovery: Lake B responds to warming and there is some decline in acid-tolerant taxa
and/or increase in acid-sensitive taxa; 4) Recovery to a novel state: Lake B responds to warming and the
abundance of acid-tolerant/acid-sensitive taxa returns to historic levels. However, regional warming has
prevented the lake from returning to its pre-acidification state. Lake A and B are again similar in species
composition, though the taxa of lake A remain characteristic of higher pH/ANC conditions. Conceptual
diagram based ideas presented in Vinebrooke et al. (2004), Hobbs et al. (2009), Battarbee et al. (2012).
97
Figure 3.2. Conceptual spatial diagram of a paleolimnological reference site study. Reference sites (white
circles), impacted sites (black circles), and sites that are neither reference nor impacted sites (grey circles)
are distributed in a region. The reference sites are sampled as part of a top-bottom study (TB) and then a
selection of reference sites are compared with a set of impacted sites in a down-core study (DC).
98
Figure 3.3. Schematic outlining the reference lake selection and characterization process followed in this
investigation. Reference lakes were identified using 6 selection criteria from 1,469 lakes included in the
Adirondack Lakes Survey. The selection criteria were: lakes had public access; pH > 6.5, ANC > 50
µeq L-1; % shoreline/watershed development < 5%; Cl < 2 mg L-1; no introduced piscivores (at time of
sampling in ALS); maximum depth > 5 m, surface areas > 4 ha. The historic watershed disturbance
experienced by each reference lake was qualitatively characterized using a digital version of an 1890 map
and quantitatively characterized using GIS methods applied to a digitized 1916 map. Using the 1890 map,
lakes were described as being part of a minimally disturbed forest in 1890 (‘A’ lakes) or a more disturbed
area (‘B’ lakes). In the quantitative assessment, lakes were scored based on the amount and type of early
watershed disturbance they experienced and sorted into disturbance classes (with Class 1 lakes being the
least disturbed). Lastly, in the disturbance classes defined using the historic watershed maps, lakes that
experienced reclamation and/or the introduction of non-native piscivores in a post-ALS fisheries survey
were demarked by a negative (-) sign. The result was a series of lake classes that represent a gradation
from least (Class 1A) to most (Class 3B-) disturbed reference lakes.
99
Figure 3.4. Map of the Adirondack Park (NY, USA) showing the locations of a set 31 Adirondack
reference lakes minimally impacted by anthropogenic stressors. Map shows the position of the Adirondack
Park in northeastern United States (enlarged area); inset shows lakes in the St. Regis Canoe area in detail –
all other reference lakes are shown as points.
100
101
Figure 3.5. PCA ordinations of ALS survey data: A) PCA of morphological & chemical variables; B) PCA of chemical variables only; and C) PCA of morphological
variables only. Morphological variables are elevation (elv, m) maximum depth (Zmax, m, log transformed), surface area (SArea, ha, log transformed), watershed area
(WArea, ha, log transformed). Chemical variables are SO42-, F-, K+ (mg L-1, sqrt transformed), DOC (dissolved organic carbon, mg C L-1, sqrt transformed), pH (field
pH measurement, sqrt transformed), Cl-, Na+, Ca2+, Fe2+, (mg L-1, log transformed), DIC (dissolved inorganic carbon, mg L-1, log transformed), TP (total phosphorus,
mg L-1, log transformed) and specific conductance (SCOND, µmhos cm-1, log transformed). Reference sites (large circles) were used to demarcate a polygon in PCA
space that contains sites that are similar in physical and/or chemical composition to the reference sites but failed to meet one or more selection criteria. These
‘polygon-sites’ are shown as grey crosses. Non-reference & non-polygon sites in the ALS are shown as points. Eigenvalues (λ) for PCA axis 1 and 2 are given.
CHAPTER 4
ADIRONDACK (NY, USA) REFERENCE LAKES SHOW A PRONOUNCED SHIFT IN
CHRYSOPHYTE SPECIES COMPOSITION SINCE CA. 1900
Abstract
Lakes in both North America and Europe are undergoing chemical recovery from
acidification and there is a pressing research need to define recovery targets for acid-impacted
species assemblages. Researchers attempting to define such recovery targets are often hampered
by two issues: 1) a lack of monitoring data; and 2) the confounding influence of multiple
stressors, especially regional stressors like climate warming. Paleolimnological studies of
minimally-disturbed reference sites can address both of these issues. By reconstructing reference
lake histories, paleolimnologists can infer if/how regional stressors like climate change influence
species assemblages independent of stressors such as acidification, eutrophication, and land-use
change. This manuscript reconstructs changes in the scaled chrysophyte (Synurophyceae and
Chrysophyceae) assemblages of 31 minimally-disturbed Adirondack (NY, USA) reference lakes
from pre-ca. 1900 to present. The reference lakes have been minimally influenced by
acidification, eutrophication, road salt seepage, and piscivore introductions, and so represent a
unique opportunity to examine the effects of regional stressors in the Adirondack Park. First, the
present-day distribution of chrysophyte taxa in the reference lakes was compared to measured
limnological variables using Redundancy Analysis (RDA). RDA found that physical/chemical
variables (summarized by PCA axis-scores) explained ~19% of the variation in chrysophytes
species assemblage. S. petersenii, a taxon known to cause taste and odour issues in freshwaters,
was common in deeper lakes with higher pH/ANC and DOC < ~5.2 mg C L-1. A top-bottom
paleolimnological analysis examined changes in chrysophyte assemblages from pre-1900 to
102
present in the reference lakes. A significant shift in species composition from pre-1900 to
present was found using both univariate (Wilcoxon signed-rank test) and multivariate
(ANOSIM) analyses. Present-day assemblages were characterized by an increased abundance of
colonial chrysophyte taxa, a trend that has been documented in other regions in North America.
This study suggests that regional warming and/or oligotrophication are influencing the species
assemblages of minimally-disturbed reference lakes, and so lakes recovering from acidification
are unlikely to return to their pre-disturbance assemblages.
Introduction
Lakes are increasingly influenced by multiple anthropogenic stressors (Keller 2009),
including introduced species, land-use change, acidification, eutrophication, and climate change.
Multiple stressors can have both synergistic and antagonistic effects on aquatic species
assemblages (Coors and De Meester 2008; Folt et al. 1999) and may lead to ‘ecological
surprises’ (sensu Paine et al. 1998) in impacted systems. As such, systems affected by multiple
stressors pose unique challenges to researchers, managers, and policy makers.
Remediative strategies for environmental issues typically focus on a single stressor, often
with an either explicitly stated or implied recovery target of a return to pre-disturbance state (e.g.
clean air legislation targeting sulphur dioxide emissions, leading to declines in acid deposition
and chemical recovery from acidification in some lakes in some regions, Garmo et al. 2014).
However, there is a growing recognition that multiple stressors may prevent systems from
returning to a pre-disturbance state, even with remediative action, a phenomenon referred to as
the ‘shifting baseline’ (Bennion et al. 2011). In this regard, climate change represents a stressor
of particular concern, given its regional nature. In a multiple stressor environment, climate
change introduces additional complexities in the physical, chemical, and biological processes of
103
lakes (Smol 2010). While there are a growing number of case studies suggesting that regional
warming may prevent lakes recovering from acidification or eutrophication from returning to
their pre-disturbance state (e.g., Arseneau et al. 2011; Battarbee et al. 2012), the effect of climate
change on recovery processes is surprisingly understudied (Verdonschot et al. 2013).
There is a pressing research need to define reasonable recovery targets for impacted sites
that account for changing climatic conditions. To address this issue, the United States
Environmental Protection Agency (USEPA) recently called for long-term monitoring at
minimally-disturbed reference sites across the United States (USEPA 2012). Reference sites (i.e.,
sites minimally impacted by multiple anthropogenic stressors, Bailey et al. 2004) are commonly
used in bioassessment surveys to define recovery targets for impacted ecosystems. Long-term
monitoring at reference sites can be used to monitor changes in baseline conditions due to
regional stressors such as climate change which are occurring independent of more local
stressors (e.g., land-use change, eutrophication, acidification, etc.).
While laudable, the approach advocated by the USEPA is limited by the fact that longterm monitoring data of aquatic ecosystems are rare and, when they exist, typically cover
relatively short periods (< 50 years). Here, paleolimnological studies can play an informative
role. Paleolimnological studies of minimally-disturbed reference sites can identify regional
changes in species composition that are occurring independent of many common stressors (for an
early example, see Battarbee et al. 2002), information that can, in turn, be used to define
recovery targets for impacted lakes in a region. For example, if a paleolimnological study of a set
of regional reference sites were to identify increases in warm-water taxa or small planktonic
algae over the past ~100-150 years (changes predicted to occur with regional warming; Winder
and Sommer 2012), a recovery target for acid-impacted species assemblages in the region should
104
reflect those changes (Chapter 3). A variety of possible methods for comparing impacted and
reference sites exist (Chapter 3), including both univariate methods (i.e. correlations between
long-term monitoring data and key taxa in reference/impacted sites Battarbee et al. 2012) and
multivariate methods (i.e. ordination techniques, Yan et al. 2004).
The goal of this manuscript is to use a top-bottom paleolimnological approach to examine
long-term changes in scaled chrysophyte assemblages (Synurophyceae and Chrysophyceae) in a
set of 31 reference lakes in the Adirondack Park (NY, USA). The reference lakes were identified
from an online database of 1,469 Adirondack lakes surveyed as part of the Adirondack Lakes
Survey (ALS) using a set of stringent selection criteria (Krester et al. 1989). The reference lakes
have been largely protected from several common stressors in the Adirondack Park, including:
acidification (i.e. decreases in lake pH and ANC), eutrophication from watershed/shoreline
development, leaching of road salt, and piscivore introductions (see Chapter 3). The reference
lakes are not, however, ‘pristine’ (i.e. lakes protected from any anthropogenic influence, Bennion
et al. 2011). Many of the sites have experienced historic watershed such as logging or fisheries
disturbance such as reclamation. These forms of historic disturbance are common in the
Adirondack Park and in other currently protected areas (Stager & Sanger 2003) and may have
long-lasting chemical/biological legacies (Martin et al. 2011).
While the reference sites may not be pristine, they have been protected from several
common anthropogenic stressors and so can be used to identify what influence, if any, regional
stressors such as climate change are having on Adirondack lakes; information that, in turn, can
be used to define recovery targets for Adirondack lakes recovering from stressors such as
acidification. Which that goal in mind, the following questions are addressed in this
investigation:
105
1) What physical/chemical variables influence the present-day distribution of chrysophyte
taxa in the reference lakes? Additionally, do historic watershed or fisheries disturbances
influence the present-day distribution of chrysophyte taxa in the reference lakes? and,
2) Has there been a significant shift in chrysophyte species composition in the reference
lakes since ca. 1900, both at the level of key individual taxa and the whole species
assemblage?
This manuscript is part of a series of papers whose goal is to highlight the important role
paleolimnological studies of minimally-disturbed reference sites can play in defining reasonable
recovery targets for lakes recovering from stressors such as acidification. This top-bottom study
is aimed at identifying whether or not regional changes in chrysophyte species composition are
occurring in minimally-disturbed Adirondack lakes and examine plausible mechanisms for any
species changes noted. Subsequent articles will provide detailed down-core analyses of species
assemblages in both reference and acidified lakes.
Methods
Study Sites: The lakes in this investigation are a set of 31 minimally-disturbed reference lakes in
the Adirondack Park. The lakes were selected from an online database of 1,469 Adirondack lakes
surveyed from 1984-1987 as part of the ALS. The lakes were chosen using a set of selection
criteria designed to identify lakes that have been minimally impacted by common anthropogenic
stressors (acidification, watershed development, road salt application, piscivore introductions).
PCA-ordination showed that the reference lakes are representative of between 24 to 36% of the
physical and/or chemical characteristics of the ALS lakes, a survey that is itself representative of
2,759 lakes in the Adirondack Ecological Zone (AEZ) (Chapter 3).
Quantitative and qualitative analyses of historic maps and fisheries data found that many
of the reference sites had experienced some form of historic disturbance since the late 1800s and
106
early 1900s (Chapter 3). An understanding of historic disturbance was used to group the
reference lakes into classes. Qualitative analysis of an 1890 map showed that 77% of the lakes
were located in minimally-disturbed forest (named ‘A’ lakes) whereas the remainder were
located in cleared or denuded watersheds (‘B’ lakes). Quantitative (GIS) analysis of a 1916 map
found that 42% of the reference lakes were listed in watersheds classed as 100% green timber
(Class 1 lakes), of which 9 were considered minimally disturbed in 1890 (Class 1A). An
additional 6 lakes (19%) had experienced at least one form of disturbance (logging, burns, or
land-clearance) in less than 50% of their watershed (Class 2), and 12 lakes (39%) had
experienced a least 1 form of historical disturbance in more than 50% of their watershed (Class
3). Forest fires were the most common form of anthropogenic disturbance experienced by the
reference lakes in 1916, followed by selective logging and agricultural clearance. In terms of
fisheries disturbance, as of 2011, 20 of the reference lakes (65%) had been reclaimed and 6 lakes
(19%) had non-native piscivores identified in their most recent post-ALS fisheries survey.
The reference sites are not ‘pristine’ (i.e., sites devoid of any anthropogenic influence,
Bennion et al. 2011); indeed, given the pervasive influence of stressors such as atmospheric
deposition and climate warming, the likelihood of identifying ‘pristine’ reference lakes is small.
Rather, the Adirondack reference lakes reflect the Adirondack Park’s history of landscape
disturbance and fisheries alterations (George 1981; McMartin 1994), disturbances that are
common in many protected areas. Despite these early disturbances, the reference sites provide us
with the opportunity to examine regional changes in species composition independent of many
common contemporary stressors. Furthermore, the fact that some (but not all) of the reference
sites have experienced either historic watershed and/or fisheries disturbance allows us to test if
these early disturbances influence the present-day distribution of species assemblages.
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Field Collection: Sediment cores were collected from the deepest basin of the study lakes in
either summer 2010 or 2011 using a 7.6 cm (ID) Glew gravity corer, and were sectioned on-site
into 0.25 cm using a vertical Glew extruder (Glew et al. 2001). Water samples were collected by
hand at ~0.5 m below the water’s surface in glass and polypropylene jars. All chemistry analyses
were performed by the Center for Environmental Systems Engineering (Syracuse University,
NY) using standard US EPA methods (Chapter 3).
Sample preparation: In a top-bottom paleolimnological investigation, a surface-sediment sample
is taken to represent the present-day species assemblage and compared to a sediment sample
taken from further down in the sediment core to represent pre-industrial times (Cumming et al.
1992b). The technique is an effective way to build a regional ‘snap-shot’ of species changes over
time. In this investigation, the ‘top’ sample is an integrated sample from 0-1 cm in the sediment
core and is estimated to represent the past 2-3 years. The ‘bottom’ sample is a 0.25 cm sediment
subsection from ~20 cm (in 4 cases where the sediment core length was < 20 cm, the bottommost
sediment interval was used; 3 cases ~17-18 cm, 1 case ~15 cm). A single 0.25 cm subsection was
analyzed at ~20 cm, rather than a 1 cm subsection, as a decrease in water content in the sediment
occurs downcore which increases the amount of ‘time’ represented per sediment subsection at
this depth. A depth of 20 cm was chosen because a survey of recently 210Pb-dated cores from the
Adirondacks suggests that supported 210Pb is reached at ~20 cm in the majority of Adirondack
region lakes (Arseneau et al. 2011, Arseneau et al. unpublished; see also Cumming et al. 1992b;
Cumming et al. 1994). Conservatively, we refer to the time period at ~20 cm in the sediment
core as ‘pre-ca. 1900’, rather than pre-industrial times (ca. 1850), as we recognize that
differences in sedimentation rate may alter the depth at which unsupported 210Pb is no longer
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present. An additional set of samples from a depth of 30 cm was analyzed in the 10 reference
lakes that had a sufficiently high concentration of chrysophyte scales to be counted to assess the
variability of sedimentary assemblages from pre-1900 (see Reproducibility study below).
Sediment samples were prepared for enumeration of chrysophyte scales using standard
methods (Battarbee et al. 2001). Briefly, sediment samples were digested in a 1:1 molar mixture
of strong H2SO4:HNO3. Samples were heated to ~70°C for 6-7 hours and were then rinsed
repeatedly until samples reached the pH of distilled water (litmus test). Samples were plated as a
series of four 100x dilutions on microscope slides using Naphrax®. Chrysophyte scales were
identified using DIC optics under oil-immersion using a 100X Fluotar objective with a numerical
aperture of 1.3. In cases where scales were sparse, a minimum of 200 scales were counted (Ginn
et al. 2010). In cases were scales were abundant, a minimum of 400 scales were counted. A small
number of pre-1900 samples were considered uncountable due to low scale concentration (i.e. <
10 scales encountered per transect). The principle taxonomic references referred to in this
investigation are: Siver (1991), Nicholls and Gerrath (1985), and Kling and Kristiansen (1983),
as well as unpublished photo reference materials. Small Mallomonas spp. scales that could not be
identified using light microscopy were group into a Mallomonas ‘small’ category (Cumming et
al. 1992a).
