opg/nserc research chair watershed biogeochemistry 2002 … · · 2005-11-20opg/nserc research...

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OPG/NSERC Research Chair Watershed Biogeochemistry 2002-2003 Workshop Report November 17

th – 18

th, 2003

April 16, 2004

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Table of Contents

Chapter 1: Introduction to the workshop …………………………………………………… 3 Peter Dillon

Chapter 2: Overview of Ontario Power Generation ……………………………………….. 8

Andy Hoffer

Chapter 3: Long-term trends in deposition and lake history ….………………………….... 12

Peter Dillon

Chapter 4: Trends in Surface Water Chemistry 1990-2001 in Europe/North America and 2004 Acid Rain Assessment results from Ontario ….…………………..…….. 15 Dean Jeffries

Chapter 5: Effects of drought-induced acidification on diatom communities in acid- sensitive Ontario lakes …………………………………………………………..… 19

Roland Hall

Chapter 6: Sources and controls of sulphur cycling in catchments in south-central Ontario

Cathy Eimers……….…………………………………………………………….... 23

Chapter 7: Sudbury Environmental Study (SES): an overview ……………………..……. 26 Bill Keller .

Chapter 8: Steady-State Water Chemistry (SSWC) and First-order Acidity Balance (FAB) models in Ontario……………………………….……………………….… 29 Julian Aherne..

Chapter 9: Model of Acidification of Groundwater in Catchments (MAGIC): overview and application in Ontario……………………………………………… 33 Julian Aherne

Chapter 10: Acidification in Ontario: monitoring and modeling …………………….……36 Shaun Watmough

Chapter 11: Drought induced pulses of S from Canadian Shield wetlands: ∂

34S and ∂18O in SO42- and DOM ………………………………………………… 40

Sherry Schiff

Chapter 12: Acidity, climate change, UV, DOC and photochemistry in lakes .. …….….. 42 Lewis Molot

Chapter 13: Overview of progress in the Research Chair Programme …………….…….. 45 Peter Dillon

Alphabetical List of Attendees …………………………………………………………… 48

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Chapter 1: Introduction to the workshop Peter Dillon

Peter Dillon opened the workshop and reviewed the rationale for the establishment of the Industrial Research Chair (IRC) in Watershed Biogeochemistry, a partnership between Ontario Power Generation (OPG), the Natural Sciences and Engineering Research Council of Canada (NSERC) and Trent University. The scientific programme of the IRC was designed to address the potential effects of increases in sulphur (S) and nitrogen (N) emissions from OPG that were anticipated to occur while nuclear generation facilities in Ontario were renovated (Fig. 1a). While fossil fuel-based power generation accounted for only 13% of OPG’s total capacity in 1996, this increased to 27% by 1999 and was anticipated to reach 35% shortly thereafter. Emissions and possibly deposition of trace metals such as mercury (Hg) also were anticipated to increase. Furthermore, there was strong evidence that climate patterns in many parts of the world including Ontario were being altered very substantially as a result of global increases in greenhouse gas concentrations.

Fig. 1a: Past, current and projected SO2 emissions from OPG (formerly Ontario Hydro) and for all of Ontario. In Canada, nearly 45% of the land area is sensitive to acid deposition, with areas situated on the Precambrian Shield, including much of Ontario, Quebec and the Atlantic Provinces, being the most vulnerable. This region is underlain by silicate bedrock and receives the highest levels of acid deposition in the country. This is also

SO2 Emissions

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the region that would most likely receive OPG’s increased emissions. Thus, the Watershed Biogeochemistry research programme focussed on sensitive lakes and watersheds in the Muskoka-Haliburton (M-H) region of south-central Ontario, where data on some of the lakes and catchments have been collected since 1980. Although the M-H area currently receives 40-45 meq/m2 of SO4 and 60 meq/m2 N per year (down from 75-80 meq/m2 in 1976-80), in the mid-1990’s only about half of the lakes studied had shown decreasing sulphate concentrations, while the remaining half had shown no positive response to changes in deposition. When parameters more important to the biota of the lakes were considered (i.e., pH and alkalinity), recovery was evident in only a small portion of the lakes. In addition, those that had shown some positive changes had improved considerably less than expected based on the change in deposition. It would appear that the responses of lakes and their watersheds to changes in acid deposition are not only variable, but are likely controlled by factors other than simple changes in acid deposition rate. The Watershed Biogeochemistry research programme was therefore designed to improve our understanding of the way in which biogeochemical processes in lakes and their watersheds, control and respond to changes in the deposition of strong acids. In addition, because it was apparent that environmental perturbations such as acid deposition do not act in isolation, but are augmented or offset by other factors, the interaction between changes in climate and acid deposition was examined. Specifically, the research programme was designed to consider the effects of climate perturbations such as summer droughts, on the chemical and biological response of aquatic ecosystems and to differentiate between those effects mediated by changes in acid deposition and those brought about by a varying climate. Four (4) projects were initiated to address these research objectives:

1) Evaluation of existing environmental data in Ontario related to surface and groundwater chemistry, to assess the likelihood that expected changes in acid deposition will result in measurable changes in water quality.

2) Measurement of the chemical and biological response of aquatic ecosystems to

changes in OPG’s sulphur emissions including: (i) measurement and modeling of the chemical response of a representative set of lakes and their terrestrial catchments, (ii) assessment of biological effects using simple indicators, and (iii) evaluation of the changes in Hg deposition related to fossil fuel use.

3) Evaluation of the interrelationships between climate change, the sulphur cycle in lakes and watersheds, and changes in emission and deposition of acid precursors, with specific reference to El Niño episodes.

4) Investigation of the role of dissolved organic carbon (DOC) in integrating the effects

of multiple stressors including acid deposition, changing climate and others.

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A summary of the projected activities and collaborators for each project is as follows:

1. Analysis of long-term trends

a) Changes in long-term chemistry Activities:

• Analysis of long-term chemical data sets for M-H (32 lakes, 24 streams)

• Comparison with trends/patterns in Norway, the Turkey Lakes, the Experimental Lakes Area (ELA) and Wisconsin

• Collaboration with EU project, RECOVER 2010, regarding recovery rates and controlling factors Collaborators: Bob Ferrier (Macaulay Land Use Research Institute, UK) Dean Jeffries (Environment Canada) Alan Jenkins (Centre for Ecology and Hydrology, UK) Bill Keller (Ontario Ministry of the Environment) Kathy Webster (University of Wisconsin, USA) Richard Wright (NIVA, Norway)

b) Paleolimnology Activities:

• Collection of sediment cores from lakes with long-term chemical records, for analysis of acid deposition and climate effects

• Collection of tree cores from gauged catchment(s), for analysis of deposition and climate effects Collaborators: Brian Cumming (Queen’s University) Roland Hall (University of Waterloo) Andrew Paterson (Ontario Ministry of the Environment) John Smol (Queen’s University)

2. Measurement and modeling of chemical and biological response

a) Chemical effects

Activities:

• 5-year measurements of lakes, streams, soil and groundwater, and wetlands at sites with existing long-term data

• Re-evaluation of forest biomass and soil chemistry at Harp and Plastic lakes

• Application of steady-state (SSW) and dynamic (FAB) models to assess critical loads

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Collaborators: Jack Cosby (University of Virginia) Arne Henriksen (NIVA, Norway) Thorjorn Larsen (NIVA, Norway) Max Posch (RIVM, Netherlands) Richard Wright (NIVA, Norway)

b) Biological effects Activities:

• Using benthic invertebrates as indicators of environmental stress/change

• Obtaining good historical data on pelagic and macro-invertebrates

• Assessing the ‘rapid bioassessment protocol’ Collaborators: Keith Somers (Ontario Ministry of the Environment)

c) Hg deposition Activities:

• Measurement of deposition at a site where previous measurements were made

• Measurement of Hg flux from catchments before, during and after El Niño events

• Measurment of Hg uptake by aquatic invertebrates Collaborators: Doug Evans (Trent University) Holger Hintelmann (Trent University) Greg Mierle (Ontario Ministry of the Environment)

3. Climate as a mediating factor in the recovery of aquatic systems

Activities:

• Analysis of coherence between meteorologic, hydrologic and chemical parameters

• Separation of natural year-to-year variability, repetitive patterns (eg. El Niño events) and long-term trends

• Analysis of relative effects of climate and the changes in S emissions on the chemistry of lakes

• Revision of MAGIC to include wetlands, redox processes and climate effects Collaborators: Jim Buttle (Trent University) Jack Cosby (University of Virginia) Mike English (Wilfred Laurier University) Bob Ferrier (Macaulay Land Use Research Institute, UK) Alan Jenkins (Centre for Ecology and Hydrology, UK) Sherry Schiff (University of Waterloo) Brit Lisa Skjelkvåle (NIVA, Norway) Richard Wright (NIVA, Norway)

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4. DOC as an integrator of the effects of multiple stresses

Activities:

• Study the interaction between acid deposition and El Niño-induced drought on DOC fluxes

• Measure DOC fluxes and hydrology in catchments with different types of wetlands Collaborators: Doug Evans (Trent University) David Lean (University of Ottawa) Lewis Molot (York University) Sherry Schiff (University of Waterloo)

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Chapter 2: Overview of Ontario Power Generation Andy Hoffer

Indroduction: OPG overview

Ontario Power Generation (OPG) is an Ontario-based company whose principal business is the generation and sale of electricity to customers in Ontario and interconnected markets (http://www.opg.com). As of December 31, 2002 OPG had a total in-service electricity generating capacity of 22,211 megawatts (MW). This consisted of two operating nuclear stations with a capacity of 5,588 MW (a third nuclear station with a capacity of 2,060 MW was laid up), six fossil-fueled stations with a capacity of 9,700 MW, 36 hydroelectric stations and 29 EcoLogoM-certified green power hydroelectric stations which together had a capacity of 6,923 MW (Fig. 2a). In 2002, fossil fuel contributed 44% of the in-service capacity of OPG which generated a total of 115.8 terawatt-hours (TWh) of electricity.

OPG also owns approximately 2.5 MW of wind power, and OPG’s share of the Huron Wind joint venture is approximately 4.5 MW of capacity. Two nuclear stations, formerly operated by OPG, are leased on a long-term basis to Bruce Power L.P.

Fig. 2a: OPG’s generating stations (Dec. 31, 2002)

During the period 1998 to 2002, the percent contribution of the operating nuclear stations to the total electricity generated by OPG decreased by ~30% (Fig. 2b) whereas fossil fuel power generation increased by 20% (Fig. 2b).

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Fig. 2b: OPG by source from 1998-2002

Fossil Emissions

In 2002, total fossil emissions of acid gas amounted to 189,500 tonnes. Comparisons of acid gas emissions since 1998 show a favourable reduction of 9,300 tonnes, or about 5%, and a decrease in the acid gas emission rate of about 17% (Fig. 2c), with a corresponding increase in the product output from the fossil plants of about 5.4 GWh, or 16% (http://www.opg.com).

Fig. 2c: Total SO2 and NOx emissions and emission rate from 1998-2002

Emission Reduction Initiatives

At the fossil stations, OPG has achieved a steady reduction in sulphur dioxide, nitrogen oxide and particulate emission rates through an integrated air management program, a strategic fuelling program, and significant, continuing investments in pollution

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prevention measures (http://www.opg.com). Investments over the past 10 years have totalled well over $1 billion. A $285M investment in Selective Catalytic Reduction (SCR) technology, announced in 2000, has and will continue to reduce emissions of undesirable nitrogen oxides (NOx), at the Lambton and Nanticoke stations. OPG will also equip the units at the Nanticoke, Lambton, and Lakeview generating stations with low-nitrogen oxide burners. Modifications to the combustion process (including continuous emission monitors and smart computer controls) and the conversion of oil-burning units to allow burning of natural gas, also have/will be attempted. To reduce sulphur dioxide (SO2) emissions, OPG is experimenting with low sulphur fuels, fuel switching and flue gas desulphurization (eg., scrubbers) at two units at Lambton. In 2002, approximately two-thirds of OPG’s energy generation was based upon nuclear and hydroelectric sources (these sources are virtually free of emissions causing smog, acid rain or global warming). OPG placed into operation our first two SCR units at Lambton, and made progress on two other units at Nanticoke, which will be in operation by early 2004. All four units are expected to reduce annual emissions of nitrogen oxides by up to 25 per cent across our entire fossil fleet. In addition, OPG is involved in major initiatives in biodiversity, clean-coal research, fuel-cell development, mercury emissions reduction and the financing of alternative energy technologies. OPG was one of the first utilities in the world to engage in international emission reduction credit trading to help in the cost-effective reduction of nitrogen oxide and carbon dioxide emissions. Our hydroelectric facilities have put in place voluntary restraints and operating limits to minimize the impacts on the natural environment, and meet the needs of competing users of these watersheds. Environmental research supported by OPG

Aquatic Research • Trent University - OPG/NSERC Chair in Watershed Biogeochemistry - effects of

acid rain and influence of climate stress on watersheds. • St. Lawrence River Institute of Environmental Sciences – OPG/NSERC Chair in

Ecotoxicology – effects of water level variations on ecosystem dynamics; sources, transport and fate of contaminants in rivers; community-based restoration of river ecosystems.

• University of Guelph – OPG/NSERC Chair in Metals in the Environment (MITE) – sources and fate of metals entering environment.

Atmospheric Chemistry

• Waterloo Atmospheric Chemistry Chair – development of air quality models to assess OPG fossil plant impacts.

Mercury emissions

• Joint programs with Canada Centre for Mineral and Energy Technology (CANMET), US Dept of Energy and Electric Power Research Institute (EPRI) to develop proven, commercially available mercury control technologies.

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• fate of mercury in the environment.

Future Plans

• Compliance with regulatory SO2 and NOx emission limits. • Future generation mix pending direction from OPG’s shareholder (Ontario

government). • Continued support of relevant research

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Chapter 3: Long-term trends in deposition and lake history Peter Dillon

Atmospheric data have been collected from 6 sites for 25 years, from 23 sites for 10 or more years and from 18 sites for 5 – 10 years. A large decrease (~40%) in SO4 deposition occurred from 1975 to 1985 as a result of SO2 reductions in emissions, but since then S deposition has declined only slightly; the large decrease in 1997-98 being a consequence of climate variables (i.e. drought conditions) (Fig 3a).

Fig 3a: Annual deposition of NO3-, SO4

2-, and NH4+ from 1975 – 2000.

Fig 3b: Sulphate concentrations in Harp and Plastic Lakes from 1981 – 1998.

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However, the decrease in SO42-

concentrations in surface waters in the Muskoka –Haliburton (M-H) area of Ontario, has been considerably less than anticipated based on the declines in atmospheric deposition (Fig. 3b). The differences are due to catchment processes, in particular storage and release of S, which are not related to changes in deposition but rather to various climate factors. Sulphate concentrations and export from catchments were particularly large following dry, warmer than average summers. Climate therefore has an important influence on SO4

2- cycling, and may obscure evidence of recovery from decreased S deposition. There is a great deal of similarity in the year-to-year patterns in SO4

2- concentrations (i.e. synchrony or coherence) among the study streams. When the data were standardized using Z-scores (i.e. annual values minus the mean, all divided by the s.d.) in order to facilitate comparisons among the streams, temporal patterns in SO4

2- concentrations were related mainly to fluctuations in the Southern Oscillation Index (SOI) and to a lesser extent, SO4

2- deposition and other indices (Table 3a). There was also a high degree of coherency in other variables including alkalinity, pH and DOC. This synchrony was apparent in data collected from several regions/zones in Norway as well; standardized SO4

2- concentrations were highly correlated to/could be predicted from the Arctic Oscillation (AO) Index and/or the North Atlantic Oscillation (NAO) (Table 3b).