Statistical Analyses
Understanding factors related to modern chrysophyte distribution: A Redundancy Analysis
(RDA) was used to examine the influence of water chemistry and lake morphometric variables
on the present-day distribution of chrysophyte taxa in the Adirondack reference lakes. RDA, a
linear direct ordination method, was selected because an initial Detrended Correspondence
Analysis (DCA) of the present-day samples for the 31 reference lakes indicated that the length of
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the longest species turnover gradient was short (< 3 s.d.) and so linear ordination techniques
were deemed appropriate (Lepš and Šmilauer 2003). Because of the relatively small number of
surface samples in this investigation (31 samples), the environmental data matrix was
summarized as PCA-scores of the first four PCA axes (Lepš and Šmilauer 2003). The normality
of environmental variables was assessed using the Shapiro-Wilks test prior to inclusion of the
environmental variables in the PCA. Non-normal environmental variables were transformed
using either square-root or log transformations, and were not included if transformation did not
result in a normal distribution. Correlations between normal input variables were all less than
0.8. All environmental variables (PCA axis scores) were used in an initial RDA to test if the
combined set could explain a significant proportion of the species variance along all canonical
RDA axes (ter Braak and Šmilauer 2012). If so, then forward selection using Monte Carlo
permutation tests was used to identify a minimum set of environmental variables that could
explain a significant proportion of the species assemblage data. Shapiro-Wilks tests and
correlation coefficients were determined using the computer program SYSTAT v. 11. PCA,
RDA (on square-root transformed species data) and forward selection procedure were performed
using the computer program CANOCO v. 5.0 (ter Braak and Šmilauer 2012). The species data
were square-root transformed to reduce the importance of the dominant taxon Mallomonas
duerrschmidtiae, a common Adirondack taxon with a wide tolerance to many environmental
variables (Cumming et al. 1992a).
Lastly, an Analysis of Similarities (ANOSIM) was used to test for a significant difference
in species composition in the present-day species assemblages of the reference lakes grouped
into different disturbance classes. ANOSIM is a non-parametric test of rank similarities that can
be used to test for a significant difference in species composition between a priori defined
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groups (Clarke 1993). The Similarities Percentages test, SIMPER, is then used to identify which
species make an important contribution to the difference seen between groups (Clarke and
Warwick 2001). In this investigation, if no significant difference is seen between different lake
classes, it can be concluded that historic watershed/fisheries disturbances have minimal influence
on the present-day chrysophyte assemblages of the reference lakes. ANOSIM was used to
compare present-day chrysophyte taxa in: Class 1 lakes (minimally disturbed, n = 13) vs. Class 3
lakes (at least 1 form of disturbance in >50% of watershed, n = 12) and non-reclaimed lakes (n =
11) vs. a randomly-selected subset of reclaimed lakes (n = 11). A randomly selected subset of
reclaimed lakes was used to ensure a balanced study design. ANOSIM (using Bray-Curtis
similarities) was performed on square-root transformed species data using the computer program
Primer v. 6.1.11.
In all of the analyses listed above (and subsequently), taxa that reached > 2% relative
abundance in more than two sediment samples were included in the analyses. This cut-off
criterion was used primarily to exclude rare taxa, as they can have a large influence in
ordinations with a small number of samples. Because the interest in this study are the general
controls on chrysophyte distribution and general trends in the reference lakes over time, we
opted to set a selection criterion that would eliminate rare taxa.
Summarizing long-term species changes: A chrysophyte-inferred pH (CI-pH) was calculated for
present-day and pre-1900 sediment intervals using an Adirondack calibration set based on 71
lakes that spanned a pH gradient from just over 4 to ~8 (Cumming et al. 1992a). CI-pH was
calculated using a ln+1 species transformation, with tolerance down-weighing and inverse
deshrinking (Cumming et al. 1994) using the program C2 v. 1.72.
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For ease of interpretation, sedimentary assemblage data were summarized using
ordination techniques. An initial DCA of present-day and pre-1900 samples for the 26 reference
lakes with sufficient scales in the pre-1900 sample found that the length of the longest gradient
was short (< 3 s.d.) and so linear ordination techniques were appropriate (Lepš and Šmilauer
2003). The sedimentary assemblages were summarized in a PCA ordination diagram using
square-root transformed species data using CANOCO v. 5.0 (ter Braak and Šmilauer 2012). A
summary diagram was composed listing the relative abundance of all the taxa in the samples,
with lakes listed by PCA-1 axis scores of a PCA with all 31 surface sediments (Appendix D).
Testing for a significant change in species composition from pre-1900 to present: Changes in
species composition from pre-1900 to present were analyzed in the reference lakes using both
univariate and multivariate techniques. Previous paleolimnological investigations have identified
recent increases in colonial chrysophyte taxa in Ontario (Flear 2011; Paterson et al. 2004;
Paterson et al. 2008) and Maritime Canada (Ginn et al. 2010), increases which the authors
suggest may be due, at least in part, to regional warming. As such, in this investigation a
Wilcoxon signed-rank test (a non-parametric analog of a paired t-test) was used to identify if
there has been a significant increase in the relative abundance of colonial taxa in the reference
lakes from pre-1900 to present. Subsequently, an ANOSIM test was used to test for a significant
difference in species composition in the overall assemblages from pre-1900 to present. Lakes
with chrysophyte concentrations too low to count in the pre-1900 interval were excluded from
the analyses.
Reproducibility study (counting variability and stability analysis): An implicit assumption of the
top-bottom analysis is that differences in species assemblages noted between present-day and
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pre-industrial (or, pre-1900) times are greater than those arising from counting variability and, as
such, can likely be attributed to some form of environmental change. To test this assumption, in
2012 quadruplicate sediment cores from the deepest basin of 4 reference lakes were collected.
Sediment samples from the core tops (0-1 cm), 20 cm, and, when possible, 30 cm were prepared
and analyzed for each core using the standard methods listed above. Prior to counting, the
sediment samples were given a non-descript identifier (sample I, sample II, etc.) to reduce
possible counter bias. Taxa that reached a relative abundance > 2% in more than 2 samples were
used to calculate Bray-Curtis (BC) similarities between groups of samples (see below).
To test the assumption that the differences between top-and-bottom reference site
samples were greater than expected counting variability, quadruplicate top samples were
compared within each lake group in a triangle similarity matrix of Bray-Curtis similarities (see
Figure 4.1). Then, the average BC similarity of all top sample comparisons was calculated
(referred to as ABC-top). This average value represents the expected similarity between samples
when variation is due primarily to counting variability (and core collection/sample preparation).
Therefore, in the top-bottom analysis of the regional reference lakes, a reference lake with a topbottom BC-similarity that is less than ABC-top is unlikely to have a difference in species
composition due to counting variability alone.
A second implicit assumption of the top-bottom approach is that the sedimentary
assemblages were relatively stable pre-industrially (or, here, pre-1900). To test this assumption,
BC-similarities were calculated for all top, 20 cm, and 30 cm samples from all replicate cores
within each lake group. Triplicate cores of sufficient length (> 30 cm) were collected from three
of the reference lakes. A Wilcoxon signed-rank test was performed to examine if a significant
difference existed between similarities calculated between top vs. 20 cm samples, and 20 vs. 30
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cm samples. Presumably, if some form of environmental change has acted on the species
assemblages over the past 100-150 years, the species assemblages should be less similar between
the top vs. 20 cm samples than the 20 vs. 30 cm samples. To further examine this assumption,
samples from a core depth of 30 cm were analyzed in all reference lakes from the regional topbottom study where sediment cores > 30 cm in length were collected. Ten lakes had a
sufficiently high concentration of scales at 30 cm to be counted. ANOSIM was used to test for a
significant difference in species composition samples from 20 and 30 cm in the 10 lakes. If no
significant difference in species composition is found between the two groups, it can be
concluded that the assemblages were relatively stable prior to 1900; or, at least, that the variation
between the time periods is less than the regional variation in species assemblages captured by
the 10 lakes.
Results
Controls on chrysophyte species distribution: PCA plots based on chemical/morphological data
captured approximately 72% of the environmental variation when summarized in 4 PCA axes
(See Figure 4.2A & 4.2B). The first PCA axis is related to ion concentration and ANC, PCA
axis-2 is related to Zmax and DOC/TN (with pH loading positively on both axis 1 and axis 2),
PCA axis-3 is related primarily to SO42-, and PCA axis-4 is related to elevation and K+.
Unfortunately, filtered TP values were only available for lakes sampled in 2011 and so TP could
not be included in the PCA. However, in the 2011 sites, TP was significantly positively
correlated with TN (r = 0.5), and so higher TN sites were also likely associated with higher TP
conditions.
The initial RDA using all environmental variables explained a significant proportion of
the species variance along all canonical axes (p = 0.002, variance explained = 24.7%). The
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forward selection procedure using Monte Carlo permutation tests found that 2 environmental
variables (PCA-axis scores) explained a significant proportion of the variation in the chrysophyte
assemblages in the present-day reference lake assemblages (PCA-2: F = 3.6, p = 0.002; PCA-1:
F = 2.8, p = 0.007). The first and second RDA axes explained approximately 19.1% of the
variance in the species assemblage data, when constrained to the PCA axis 1 and 2 scores (See
Figure 4.2C). The main separation in the redundancy analysis fell along a pH gradient and was
between lakes characterized by a higher relative abundance of S. petersenii and lakes
characterized by a higher relative abundance of S. sphagnicola. Lakes with a higher relative
abundance of S. petersenii were associated with higher PCA-1 & PCA-2 scores (i.e. deeper lakes
with higher pH values and lower DOC values). In general, lakes characterized with a higher
relative abundance of S. sphagnicola were associated with lower pH values. While some lakes
with a high relative abundance of S. sphagnicola were associated with low PCA-2 scores (i.e.
shallower, high DOC lakes; e.g. Streeter Lake), others were associated with high PCA-2 scores
(e.g. Deer Pond).
ANOSIM did not detect a significant difference in the present-day chrysophytes
assemblage between lakes that have been reclaimed (R = 0.02, p = 0.32) or that had experienced
early watershed disturbance (Class 1 vs. Class 3: R = 0.001, p = 0.41). The ANOSIM tests had a
large enough sample size to compute a p < 0.05 if a significant difference existed between
groups (# permutation >> 999, Clarke and Warwick 2001). However, the similarity within each
group was low for each comparison (reclaimed: 66.2%, non-reclaimed: 51.2%; Class 1: 59.6%,
Class 3: 56.5%), due to the regional variation in chrysophyte assemblages. It may be that
chrysophyte assemblages are influenced by reclamation/early disturbance but that the sample
sizes used in the ANOSIM tests (n = 11-13) are too low to detect that influence, given the
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variability between sites. Thus, while it is possible that reclamation/early disturbance have no
influence on present-day chrysophyte assemblages, a more conservative conclusion is that the
influence of the stressors is small and less than the regional variability captured by
approximately one third of the reference sites.
Long-term changes in species composition: The first two axes of the PCA ordination explained
approximately 46% of the species variation in the present-day and pre-1900 assemblages of 26
Adirondack reference lakes (See Figure 4.3). In general, the relative abundance of colonial
chrysophyte taxa has increased in a majority of the reference lakes from pre-1900 to present (See
Figure 4.3 & Appendix D). The only colonial chrysophyte that generally had a higher relative
abundance in the pre-1900 samples was Synura curtispina. Unicellular taxa such as M.
duerrschmidtiae, Mallomonas pseudocoronota, Mallomonas crassisquama, and Mallomonas
lychenensis have generally declined in relative abundance from pre-1900 to present. In contrast,
Mallomonas punctifera appears to have increased since the 1900s, as has Mallomonas elongata.
Mallomonas hamata, an acid-tolerant chrysophyte species (Cumming et al. 1992a), was not
abundant in the reference lake dataset (i.e. relative abundance < 5%), although it has shown a
small increase in relative abundance in some lakes. Lastly, CI-pH increased, decreased, or
remained unchanged between the pre-1900 and present-day samples (See Appendix D). Only
two lakes (Middle Branch Lake and Bass Lake) showed a decline in CI-pH greater than the
RMSEP (0.51), suggesting that the majority of the reference lakes are unlikely to have acidified.
Wilcoxon signed-rank test found that the relative abundance of colonial chrysophytes
was significantly higher in the present-day than pre-1900 samples (Z = 3.9; p < 0.0001) (See
Figure 4.4). ANOSIM also showed a significant shift in species composition from pre-1900 to
present (R = 0.12; p = 0.002). The low R-coefficient is due to the low similarity between lakes
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within each time period (average similarity present-day: 62.3; average similarity pre-1900: 62.1).
However, the fact that there is a significant difference between time periods suggests that the
change in species composition across time periods is greater than the variability in species
assemblages seen spatially in the Adirondack Park in pre-1900 or present-day samples. The
SIMPER test found that recent increases in the relative abundances S. petersenii and S.
sphagnicola in the present-day samples make the largest contribution to the difference in species
composition between the two time periods (11.2% and 10.5% contribution, respectively) (See
Table 4.1).
Reproducibility & stability analysis: Quadruplicate cores from 4 reference lakes found that the
average BC similarity of replicate top samples (ABC-top) was 89.0% (range: 82.5% to 94.7%).
In comparison, all BC-similarities calculated between the 26 top and bottom reference lake
samples were less than the ABC-top value (average: 57.7%; range: 17.2% to 83.5%) and only
one top-bottom comparison had a similarity greater than the minimum value calculated between
any replicate top samples (82.5%). Thus, the species changes documented between top and
bottom samples in the reference lake study are generally greater than expected differences due to
counting variability alone.
Wilcoxon signed-rank test showed that the BC similarities were lower between replicate
top and 20 cm than replicate 20 and 30 cm samples (Z = 3.4, p < 0.0001). Similarly, ANOSIM
did not find a significant difference between the 20 cm and 30 cm samples of 10 top-bottom
reference lakes (R = -0.09, p = 0.95), suggesting that the assemblages were relatively stable
between 20 and 30 cm or, at least, that the variability between the two time periods was less than
the regional variability in species assemblages captured by the 10 reference sites.
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Discussion
Controls on present-day chrysophyte distribution in reference lakes: Even though the pH
gradient in the reference lakes was relatively narrow (~5.9 to 6.8), the pH/ANC gradient captured
by PCA axes 1 and 2 was tightly correlated with RDA axis-1. Numerous studies have shown the
importance of pH for determining chrysophyte species distribution, both in the Adirondacks and
elsewhere (Cumming et al. 1992a; Hyatt et al. 2010; Paterson et al. 2001; Siver and Hamer
1989). In this investigation, lakes with higher pH values and deeper maximum depths (higher
PCA-1 & PCA-2 scores) were characterized by a higher relative abundance of S. petersenii.
Interestingly, Flear (2011) found a similar separation in a study of 40 minimally-disturbed lakes
in the Experimental Lakes Area. Deep, high-pH lakes were characterized by a higher relative
abundance of S. petersenii and shallower, low-pH lakes characterized by a higher relative
abundance of S. sphagnicola, M. acaroides, and M. punctifera. In the Adirondack reference
lakes, S. sphagnicola was abundant both in shallower, high DOC sites (low PCA-2 scores) and
deeper, low DOC sites, suggesting that pH rather than DOC/depth per say was more important in
determining the abundance of S. sphagnicola. In contrast, a close examination of the DOC data
reveals that S. petersenii was less abundant (< ~10% relative abundance) in lakes with a DOC
concentration greater than 5.2 mg C L-1, which corresponds to a photic zone depth of ~5.2 m
(Bukaveckas and Robbins-Forbes 2000) (See Appendix E). Therefore, high DOC levels may
exclude S. petersenii from some reference lakes.
The fact that the RDA only explained approximately 19.1% of the variation in the
present-day chrysophyte assemblages suggests that one or more important variables were not
included in the analysis. Lake thermostructure, for example, is known to influence chrysophyte
species composition. Some chrysophyte taxa are considered warm-water species (M. punctifera)
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and during long periods of thermal stratification, large metalimnetic populations of both colonial
(Synura, Chrysosphaerella) and unicellular (Mallomonas caudata, M. acaroides, M.
crassisquama) taxa can occur (Nicholls 1995; Sandgren 1988; Siver 1995; Siver 2003). The fact
that maximum depth was important for structuring chrysophyte assemblages in the reference
lakes suggests that the thermostructure of lakes is likely an important control on chrysophyte
distribution, as shallower lakes are often associated with higher surface-water temperatures and
may be more prone to summertime mixing (Scheffer 2004). Biological factors may also be an
important control on chrysophyte species distribution. Large colonial chrysophytes likely have a
size refuge from small herbivores but may be susceptible to predation by large daphniids
(Sandgren and William 1995). A future paleolimnological investigation of the reference lakes
will include cladoceran zooplankton as paleo-indicators and so it will be possible to determine
what influence, if any, the presence/abundance of Daphnia has on chrysophyte species
distributions in the reference lakes.
The Adirondack reference lakes have experienced a significant shift in species
composition from pre-1900 to present, with the present-day samples characterized by an
increased abundance of colonial taxa (detailed discussion below). The RDA results suggest that
both physical and chemical characteristics may control the distribution of chrysophyte taxa. Of
particular interest are the controls on S. petersenii, a taxon that has been shown to cause taste and
odour issues in lakes and reservoirs (Nicholls and Gerrath 1985; Watson et al. 2001). The results
of this investigation suggest that S. petersenii is more common in deeper lakes with higher pH
values and DOC concentrations < ~5.2 mg C L-1. Nuisance S. petersenii blooms are thus unlikely
to be a concern in the many Adirondack lakes which are small (SA < 4 ha, Zmax < 5 m), relatively
acidic (pH < 6), and/or higher in DOC concentrations (DOC > 5 mg C L-1) (Krester et al. 1989).
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However, lakes which support sportfish like lake trout are often fairly large, deep, and
characterized by higher ANC and hence may provide suitable habitat for S. petersenii. Increases
in colonial chrysophyte taxa (and S. petersenii in particular) in these sites may therefore pose a
management concern (Paterson et al. 2004), and so it is important to discuss the possible causal
mechanisms behind the increases in colonial chrysophyte taxa noted in this investigation. Future
top-bottom and down-core paleolimnological studies of the reference lakes will build on the
discussion this manuscript provides by employing additional paleo-indicators and comparing
long-term changes in species composition with long-term monitoring data.
Changes in chrysophyte composition from pre-1900 to present: The reproducibility analysis
suggests that changes between the present-day and pre-1900 assemblages of the reference lakes
are unlikely to be due to counting variability. Similarly, the stability analysis suggested that the
change in chrysophyte species assemblages between the present day and 20 cm sample was
greater than changes earlier in the record (between 20 and 30 cm). Thus, we believe that the shift
in chrysophyte composition between 20 cm and the present-day samples in the top-bottom study
is interpretable and merits investigation.
Both univariate and multivariate analyses found that there has been a significant increase
in the relative abundance of colonial taxa from pre-1900 to present, along with increases in some
unicellular taxa in some lakes (M. punctifera, M. elongata). Because of the regional distribution
of the reference lakes, we can conclude that a regional rather than local stressor is likely
responsible for the changes in chrysophyte species composition. Furthermore, because of the
reference lake study design, we can conclude that these increases are not due to: acidification,
eutrophication (from shoreline or watershed development), salinization, or introduced piscivores.
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ANOSIM results also suggest that both historic fisheries disturbance and watershed disturbance
have minimal influence on the present-day chrysophyte taxa in the reference lakes.
Given the large impact atmospheric deposition has had in the Adirondack region, it is
important to consider whether or not the changes in chrysophyte species composition noted in
this investigation could be due to long-term effects of acid deposition. CI-pH reconstructions
suggest that the majority of the reference lakes are unlikely to have acidified (i.e. experienced
decreases in pH/ANC) but atmospheric deposition may have effects outside acidification.
Increased leaching of cations with acid deposition is unlikely to be an important driver of the
species changes noted from pre-1900 to present as chrysophyte taxa are broadly tolerant of a
range of cations, though increased long-distance transport and deposition of certain trace metals
may be important (i.e. Se, a micronutrient shown to stimulate chrysophyte blooms, Sandgren
1988). Long-term exposure to acid deposition can also increase the ionic strength of soil
solutions, decreasing DOC flux to nearby waterbodies (see Monteith et al. 2007 and citations
therein). Decreases in DOC flux, in turn, would increase light penetration in the water column
and reduce suboptimal low-light conditions for metalimnetic colonial chrysophyte populations
(Healy 1983). However, while S. petersenii may have benefitted in lakes where DOC decreased
below 5.2 mg C L-1, S. sphagnicola and S. echinulata appear to be DOC-indifferent and so
changes in DOC concentrations cannot explain increases in those taxa. Finally, long-term
exposure to acid deposition may decrease the flux of phosphorus to lakes by increasing the
adsorption of TP in watershed soils (Eimers et al. 2009). Furthermore, logging and subsequent
forest regrowth may also decrease TP export to nearby lakes, and many of the reference sites
experienced early logging disturbances (Hall and Smol 1996). The fact that M. lychenensis and
S. curtispina decreased in abundance from pre-1900 to present in some lakes suggests that total
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phosphorus (TP) concentrations may have decreased in those sites, as both taxa are associated
with higher lake trophic status (Siver and Marsicano 1996; Siver 1995). Furthermore, in low-TP
conditions, the metalimnetic position of colonial chrysophytes may provide them a competitive
advantage by allowing access to nutrient-rich hypolimnetic waters (Nicholls 1995). Interestingly,
though unicellular taxa generally declined in relative abundance from pre-1900 to present, a few
lakes showed increases in the relative abundance of M. caudata and M. elongata, both of which
can form metalimnetic blooms (Fee et al. 1978; Siver 2003). Thus, oligotrophication may have
contributed to the changes in chrysophyte assemblages noted in the reference lakes.
Regional warming may have also contributed to the changes in species composition noted
in the reference lakes from pre-1900 to present. The Adirondack region has warmed are a rate of
1.3ºC per 100 years (Jenkins 2010), a warming trend associated with an increased ice-free period
in Adirondack lakes (Beier et al. 2012). The increases in M. punctifera (a warm-water taxon,
Siver 1991) noted in some reference lakes may therefore be due to warming surface waters.
Similarly, an extended ice-covered season may benefit taxa like S. sphagnicola which persists in
the water column through summer and fall and is lost at ice-on (Siver and Hamer 1992). Longer
ice-free periods can also lead to a longer period of summer stratification (Vincent 2009), which
may provide motile algae like chrysophytes a competitive advantage over non-motile algae like
diatoms because they can seek out optimal temperature/nutrient/light conditions in the water
column (Winder and Sommer 2012). At the assemblage level, longer periods of stratification
may provide a competitive advantage to colonial chrysophyte taxa over unicellular ones because
large flagellates tend to be superior swimmers, with faster swim velocities and larger migration
amplitudes (Sommer 1988). Longer periods of stratification can also result in increased nutrient
limitation in the epilimnion (Vincent 2009), providing chrysophytes that form metalimnetic
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blooms with a competitive advantage because of their access to nutrient-rich hypolimnetic
waters. Thus, the increases in colonial taxa noted in this investigation, along with increases in
certain unicellular taxa like M. punctifera, are likely related at least in part to regional warming.
It is possible that the long-term effects of regional warming and/or acid deposition may
have also resulted in changes in biotic conditions that could have contributed to the increased
abundance of colonial chrysophytes. Large colonial chrysophytes likely have a size refuge from
small zooplankton (Bosmina, calanoid and cyclopoid copepods) but can be predated upon by
large Daphnia taxa (i.e. > 1 mm in size) (Sandgren and William1995). It has been hypothesized
that regional warming may benefit small zooplankton (Daufresne et al. 2009). Furthermore,
long-term exposure to acid deposition may lead to calcium depletion in watershed soils (Likens
et al. 1996), a phenomenon that can negatively affect Ca-sensitive Daphnia (Jeziorski et al.
2008). A subsequent paleolimnological study of cladoceran assemblages in the reference lakes
will determine whether or not cladoceran assemblages have undergone a significant shift in
species composition from pre-1900 to present and what link, if any, those changes may have to
changes in colonial chrysophyte abundance in Adirondack lakes.
Importance of reference lake results for recovery studies: Arguably, the most important finding
of this top-bottom study is simply that minimally-disturbed reference lakes in the Adirondack
Park show a pronounced shift in species composition from pre-1900 to present, likely due to
climate warming and/or oligotrophication. Thus, it should be expected that lakes in the area
recovering from acidification will not return to their pre-disturbance state. Rather, the recovery
endpoints for acidified lakes will be one that is different from their pre-disturbance condition – a
state characterized by an increased abundance of colonial and/or warm-water chrysophyte taxa.
By repeating the top-bottom analysis of the reference sites with other paleo-indicators (diatoms,
123
cladocera, chironomids, etc.), we will be able to determine whether or not such shifting baselines
are occurring in a suite of aquatic organisms. Moreover, by pairing both acidified and reference
lakes in subsequent paleolimnological studies, we will be able to determine whether or not the
recovery trajectories of impacted sites suggest that they are moving towards a novel ‘recovered’
condition similar to the reference sites (Chapter 3). While a growing number of studies have
shown chemical recovery from acidification, lakes are increasingly influenced by multiple
anthropogenic stressors, making the return of biological assemblages to a pre-disturbance state
unlikely. Paleolimnology can and should play an important role in recovery studies because
paleolimnological techniques can both identify species changes in impacted sites and, as we have
illustrated, be applied in reference lakes to identify shifting baselines and define reasonable
bounds for recovery targets.
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Tables & Table Captions
Table 4.1. SIMPER test results showing taxa that contributed to the significant difference in species
composition between top and bottoms samples. ANOSIM identified significant differences between top
(present-day) and bottom (pre-1900) samples in the 26 reference lakes (ANOSIM R = 0.12, p = 0.002).
Taxa that made a > 5% contribution (% Cont.) to the dissimilarity between groups are listed. Average
square-root relative abundance (%) of the species in the top (Av √RA TOP) and bottom (AV √RA
Bottom) groups are provided, with the non-square root transformed average relative abundance given in
brackets.
Species
S. petersenii
S. sphagnicola
M. duerrschmidtiae
M. ‘small’
M. pseudocoronata
S. echinulata
M. crassisquama
Chrysosphaerella spp.
M. acaroides
Av √RA
TOP (%)
3.4 (16.8)
3.2 (15.3)
3.7 (15.2)
2.9 (11.0)
1.2 (2.7)
2.6 (8.5)
2.4 (7.3)
1.5 (3.8)
1.2 (2.6)
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Av √RA
Bottom (%)
2.02 (7.0)
2.4 (8.9)
5.4 (31.5)
2.9 (11.6)
2.3 (8.9)
1.6 (4.4)
2.7 (8.5)
0.8 (1.3)
1.3 (2.6)
% Cont.
11.2
10.5
9.6
8.4
8.1
7.3
6.1
5.5
5.3
Figure & Figure Captions
Figure 4.1. Schematic showing the paleolimnological study design. The study consists of
two parts, a top-bottom study (a) and a reproducibility study (b). The reproducibility study
consists of two parts, an analysis of counting variability (i) and a stability analysis (ii). i)
Counting variability. The Bray-Curtis (BC) similarities between top (present-day) samples
from 4 replicate cores are computed separately for 4 reference lakes. An average BCsimilarity (ABC-top) is then computed using all comparisons from all lakes. ii) Stability
analysis. BC-similarities are calculated for all comparisons between top vs. 20 cm samples
and 20 vs. 30 cm samples from 3 replicate cores and are computed separately for 3
reference lakes. A Wilcoxon signed-rank test is then performed to compare the BC
similarities from top vs. 20 cm samples and 20 vs. 30 cm samples from all lakes.
131
132
PCA λ4 = 0.10
PCA λ2 = 0.21
PCA λ3 = 0.14
RDA λ1 = 0.14
Figure 4.2. Principal Components Analysis (PCA) of environmental variables and Redundancy analysis (RDA) using PCA-axes scores as input. A) Biplot of
PCA axes 1 & 2. B) Biplot of PCA axes 3 & 4. Input variables are: pH, acid neutralizing capacity (ANC, μeq L-1), K, Na, SO42- (μmol L-1), Cl-, Mg2+ (μmol L1
, log transformed), dissolved organic carbon (DOC, μmol C L-1, log transformed), total nitrogen (TN, μmol L-1, log transformed), Elevation (elv, m), and
maximum depth (Zmax, m, log transformed). C) RDA of present-day samples from 31 reference lakes. PCA axes (environmental variables) are open arrows,
species are closed arrows, and samples are shown as points. Eigenvalues (λ) are given.
PCA λ1 = 0.27
RDA λ2 = 0.05
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PCA λ2 = 0.22
PCA λ2 = 0.22
PCA λ1 = 0.24
PCA λ1 = 0.24
Figure 4.3. Principal Components Analysis (PCA) of chrysophyte assemblages in top (present-day) and bottom (pre-1900) samples in 26 reference lakes.
The PCA is summarized in multiple panels to reduce crowding in biplots. A) Species vectors. B) Present-day (filled circle) and pre-1900 (empty square)
samples. Dashed lines show trajectory of species change in ordination space from pre-1900 to present. C) Present-day samples only (labelled). Eigenvalues
(λ) are given.
PCA λ1 = 0.24
PCA λ2 = 0.22
Figure 4.4. Total relative abundance (%) of colonial chrysophyte taxa in top (present-day)
and bottom (pre-1900) samples in 26 reference lakes. Dashed line represents a one-to-one
line. All samples above the grey dashed line have shown an increase in the total relative
abundance of colonial chrysophytes from pre-1900 to present.
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CHAPTER 5
UNDERSTANDING BIOLOGICAL RECOVERY FROM ACIDIFICATION: APPLYING
A REFERENCE SITE FRAMEWORK IN THE ADIRONDACKS (NY, USA)
Abstract
Since the implementation of the US Acid Rain Program (US ARP), lakes in some areas
of the United States have begun to undergo chemical recovery from acidification (e.g. increases
in pH/ANC, decreases in aluminium concentrations). Chemical recovery is expected to promote
biological recovery from acidification. However, it is often difficult to assess biological recovery
due to a lack of long-term monitoring data. Furthermore, climate warming may prevent
recovering lakes from returning to their pre-acidification state, even with management
intervention. In such a scenario, a reasonable recovery target can be defined by comparing
impacted species assemblages to those of higher pH/ANC reference lakes that have experienced
limited effects of acid deposition. These reference systems can be used to identify regional
changes in species assemblages. In this study, we use paleolimnological techniques within a
reference site framework to provide evidence that biological recovery from acidification is
occurring in two Adirondack (NY, USA) lakes undergoing chemical recovery. A set of three
acid-insensitive reference lakes were used to compare and contrast recent changes in species
composition in the acidified lakes and to identify the possible influence of climate warming on
Adirondack lakes. The two lakes recovering from acidification showed a significant shift in
chrysophyte species composition after the 1995 implementation of the US ARP. Declines in
acid-tolerant chrysophyte taxa indicate a modest biological recovery. However, both the
acidified lakes and reference lakes show recent increases in the relative abundance of colonial
chrysophytes similar to patterns documented elsewhere in North America. A shift toward novel
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chrysophyte assemblages since ca. 1970-1980 is illustrated in non-metric multidimensional
scaling (nMDS) ordinations. The relative abundance of colonial taxa was significantly correlated
with mean annual air temperature and ice-cover measures in two reference lakes. While modest
biological recovery has occurred in Adirondack lakes following a decrease in acid deposition, the
chrysophyte assemblages do not appear to be returning to their pre-acidification state, likely due
to recent climate warming and/or oligotrophication.
Introduction
Acid deposition has impacted aquatic ecosystems across North America. Declines in
pH/ANC (acid neutralizing capacity), increases in toxic inorganic monomeric aluminium, and
other chemical and biological changes associated with long-term exposure to acid deposition
result in the progressive loss of acid-sensitive taxa and a general simplification of aquatic
foodwebs (Lovett et al. 2009). With the implementation of Title IV of the 1990 Clean Air Act
Amendments (CAAA) and the 1995 implementation of the US Acid Rain Program (ARP),
sulphur dioxide (SO2) emissions in the United States from sources governed by the US ARP
have decreased 64% below their 1990 levels as of 2009 (Burns et al. 2011). In addition, there has
been an ~70% decrease in nitrogen oxide (NOx) emissions from 1990 to 2009 as a result of the
1990 CAAA and the U.S. Environmental Protection Agency Nitrogen Budget Program (Waller
et al. 2012). Consequently, many lakes in the United States have begun to undergo chemical
recovery from acidification (Garmo et al. 2014; Skjelkvåle et al. 2005).
In acid-impacted regions, it is expected that chemical recovery will lead to biological
recovery from acidification, characterized by a decline in the abundance of acid-tolerant species
and an increase in the abundance of acid-sensitive species as lake pH and ANC increase over
time (e.g. Driscoll et al. 2001). However, studies attempting to identify biological recovery are
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often hampered by a lack of long-term biological monitoring data. Fortunately, this problem can
be addressed using paleolimnological techniques. Paleolimnological studies infer the
environmental histories of lakes using the physical, chemical, and biological characteristics in
dated sediment cores (Smol 2008). Paleolimnological techniques can be used to both identify the
onset of acidification and provide evidence of biological recovery (Battarbee 1999).
In paleolimnological studies, a lake’s recovery target is usually defined as its predisturbance condition (e.g. ca. 1850, Battarbee et al. 2011). However, a lake responding to
multiple stressors over the recovery period is unlikely to return to its pre-disturbance state, even
with management intervention (the ‘shifting baseline’ phenomenon, Hobbs et al. 2009; Bennion
et al. 2011). In such a scenario, reasonable recovery targets can be defined by comparing
impacted species assemblages to those of minimally-disturbed reference sites protected from
multiple anthropogenic stressors (Bailey et al. 2004). Pairing acidified lakes with reference
systems allows researchers to distinguish between species responses to acidification and
chemical recovery versus responses to regional stressors such as climate change (Chapter 3). In
this framework, the reference sites do not serve as ‘controls’ for the impacted sites in the
traditional sense (i.e. sites with identical species composition, morphology, etc., Downes et al.