Table 3a: SO4 predictions (using backwards elimination stepwise multiple regression) for 1 group of study lakes based on various suites of parameters

Predictor

suite r F p Important Predictors

6-lake set

SOI indices 0.71 3.63 0.03 annual (lag1), annual (lag2), spring, summer (lag1)

NAO indices 0.57 3.92 0.04 annual (lag0), fall (lag1)

AO indices none

SO4 deposition 0.72 18.1 0.0005 fall (lag1)

combined 0.83 7.61 0.002 spring SOI, annual NAO (lag 1), fall NAOI, fall SO4 deposition

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Table 3b: Results of stepwise regression analysis of average (z-scored) trend in SO42-

concentration of all lakes in each of the pre-defined 10 regions in Norway with the various climate indices. Only models that were significant at the p<0.05 level are shown.

region number of lakes

no. of years model parameters r2 intercept coefficient P

1 2 14 AOfall 0.40 -0.052 0.40 0.015

2 16 14 AOfall 0.40 -0.083 0.43 0.015

3 3 14 AOfall 0.33 -0.086 0.42 0.031

4 17 14 AOfall 0.31 -0.120 0.32 0.038

5 14 14 no model

6 4 14 no model

7 6 14 NAOspring 0.53 0.212 0.15 0.032 NAOall 0.17 0.021

8 10 14 NAOannlag0 0.58 -0.146 0.05 0.024 AOfall

0.27 0.029

9 6 14 NAOspring 0.33 -0.110 0.14 0.030

10 22 14 no model

fall = average for Sept-Nov; winter = average for Dec-Feb; spring = average for Mar-May; summer = average for Jun-Aug; annlag0 = annual average for same year; annlag1 = average value for preceeding year; annlag2 = annual average 2 years prior

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Chapter 4: i) UN-ECE International Cooperative Programme (ICP) on Assessment and Monitoring of Acidification of Rivers and Lakes: Trends in

surface water chemistry 1990-2001 ii) 2004 Acid Rain Assessment results from Ontario

Dean Jeffries

i) UN-ECE International Cooperative Programme (ICP) on Assessment and

Monitoring of Acidification of Rivers and Lakes: Trends in surface water chemistry

1990-2001

The United Nations-Economic Commission for Europe (UN-ECE) is a catalyst in the international work aiming at reducing transboundary air emissions. The International Co-operative Programme on Assessment and Monitoring of Rivers and Lakes (ICP) monitors the effects of acid rain on water and water courses. The programme was started in 1984, and 27 countries (25 European plus Canada and the USA) participate and supply data to the programme's central database at the Norwegian Institute for Water Research (NIVA). The results provide a joint understanding of the problems of long-range transboundary air pollutants, regardless of international boundaries, as well as it assists the process of harmonisation of methods. The project also demonstrates the benefit of having international agreements to reduce emissions of pollutants.

The ICP collaborators are currently working on the 15-year report that will summarize trends in the water chemistry data for the 12-year period 1990-2001 collected at 189 monitoring stations. The parameters being examined included ‘driving’ variables (SO4

2- and NO3

-), ‘acidity’ variables (Gran alkalinity and pH) and ‘mitigating’ variables (base cations, Ca + Mg, and DOC/TOC). In order to be included in the trend analysis, the sites must meet all the following requirements:

• data available for at least 7 out of 12 years

• data available for all ‘response’ variables (SO42-, NO3

-, base cations) and at least one ‘recovery’ variable (alkalinity, pH)

• sites sensitive to acidification (alkalinity < 200 µeq/L)

• sites with undisturbed catchments

• no major internal sources of sulphate A slightly different method from past trend assessments is being used to analyze the data. Regression slopes were calculated for each water site, and then individual site results were grouped into regions and examined for trends in recovery from and response to changes in acid deposition. The 6 European regions identified were the Alps, East Central Europe (ECE), North Nordic (NoN), South Nordic (SoN), UK/Ireland (UKI) and West Central Europe (WCW), and the 6 North Amercian regions were Maine/Atlantic (Atl), Vermont/Québec (VtQue), Adirondacks (Adk), Appalachians (App), Upper Midwest (MidW) and the Blue Ridge Mountains (BR). The most significant finding in this regional trend analysis is the almost universal decrease in SO4

2– concentrations in lakes and streams throughout Europe and North

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America. Only one region in this analysis failed to show a significant SO42– decrease, and

this is a region (the Virginia Blue Ridge) where soil characteristics make a SO42–

decrease unlikely. In Europe, regional rates of SO42– decline ranged from ca. -1 µeq/L/yr

in the U.K. and Ireland and the Northern Nordic region, to more than -6 µeq/L/yr in the Southern Nordic region. Rates in central Europe were intermediate, with both East- and West-Central Europe exhibiting regional SO4

2– declines of ca. -4 µeq/L/yr. All of these changes represent ecologically significant declines in this acid anion, and are consistent with declines in rates of S deposition in Europe. In North America, rates of SO4

2– decline ranged from ca. -1 µeq/L/yr in the region with the lowest rates of S deposition (Maine and Atlantic Canada), to more than -2 µeq/L/yr in the Adirondack Mountains, Appalachian Mountains, and the Upper Midwest (U.S. and Canada). Rates of sulphate decline in surface waters follow the same geographic pattern, but are lower than, trends in regional sulphate deposition. Over the 12-year period, ICP sites show decreasing NO3

- concentrations in the Adirondack Mountains, Appalachian Mountains and the Virginia Blue Ridge (all in North America), and increasing concentrations in the Alps. No other regions exhibit clear patterns. Although, Gran alkalinity increased slightly in most of the regions, the overall trend was slightly negative (approximately 1-2 µeq/L/yr) in the Alps (Europe) and Atlantic (North America) regions. All of the ICP regions show tendencies toward decreasing Ca+Mg. When the rate of Ca+Mg decline is equal, or nearly equal, to the rate of SO4

2– and NO3– decline, then

chemical improvement (increasing alkalinity and pH) is impeded. In the European regions, most rates of Ca+Mg decrease are smaller than those for SO4

2–. One important exception is the U.K. and Ireland, where rates of rates of SO4

2– and Ca+Mg decline were nearly equal for the time period 1990-2001—this has important implications for improvements in acidity in this region. One North America region (Maine and Atlantic Canada) exhibited stronger decreasing trends in Ca+Mg than in SO4

2– which accounts for the negative alkalinity trend observed for that region. Three North America regions showed significant improvement in Gran alkalinity (Vermont/Quebec, Adirondacks and Appalachians), and two exhibited no change (Upper Midwest and Virginia Blue Ridge).

Dissolved organic carbon is of great interest in any analysis of surface water recovery, because it is an indicator of natural organic acidity. All but one of the ICP regions exhibited positive slopes for DOC, and six of these were considered significant. Overall, the results suggest an almost universal increase in the importance of organic acids. Rates of sulphate decline are smaller in surface waters than in deposition for most regions indicating a lagged response (Figure 4a). This may reflect desorption of S that has accumulated in catchment soils over the past century due to atmospheric deposition. Desorption of stored S has the effect of damping the trends in surface water sulphate and slowing the rate of decline. One exception to the pattern in North America is in the Upper Midwest region of the U.S. (Figure 4a), where most lakes are seepage lakes. Here, the soils play only a minor role in controlling sulphate concentrations, and declines in lake sulphate concentrations are driven by dilution, following the drought that affected this area in the late 1980s and early 1990s.

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Figure 4a: Comparison of trend slopes for SO42– in precipitation (left box -

shaded) and SO42– in surface waters (right box - clear) for the period 1990-

2000 in acid sensitive regions of North America and Europe. North American deposition data are for trends in wet deposition concentrations. European deposition data are for trends in combined wet and dry deposition. Each box shows the range (25th to 75th percentiles, with line at median) of slopes; confidence intervals indicate 10th and 90th percentiles; dots indicate 5th and 95th percentiles.

ii) 2004 Acid Rain Assessment results from Ontario

A national Acid Rain Assessment is currently in preparation. The aquatics component will comprise Chapters 6, 7 and 8. Chapter 6 will focus on chemical and biological

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effects including trends, Chapter 7 will focus on ‘recovery’, including observations and predictions and Chapter 8, aquatic and terrestrial critical loads and exceedances.

To detect chemical trends in Ontario lakes, the non-parametric statistical procedures have been used to analyse data collected from 615 lakes over the 1990 – 2001 time period and with the same variables (SO4

2-, NO3- and base cations and alkalinity or pH) as in 15-year

UNECE ICP-Waters report. In Ontario, the lakes, located primarily in the M-H., Sudbury and Algoma regions of Ontario (with a few located in NW Ontario) are being/have been investigated, with the shortest record = 8 years and the median record = 11 years. Data sources include Environment Canada (EC), the Canadian Wildlife Service (CWS), MOE, and DFO. The sample frequency in the individual data sets is quite variable. The occurrence of significant +ve and –ve trends are summarized in the following table. The trend magnitudes are estimated by the median linear regression slopes. Table 4a: Summary of significant chemical trends (p<0.05) in 615 Ontario lakes within the period 1990-2001 detected using non-parametric statistical procedures.

n % Slope

Increasing Trends 5 1 4.65

Decreasing Trends 162 26 -5.21

No Significant Trends 447 73













Increasing Trends 41 7 139

Decreasing Trends 17 3 -234


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Chapter 5: Effects of drought-induced acidification on diatom communities in acid-sensitive Ontario lakes

Roland Hall

(from Faulkenham, Hall, Dillon and Karst-Riddoch. 2003. Limnol. Oceanogr.

48:1662-1673)

Introduction

Despite the implementation of sulfur emission control programmes in North America and a 60% decrease in bulk sulphate deposition in eastern Ontario since the late 1970s (Fig. 3a), the chemical recovery of acidified lakes is often not as expected due to interactions between climatic events and previously deposited acids that have been stored in the watershed, specifically in wetlands. Wetlands play an important role in mediating drought-related reacidification and delaying chemical recovery of acid-sensitive lakes. During cool, wet years, much of the sulphate deposited from the atmosphere is reduced and stored as elemental S, but during subsequent dry years, the stored S becomes oxidized to sulphate (as the water table is drawn down) which is readily exported to lakes and streams when wetter conditions resume (see Chapter 6). While wetland-mediated reacidification has been linked to declines in pH and increases in sulphate concentrations in acid-sensitive lakes, the effects of this phenomenon on aquatic biota remain largely unknown. Thus, in order to evaluate whether wetland-mediated interactions between climatic variability (specifically drought) and acid deposition exert a strong control on algal communities, we compared changes in diatom assemblages in 200-yr-long sediment cores collected from two lakes with similar basin characteristics but different proportions of wetland area (0% in Blue Chalk, BC and 4.4% in Chub, CB). Comparisons of past changes in diatom assemblages were used to assess whether the influence of wetlands in the CB Lake catchment resulted in more variable conditions than in BC Lake. We also explored whether water residence time could account for differences between the two lakes. In addition, canonical ordination-based variance partitioning analysis was used to quantify relationships between annually-resolved sedimentary analyses of diatoms and 20-year records (1977-1997) of acid deposition, climatic factors and water chemistry changes and their interactions, in order to test whether wetlands are important in regulating diatom communities in acid-sensitive lakes. Results

There were marked differences in temporal water chemistry trends between CBLake and BC Lake (Fig. 5a). Mean ice-free concentrations of DOC, SO4

2- and pH were more variable in CB than in BC Lake. Distinct changes in [SO4] and pH occurred in CB Lake after droughts which corresponded to declines in the Southern Oscillation Index (SOI). In CB Lake, DOC concentrations were higher, pH was lower and both were more variable than in BC Lake. Similarly, diatom assemblages were significantly more variable in CB Lake over the past ~200 yr (Fig. 5b). In CB Lake, increases in Fragilaria sp. 5 PIRLA, which coincided with

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Figure 5a: Water chemistry trends in Chub (crosses) and Blue Chalk (squares) Lakes.

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reduced abundance of Rhizosolenia eriensis, tended to occur during years follow droughts. However, similar short-term fluctuations were not observed in BC Lake; rather there was a pattern of declining abundances of Asterionella ralfsii v. americana and increasing A. formosa, consistent with recovery from acidification since the late 1970s. Figure 5b: Diatom assemblages in the sediments of Chub (top) and Blue Chalk (bottom) lakes.

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ra p

ergl

abra

v. f

luor

inia

e

10

Aulac

osei

ra sub

arct

ica

10

Cyc

lote

lla b

odan

ica

v. le

man

ica

10

Cyc

lote

lla ste

llige

ra

20

Eunot

ia n

aegl

ii

20

Eunot

ia zas

umin

ensis

10

Fragi

laria

capu

cina

v. g

raci

lis

10

Fragi

laria

nana

na

10

Rhizo

sole

nia

cf. e

rien

sis

20

Tabel

laria

flocc

ulos

a str

. IIIp

20 40 60

Achna

nthe

s m

argi

nula

ta

10

Anom

oeon

eis br

achy

sira

10

Total

Eun

otia

10

Fragi

lara

con

stru

ens v.

ven

ter

20

Fragi

lara

sp. 5

PIR

LA

20

Frustul

ia rho

mbo

ides

10

Tabel

laria

flocc

ulos

a str.

IV

10

% P

lank

toni

c

40 60 80 100

% Abundance

Planktonic Benthic

2000

1986

1965

1937

1902

1863

1817

Age

(yr)

P la n k to n ic B e n th ic

% A b u n d a n c e

Ast

erio

nella

form

osa

2 0

Dep

th (

cm)

0

5

1 0

1 5

2 0

2 5

3 0

3 5

4 0

Ast

erio

nella

ralf

sii v . a

mer

icana

1 0

Cyc

lote

lla b

odanica

v. lem

anica

1 0

Cyc

lote

lla k

uetzi

ngiana

v. r

adiosa

1 0

Cyc

lote

lla s

tellig

era

2 0 4 0 6 0

Fra

gilari

a capuci

na v. g

raci

lis

1 0

Fra

gilari

a nanana

1 0

Tabel

lari

a flocc

ulosa s

tr. I

IIp

1 0

Ach

nanthes

leva

nderi

1 0

Ach

nanthes

min

utiss

ima

1 0

Fra

gilari

a const

ruen

s v. ven

ter

1 0

Navi

cula

sem

inulu

m

1 0

Navi

cula

subm

inusc

ula

1 0

Navi

cula

vitio

sa

1 0

Pin

nulari

a bic

eps v. p

usilla

1 0

% P

lankto

nic

4 0 6 0 8 0 1 0 0

1 9 8 1

1 9 5 4

1 9 1 4

1 8 6 9

1 8 0 3

2 0 0 0A

ge (y

r)