2002) but are, however, similar to the impacted sites in general characteristics (e.g. deep
oligotrophic lakes with predominently forested catchments). As minimally-disturbed sites, the
reference lakes serve as ‘sentinel’ systems that allow the detection of species’ responses to recent
warming (USEPA 2012). Despite these benefits, reference sites are under-utilized in recovery
research (Jones and Schmitz 2009).
The Adirondack region represents a unique opportunity to highlight the benefits of using
paleolimnological techniques to study biological recovery within a reference-site framework.
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The Adirondacks have been heavily impacted by acid deposition (Driscoll et al. 1991) and the
acidification history of the area has been well-documented in previous paleolimnological studies
(Cumming et al. 1992b; Cumming et al. 1994). Following declines in acid deposition, many
lakes in the region are undergoing chemical recovery from acidification (Driscoll et al. 2007;
Waller et al. 2012). However, while chemical recovery may be progressing, the region is also
experiencing the effects of climate change. In the Adirondacks, mean annual temperature has
increased at a rate of 1.3°C per 100 years, a warming trend that has been most pronounced since
the 1970s (Jenkins 2010) and is coincident with a shortened ice-covered period on lakes (Beier et
al. 2012).
To examine the possible influence of regional stressors like climate change on
Adirondack lakes, Chapter 3 identified a regional set of 31 minimally-disturbed reference lakes
from an online database of 1,469 lakes. The sites were chosen using a set of stringent selection
criteria and have been largely protected from: acidification, eutrophication from shoreline or
watershed development, road salt seepage, and piscivores introductions. The reference lakes
were studied in a top-bottom paleolimnological study, a comparison of species assemblages from
the present day and pre-1900 times (Cumming et al. 1992b). The top-bottom analysis lakes
revealed that scaled-chrysophyte assemblages (algae belonging to the Synurophyceae and
Chrysophyceae) showed a significant shift in species composition from pre-1900 to present,
including an increased abundance of colonial taxa, taxa that form metalimnetic blooms, and
warm-water species (Chapter 4). Similar changes have been documented elsewhere in North
America (Paterson et al. 2001; Ginn et al. 2010; Flear 2011) and have been linked, at least in
part, to recent warming. Because of the regional reference lake study design, were able to
conclude that the changes in species composition noted were not due to acidification (decreases
138
in pH/ANC), eutrophication, road salt seepage, historic watershed disturbance, and fisheries
alterations. We concluded that the species changes were likely due to the long-term effects of
regional warming or oligotrophication. Moreover, we suggested that since Adirondack reference
lakes had undergone a significant shift in species composition over time, lakes recovering from
acidification are unlikely to return to their pre-disturbance state.
This manuscript builds on the work of Chapter 4 and examines long-term changes in
species composition in a subset of the regional reference lakes, pairing them with a set of lakes
recovering from acidification. We hypothesized that the lakes recovering from acidification
would show some evidence of biological recovery (i.e. declines in the abundance of acid-tolerant
taxa) but would also show novel changes in species composition similar to those documented in
the reference lakes. The following research questions are addressed:
1) Do the acidified lakes show a significant shift in species assemblage after the 1995
implementation of the US ARP? If so, is the change consistent with biological recovery
from acidification?
2) Are the recent (post-1995) species assemblages of the acidified lakes moving towards
their pre-acidification state? How do the changes in species composition compare with
those noted in the subset of reference lakes? Can reference sites be used to extrapolate
probable endpoints in the recovery trajectories of the acid-impacted sites? and,
3) Is the relative abundance of colonial chrysophytes correlated with climate variables such
as mean annual air temperature or the duration of ice cover?
In addition to providing information on the ecological dynamics of biological recovery
from acidification (a subject that is not well documented, Verdonschot et al. 2013), this study
will illustrate the advantages of pairing paleolimnological techniques within a regional referencesite framework to untangle the effects of two important stressors on aquatic systems:
anthropogenic acidification and climate change.
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Methods
Site selection rationale: Sediment cores from five lakes were collected and examined in this
investigation: two from acidified lakes and three from reference lakes. The two acidified lakes
(South Lake, N 43°30’35”, W 74°52’35”; and Queer Lake, N 43°48’49”, W 74°46’38”) were
chosen because previous paleolimnological investigations demonstrated that the lakes acidified
under high loads of sulphate and nitrate deposition (South Lake ca.1930-1950, Cumming et al.
1994; Queer Lake in the late 1970s, Charles et al. 1990). Long-term chemical data indicate that
both lakes are undergoing chemical recovery from acidification (See ‘chemistry trends’ below
and Figure 5.1A).
Three reference lakes were chosen from a regional set of 31 reference lakes (Chapter 3).
Moss Lake (N 43°46’52”, W 74°51’11”) and Arbutus Lake (N 43°58’58”, W 74°14’09”) were
included because they have been subject to long-term monitoring since the early 1980s (See
‘chemistry trends’ below and Fig 4.1B). Wolf Lake (N 44°01’42”, W 74°13’16”) was included
because it is an Adirondack ‘Heritage lake’ (Stager and Sanger 2003) and is considered one of
the most ecologically intact lakes in the Adirondacks (Beier et al. 2012).
Site descriptions: Detailed descriptions of Arbutus Lake, Moss Lake, South Lake, and Queer
Lake are provided by Roy et al. (2011) and a detailed description of Wolf Lake is provided by
Stager and Sanger (2003) (see also Chapter 3). All of the study sites are relatively large (> 40
ha), deep (max depth ~8-21 m), and oligotrophic (TP < 10 μg L-1) drainage lakes located at
elevations greater than 500 m. The lakes’ watersheds are predominantly forested with either a
mixed deciduous-coniferous forest or northern hardwood forest. The lakes all have fish
communities including both piscivorous and planktivorous fish and all the lakes, except for Wolf
140
Lake, have been stocked with brook trout and/or lake trout in the past (Appendix F). The
watersheds of Arbutus Lake and Wolf Lake are protected within the Huntington Wildlife Forest,
a property owned by the State University of New York College of Environmental Science and
Forestry (SUNY-ESF) that was established in 1932. The watersheds of Moss Lake, Queer Lake,
and South Lake have been protected since the early 1970s in ‘wild forest’ or ‘wilderness’ areas
in the Adirondack Park where logging and watershed development are prohibited. The
watersheds of Moss Lake and Queer Lake lie entirely within protected areas while 56% of the
South Lake watershed is protected (Appendix F). The lakes have, however, all experienced
forms of early disturbance, primarily from selective logging prior to the 1920s and forest
‘blowdown’ from a windstorm in 1950. The watershed of South Lake is the most disturbed as it
experienced heavy forest blowdown in 1950 and currently has cottages on its lakeshore
(Appendix F).
Chemistry trends: Though not the main focus of this investigation, we summarize long-term
chemistry trends in both the acidified and reference lakes to contextualize any drivers of
biological change in these lakes. Detailed field and analytical methods used in the collection and
analysis of long-term chemistry data are available in Driscoll and van Dreason (1993).
South Lake had a pre-industrial chrysophyte-inferred (CI) pH of ~5.8 (Cumming et al.
1994) and Queer Lake had a pre-industrial diatom-inferred (DI) pH of ~6.4 (Charles et al. 1990).
The CI-pH of South Lake began to decline ca. 1930-1950 and was near or below 5.0 in the early
1980s. The DI-pH of Queer Lake began to decline in the late 1970s and decreased below 5.5 in
the early 1980s. In 1984, the measured pH of South Lake was 5.2 and ANC was 2.9 μeq L-1
(Roy et al. 2011). In 1986, the measured pH of Queer Lake was 5.5 and ANC was 1.9 μeq L-1
(Roy et al. 2011). From 1992-2010, chemistry records show that both lakes experienced a
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significant decline in the concentration of strong acid anions (sulphate and nitrate), inorganic
monomeric aluminium, and specific conductance (Figure 5.1A & Appendix G). Likewise, both
lakes have experienced significant increases in ANC, pH, and dissolved organic carbon (DOC)
over the same period (Figure 5.1A & Appendix G).
The reference lakes also show long-term changes in water chemistry. From 1982-2010 in
Moss Lake and from 1983-2010 in Arbutus Lake, there was a significant decline in the
concentration of strong acid anions and specific conductance (Figure 5.1B & Appendix G). Moss
Lake has also shown a significant increase in pH, ANC, and DOC since 1982. The increase in
pH and ANC in Moss Lake appears to be due to a decline in the strength of episodic
acidification. Long-term water chemistry data are not available for Wolf Lake. However,
sporadic pH measurements show that Wolf Lake had a pH between 6.5 and 7.0 from 1950-1997
(Stager and Sanger 2003).
Sediment collection and preparation: Sediment cores were collected from the deepest basin of
the study lakes in May 2008 (South Lake, Moss Lake), November 2009 (Queer Lake), and May
2010 (Wolf Lake, Arbutus Lake). Sediment cores (~20-30 cm in length) were collected using a
Glew gravity corer with an internal diameter of 7.6 cm and were sectioned on site into 0.25-cm
intervals using a modified Glew vertical extruder (Glew et al. 2001).
Scaled chrysophytes (Synurophyceae and Chrysophyceae; hereafter referred to as
‘chrysophytes’) are used in this investigation because they are sensitive to pH/aluminum levels
and because their characteristic siliceous scales are well-preserved in lake sediments (Cumming
et al. 1992a). Slides were prepared for chrysophyte enumeration following the methods of
Battarbee et al. (2001). Briefly, 0.2-0.3 g of wet sediment was transferred to a labelled glass vial
and digested in approximately 20 mL of concentrated HNO3-H2SO4 to isolate the siliceous
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chrysophyte scales. Samples were heated in a hot water bath to ~70°C for 6 hours and then
rinsed 8-9 times until the samples were acid-free (litmus test). The samples were plated as a
series of four 100% dilutions and mounted on microscope slides using Naphrax®.
Activities of 210Pb, 137Cs, and 214Bi were measured following the methods of Schelske et
al. (1994). Briefly, ~20 sediment intervals per core were freeze-dried using a VirTis Advantage
freeze-drier (SP Industries, Inc.). Freeze-dried sediment was added to polypropylene tubes,
capped with 2-Ton epoxy, and kept at room temperature for at least two weeks before 210Pb,
137
Cs and 214Bi activities were measured in a high-purity germanium coaxial well-detector with a
4-mm active well depth with an internal diameter of 15.5 mm (EG& G Ortec) attached to a
DSPEC digital spectrometer. Sediment age was calculated using the constant rate of supply
(CRS) model (Appleby and Oldfield 1978). In cases where sediment intervals were inferred to
represent < 1 year, species counts were combined in adjacent intervals to provide a minimum
yearly resolution throughout the profile.
Chrysophyte scales were enumerated on a Leica DMBR microscope using differential
interference contrast (DIC) optics. A minimum of 400-500 chrysophyte scales were counted for
samples from South Lake, Queer Lake, Moss Lake, and Wolf Lake. A minimum of 200 scales
were counted for samples from Arbutus Lake, as the abundance of scales was low prior to 1980.
The principle taxonomic references used in this investigation were: Kling and Kristiansen
(1983), Nicholls and Gerrath (1985), Siver (1991), and unpublished reference materials. A
Mallomonas ‘small’ category was used to group small Mallomonas taxa that are not easily
distinguished using light microscopy (Cumming et al. 1992a).
Statistical analyses – understanding long-term changes in chrysophyte assemblages: Unless
otherwise stated, chrysophyte taxa that reached a relative abundance > 2% at least twice in the
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sedimentary profile were included in the following analyses. CI-pH values were generated for
the acidified and reference lakes from a calibration set of 71 Adirondack lakes (Cumming et al.
1992a). CI-pH values were calculated based on a weighted-averaging model with tolerance
down-weighting and inverse deshrinking with a ln (x+1) species transformation (Cumming et al.
1994). Stratigraphies of the relative abundance of chrysophyte taxa were created using the
computer program C2 v. 1.7.2 (S. Juggins, unpub. program, 2011) and are included in an
appendix (Appendix H).
Ordination techniques were used to summarize long-term changes in species composition
in the acidified and reference lakes. Non-metric multidimensional scaling (nMDS) plots were
used to summarize long-term (pre-1900 to present) changes in species assemblages in the
acidified and reference lakes. nMDS is based on a ranked matrix of Bray-Curtis (BC) distances
(dissimilarities), the twin concept to BC similarity (BC similarity = 1 – BC dissimilarity).
Because nMDS is based on a ranked distance matrix, the technique avoids the assumptions other
ordination methods like PCA make about the underlying structure of the species data. nMDS
simplifies multidimensional species data into two (or three) dimensions and sedimentary
assemblages with a similar species composition will plot close together in the ordination
diagram. Thus, if recent species assemblages are in close proximity to the pre-acidification
assemblages in the ordination diagram, it can be inferred that the recent assemblages are moving
towards their pre-acidification state. While both PCA and nMDS are appropriate for illustrating
long-term trends in species assemblages, nMDS was used in this investigation because the
ordination is based directly on an underlying dissimilarity matrix. This consideration was
important for an ordination used to illustrate how similar the pre-1900 species assemblages of the
acidified lakes were to the pre-1900 species assemblages of a set of regional reference lakes (See
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Statistical analyses – identifying a return to pre-disturbance state and recovery endpoints). PCA
ordinations of long-term species changes in both the acidified and reference lakes are provided in
an Appendix for comparison (see Results below).
When designating reference periods and recovery targets in paleolimnological studies, it
is important to consider factors such as landscape disturbance (Renberg et al. 2009). Known
watershed disturbances in the acidified lakes and reference lakes were minimal prior to ca. 1900
(Appendix F). As such, we defined four time-periods in the nMDS for each lake: pre-disturbance
(pre-1900), early disturbance (1900 to pre-1950), acidification (1950 to 1995), and post-ARP
(post-1995 to present). The sedimentary assemblages of the reference lakes were grouped in the
same time periods to allow for comparison of species changes between the reference and
acidified lakes. nMDS plots were created using BC distances of square-root transformed species
data. Permutations tests (999 permutations) were used to identify the nMDS ordinations
summarized in 2 axes with the lowest stress (stress equation 2; primary distance measures).
nMDS ordinations were performed on square-root transformed species data using the computer
program CANOCO v. 5.0 (ter Braak & Šmilauer 2012).
Statistical analyses – quantifying biological recovery: An Analysis of Similarities (ANOSIM)
(Clarke 1993) was used to determine if the species assemblages in the acidified lakes showed a
significant shift in species composition after the 1995 implementation of the US ARP in
comparison to a time period of equal length prior to 1995 (details below). ANOSIM is a nonparametric test of ranked similarities that tests for a significant difference in species
compositions between a priori defined groups in multivariate data. Subsequently, the similarities
percentage test (SIMPER) can be used to identify which taxa contribute to the differences in
species composition seen between groups.
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In each core, two ~ten-year time periods were identified prior to and after the 1995
implementation of the US ARP. An analysis of long-term monitoring data from Adirondack
lakes showed no systematic increase in pH or ANC from 1982-1994 (Driscoll et al. 1995). Thus,
in the absence of pre-1992 long-term chemical monitoring data in South Lake and Queer Lake,
the ANOSIM framework in this investigation attempts to designate two time periods: one period
when the lakes were still acidic (Period 1, pre-1995) and one period when they are undergoing
chemical recovery (Period 2, post-1995). We interpreted a decline in the relative abundance of
Mallomonas hamata and Mallomonas hindonii (acid-tolerant taxa associated with low pH and
high concentrations of inorganic monomeric aluminium, Cumming et al. 1992a), and a
corresponding increase in taxa with higher pH optima in Period 2 as evidence of biological
recovery from acidification. The approach used here is conservative in that it recognizes that
210
Pb dating has errors that are difficult to quantify (Binford 1990). By delineating two ten-year
time periods and treating sedimentary intervals within each time period as replicates, major
trends in species composition may be identified. ANOSIM (using BC similarity coefficient) and
SIMPER were performed on square-root transformed species data using the computer program
PRIMER v. 6.1.11.
Statistical analyses – identifying a return to pre-disturbance state & recovery endpoints: By
plotting the species assemblages of the acidified sites in nMDS ordination space, we can identify
how the assemblages have changed over time and identify if the recent post-1995 assemblages
have moved back towards a pre-disturbance (pre-1900) condition. While informative, this
approach is somewhat limited as it is only able to identify species changes that have occurred in
South Lake and Queer Lake up until the present and does not identify probable recovery
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endpoints (i.e., what species assemblages are likely to occur in the lake when chemical recovery
is complete?)
We can expand on the long-term nMDS-ordination analysis for the acidified lakes by
comparing the species trends in the acidified lakes to those documented in the regional set of 31
reference sites. First, we actively plotted the pre-1900 sedimentary assemblages of the acidified
lakes with all pre-1900 samples from the regional reference lakes (the ‘bottom’ samples from the
top-bottom analysis; 26 of 31 reference lakes had a sample from pre-1900). This allows us to
identify which reference lakes were most similar in species composition to the acidified lakes in
pre-disturbance (pre-1900) times. In a subsequent ordination, we actively plot the pre-1900 and
present-day samples of a subset of the reference lakes that were similar in species composition to
the acid-impacted lakes with the long-term record of each acidified lake. Active plotting is
necessary in these ordinations so that taxa present in either the acidified lakes but absent from the
reference lakes (such as M. hindonii) or vice versa contribute to the ordination. The ordination
space inhabited by the present-day assemblages of the reference lakes represents a hypothetical
‘endpoint’ for biological recovery in the acidified lakes – a space that may be both different from
the pre-disturbance condition and the present-day species composition of the acidified.
The assumption underlying the above approach is that the species assemblages of the
acidified and reference lakes were similar pre-1900 because the physical/chemical/biological
conditions of the lakes were similar. Thus, changes in the reference lakes over time can be used
as a proxy of changes in the acid-impacted sites. Presumably, reference sites that were more
similar to the acid-impacted sites pre-1900 provide better proxies for expected changes in the
recovering acidified lakes (i.e. sites 70-80% similar are better analogues than sites 40-50%
similar). Taxa that made a > 2% contribution to more than two samples in the combined dataset
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of the regional reference lakes and South Lake/Queer Lake were included in the ordination. The
two reference lakes that had the highest average BC similarity with the pre-1900 samples from
each acid-impacted lake in the first nMDS ordination were included in the second nMDS
ordination. nMDS plots were created in CANOCO v. 5.
Statistical analyses - correlations with climate variables: A Mann-Kendal correlation test (Helsel
and Hirsch 2002) was used to determine if there was a relationship between total relative
abundance of colonial chrysophytes and selected climate variables. The Mann-Kendall test is a
non-parametric test that identifies if a monotonic trend exists between two data series (i.e.
generally increasing or generally decreasing). Note that a ‘strong’ Mann-Kendall correlation
typically has a smaller correlation coefficient than an equivalently strong Pearson correlation
coefficient (i.e. R = 0.9 ≈ τ = 0.7, Helsel and Hirsch 2002). The correlation between total relative
abundance of colonial taxa and ice-on date, ice-off date, and ice-cover duration were examined
for Arbutus Lake and Wolf Lake from ca. 1970 to present (the only sites with long-term icecover data). Only years coinciding with the 210Pb-inferred dates of the sedimentary intervals were
included in the temperature and ice cover data. The length of the ice-covered season has been
shown to influence the timing and composition of spring phytoplankton blooms (Gerten and
Adrian 2002). Therefore, in this investigation we compared ice-duration data to sedimentary
intervals whose 210Pb-inferred dates matched the year in which ice-off occurred. Mann-Kendall
tests were performed using the computer program of Helsel et al. (2006). Ice cover data were
provided by Dr. C. Beier (Adirondack Ecological Center, SUNY- ESF). Quality-controlled
mean-annual air-temperature data were provided by the United States Historical Climatology
Network (Menne et al. 2009) and were from the meteorological station located within the
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Adirondack Park nearest to each lake: Station 304102 Arbutus Lake (~ 26 km) and Wolf Lake (~
30 km).
Results
210
Pb activities and dates: All lakes showed exponential decay of unsupported 210Pb with
cumulative dry mass in the core (Figure 5.2). 210Pb activity in the Wolf Lake sediment core did
not reach background activity and so unsupported 210Pb activity was estimated using background
levels from a previously collected sediment core (Stager and Sanger 2003). 137Cs profiles did not
show a pronounced peak in activity, which is common in the highly-organic sediment typical of
Adirondack lakes (Heit and Miller 1987). Standard errors of inferred-dates were low (< 1 year)
in recent (post-1950) sediment intervals. Age-depth models were calculated based on a quadratic
equation for South Lake, Moss Lake, and Arbutus Lake. For Queer Lake and Wolf Lake, a cubic
equation was used as quadratic equations under-estimated the age of sediment at depths < 6-7
cm.
Long-term changes in species composition
Acidified lakes: In general, the acidification histories of South Lake and Queer Lake are
similar to previously published results (Cumming et al. 1994, Charles et al. 1990). The CI-pH of
South Lake began to decline ca. 1940-1950 and reached its lowest values in the 1980s (~5.0).
Since the mid-1990s, there has been a modest increase in CI-pH, though values remained low
(<5.5) (See Figure 5.3). In Queer Lake, the CI-pH began to decline in the 1970s, decreased
below 6.5 in the 1980s, and has shown little change since.
Stress tests of nMDS biplots of long-term species assemblage data from South Lake and
Queer show that the 2-axes of the nMDS provide a good approximation of the multivariate data
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(stress = 0.10; stress = 0.07, respectively) (See Figure 5.4 – species codes listed in Appendix I).
The nMDS plots showed a similar separation between time-periods as PCA plots made on
square-root transformed species data (See Appendix J). In both lakes, the pre-disturbance (pre1900) and early disturbance (1900 to pre-1950) species assemblages were similar in
composition, though some pre-1950 assemblages in South Lake had a high relative abundance of
S. echinulata, a colonial chrysophyte (See Figure 5.4 & Appendix H, Figure H1). The pre-1950
assemblages in both lakes are characterized by a higher relative abundance of taxa with higher
pH optima such as M. acaroides (pH optimum = 5.9, Cumming et al. 1992a) and M.
pseudocoronata (pH optimum = 7.2, Cumming et al. 1992a) and a lower relative abundance of
the acid-tolerant taxa M. hamata and M. hindonii. In South Lake, M. hamata and M. hindonii
reached their highest relative abundance from the 1980s to the early 1990s and then showed a
subsequent decline while S. echinulata increased again post-1970. In Queer Lake, M. hamata
increased in the late 1970s, peaked in the late 1980s, and then showed a subsequent decline. In
contrast, Mallomonas punctifera and the colonial chrysophytes Synura petersenii, S. echinulata
and Chrysosphaerella spp. increased in relative abundance in the 1980s.
Reference Lakes: The reference lakes all maintained a CI-pH greater than 6.5 over the
sedimentary record, suggesting that the lakes have not acidified over time (See Figure 5.3).
Stress tests of nMDS biplots of long-term species assemblage data from Moss Lake, Arbutus
Lake, and Wolf Lake show that the 2-axes of the nMDS provide a good approximation of the
multivariate data (stress = 0.08; stress = 0.15; stress = 0.17, respectively) (See Figure 5.4 –
species codes listed in Appendix I). In general, the pre-disturbance and the early disturbance
periods in each lake are similar in species composition, although, similar to South Lake, some
years in the early disturbance period of Moss Lake are characterized by an increased relative
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abundance of a colonial chrysophyte (Moss Lake: S. sphagnicola; South Lake: S. echinulata)
(See Figure 5.4 & Appendix H, Figure H2). Similar to the acidified lakes, all three reference
lakes show an increasing relative abundance of colonial chrysophyte taxa since the 1970s and
1980s. The post-1995 assemblages of Moss Lake are generally characterized by an increased
relative abundance of S. echinulata, M. punctifera, and S. spinosa, though some intervals still
have a high relative abundance of the unicellular taxon M. duerrschmidtiae. The post-1995
assemblages of Arbutus Lake are characterized by an increased relative abundance of the
colonial taxa S. sphagnicola and S. petersenii, though the relative abundance of S. petersenii has
decreased since the early 2000s. Lastly, the post-1995 species assemblages of Wolf Lake are
characterized by a decline in the relative abundance of M. pseudocoronata and an increase in the
relative abundance of S. petersenii, though some of the pre-1950 species intervals had a similar
relative abundance of S. petersenii.
Species changes in the acidified lakes post-1995 US ARP: The ANOSIM results showed a
significant shift in species composition post-1995 in both South Lake (R = 0.83, p = 0.001) and
Queer Lake (R = 0.68, p = 0.002) (See Table 5.1). The SIMPER post-hoc test revealed that
recent declines in the relative abundance of acid-tolerant species contributed to the dissimilarity
between Periods 1 and 2 in South Lake (M. hamata, 24.8% and M. hindonii, 12.7%) and Queer
Lake (M. hamata, 7.6%). However, in Queer Lake the recent increase in the relative abundance
of S. petersenii and M. punctifera contributed more to the dissimilarity between periods (25.0%
and 9.8%, respectively) than M. hamata.
Return to pre-disturbance state and recovery endpoints: The post-1995 assemblages of South
Lake plot intermediately on nMDS axis-1 between the pre-1900/pre-1950 samples and the
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majority of the samples in the acidification time period (1950 to 1995), suggesting that the lake
may be moving towards a pre-acidification assemblage (Figure 5.4). The intermediate placement
is due to the persistence of M. hamata and M. hindonii and the fact that S. echinulata showed a
peak in relative abundance in the early 1900s. In contrast, the post-1995 species assemblages of
Queer Lake appear to be moving towards a unique chrysophyte assemblage, characterized by an
increased relative abundance of S. petersenii, M. punctifera, S. echinulata, and Chrysosphaerella
spp. Interestingly, the species assemblages of the three reference lakes also show a separation in
ordination space between their recent and pre-1950 species assemblages, suggesting that they are
also moving towards a novel chrysophyte assemblage characterized by an increased abundance
of colonial chrysophytes (S. echinulata, S. sphagnicola, and/or S. petersenii) and an increased
abundance of M. punctifera.
The nMDS plot of the pre-1900 samples from South Lake and 26 regional reference
lakes found that Nellie Pond (NellP) and Round Pond (RounP) were most similar to South Lake
species assemblage pre-1900 (BC similarity 73% and 70%, respectively) (See Figure 5.5A).
Both lakes have shown an increase in the relative abundance of colonial chrysophytes over time
(S. sphagnicola, NellP; Chrysosphaerella spp., RounP). The present-day assemblages of the
reference lakes are located in a different part of the ordination diagram than the recent
assemblages of South Lake (See Figure 5.5B).
The nMDS of pre-1900 samples from Queer Lake and the regional reference lakes
showed that Wolf Lake (WolfL) and Moss Lake (MossL) were most similar to the pre-1900
Queer Lake assemblages (78% and 75%, respectively) (See Figure 5.6A). Because both lakes
had long-term profiles created in this investigation, long-term changes in both lakes are
compared separately with long-term changes in Queer Lake. The long-term nMDS plots show
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that Queer Lake, Moss Lake, and Wolf Lake are moving towards novel chrysophyte
assemblages. Wolf Lake has shown an increase in S. petersenii over time along with a decrease
in abundance of the unicellular taxon M. pseudocoronata, though the abundance of S. petersenii
is less than its current abundance in Queer Lake (See Figure 5.6B). The abundance of S.
petersenii is also lower in Moss Lake, but Moss Lake has also shown an increase in S. echinulata
and M. punctifera over time (generally < 2% and < 5% relative abundance in Queer Lake) (See
Figure 5.6C).
Correlations with air temperature/ice cover: The total relative abundance of colonial
chrysophyte taxa was significantly positively correlated with mean annual air temperature and
measures of ice-cover duration in Arbutus Lake and Wolf Lake (Table 5.2). The total percent
relative abundance of colonial chrysophytes was positively correlated with ice-on date in
Arbutus Lake (i.e. years with a later freeze-up date were associated with a higher relative
abundance of colonial chrysophytes) and negatively correlated with ice-cover duration in Wolf
Lake (i.e. years with shorter ice-cover periods were associated with a higher relative abundance
of colonial chrysophytes).
Discussion
Biological recovery from acidification and novel species assemblages
Both South Lake and Queer Lake are undergoing some biological recovery from
acidification. ANOSIM documented a significant shift in species composition in both lakes after
the 1995 implementation of the US ARP associated with a decline in the relative abundance of
acid-tolerant taxa (M. hamata and M. hindonii, South Lake; M. hamata, Queer Lake). However,
neither South Lake nor Queer Lake showed recent increases in taxa with higher pH-optima such
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as M. acaroides or M. pseudocoronata. The pH of South Lake decreases below 5.0 most years
and the pH of Queer Lake has only begun to reach consistently above 6.0 since 2001. Multiple
studies have suggested that the pH ‘benchmarks’ of 5.0 and 6.0 must be maintained for acidsensitive species to persist in recovering lakes (Lovett et al. 2009). As such, recovery is likely to
be limited until further improvements in water chemistry.
The nMDS ordination of the South Lake samples alone suggested that South Lake was
moving towards a chrysophyte assemblage more characteristic of pre-acidification conditions.
However, an nMDS plot comparing the long-term changes in chrysophyte species composition
in South Lake to two reference sites that had a similar species composition pre-1900 suggested
that South Lake may not return to its pre-disturbance condition. Instead, the lake may show an
increased abundance of colonial chrysophyte taxa currently rare or absent from South Lake (S.
sphagnicola, currently < 1% and/or Chrysosphaerella, currently absent). Queer Lake appears to
be moving towards a novel assemblage characterized by an increased relative abundance of
colonial chrysophytes relative to pre-disturbance times, a trend noted in all three reference lakes.
The comparative nMDS plots with Moss Lake and Wolf Lake suggest that the relative
abundance of S. petersenii may eventually decline in Queer Lake and that other colonial taxa
and/or M. punctifera may increase in relative abundance in the lake as recovery progress. In the
study of 31 Adirondack reference lakes, S. petersenii was most abundance in lakes with DOC
concentrations less than 5.2 mg C L-1 (Chapter 4). DOC is currently increasing in Queer Lake
and if its concentration surpasses 5.2 mg C L-1, this could lead to a decrease in S. petersenii, as
suggested by the comparative nMDS plots with Moss Lake and Wolf Lake.
As increases in colonial chrysophytes were documented in both the acidified lakes and
the reference lakes which have maintained a CI-pH of ~6.5 or greater throughout the sedimentary
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record, the increased abundance of colonial chrysophytes does not appear to be related to
acidification and chemical recovery. Because both the reference and acidified sites have been
subject to long-term chemical monitoring, we are also able to dismiss other changes in water
chemistry variables as possible drivers of the changes in chrysophyte assemblages. Chrysophyte
taxa are known to be sensitive to specific conductance (Siver 1993). However, the declines in
specific conductance measured in the study lakes are within the tolerance range of most taxa and
so are unlikely to explain the recent increases in colonial chrysophytes (Siver 1993). The total
relative abundance of colonial chrysophyte taxa increased in all sites, regardless of whether the
lakes have shown an increase in DOC (Moss Lake, South Lake, Queer Lake) or no directional
change in DOC concentrations over time (Arbutus Lake). However, as mentioned above, DOC
may have an important influence of the relative abundance of S. petersenii. Indeed, in Arbutus
Lake there has been a decrease in the relative abundance of S. petersenii since the early 2000s, a
time period typically characterized by DOC concentrations greater than 5 mg C L-1. In contrast,
as both S. echinulata and S. sphagnicola are likely DOC-indifferent in the reference lakes
(Chapter 4), changes in DOC concentration cannot explain recent increases in the total relative
abundance of colonial chrysophytes noted in both the acidified and reference lakes.
Chapter 4 suggested that increases in colonial chrysophytes in the regional set of 31
Adirondack reference lakes were due to regional warming and/or oligotrophication due to
increased adsorption of TP in watershed soils with long-term exposure to acid deposition and/or
logging followed by forest regrowth. Similar conclusions have been reached in other North
American studies (Paterson et al. 2001, Paterson et al. 2008, Ginn et al. 2010, Hyatt et al. 2010,
Flear et al. 2011). With longer ice-free periods and longer periods of thermal stratification,
flagellated algae like chrysophytes may have a competitive advantage over non-flagellated taxa
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like diatoms because they have an ability to seek out optimal light/nutrient conditions in the
water column (Winder and Sommer 2012). At an assemblage level, colonial taxa may have a
competitive advantage over unicellular ones with longer periods of stratification, as large
flagellates tend to be superior swimmers (Sommer 1988). Of course, the effects of regional
warming could be interactive with oligotrophication. In low-nutrient conditions, colonial
chrysophytes may be favoured because they form blooms at or below the thermocline where
nutrient conditions are more favourable (Fee 1976). Furthermore, phagotrophic colonial taxa like
Chrysosphaerella may have an advantage because they can consume alternative sources of
nutrients (Holen and Boraas 1995).