22

Variance partitioning analysis of (approximately) annually resolved sedimentary diatom analyses (1977-1997) indicated that the effects of water chemistry (i.e., ice-free mean alkalinity, conductivity, and Ca, DOC and total phosphorus concentrations), independent of acid deposition and climatic factors, accounted for 25% of the variation in the diatom assemblages in Chub Lake. Acid deposition (i.e., annual SO4

2-, Jan NO3

-, May NO3- and Aug NO3

- deposition in mg/m2) and climatic factors (i.e., mean April temp., mean May temp., mean Nov. temp., Jan. precip. and Aug. precip.) accounted for similar, significant amounts of the variation in the diatom communities (22% and 24%, respectively). Complex interactions among all three factors explained an additional 10% of the variation in the diatom community at CB Lake, but only 1% at BC Lake. Conclusions

The results suggest that wetland-mediated interactions among climatic factors and acid deposition can lead to chemical changes that exert substantial effects on the diatom communities at Chub Lake. The study demonstrates the striking differences in variability and composition (over a period of 20 – 200 years) of the primary producer communities in the two lakes, which result primarily from differences in the area of contributing wetland in each catchment. An important implication is that current estimates of rates of biological recovery in acidified lakes, which fail to account for wetland-mediated re-acidification, may be erroneous and over-optimistic.

23

Chapter 6: Sources and controls of sulphur cycling in catchments in south-central Ontario

Catherine Eimers

There are two ways to evaluate the effects of changing SO42- deposition on surface

waters; the first is by examining changes in SO42- concentrations in lakes and streams

and the second is by using catchment mass balances (i.e., an accounting of SO42-

inputs and outputs), which can provide insights into biogeochemical processes An examination of long-term (20+ years) surface water SO4

2- concentrations revealed that the response of lakes and streams to declining SO4

2- has been varied. While SO42-

concentrations have generally declined in most surface waters (Fig. 6a), the magnitude of the decline has been substantially less than the change in deposition. Furthermore large inter-annual variations in annual average SO4

2- concentrations and export were synchronous among all catchments (Figure 6a), despite considerable diversity in soil depth, wetland-coverage, forest type, etc. among sites. These temporal patterns in SO4

2- levels and export were not related to changes in deposition but rather to various climate factors. Specifically, SO4

2- concentrations and export from catchments were particularly great following dry, warmer than average summers, when stream flow ceased for up to several weeks at a time such as the El Niño years of 1983/84 and between 1987/88 and 1990/91, inclusive. Climate therefore has an important influence on SO4

2- cycling, and may obscure evidence of recovery from decreased sulphate deposition.

Figure 6a: Pattern of mean annual SO4

2- concentration in 8 lakes (top) and 10 streams (bottom) in the M-H region of Ontario.

0

2

4

6

8

10

12

14

16

18

1982

1984

1986

1988

1990

1992

1994

1996

1998

SO

42-

(mg

/L)

PC1

RC1

BC1

PC1-08

HP4

HP5

CB1

CB2

DE10

CN1

0

2

4

6

8

10

1982

1984

1986

1988

19901992

19941996

1998

SO

42

- (m

g/L

)

PC

RC

HP

BC

DE

HY

CN

CB

24

It appears that in wetland-dominated catchments, such as the main inflow to Plastic Lake (PC1), S mobility is largely controlled by variations in wetland water table height; retention occurs under wet conditions when water tables are high, whereas during warm dry periods, when water tables decline, oxidation of stored S compounds in peat results in enhanced SO4

2- mobilization.

In predominantly upland catchments, the source of climate-related SO42- release is less

clear. We hypothesized that riparian zones and/or minor wetland areas were contributing to climate-related SO4

2- release and used changes in stable S-isotope composition to identify the source(s) of SO4

2- exported in PC1 stream water. Sulphur is comprised of four stable isotopes, although the majority is 32S and 34S. Because microbial SO4

2- reduction is selective for lighter 32S, peat becomes enriched in 32S and

residual SO42- is enriched in 34S (i.e. δ34S is more positive during anoxic/wet periods);

oxidation (drought) of reduced S releases 32S-enriched SO42- and so δ34S is more

negative following dry periods.

Deposition Throughfall Foliage Soil Leachate PC1-08 (upland) PC1 (wetland outflow) Peat

δ34S‰

Fig. 6b: δ34S‰ isotope ratios at Plastic Lake.

Higher δ34SO4 ratios (Fig 6b) in the swamp outflow (average +8.6 ± 2.6‰) compared to the upland-draining inflow (average +5.7 ± 0.7‰) indicated that dissimilatory sulphate reduction was at least partly responsible for net SO4

2- retention in the swamp. In contrast, the isotopic composition of SO4

2- in upland stream water was relatively constant over time (CV 12%) and was similar to bulk deposition (average +5.1± 0.6‰), despite large variations in water table height (-47cm to +45cm) in the streambed over the same period. These data indicate that redox reactions are not an important component of the upland S cycle. Thus we rejected our initial hypothesis

-2 -1 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

25

that relatively organic S-rich streambed soil was functioning as a ‘mini wetland’ and supplying net SO4

2- export from the upland catchment.

Sulphate input-output budgets indicated that in PC1 as well as in the majority of the M-H catchments, SO4

2- export in stream water was greater than inputs in deposition (i.e. net export) in nearly all years of record, signifying an unidentified, internal source of SO4

2-. This observation is not unique to the M-H region and has been reported in other catchments in eastern Canada (eg. Turkey Lakes). Factors which trigger the release of SO4

2- (eg., temperature, moisture) including the effects of decreased SO42-

deposition and changing climate and N-deposition are critical and it is difficult to determine which of the potential mechanisms for SO4

2- export (i.e., desorption of SO4

2- from the inorganic adsorbed pool or mineralization/oxidation of SO42- from the

organic pool) is most important. Laboratory experiments indicated that peat was the primary source of SO4

2- in wetland outflows following drought whereas in ‘upland’ soils, higher SO4

2- concentrations following fluctuations in soil moisture may have resulted from enhanced mineralization due to drying and rewetting of the forest floor. Results from other experiments showed that soils adsorbed additional SO4

2- when SO4

2- concentrations in soil solution were increased, and conversely released stored SO4

2- when SO42- concentrations in soil solution were lowered. Extrapolating these

results to the field suggests that mineral soils will respond to decreases in SO42-

deposition by releasing previously adsorbed SO42-. The magnitude of organic S

storage (~580 kg S/ha) relative to inorganic SO42- storage (290 kg S-SO4/ha) at PC1-08

indicates that this pool could also be an important source of export over the long-term since the experimental data further indicate that SO4

2- release from organic S may be responsive to climate perturbations.

We conclude that net SO4

2- export in upland catchments is due to a combination of desorption from the adsorbed SO4

2- pool and mineralization of organic S compounds. At PC1-08, the pools of adsorbed SO4

2- and organic S in the soils are large relative to the magnitude of net SO4

2- export currently measured (~6 kg S-SO4/ha/yr), indicating that soils may release excess SO4

2- for years to come. Since net SO42- export can cause

continued net losses of base cations (BC), the duration of the SO42- release may be

critical in catchments with low BC pools. Furthermore, if net SO42- export is

substantial, critical load models based solely on deposition will inaccurately predict the impact of declining deposition.