Unfortunately, the TP records from the study sites in this investigation were too short to
be analyzed (< 5 years) and so future studies should examine long-term trends in TP
concentrations in the Adirondacks and their potential influence on the recovery of algal
assemblages. This study does however suggest a link between regional warming and changes in
chrysophyte species composition in Adirondack lakes. Both mean annual temperature and
measures of ice cover duration were significantly correlated with the total relative abundance of
colonial chrysophytes in Arbutus Lake and Wolf Lake. Colonial chrysophytes can be found in
both cool and warm waters (Siver 2002) and so colonial taxa are likely responding to indirect
factors associated with warming such as ice cover duration or thermal stratification rather than
directly to air or water temperature. In contrast, the fact that M. punctifera increased coincidently
with colonial taxa in Moss Lake and Queer Lake likely reflects warming water temperatures as it
is a warm-water chrysophyte taxon (Siver 1991). In Arbutus Lake, the relative abundance of
colonial chrysophytes was positively correlated with ice-on date. S. sphagnicola, the main
colonial chrysophyte in Arbutus Lake, typically blooms in the water column from the summer to
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early autumn and is lost with the appearance of ice cover (Siver and Hamer 1992). Thus,
delaying the onset of the ice-covered period would allow this taxon to persist in the water
column for longer periods. In contrast, S. petersenii, the principal colonial chrysophyte in Wolf
Lake, is a dominant taxon just after ice off, though it can remain in the water column throughout
the year (Siver and Hamer 1992). The relatively modest increase in S. petersenii noted in Wolf
Lake may therefore be due to the fact that the ice-off date for Wolf Lake has changed little since
the 1970s (Beier et al. 2012). This investigation thus suggests that the response of taxa to
regional warming may be affected by their seasonal occurrence in the water column.
Given the findings of this and other studies (e.g., Paterson et al. 2004, Flear 2011), it is
unlikely that lakes recovering from acidification in the Adirondacks and elsewhere will return to
their pre-disturbance state. There is, however, a general need for more research into the factors
that control algal species composition and seasonal dynamics in freshwater ecosystems. Without
more detailed information, it is difficult to project phytoplankton responses to warming beyond
general trends (e.g. increases in flagellated algae, Winder and Sommer 2012). Such projections
would be highly beneficial, as some taxa that are likely to gain a competitive advantage with
warming conditions are also known to cause taste and odour issue in freshwaters, an important
management concern (e.g. S. petersenii, Nicholls and Gerrath 1985). The nMDS plots created in
this investigation were an attempt to predict what possible recovery endpoints could be expected
for the acidified lakes. This relatively simple approach could certainly be expanded upon by
modelling work of the reference/impacted sites that accounted for the
physical/chemical/biological effects of climate change.
157
Building context: a long-term perspective of chrysophyte assemblage change
Using paleolimnological techniques to define biological recovery enables us to
understand recent increases in colonial chrysophyte taxa in a historical context. This study
revealed two long-term patterns in colonial chrysophyte species abundance. In Queer Lake,
Arbutus Lake, and Wolf Lake, there was little change in the relative abundance of colonial
chrysophyte taxa until ca. 1970-1980. In contrast, in Moss Lake and South Lake there were pre1950 peaks in the abundance of colonial chrysophytes (South Lake: early 1900s; Moss Lake: late
1940s), followed by subsequent increases post-1970. Both trends have been noted in previous
paleolimnological investigations (Cumming et al. 1994) and merit further discussion.
As pre-1950 peaks in colonial taxa were noted both in acidified and reference lakes, it does not
appear that the trend is the result of an early onset of acidification. Northern New York
experienced a modest warming from the early 1900s to the mid-20th century, followed by a rapid
warming from the 1970s to present (Jenkins 2010). Thus, the mechanisms related to warming
discussed above may have promoted pre-1950 increases in colonial chrysophytes. Alternatively,
Davis et al. (2006) suggested that colonial chrysophytes increased in two New Hampshire lakes
due to watershed disturbance from logging and development, early stressors experienced by
South Lake and Moss Lake. However, Paterson et al. (1998) found that logging had little impact
on chrysophyte assemblages. It is therefore unclear what role, if any, disturbance had in
promoting early peaks in colonial chrysophytes in Adirondack lakes. Future studies should
examine possible relationships among climate warming, landscape disturbance, and changes in
algal assemblages. However, it must be emphasized that watershed disturbance cannot explain
recent increases in colonial chrysophytes as the lakes examined in this investigation have
158
experienced little watershed disturbance since the 1930s (Wolf Lake and Arbutus Lake) and the
1970s (Moss Lake, South Lake, Queer Lake).
Recovery in a multiple stressor environment
Modest biological recovery from acidification is occurring in Adirondack lakes, though
further recovery may be limited until key pH benchmarks are met. The acidified and reference
lakes examined in this investigation appear to be moving towards a novel chrysophyte
assemblage, characterized by recent (post-ca. 1970-1980) increases in colonial chrysophyte taxa
and warm-water taxa. The recent increases in colonial chrysophytes cannot be explained by
chemical recovery from acid deposition alone as they were noted in both acidified and nonacidified lakes. Rather, the increases may be related to recent climate warming and/or
oligotrophication. Additionally, the pre-1950 peaks in colonial chrysophyte abundance noted in
some lakes may be related to early warming and/or watershed disturbance. However, disturbance
cannot explain recent increases in colonial chrysophytes as the lakes in this investigation have
been protected from watershed disturbance since the 1930s and 1970s.
This study illustrates that applying paleolimnological techniques within a reference lake
framework provides researchers with powerful tools to contextualize species responses to
multiple stressors, a growing research need as lakes become increasingly impacted by multiple
stressors in North America (Keller 2009) and elsewhere. The US EPA recently recognized that
current SO2/NOx emissions standards do not adequately protect sensitive ecosystems from acid
deposition (Burns et al. 2011). Modelling studies have suggested that many Adirondack lakes
will have a pH less than 6.0 and ANC values less than 50 μeq L-1 by 2050 (Chen and Driscoll
2005) and that, without further reductions in sulphate deposition, many watersheds will
experience deposition greater than target loads for ANC of 50 μeq L-1 by 2100 (Sullivan et al.
159
2012). As such, biological recovery will likely be a protracted process, during which aquatic
communities will be influenced by multiple anthropogenic stressors. Shifts towards novel species
assemblages such as the recent increases in colonial chrysophytes documented in this
investigation are likely to be prevalent as recovering communities in North America may pose
unexpected management concerns, such as taste and odour issues.
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Tables & Table Captions
Table 5.1. Analysis of Similarity (ANOSIM) and Similarities Percentage (SIMPER) test results assessing whether
chrysophyte assemblages have changed between two a priori defined time periods (Period 1 and Period 2). Time
periods pre-1995 (Period 1, P1) and post-1995 (Period 2, P2) are listed for each lake with the number of samples per
period (n). ANOSIM results (test statistic, R, and p-value) are given. ANOSIM results significant at p < 0.05 are
highlighted in bold. SIMPER post-hoc test results are provided. Taxa that made a > 5% contribution (% Cont.) to the
dissimilarity between groups are listed. The average square-root relative abundance (%) of the species in Period 1
(Av √RA P1) and Period 2 (AV √RA P2) is provided, with the non square-root transformed
Time Periods
ANOSIM
results
R
p
P1
P2
South
Lake
1993-1981
(n = 7)
2008-1998
(n = 10)
0.828
0.001
Queer
Lake
1993-1981
(n = 8)
2009-1997
(n = 7)
0.682
0.002
SIMPER Results
Species
M. hamata
M. hindonii
M. duerrschmidtiae
S. echinulata
S. sphagnicola
M. acaroides
M. caudata
M. ‘small’
C. synuroides
S. petersenii
M. caudata
M. punctifera
Chrysosphaerella
M. duerrschmidtiae
S. echinulata
M. hamata
M. ‘small’
M. pseudocoronata
166
Av √RA
P1 (%)
4.6 (26.2)
2.0 (4.2)
6.8 (46.2)
4.2 (17.6)
0.8 (0.6)
1.3 (1.7)
2.3 (5.3)
1.0 (1.0)
1.1 (1.2)
2.6 (6.8)
3.5 (12.3)
1.6 (2.6)
1.0 (1.0)
8.1 (65.6)
0.9 (0.8)
2.1 (4.4)
1.2 (1.4)
0.8 (0.6)
Av √RA
P2 (%)
3.2 (10.2)
1.4 (2.0)
7.5 (56.3)
4.7 (22.1)
0.2 (0.04)
1.0 (1.0)
2.1 (4.4)
0.9 (0.8)
1.2 (1.4)
4.1 (16.8)
2.8 (7.8)
2.2 (4.8)
1.5 (2.3)
7.6 (57.8)
1.4 (2.0)
1.7 (2.9)
0.9 (0.8)
0.6 (0.4)
% Cont.
24.8
12.7
12.3
10.4
8.2
7.2
7.0
6.8
5.8
25.0
12.3
9.8
9.5
9.2
8.6
7.6
7.5
7.2
Table 5.2. Kendall τ correlation test results. Correlations between the total relative abundance of colonial
chrysophyte taxa (% colonial) and mean annual air temperature (TMEAN, in °C), ice cover parameters (ICE ON,
ice-on date; ICE OFF, ice-off date; ICE DUR, ice cover duration) are provided for Arbutus Lake and Wolf Lake.
Correlation coefficient (τ), p-value (p), and sample size (n) are provided. Correlations significant at p < 0.05 are
highlighted in bold (positive correlation, positive τ; negative correlation, negative τ).
% colonial vs. TMEAN
(ca. 1970-present)
% colonial vs. ICE ON
(ca. 1970-2007)
% colonial vs. ICE OFF
(ca. 1970-2007)
% colonial vs. ICE DUR
(ca. 1970-2007)
Arbutus Lake
τ
p
n
0.28
0.03
32
τ
0.28
0.31
0.02
27
0.08
0.61
25
0.03
0.87
27
-0.25
0.07
26
0.06
0.71
25
-0.41
0.01
24
167
Wolf Lake
p
n
0.03
30
Figures & Figure Captions
Figure 5.1. Long-term chemistry trends in acidified and reference lakes. Chemistry data for A: South Lake (●)
and Queer Lake (●) are from 1992-2010 and, B: Moss Lake (■) and Arbutus Lake (■) are from 1982/1983-2010.
Sulphate and nitrate (SO42- + NO3-, μeq L-1), pH, acid neutralizing capacity (ANC, μeq L-1), inorganic
monomeric aluminium (Ali, μg L-1), dissolved organic carbon (DOC, mg L-1) and specific conductance (Sp.
Cond., μS cm-1) are shown.
168
Figure 5.2. 210Pb and 214Bi activities in sediment cores from acidified and reference lakes.
Unsupported 210Pb (●) and 214Bi (■) activities (Bq per gram dry sediment, Bq g-1) plotted against
cumulative dry mass (g cm-2) in sediment cores from acidified lakes, A: South Lake (open symbols)
and Queer Lake (black symbols) and reference lakes, B: Moss Lake (grey symbols), Arbutus Lake
(open symbols), and Wolf Lake (black symbols). Dates are inferred using the constant rate of supply
Model (CRS). CRS-Inferred dates are plotted against sediment interval mid-point (cm) in sediment
cores from acidified lakes, C: South Lake (○) and Queer Lake (●) and reference lakes, D: Moss Lake
(▲), Arbutus Lake (○), and Wolf Lake (●). Age-depth models for Moss Lake, South Lake, and
Arbutus Lake were based on a simple quadratic function (solid lines). Queer Lake and Wolf Lake were
fitted with a cubic functions (solid lines), as quadratic equations (dashed lines) appeared to
underestimate sediment-age in the top 6-7 cm of the core. Error bars represent the standard error.
169
170
Figure 5.3. Total relative abundance of colonial chrysophytes (% colonial) and chrysophyte-inferred pH (CI-pH) values for two acidified lakes
(South Lake and Queer Lake) and three reference lakes (Arbutus Lake, Wolf Lake, Moss Lake). Acidified and reference lakes are separated by
a grey, dashed line. Samples are arranged by 210Pb-inferred date. % Colonial data CI-pH data are presented at an ~ 5 year resolution post-1970,
and an ~10-20 year resolution is present pre-1970 for all lakes. CI-pH is based on a calibration set of 71 Adirondack lakes (Cumming et al.
1992a). CI-pH was calculated by a weighted-averaging function with tolerance down-weighting, inverse deshrinking and a ln (x + 1) species
transformation (r2 boot = 0.75, RMSEP = 0.51).
Figure 5.4. Non-metric Multidimensional Scaling (nMDS) plots for acidified (A, South Lake; B, Queer Lake)
and reference lakes (C, Moss Lake; D, Arbutus Lake; E, Wolf Lake). Species assemblages are divided into 4
time periods: pre-disturbance (pre-1900, □), early disturbance (1900 to pre-1950, Δ), acidification (1950 to
1995, ♦), and post-1995 implementation of the US Acid Rain Program (post-1995, ●). Species codes are
provided in Appendix I. Species vectors are plotted passively in the ordination space. Stress values are
provided.
171
b
nMDS-2
nMDS-2
a
Stress = 0.09
Stress = 0.17
nMDS-1
nMDS-1
Figure 5.5. non-metric Multidimensional Scaling (nMDS) plots relating to South Lake species assemblages. A) pre-1900
South Lake assemblages (shown as black stars) in comparison to all pre-1900 samples from a set of 26 Adirondack reference
lakes. The two reference sites with the highest average similarity in comparison with the pre-1900 assemblages of South Lake
are shown as grey squares. B) Long-term changes in chrysophyte species assemblages in South Lake (black circles). Note that
data are presented at a sub-decadal resolution post-1970 and an approximately 10-20 resolution pre-1970 to reduce crowding
in ordination plot. Present-day (‘top’; T samples) and pre-1900 (‘bottom’; B samples) samples for two reference lakes (NellP
and RounP; white circles) are shown. Species codes are provided in Appendix I. Samples are plotted actively in the ordination
space and species vectors are plotted passively in the ordination space. Stress values are provided.
172
a
b
nMDS-2
nMDS-2
nMDS-2
c
Stress = 0.17
nMDS-1
Stress = 0.07
nMDS-1
Stress = 0.07
nMDS-1
Figure 5.6. Non-metric Multidimensional Scaling (nMDS) plots relating to Queer Lake species assemblages. A) pre-1900 Queer Lake assemblages (shown as black
stars) in comparison to all pre-1900 samples from a set of 26 Adirondack reference lakes. The two reference sites with the highest average similarity in comparison
with the pre-1900 assemblages of Queer Lake are shown as grey squares. B) Long-term changes in chrysophyte species assemblages in Queer Lake (black circles)
versus Wolf Lake and C) Moss Lake (white circles). Note that data are presented at a sub-decadal resolution post-1970 and an approximately 10-20 resolution pre1970 to reduce crowding in ordination plot. Species codes are provided in Appendix I. Samples are plotted actively in the ordination space and species vectors are
plotted passively in the ordination space. Stress values for the nMDS ordinations are provided.
173
CHAPTER 6 – GENERAL DISCUSSION
General Discussion
The 1990 Clean Air Act Amendments are arguably some of the most successful pollution
control measures ever implemented in the United States. With declines in sulphate emissions,
some areas in the United States are undergoing chemical recovery from acidification (Garmo et
al. 2014; Skjelkvåle et al. 2005). As such, there is a pressing need to define recovery targets for
acid-impacted lakes. Researchers trying to define recovery targets from acidification must
overcome two issues: 1) a lack of long-term monitoring data, and 2) the confounding influence
of multiple stressors. The difficulty posed by a lack of long-term monitoring data can be
addressed by applying paleolimnological techniques in recovery studies and regional reference
sites can be used to clarify the problem of multiple stressors.
This dissertation aimed to bridge the paleolimnological and bioassessment literature by
applying paleolimnological techniques in a study of a regional set of minimally-disturbed
reference lakes in the Adirondacks (NY, USA). The main goal of this dissertation was to use the
regional reference sites to track shifting baselines due to regional stressors like climate change
and use that information to define recovery targets for lakes recovering from acidification in the
Adirondack Park. Chapter 3 introduced the conceptual framework underlying this project and
critically evaluated the 31 reference sites using both qualitative and quantitative approaches;
Chapter 4 provided the results from a top-bottom study examining shifting baselines in
Adirondack reference lakes, and; Chapter 5 provided an application of the reference site
approach to defining recovery targets by pairing a set of reference lakes and a set of acidimpacted lakes recovering from acidification.
Chapter 3 outlined the conceptual framework applied in this investigation. The premise
behind this project was that a set of minimally-disturbed reference sites embedded in a region
174
impacted by multiple stressors could be used to track shifting baselines due to regional stressors
like climate change; information that could, in turn, be used to define reasonable recovery targets
for lakes recovering from acidification. For example, if the regional reference sites exhibited
increases in warm-water taxa or motile taxa, changes predicted to occur with regional warming
(Winder and Sommer 2012), those changes should be included in recovery targets for acidified
lakes in the region. In essence, acidified lakes would be expected to recover to a ‘novel’ state
similar to that of the reference sites with chemical recovery, rather than return to a predisturbance state. The approach adopted in this investigation is powerful because it bases
recovery targets on historical data, a long-term temporal perspective, and the use of reference
sites, three components Gray and Arnott (2009) argued would be ideal for studying recovery and
defining recovery targets.
A set of stringent selection criteria were used to identify a set of Adirondack lakes that
have been protected from: acidification, eutrophication (from shoreline and watershed
development), road salt seepage, and piscivores introductions. However, a critical evaluation of
the reference sites using both qualitative and quantitative methods revealed that most of the
reference lakes had experienced some form of historic fisheries or watershed disturbance,
highlighting the danger of assuming that a site’s present-day condition is characteristic of its
earlier history (Leira et al. 2006). Ordination methods found that the reference sites were
representative of between 24-36% of the chemical and/or physical characteristics of the ALS
lakes. The need to make meaningful comparisons between reference lakes and impacted sites
was discussed, as was the possibility of implementing similar frameworks in other regions.
In Chapter 4, a top-bottom paleolimnological investigation of the reference lakes
demonstrated that a shifting baseline effect was occurring in Adirondack reference lakes. Both
175
univariate and multivariate analyses showed that there has been a significant shift in species
composition, with present-day sediments characterized by an increased relative abundance of
colonial chrysophyte taxa. The changes in species composition noted in the reference lakes from
pre-1900 to present were greater than expected from counting variability, and so merited further
investigation. Because of the reference lake study design, it was possible to conclude that the
changes in species composition were not due to acidification (decreases in pH/ANC),
eutrophication (from watershed or shoreline development), road salt seepage, and piscivores
introductions. Furthermore, ANOSIM results found that historic watershed disturbance and
fisheries disturbance had minimal influence on chrysophyte species composition in the presentday samples. The increases in colonial taxa noted in the investigation were likely due to either
regional warming and/or oligotrophicaion resulting from increased adsorption of TP in
watershed soils with long-term exposure to acid deposition and/or logging followed by
subsequent forest regrowth. Both regional warming and oligotrophication could favour colonial
chrysophytes and indeed would likely have interactive effects (e.g. a longer period of thermal
stratification leading to nutrient depletion in the epilimnion). Regardless of the causal
mechanism(s) for the species changes noted, the results of the top-bottom investigation suggest
that lakes recovering from acidification in the region are unlikely to return to a pre-disturbance
state. Rather, the lakes are likely to show an increased abundance of colonial taxa and/or warmwater species as recovery progresses.
While an increase in colonial taxa from pre-1900 to present was seen in a majority of the
reference lakes, both physical and chemical factors influence the distribution of chrysophyte taxa
in Adirondack lakes. An RDA of the reference sites found that S. petersenii occurs in deeper
lakes with higher pH values and DOC values less than 5.2 mg C L-1. In contrast, S. sphagnicola
176
occurred in sites with lower pH values, both in low and high DOC sites. Because many
Adirondack lakes are shallow sites with pH values less than 6 (Kreseter et al. 1989), S. petersenii
may not be widespread in Adirondack surface waters. Taste and odour issues caused by S.
petersenii may, however, pose a problem in recreationally-important large, deep, alkaline lakes
that are home to important sportfish taxa like lake trout (Salvelinus namaycush).
Chapter 5 provides an application of the reference site approach that pairs a set of three
reference sites with two lakes (South Lake & Queer Lake) which acidified and are undergoing
chemical recovery from acidification. ANOSIM found that both of the acidified lakes had
undergone a significant shift in species composition since the 1995 implementation of the US
ARP. SIMPER found that the change in species composition was due both to a decline in the
relative abundance of acid-tolerant taxa (M. hamata, M. hindonii), as well as increases in
colonial chrysophyte species. Plotting the long-term species trajectories of the acid-impacted
sites with two reference sites that had similar species assemblages pre-1900 in an nMDS
ordination revealed that the probable recovery endpoints for South Lake and Queer Lake include
an increased abundance of colonial taxa and/or warm-water taxa, rather than a return to predisturbance state.
The results of Chapter 5 support statements made in Chapters 3 and 4 that acid-impacted
lakes in the Adirondacks are unlikely to return to a pre-disturbance state. Paleolimnological
reconstructions from all five study lakes showed an increase in colonial chrysophyte taxa since
ca. 1970-1980. The fact that increases in colonial taxa occurred both in lakes that acidified and
reference sites again highlights that the changes are not due to acidification and recovery.
Similarly, the increases in colonial taxa were unlikely due to declines in specific conductance
and though DOC may influence the abundance of S. petersenii, increases in DOC cannot explain
177
increases in the total relative abundance of colonial chrysophytes. Increases in colonial taxa were
correlated to ice-cover measures and mean annual air temperature in two reference lakes where
long-term monitoring data were available, linking increases in colonial taxa to regional warming,
though as noted in Chapter 4, these changes may also be attributable to oligotrophication. Pre1950 peaks in colonial chrysophyte abundance noted in some lakes may be related to early
warming and/or watershed disturbance but disturbance cannot explain recent increases in
colonial chrysophytes as the study lakes have been protected from watershed disturbance since
the 1930s (Wolf Lake and Arbutus Lake) and 1970s (South Lake, Queer Lake, Moss Lake).
This dissertation has made several important contributions to the ecological literature,
including improving our understanding of biological recovery from acidification (a subject that is
not well documented, Verdonschot et al. 2013), and our understanding of the influence of
climate warming on algal assemblages, a topic of growing research interest (Winder and Sommer
2012). Perhaps most globally, this dissertation provides an effective framework for examining
the effects of multiple stressors on biological recovery in lakes that can be applied to other study
other stressors in other regions. The application of this approach in the Adirondacks relied on
three factors: the availability of data from a large synoptic survey, previous
limnological/paleolimnological investigations to provide a basis for the reference lake selection
criteria used, and historical watershed and fisheries disturbance data. Datasets like the National
Lakes Assessment in the United States (USEPA 2011) or the European Diatom Database
(Battarbee et al. 2014) highlight the growing availability of regional survey data. In cases where
monitoring programs have already been established, including a paleolimnological component in
such studies will provide valuable information about how both impacted and reference lakes
have changed over time, tracking the ‘shifting baseline’ phenomnenon in response to regional
178
warming (USEPA 2012). However, the paleolimnological approach does have certain
limitations. First, certain taxa of interest (i.e. soft-bodied zooplankton like copepods) may not be
preserved in the sedimentary record, and so ideally the paleolimnological perspective should be
coupled with long-term biological monitoring data (Gray & Arnott 2009). Second, the
paleolimnological approach does require specialized equipment and knowledge, which can make
the implementation of paleolimnological studies difficult (i.e., depending on the approach:
taxonomic knowledge of subfossil species assemblages, dating techniques, multivariate data
analysis, etc.). While the use of paleolimnological data in the National Lakes Assessment in the
United States (USEPA 2011) and studies related to the implementation of the European Water
Framework Directive (Bennion et al. 2011) highlight that there is a growing use of
paleolimnological techniques in limnological assessment studies, long-standing methods like the
use of indices of biotic integrity are more widely used by lake managers (Soranno et al. 2011).
While such indices may be simpler to interpret than the multivariate species assemblage data
commonly used in paleolimnological studies, they may also mask important ecological
information. Likewise, if transfer functions are used in paleolimnological work to reconstruct
inferred pH (or TP, etc.) values, such changes should be interpreted within the context of overall
species assemblage change so that important ecological information is not overlooked (Quinlan
et al. 2008). Lastly, despite the utility of the reference site design, there can be complexities
when interpreting long-term species changes. For example, in Chapter 5 both early watershed
disturbance and regional warming could have contributed to pre-1950 increases in the relative
abundance of colonial chrysophytes in some lakes. Such complexity illustrates the importance of
building a long-term perspective of land-use and land-cover change (Renberg et al. 2009).
However, if early watershed disturbances have a lasting chemical/biological legacy in lakes, this
179
complexity is arguably present in most aquatic bioassessment work, regardless of whether or not
a paleolimnological approach is applied (Martin et al. 2011). Having long-term
paleolimnological data can, at least, allow researchers and lake managers to test for the effect of
early disturbances (vs. regional warming) on historic and present day species assemblages, if a
suitable experimental design is implemented (see example below).
Overall, by combining the long-term perspective of paleolimnological studies with a
rigorous reference site study design, this framework was used to both identify shifting baseline
conditions in reference lakes and define reasonable recovery targets for impacted lakes, a critical
research and management need. The results of this project also suggest several interesting
possibilities for expansions that could be made to this work, which are discussed below.
Future Directions
Expanding the reference-lake top-bottom study: The first obvious expansion point for this project
would be to include additional paleolimnological proxies in top-bottom studies of the reference
sites to examine if shifting baselines are occurring in other species groups in Adirondack lakes.
A group of particular interest would be the cladoceran zooplankton. While zooplankton may be
sensitive to climate warming (Moore et al. 1996; Daufresne et al. 2009), the response of
cladoceran to climate change is likely to be complex in regions influenced by acid deposition.
Long-term exposure to acid deposition can lead to calcium depletion in watershed soils (Likens
1996). Many large daphniids are Ca-sensitive organisms because calcium is an integral structural
component of their carapaces (Jeziorski and Yan 2006). While calcium-depletion in the
Adirondack region could lead to a reduced abundance of Daphnia in impacted lakes (Jeziorski et
al. 2008), it can be difficult to disentangle the effect of low calcium levels from acidification.
Low pH values frequently co-occur with low Ca levels and most daphniids are pH-sensitive
180
organisms (Jeziorski et al. 2012). Here, this project can make a valuable contribution because the
lakes span a calcium gradient (< 1 mg L-1 to > 5 mg L-1) and the majority of lakes are unlikely to
have acidified. Moreover, the selection criteria used to identify the reference lakes included a
depth criterion to increase the probability that daphniids would occur in the sites (see Chapter 3).
This project could offer an opportunity to disentangle a Ca-effect from a low-pH effect and
clarify our understanding of cladoceran responses to regional warming.
Identifying the influence of oligotrophication on Adirondack lakes: The results of Chapter 4 and
Chapter 5 suggest that climate warming and/or oligotrophication as a long-term consequence of
acid deposition could have promoted the increases in colonial chrysophyte relative abundance
noted in the lakes in this investigation. While there is evidence that regional warming is having
an influence on Adirondack lakes (for example, by reducing the length of the ice-covered period,
Beier et al. 2012), there is limited information on long-term TP trends in the region. Diatominferred TP reconstructions in the reference lakes could be used to identify whether or not
oligotrophication is occurring in Adirondack lakes. Dixit et al. (1999) identified that the
proportion of Adirondack lakes classified as oligotrophic or mesotrophic did not change from
pre-industrial times to present, though lakes within each category showed both increases and
decreases in diatom-inferred TP. However, the diatom flora may not be sufficiently sensitive to
track ecologically-important but subtle changes in TP. For example, while a reduction in TP
levels from 4 μg L-1 to 2 μgL-1 would likely be ecologically significant (a 50% reduction in
available TP), each state would be characterized by ultra-oligotrophic diatom taxa and the
change in inferred-TP could be within the error of the inference method (e.g. ln TP RMSEboot =
0.79, Dixit et al. 1999). For this reason, modelling studies may be needed to identify what
influence, if any, oligotrophication from long-term exposure to acid deposition is having on
181
Adirondack lakes. Mesocosm studies could also be undertaken to evaluate chrysophyte species
responses to both oligotrophication and changes in stratification patterns (e.g. nutrient-reduction
trials, vs. experimental mixing/stratification manipulations).
Disentangling the effects of early logging vs. early warming: In Chapter 5, both a reference site
(Moss Lake) and acid-impacted lake (South Lake) show a pre-1950 peak in colonial chrysophyte
abundance, a trend noted in previous paleolimnological studies of the region (Cumming et al.
1994). Such early increases in colonial chrysophyte may have been due to early warming pre1950 and/or early watershed disturbance. The effect of these two stressors could be disentangled
using the reference site approach and a balanced study design. A detailed paleolimnological
study of the four Class 1A lakes could be compared with four suitable Class 2 or Class 3
reference lakes that are known to have experienced early watershed disturbance (see Chapter 3).
Ideally sampling with an equal temporal resolution, a comparison of the long-term changes in
species assemblages in these lakes types would help identify what influence, if any, early
warming and/or watershed disturbance had on chrysophytes assemblages.
Chrysophyte species change at a Holocene-level temporal resolution: The results of this
investigation suggest that Adirondack lakes have undergone a significant shift in species
composition over the past 100-150 years, likely due to regional warming and/or
oligotrophication. However, the reference lake analysis could be expanded to include a
Holocene-level chrysophyte assemblage reconstruction. Such analysis would provide a more
complete view of the temporal variability of chrysophyte species assemblages. Moreover, an
analysis of time periods that were on average warmer than the present-day like the mid-Holocene
could provide a perspective on the likely species composition of Adirondack lakes in a
182
significantly warmer environment. In 2012, a piston core whose sedimentary record covers the
entire Holocene was collected from Wolf Lake. Analysis of this core will provide what is, to our
knowledge, the first low-elevation Holocene-perspective on an Adirondack Lake. Analysis of the
Wolf Lake core, and other Holocene records from the Adirondacks, will be able to provide us
with a critical insight not only into millennial-level changes in species composition in
Adirondack lakes, but also changes in regional vegetation and precipitation patterns. As always,
the long-term perspective of paleolimnological work will be critically important in evaluating the
effects of ongoing global environmental change on aquatic ecosystems.
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185
APPENDICES
Appendix A - Figure A1. Histogram plots of Adirondack Lakes Survey (ALS) variables. Morphological
variables are elevation (elv, m) maximum depth (max depth, m), surface area (Area, ha), and watershed area
(WArea, ha). Chemical variables are field pH, SO42-, F-, Cl-, Na+, K+, Ca2+, Fe2+ (mg L-1), DIC (dissolved
inorganic carbon, mg L-1), DOC (dissolved organic carbon, mg C L-1) TP (total phosphorus, mg L-1) and
specific conductance (SCONDUCT, µmhos cm-1). Transformations are listed in histogram plots. Data were
log (x + 1) transformed if minimum value in ALS survey > 1; log (x + ½ minimum value) if 1 > minimum
value > 0.1; log (x + 0.001) if smallest non-zero recorded value > 0.001; and, log (x + 0.0001) if smallest nonzero recorded value > 0.0001. If a variable included negative values after log transformation, a constant (the
absolute value of the minimum log-transformed data) was added to the log-transformed results.
186
Appendix B - Table B1. List of variables measured in 2010 and 2011 water chemistry samples from reference lakes.
Methods and instruments used are given.
Analyte
pH
Method
EPA 150.1
ANC
Modified Gran analysis
technique
Cations (Na, Mg, P, K, Ca, Si, Mn,
Fe, Zn)
Anions (F, Cl, SO4, NO3)
Total Nitrogen
EPA 200.8
8
Dissolved Organic Carbon
Dissolved Inorganic Carbon
EPA 300.0
EPA 415.3
(Modified for TN)
EPA 415.3
EPA method 160.2
Monomeric Aluminum
McAvoy et al. 19928
Instrument
Brinkman Metrohm 799 DMS Titrino and the
Brinkmann Metrohm 748 autosampler; and the
Brinkmann Titrino
Brinkman Metrohm 799 DMS Titrino and the
Brinkmann Metrohm 748 autosampler; and the
Brinkmann Titrino
ICP-MS, PE ELAN 6000
Dionex DX-500 ion chromatography system
Teledyne Tekmar Apollo 9000 analyzer with TN
module
Teledyne Tekmar Phoenix analyzer
Teledyne Tekmar Apollo 9000 analyzer with TN
module
Bran & Luebbe Auto Analyzer 3
McAvoy DC, Santore RC, Shosa JD, Driscoll CT. (1992) Comparison between pyrocatechol violet and 8hydroxyquinoline procedures for determining aluminum fractions. Soil Science Society of America Journal
56: 449-455.
187
Appendix C - Table C1. Location information for 31 Adirondack reference lakes. Lake name, code,
latitude and longitude, and Adirondack Park Agency (APA) land classification are provided.