26

Chapter 7: Sudbury Environmental Study (SES): an overview Bill Keller

The Cooperative Freshwater Ecology Unit (Coop Unit) represents a long-term collaboration among government agencies/universities including the Ontario Ministry of Natural Resources (MNR), the Ontario Ministry of the Environment (MOE) and Laurentian University (LU). Through the Coop Unit, MOE conducts both an extensive and an intensive sampling program, a continuation of studies initiated in the 1970’s under the Sudbury Environmental Study. The Extensive Monitoring Program began with 209 lakes (in 1974-1976) which grew to 250 lakes (in 1981-1983) for which water chemistry and zooplankton samples (in 1981, 1988 and 2003) have been taken with the intention of identifying and determining the scope of problems related to the Sudbury smelters (eg., acidification, metal contamination). Within this data set is a subset of 44 lakes located in northeastern Ontario, which had pH < 5.5 in the early 1980’s and which are still sampled once annually to monitor recovery from acidification and metal contamination. Twelve other Intensive Monitoring lakes are sampled more frequently (eg., monthly through the ice-free season); in addition to water chemistry and zooplankton samples, phytoplankton, some benthic invertebrates, oxygen, temperature and Secchi depth also are monitored. In the extensive lake data set, the % of lakes having pH < 5.0 has declined substantially since 1981, while the number with pH > 6 has increased (Fig. 7a). Ni concentrations in the water also have declined with levels in 2003 below the Provincial Water Quality Objective of 25µg/L, except for those lakes within ~20 km of Sudbury. On the other hand, Ni concentrations in the sediments of lakes sampled in the 1990s, as far as 50 km from Sudbury, were still above the Provincial Sediment Quality Guideline of 75 µg/g. In 2002, SO4

2- concentrations in the lakes ranged between ~100 and 400 µeq/L depending on the distance from Sudbury, with levels declining approximately exponentially as distance increased. While the distance relationship is still evident, SO4

2- concentrations are now much lower than in the past. Unfortunately, the declining Ni and SO4

2- concentrations and increasing pH have been accompanied by a general decline in Ca concentrations. In George Lake, for example, [Ca]’s have declined by ~50% since 1970 (Fig. 7b).

27

Figure 7a: Changes in pH in the 44 extensive lake data set since 1981. The y-axis represents the percentage of lakes studied. Figure 7b: Measured Ca concentrations in George Lake since 1970. The dashed horizontal line represents the inferred (from diatom fossils in the sediments) pre-industrial [Ca] in the lake.

0%

20%

40%

60%

80%

100%

1981 1985 1990 1995 2000 2003

>6.0

5.5-6.0

5.0-5.5

<5.0

0

50

100

150

200

250

1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002

Year

Measure

d C

a (ueq/L

)

28

Calcium is important in lakes for several reasons. Many organisms, such as crustacean zooplankton and crayfish require Ca for their shell; thus changes in Ca levels may affect community structure. Furthermore, Ca may modify the effects of UV-B radiation, acidity and metals on aquatic biota. While zooplankton species richness has generally increased in Sudbury area lakes since 1980 with chemical recovery, the zooplankton communities in these lakes are still different than those found in M-H area lakes that have not undergone such extensive acidification and metal contamination. Both biological resistance to recovery due to altered food webs and insufficient time for recolonization are factors that appear to be affecting the biological recovery of Sudbury area lakes. It is also evident that significant weather events such as drought, corresponding to El Niño years of 1983/84 and 1987/88 to 1990/91, inclusive, are having a considerable impact on lake recovery. pH, bottom water temperature and water clarity (i.e., Secchi depth) all have been found to be influenced by this weather phenomenon. Wind has also been shown to be important in explaining lake thermal structure. A preliminary model was therefore developed on one long term study lake, Whitepine Lake, to predict depth (m) of the 10oC isotherm, whereby: Depth at 10˚C = -3.307 *[DOC] + 0.0032*(Wind Days) + 13.773 (7.1) (r2 = 0.75) This initial model was then tested on nearby Aurora Whitepine Lake, with excellent results (r2 = 0.83).

29

Chapter 8: Application of the Steady-State Water Chemistry (SSWC) and First-Order Acidity Balance (FAB) Models in Ontario

Julian Aherne

Critical load (CL) has been defined as “a quantitative estimate of an exposure to one or more pollutants below which significant harmful effects on specified sensitive elements of the environment do not occur according to present knowledge”. It can be used to: 1) evaluate the adequacy of existing and future sulphur (S) emission control programmes, 2) quantify the need for further reductions in S and nitrogen (N) depositions, 3) determine the potential impact of forest harvesting on lake chemistry and critical loads estimates. At present two steady-state models are widely used for calculating CLs in surface waters, the Steady-State Water Chemistry (SSWC) model and the First-order Acidity Balance (FAB) model. The SSWC model requires weighted annual mean water chemistry, or an estimate thereof, and annual runoff to calculate critical loads of acidity, CL(A). The FAB model allows the simultaneous calculation of critical loads of acidifying S and N deposition. In addition, it takes into account in-lake retention of S and N and also processes (uptake, immobilisation and denitrification) in the terrestrial catchment soils.

0 100 km

CANADA

USA

Georgian Bay

Muskoka

Nipissing

Haliburton

Parry Sound

Muskoka river catchment

Figure 8a: Study area showing the lakes (n = 1469) in south-central Ontario used for the SSWC assessment and lakes (n = 285) in the Muskoka River Catchment used for the FAB assessment.

30

Steady-State Water Chemistry (SSWC) model

The SSWC model was used to determine the critical loads of acid deposition, CL(A), in 1469 lakes in five regions in south-central Ontario (Figure 8a). The majority of the lakes in four of the regions (i.e., Muskoka, Haliburton, Nipissing, and Parry Sound) are sensitive to the deposition of strong acids because of their geological setting. The fifth set of lakes, the southern lakes, are located along the southern boundary of the Precambrian Shield and were suspected to be much less sensitive to acid deposition. The CLs are <50 meq/m2/yr in 28% of the lakes, 50–100 meq/m2/yr in 54% of the lakes and >100 meq/m2/yr in only 18% of the (mostly southern) lakes with a mean CL(A) for all lakes of ~100 meq/m2/yr. Critical loads calculated for a subset of 285 lakes in the Muskoka River (M-R) catchment were generally lower than in the region as a whole (Fig. 8b).

0

5

10

15

20

25

30

35

40

0 10 20 30 40 50 60

Regional (n = 1370)

Muskoka (n = 285)

Cu

mu

lative

dis

trib

uti

on

(%

)

Critical loads of acidity (meq/m2/yr)

0

20

40

60

80

100

0 50 100 150 200

Figure 8b: Cumulative distribution function for critical loads of acidity ‘regionally’ versus the Muskoka River catchment lakes. The dotted line represents annual average sulphate bulk deposition for the period 1995–1999 (41 meq/m2/yr). The SSWC model indicates that most lakes that had critical load exceedances in the 1970s and early 1980s now experience deposition at or near their critical load. However, further S emission reductions may still be warranted. All lakes are protected, if sulphate (SO4

2–) deposition is below the minimum of the CL(A) values. However, to ensure that a sufficient percentage of lakes is protected, the pentile (95% protection limit) is the commonly accepted level. The pentile CL(A) is 34 meq/m2/yr for all regions and 30 meq m2/yr for the M-R catchment; both values are similar to previous critical load studies in central Ontario (i.e., 33 meq/m2/yr).

31

First-order Acidity Balance (FAB) model

The FAB model was used to estimate the critical loads of acid deposition for the subset of lakes in the Muskoka River (M-R) catchment. The advantage of the FAB model is that it can provide estimates of the reductions needed for S and/or N deposition. In general, these 285 study lakes are slightly smaller and more acid sensitive than the regional data set (M-R lakes; median lake area = 25.5 ha, runoff = 510 mm, Ca2+ = 132.1 µeq/L; regional data-set: lake area = 28.1 ha, runoff = 479 mm, Ca2+ = 144.7 µeq/L). The simultaneous treatment of S and N does not allow the calculation of a single CL value; rather the concept of a Critical Load Function (CLF) must be introduced. The CLFs for each of the 285 study lakes have been aggregated by computing percentile functions (Figure 8c). All combinations of S and N deposition lying below the n-th percentile function protect n-th percent of the lakes. Thus ~55% (i.e., 133 lakes) of the lakes exceed CL (45% are protected) based on current S and N deposition (see cross in Figure 8c). Unlike the SSWC method, there is no unique amount of S and N to be reduced to reach non-exceedance. For example, from current deposition, 95% of the lakes can be protected (5% line in Fig. 8c) by reducing only S deposition to ~10 meq/m2/yr; alternatively the pentile CLF may be reached by reducing a combination of S and N deposition. However, the pentile CLF cannot be reached by N reductions alone.

0

10

20

30

40

50

60

70

80

0 50 100 150 200

Nitrogen deposition (meq/m2/yr)

Su

lph

ate

de

po

sitio

n (

me

q/m

2/y

r)

5% 50%10% 40%

n = 285

0% 20% 30% 60%

70%

80%

90%

95%

Figure 8c: Percentile distributions of the Critical Loads Functions for the study lakes in the Muskoka River catchment. The annual average bulk deposition of sulphate and nitrogen (41 meq/m2/yr and 63 meq/m2/yr, respectively) for the period 1995–1999 is indicated by the cross.

32

The FAB model can be used to predict the effects of changing SO4

2- and N deposition on CLs (Fig. 8d). For example, in the year 2010, if sulphate deposition were to decrease by 70% with no corresponding decrease in N deposition, then 95% of the lakes would be protected. However, a 70% decrease in N deposition would also require ~25% decrease in SO4

2- deposition to attain the pentile CLF. It is evident, therefore, that in these lakes, S reductions are more effective than N reductions in protecting our surface waters.

0

20

40

60

80

100

0 20 40 60 80 100

Nitrogen reduction (%)

Su

lph

ate

re

du

cti

on

(%

)

Year 2010

60% 70% 80% 90%

100%

85%75%

95%

Figure 8d: Predicted percentile distributions of the Critical Loads Functions for the study lakes in the Muskoka River catchment based on varying decreases in S and/or N deposition.

33

Chapter 9: Application of the Model of Acidification of Catchments (MAGIC) in Ontario

Julian Aherne

The Model of Acidification of Groundwater in Catchments (MAGIC) is a lumped-parameter model of intermediate complexity, developed to predict the long-term effects of acidic deposition on soil and surface water chemistry. The essential data for the regional site-specific application of the model are time-series of annual atmospheric deposition (wet plus dry), physical and chemical characteristics of the soils and lakes, and observed soil and lake chemistry for each study lake.

MAGIC was applied to 25 lakes (all located in the District of Muskoka and the Counties of Haliburton and Nipissing in south-central Ontario) having 10+ years of monitoring data. The majority (17) are headwater lakes; as a result catchment to lake area ratios are relatively high and water replenishment rates are long (median ~2 years). The model was calibrated using observed (or ‘target’) values of surface water and soil chemistry for a specified period. The 8 site-measured target variables were surface water concentrations and soil exchangeable fractions for base cations, calcium (Ca2+), magnesium (Mg2+), sodium (Na+) and potassium (K+). Multiple calibrations were performed on simulations run from 1850 to 1997 using historical deposition sequences. Once calibrated, soil and lake water chemistry data for each lake were simulated for the period 1850–2050. Lake water sulphate (SO4

2–) concentrations follow the same trends as deposition with large changes in concentrations over time (Figure 9a). Spatially there is very little variation in concentration ranges among lakes. In 1875, SO4

2– was <25 µeq/L in all lakes, whereas by 1975 the majority of lakes had concentrations >175 µeq/L. In recent years, reductions in acid deposition have resulted in recovery (mode = 100–125 µeq/ L), which is predicted to further improve by 2016 (mode = 50–75 µeq/L). However, the proposed emission reductions will not return [SO4

2–] to pre-acidification levels. The sum of base cations (SBC) follows a similar pattern to SO4

2–; however, spatially there is a greater variation in values among lakes, reflecting the difference in catchment weathering rates (Figure 9a). In 1875, the modal value was ~100–150 µeq/L, which increased to 250–300 µeq/L by 1975. This reflects the increased export of SBC from the soil exchangeable pool under high levels of acid deposition. The reduction in acid deposition by 1997 resulted in a modal value of 150–250 µeq/L. By 2016, the SBC are predicted to almost completely recovery to pre-acidification concentrations. This suggests that soil exchangeable pools in the region have become depleted of base cations and have a reduced buffering capacity against acid deposition. Modal ANC decreased from 100–125 µeq/L in 1875, to 50–75 µeq/L in 1975 and partially recovered to 75–100 µeq/L in 1997. Although future predictions show the modal value at 50–75 µeq/L in 2016, there is a continued slight recovery as a greater percentage of lakes had ANC >50 µeq/L (88% as opposed to 72% in 1997). The absence of a significant recovery is due to the reduction in the SBC in 2016. The pH values show a similar pattern to ANC; historically high pH values shift to lower values during peak

34

depositions and gradually recover under emission reductions. In 1875, the modal pH was 6.75–7.00. Under 1975 deposition, pH values show a bimodal distribution with one peak at 5.50–5.75 and another at 6.00–6.25. The lower values represent lakes that have limited buffering under the higher acid deposition load. By 1997, the bimodal distribution disappeared and there was gradual recovery (mode = 6.25–6.50), with further recovery predicted by 2016 (mode = 6.50–6.75).

25 75 125 175

ANC (µmol L )c

–1

2

6

7

5

3

2

4

5

10

3

2

1

2

5

9

1

2

1

3

8

6

5

1 1 1

5

SBC (µmol L )c

–1

3

11

9

1 1

3

7

9

4

2

2

10

2

1

10

50 150 250 350

1

11

2

1

10

18

75

2016

1997

1975

25

1

5

2

22

2

18

25 75 125 175

21

4

0

25

50

75

100

0

25

50

75

100

0

25

50

75

100

5.50 6.00 6.50 7.00

pH

0

25

50

75

100

Cum

ula

tive fre

quency (%

)

4

43

8

4

1

6 6

9

3

3

11

1

6

6

3

7

15

Fre

quency (

n)

Figure 9a: Predicted distribution (histograms and cumulative frequencies) of sulphate, sum of base cations (SBC), acid neutralising capacity (ANC), and pH concentrations in the 25 study lakes during the years 1875, 1975, 1997 and 2016. Several refinements or additions to MAGIC have been proposed or implemented through the years as a result of the many applications of the model. The latest version of MAGIC (7.77) includes a wetland compartment that incorporates redox processes driven by climate events. The MAGIC ‘wetland’ module must be run on a monthly time step as redox dynamics are driven by the monthly discharge (Q) thresholds, i.e., annual resolution data are not sensitive enough to define periods of drought. When wetland

35

monthly Q is less than the specified Q threshold, only S oxidation occurs; S reduction can be specified to occur at all times or only when discharge is greater than the Q threshold.

0

10

20

30

40

50

60

1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997

Su

lph

ate

(m

eq

/m2/y

r)

Figure 9b: Comparison of simulated (line) and observed (circle) monthly average stream-water fluxes (meq/m2/yr) of sulphate at Plastic Lake (PC1), Jan. 1980–Dec. 1997. The output of the calibrated model compares well with observed monthly data (Figure 9b; R2 = 0.91). The wetland retains SO4

2– during non-drought years and then exports SO4

2– during drought years (1983, 1987, 1988, 1989 and 1990). The model simulations agree with previous studies that demonstrate the importance of the Plastic Lake (PC1) wetland in controlling the runoff chemistry. The results of the simulation clearly demonstrate that drought years have a substantial effect on wetland dynamics. The dynamic perspective of acidification complements previous steady-state modelling approaches used to evaluate the adequacy of existing emission controls in south-central Ontario. Model simulations predict clear links between lake acidification and the deposition of strong acids. During the past two decades, decreased SO4

2– deposition has resulted in significant recovery, which is predicted to further improve by 2050. Under the current deposition scenario, lake water concentrations in 2050 will have recovered to levels predicted for the early 1900s. However, although regional simulations predict that base cation losses have decreased in recent years, soils in the region will continue to acidify with Ca losses dominating depletion of the exchangeable pool.