Lake Name
Arbutus Lake
Bass Lake
Bessie Pond
Boottree Pond
Cascade Lake
Cascade Pond
Challis Pond
Clamshell Pond
Clear Pond
Copperas Pond
Deer Pond
Eagle’s Nest Lake
East Pine Pond
Fish Pond
Grizzle Ocean
Gull Pond
Island Pond
Little Fish Pond
Long Pond
Lower Sargent Pond
Lydia Pond
Middle Branch Lake
Moss Lake
Nellie Pond
Rock Pond
Round Pond
Sampson Lake
Streeter Lake
Upper Spectacle Pond
Wilcox Lake
Wolf Lake
Code
ArbL
BassL
BessP
BootP
CascL
CascP
ChalP
ClamP
CleaP
CoppP
DeerP
EagNL
EPineP
FishP
GrizO
GullP
IslaP
LitFP
LongP
LSargP
LydiP
MidBL
MossL
NellP
RockP
RounP
SampL
StreL
UpSP
WilcL
WolfL
Latitude
43º 59’ 14’’
43º 58’ 25’’
44º 22’ 51’’
44º 14’ 24’’
43º 47’ 21’’
43º 49’ 46’’
43º 59’ 00’’
44º 23’ 00’’
44º 33’ 30’’
44º 19’ 45’’
44º 13’ 47’’
43º 45’ 50’’
44º 20’ 23’’
44º 23’ 42’’
43º 49’ 22’’
43º 49’ 51’’
43º 40’ 50’’
44º 23’ 42’’
44º 18’ 59’’
43º 51’ 34’’
44º 23’ 45’’
43º 41’ 52’’
43º 46’ 52’’
44º 23’ 02’’
44º 11’ 26’’
44º 07’ 26’’
43º 34’ 43’’
44º 06’ 39’’
43º 48’ 46’’
43º 23’ 59’’
44º 01’ 01’’
Longitude
74º 14’ 28’’
73º 40’ 39’’
74º 23’ 14’’
74º 39’ 20’’
74º 48’ 46’’
74º 26’ 12’’
73º 40’ 14’’
74º 21’ 54’’
74º 46’ 21’’
73º 53’ 54’’
74º 39’ 47’’
74º 43’ 42’’
74º 24’ 44’’
74º 22’ 42’’
73º 35’ 51’’
73º 42’ 29’’
73º 39’ 17’’
74º 23’ 04’’
73º 55’ 45’’
74º 34’ 02’’
74º 23’ 46’’
75º 06’ 08’’
74º 51’ 11’’
74º 23’ 28’’
74º 18’ 10’’
73º 43’ 57’’
74º 34’ 33’’
75º 04’ 17’’
73º 42’ 17’’
74º 09’ 26’’
74º 13’ 13’’
188
APA classification
Private (Huntington Wildlife Forest)
Hammond Pond Wild Forest Area
St. Regis Canoe Area
Private (Massawepie Scout Camp)
Pigeon Lake Wilderness Area
Blue Ridge Wilderness Area
Hammond Pond Wild Forest Area
St. Regis Canoe Area
White Hill Wild Forest
Sentinel Range Wilderness
Private (Massawepie Scout Camp)
Moose River Wild Forest River
Saranac Lakes Wild Forest Area
St. Regis Canoe Area
Pharaoh Lake Wilderness Area
Pharaoh Lake Wilderness Area
Lake George Wild Forest Area
St. Regis Canoe Area
McKenzie Mountain Wilderness Area
Sargent Ponds Wild Forest Area
St. Regis Canoe Area
Ha-De-Ron-Dah Wilderness Area
Moose River Wild Forest Area
St. Regis Canoe Area
High Peaks Wilderness Area
Dix Mountain Wilderness Area
West Canada Lakes Wilderness Area
Aldrich Pond Wild Forest Area
Pharaoh Lake Wilderness Area
Wilcox Lake Wild Forest Area
Private (Huntington Wildlife Forest)
189
Appendix D - Figure D1. Summary diagram of species assemblages in 31 Adirondack reference lakes. Data are provided as relative abundance (%). Top
(present-day) assemblages are show in solid black bars and bottom (pre-1900) assemblages are show in open bars. Samples are organized in descending
order based on PCA-1 axis scores of 31 present-day samples (data not shown). Chrysophyte species are split between colonial and unicellular taxa.
Chrysophyte-inferred pH (CI-pH) values are provided. Note that 5 reference lakes had scale concentrations too low to count in the pre-1900 sample and so
only a present-day sample is provided.
Appendix E – Figure E1. Relationship between colonial taxa and DOC in 31 reference lakes.
Relative abundance (% RA) of 3 colonial taxa (S. petersenii, S. sphagnicola, S. echinulata)
plotted against DOC (mg C L-1) and depth of 1% PAR irradiance (m). Depth 1% PAR based
on Bukaveckas and Robbins-Forbes (2000). Gray dotted line denotes a DOC concentration >
5.2 (DOC 5.3 mg C L-1), which corresponds to a 1% PAR depth < 5.2 m (5.1 m). Note
decreased abundance of S. petersenii when DOC > 5.2 mg C L-1 whereas S. sphagnicola and
S. echinulata occur at > 10% RA when DOC > 5.2 mg C L-1.
190
Appendix F - Table F1. Stocking history, recent netting survey results, land use characteristics, and watershed disturbance histories for South Lake, Queer Lake,
Moss Lake, Arbutus Lake, and Wolf Lake. See footnotes for citation information. Stocking history is summarized and recent netting survey results are provided
with species listed alphabetically by scientific name. Current land use information (e.g. forest preserve, cottage development, etc.) is provided. Disturbance history
includes information on logging events, and watershed impact of large windstorms in 1950 and 1995. An ice storm in 1998 and forest fires in 1903 and 1908 had no
known impact on the study sites.
Lake
South
Lake9
Queer
Lake10
Moss
Lake11
Stocking History
Stocked with lake trout, brook
trout, and other piscivorous
fish from 1993-2006
Date
2002
Stocked with brook trout from
1942 to 1975 and from 19952006; currently managed as a
cold-water fishery
1999
Earliest records of stocking of
brook trout and lake trout in
1898. Stocked with lake trout
and brook trout from 19861998.
2000
Netting Survey
Survey Result
- Couesius plumbeus
- Ictalurus nebulosus
- Phoxinus eos
- Rhinichthys atratulus
- Salmo salar
- Salvelinus fontinalis
- Semotilus atromaculatus
- Catostomus commersonii
- Fundulus diaphanus
- Ictalurus nebulosus
- Lepomis gibbosus
- Notemigonus crysoleucas
- Salvelinus fontinalis
- Salvelinus namaycush
- Catostomus commersonii
- Ictalurus nebulosus
- Lepomis gibbosus
- Luxilus cornutus
- Micropterus salmoides
- Notemigonus crysoleucas
- Perca flavescens
- Salvelinus fontinalis
- Semotilus atromaculatus
- Umbra limi
Watershed land-use
Disturbance History
Logging History
Windstorms
Pre-1890: some denuded
- 1950: 50-100%
area
blowdown in 65.8%
of watershed; 25-50%
1890-1920: logged for
in 34.2% of the
spruce & hardwoods
watershed
1916: 100% of the
- 1995: No impact
watershed either virgin or
2nd growth forest
No known history of
- 1950: minimal impact
logging
- 1995: 0-30% change
1916: 100% of the
in tree crowns in
watershed either virgin or
100% of watershed
2nd growth forest
- 56% protected ‘wild forest’
preserve since 1972
- 40% privately owned with
little or no development
- 4% developed, with cottages,
private camps, and a gravel
road on the NW lakeshore
- 100% of watershed is
protected ‘wilderness’
preserve since 1972
- Primitive campsites along
lakeshore; no cottages
-
- A road paved bisects the
watershed (originally built
1895-1896)
- 1923-1973: 612 acre camp
surrounding lake
- Watershed protected in a
‘wild forest’ and ‘wilderness’
preserve since 1973
- Primitive campsites along
lakeshore; no cottages
- Pre-1890: no known
logging
- 1916: 93.7% of the
watershed described as
either virgin or 2nd growth
forest
9
-
-
- 1950: 50-100%
blowdown in 13.2%
of watershed; 25-50%
blowdown in 3.5% of
watershed
- 1995: 0-30% change
in tree crowns in
93.2% of watershed;
30-60% change in
6.8% of watershed
Stocking history, netting survey data, land-use, windstorms: Roy et al. 2011; logging history: Roy et al. 2011, Sullivan et al. 1999, McMartin 1994
Stocking history, netting survey data, land-use, windstorms: Roy et al. 2011; logging history: Roy et al. 2011, Sullivan et al. 1999, McMartin 1994
11
Stocking history, netting survey data, windstorms: Roy et al. 2011; land-use: Roy et al. 2011, McMartin 2004, Charles et al. 1987; logging history: Roy et al. 2011, McMartin
1994
10
191
Table F1. Continued
Arbutus
Lake12
Stocking history is unknown;
lake treated with rotenone in
1973 and stocked with brook
trout (dominant fish taxa were
brown bullhead, yellow
perch, bass, and ‘suckers’ –
species not specified)
2001
-
Ictalurus nebulosus
Phoxinus eos
Rhinichthys atratulus
Salvelinus fontinalis
Wolf
Lake13
- No known history of
stocking; repeated surveys
until 1980s found no nonnative fish taxa
1997
-
Salvelinus fontinalis
Ictalurus nebulosus
Semotilus atromaculatus
Lepomis auritus
Catostomus commersonii
Other minnows (species
unspecified)
- Private lodge & camp in late
1800s
- Part of a private research
property established in 1932;
2000 acres around Arbutus
Lake added in 1939
- Watershed largely
undeveloped
- Cabins & unpaved roads on
southern lakeshore
- Part of a private research
property established in 1932
- Watershed largely
undeveloped
- Unpaved road (built in late
1930s) within 200-300 m of
shoreline; 1 cabin near
lakeshore (abandoned)
- Pre-1890: no known logging
- 1916: 87.9% of watershed
classified as either virgin or
2nd growth forest.
- Watershed logged in late
1910s
- 1960-1961: 162 ha of
softwoods logged
- 1950: no impact
- 1995: no impact
- Pre-1890: no known logging
- Selective logging between
1957-1991
- 1950: minimal
impact
- 1995: no impact
Literature Cited:
Charles DF., Whitehead DR, Engstrom DR, Fry BD, Hites RA, Norton SA, Owen JS, Roll LA, Schindler SC, Smol SP, Uutala AJ, White JR, Wise RJ. (1987)
Paleolimnological evidence for recent acidification of Big Moose Lake, Adirondack Mountains, N.Y. (USA). Biogeochemistry 3:267-296.
Chen LM, Driscoll CT. (2004) Modeling the response of soil and surface waters in the Adirondack and Catskill regions of New York to changes in atmospheric
deposition and historical land disturbance. Atmospheric Environment 38:4099-4109.
Jenkins J., Keal A. (2004) The Adirondack atlas: a geographic portrait of the Adirondack Park. Syracuse University Press, Syracuse, New York, 296 p.
McMartin B. (1994) The great forest of the Adirondacks. North Country Books, Utica, New York, 254 p.
McMartin B. (2004) The privately owned Adirondacks. Lake View Press, Caroga, New York, 304 p.
Roy K., Houck N, Hyde P, Cantwell M, Brown J. (2011) The Adirondack long-term monitoring lakes: a compendium of site descriptions, recent chemistry and
selected research information. New York State Energy Research and Development Authority, Albany, 298 p
Stager JC, Sanger T. (2003) An Adirondack "heritage lake". Adirondack Journal of Environmental Studies 10:6-10.
Sullivan TJ, Charles DF, Bernert JA, McMartin B, Vache KB, Zehr J. 1999. Relationship between landscape characteristics, history, and lakewater acidification
in the Adirondack Mountains, New York. Water, Air, and Soil Pollution 112:407-427.
12
Netting survey data, windstorms: Roy et al. 2011; stocking history: C.L Demers & R.J. Pearl, personal communication, 2010; land-use: Roy et al. 2011, McMartin 2004; logging
history: Roy et al. 2011, Chen & Driscoll 2004, McMartin 1994
13
Stocking history and netting survey data: Stager & Sanger 2003; land-use: Stager & Sanger 2003, McMartin 2004; logging history: Stager & Sanger 2003, McMartin 1994;
Windstorms: Stager & Sanger 2003, Jenkins & Keal 2004
192
Appendix G – Table G1. Seasonal Kendall τ results for long-term monthly chemistry data. The Seasonal Kendall
τ is a non-parametric test that identifies if a data series shows a monotonic trend overtime (i.e. generally
increasing or generally decreasing) (Helsel and Hirsch 2002). The test is recommended for data with serial
dependence, such as monthly water chemistry measurements. Tests were performed using the computer program
of Helsel et al. (2006). Tests were run on long-term chemistry data from South Lake (1992-2010), Queer Lake
(1992-2010), Moss Lake (1982-2010), and Arbutus Lake (1983-2010). Chemistry variables are: sum acid anions
(SO42- + NO3-, in μeq L-1), acid neutralizing capacity (ANC, in μeq L-1), pH, inorganic monomeric aluminum (Ali,
in μg L-1), dissolved organic carbon (DOC, in mg C L-1), and specific conductance (Sp. Cond., in μS cm-1). The
correlation coefficient (τ), test statistic (S), and p-value adjusted for seasonal dependence of chemistry data (p adj)
are provided. Chemistry variables that show a significant (p adj < 0.05) monotonic trend overtime (generally
increasing, positive τ; generally decreasing, negative τ) are highlighted in bold.
Lake
South Lake
Queer Lake
Moss Lake
Arbutus Lake
τ
-0.81
0.53
0.52
-0.47
0.32
-0.73
-0.84
0.63
0.59
-0.45
0.37
-0.80
-0.82
0.22
0.16
-0.01
0.20
-0.56
-0.72
-0.01
-0.01
0.03
0.07
-0.53
Variable
SO42- + NO3ANC
pH
Ali
DOC
Sp. Cond.
SO42- + NO3ANC
pH
Ali
DOC
Sp. Cond.
SO42- + NO3ANC
pH
Ali
DOC
Sp. Cond.
SO42- + NO3ANC
pH
Ali
DOC
Sp. Cond.
S
-1574
1035
1006
-908
617
1425
-1602
1193
1117
-854
702
-1522
-3666
989
695
-53
858
-2499
-2999
-33
-34
133
274
-2221
padj
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
<0.001
0.002
0.020
0.845
0.003
<0.001
<0.001
0.917
0.932
0.563
0.427
<0.001
Literature Cited:
Helsel DR, Hirsch RM (2002) Statistical methods in water resources techniques of water resources investigations,
book 4, chapter A3. U.S. Geological Survey, pp. 209-220.
Helsel DR, Mueller DK, Slack, JR. (2006) Computer program for the Kendall family of trend tests. U.S. Geological
Survey Scientific Investigations Report 2005-5275, U.S. Geological Survey, Reston, Virginia, 4 p.
193
Appendix H – Figure H1. Stratigraphies of chrysophyte relative abundance in acidified lakes. Long-term (preca. 1900-present) stratigraphies of the relative abundance of chrysophyte taxa (%) are provided for South Lake
(A) and Queer Lake (B). pH optimum is given in brackets after species name (Cumming et al. 1992a).
Chrysophyte-inferred pH (CI-pH) is based on a calibration set of 71 Adirondack lakes (Cumming et al. 1992a).
CI-pH calculated by a weighted-averaging function with tolerance down-weighting, inverse deshrinking and a ln
(x + 1) species transformation (r2 boot = 0.75, RMSEP = 0.51). The total relative abundance of colonial chrysophyte
taxa (% Colonial) is shown.
194
Appendix H – Figure H2. Stratigraphies of chrysophyte relative abundance in reference lakes. Long-term
(pre-ca. 1900-present) stratigraphies of chrysophytes relative abundance (in %) are provided for Moss Lake
(A), Arbutus Lake (B), and Wolf Lake (C). pH optimum is given in brackets after species name (Cumming et
al. 1992a). Chrysophyte-inferred pH (CI-pH) is based on a calibration set of 71 Adirondack lakes (Cumming et
al. 1992a). CI-pH calculated by a weighted-averaging function with tolerance down-weighting, inverse
deshrinking and a ln (x + 1) species transformation (r 2 boot = 0.75, RMSEP = 0.51). The total relative abundance
of colonial chrysophyte taxa (% Colonial) is shown (note change of scale for Arbutus Lake and Wolf Lake).
195
Appendix I – Table I1. List of taxon codes used in nMDS
ordination diagram. Taxa are listed alphabetically by name and a
code for each taxon is given. The taxon’s form (either unicellular
or colonial) is also provided.
Taxon
Chrysodidymus synuroides
Chrysosphaerella spp.
Mallomonas acaroides
Mallomonas caudata
Mallomonas crassisquama
Mallomonas duerrschmidtiae
Mallomonas elongata
Mallomonas hindonii
Mallomonas hamata
Mallomonas lychenensis
Mallomonas pseudocoronata
Mallomonas punctifera
Mallomonas ‘small’
Synura echinulata
Synura petersenii
Synura spinosa
Synura sphagnicola
Synura uvella
Synura sp
Code
C syn
Chrysos
M acar
M caud
M crass
M duerr
M elong
M hind
M hama
M lych
M pseudo
M punct
M small
S echin
S peter
S spin
S spag
S uvella
S sp
14
Form
14
Colonial
15
Colonial
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
Unicellular
16
Unicellular
Colonial
Colonial
Colonial
Colonial
Colonial
Colonial
Chrysodidymus synuroides forms 2-celled colonies
Likely includes both Chrysosphaerella longispina and Chrysosphaerella brevispina
16
Group of Mallomonas scales too small to distinguish between using light-microscopy
15
196
Appendix J – Figure J1. Principal Components Analysis (PCA) results for acidified and reference
lakes. Long-term (pre-ca. 1900-present) PCA results of chrysophyte species assemblages in South
Lake (A), Queer Lake (B), Moss Lake (C), Arbutus Lake (D), and Wolf Lake (E) are provided. Species
assemblages are divided into 4 time periods: pre-disturbance (pre-1900, □), early disturbance (1900 to
pre-1950, Δ), acidification (1950 to 1995, ♦), and post-1995 implementation of the US Acid Rain
Program (post-1995, ●).
197