36

Chapter 10: Acidification in Ontario: monitoring and modeling Shaun Watmough

Understanding and quantifying biogeochemical processes in forest soils receiving elevated deposition of sulphur (S) and nitrogen (N) (i.e., acid rain) is vital for predicting both the short- and long-term impacts of acid deposition on forests and lakes. Acid deposition falls onto forest soils, and mobile acid ions such as SO4

2+ and NO3

- may pass through the soil into streams. However, in soils, H+ ions in solution are replaced by base cations (BC) leading to an increase in the pH of water percolating through soil. As a result, lakes are protected to some extent against acidification, but if losses of BC from soils exceed the natural replenishment from mineral weathering and deposition, soils will acidify. As soils acidify, the level of exchangeable BC falls, and their ability to buffer the acid deposition diminishes. In areas that are harvested, these impacts will be much more pronounced because large quantities of BC will also be removed from the site as timber. Ultimately, declining exchangeable BC pools in soils could lead to a decline in forest health, increased metal toxicity and further acidification of surface waters. Key processes that need to be understood are 1) inputs of BC from weathering and deposition; 2) the behaviour of S and N in forest soils; 3) the response of biota to declining BC levels and increased metal levels in soils; 4) the impact of harvesting on forest biogeochemistry and 5) the impact of climate changes on forest biogeochemistry. This information can then be used to calculate more accurate critical loads (acceptable level of acid deposition that will not damage the ecosystem) for forests in Ontario. Mass balance calculations for base cations, S and N were previously conducted at 7 forested catchments in Muskoka-Haliburton (M-H) that are particularly sensitive to acid deposition because of shallow soils, low base saturation and high levels of acid deposition. Atmospheric inputs included bulk deposition, which was measured (i.e., wet + dry deposition) plus an estimate of additional dry deposition resulting from canopy savaging. Export in streams was determined from continuous flow measurements (weirs) and frequent sampling for chemistry. Soil measurements were used to determine biomass nutrient pools and BC mineral weathering rates, which were estimated from: 1) the level of Zr enrichment in the soil, 2) PROFILE - a computer model using current soil characteristics, and 3) correlations between mineralogy and climate. The main results of the mass balance studies were that: 1) soils in M-H are still acidifying despite reductions in S deposition, particularly in the biologically important upper soil horizons, 2) losses of Ca are of greatest concern; Ca-losses from soil over the period 1982-1999 represent up to 70% of the exchangeable (plant available) Ca pool measured in 1999/2000 (Fig. 10a), 3) losses of Ca occur through both leaching and harvesting, 4) our predictions of BC weathering are reliable.

37

0

200

400

600

800

1000

1200

1400

1600

PC1 CB1 RC4 HP3A HP4 HP6 HP6A

Sub-Catchment

Ca (

kg

/ha

)

Ex-Ca in 1999/2000 Ca loss (1982-1999)

Figure 10a: Estimated exchangeable Ca (1999/2000) and total Ca-losses from soils in the seven study catchments over the 17-year period (1982-1999).

Critical Loads (CLs) of acidity for forest soils

Information from catchment studies were used to estimate CLs of acidity for forest soils at the seven intensively studied catchments. Critical loads are estimates of acceptable acid deposition that will not damage biota. By definition they allow soils (and lakes) to acidify until the critical chemical criteria are reached (see Chapter 8). Thus, losses of Ca (Fig. 10a) could still occur from forest soils even if the CL is exceeded, provided the soil chemistry does not fall below the critical chemical threshold. The CL, i.e., CL(S + N) for forest soils, is calculated as: CL(S + N) = CL(S) + CL (N) = BCdep + BCwe - BCup + Nup + Ndnit + Nimm +Alkle(crit) where (all units in eq/ha/yr), BCdep = base cation (Ca, Mg, K, Na) deposition, BCwe = base cation weathering, BCup = net base cation uptake, Nup = net nitrogen uptake, Ndnit = net denitrification rate, Nimm = net nitrogen immobilization rate in soil (assumed to be 2 kg/ha/yr=143 eq/ha/yr), and Alkle(crit) = critical alkalinity leaching rate, which was determined using a BC to aluminum (Al) ratio of 10. Table 10a: Calculated critical loads for acidity (includes tree harvesting) and exceedances (assuming a current deposition of S + N = 1030 eq/ha/yr) at 7 forested catchments in south-central Ontario.

CL (eq/ha/yr) Exceedance (eq/ha/yr)

PC1 666 364

HP4 1020 10

HP6 684 346

HP6A 776 254

HP3A 808 222

CB1 688 342

RC4 831 199

38

The results (Table 10a) suggest that reductions in S and/or N deposition of between 10 and 364 eq/ha/yr (~30 % of present deposition) are needed to prevent the critical chemical limit in soil from being exceeded at the 7 study catchments. Soils will acidify further until these chemical limits are reached and Ca will continue to be lost. These calculations assume that current bulk S + N deposition in south-central Ontario is ~1030 eq/ha/yr; however total S and N deposition could be 20 – 30% higher depending upon unmeasured contributions from dry deposition. Furthermore, setting a critical chemical criterion based on a BC:Al ratio of 10 may not protect forests from negative biological effects such as declining sugar maple growth and foliar [Ca]:[Mg]s. Thus setting the critical chemical limit as an ANC (Acid Neutralizing Capacity) = 0 µeq/L may be more realistic. Given the effect of climate variations on SO4

2- and NO3- export (see Chapter 6), and the fact that SO4

2- export is much greater than inputs in bulk deposition (Chapter 6) and also that mass balance studies have demonstrated that most of the N is currently retained in the catchments, then SO4

2- leaching losses may have to be reduced by up to 79% to protect forests.

Critical load estimates for forests have been expanded to cover the whole of Ontario using CANSIS as the base layer. Weathering rates were calculated for over 1400 soil polygons using weighted average values for soil properties (eg., texture, depth, bulk density), bedrock geology and temperature, while values for uptake/harvesting/land cover, deposition and runoff/N dynamics were calculated directly. These calculations indicate that the CL is exceeded at over 40 million hectares in Ontario (Figure 10 b).

exceedance > 100exceedance 0–100no exceedance 0–100no exceedance > 100

Exeedance (eq/ha/yr)

Figure 10b: Exceedance of the critical loads of acidity (S + N) for forests in Ontario.

39

Linking changes in soil chemistry to surface waters

The potential impacts of harvesting on lake chemistry and thus the CL were determined using a modified version of the SSWC model, which included various harvesting scenarios (i.e., no harvesting, wood only, stem only (wood + bark), and whole-tree harvest). In a survey of more than 1300 lakes, the CL was exceeded by bulk SO4

2- deposition in only 9% of the lakes if harvesting did not occur. However, the percentage increased to 23%, 56% and 72% under forest harvesting scenarios that assumed wood only, stem only or whole-tree harvesting, respectively. This increase in the exceedance of CLs is due to the much lower BC concentrations in lakes that result from BC removals during harvest. For example, only 0.3% of lakes will have Ca2+ concentrations <50 µeq/L if harvesting does not occur, whereas 52% of lakes are predicted to have Ca2+ concentrations <50 µeq/L if whole-tree harvesting occurs. It is apparent that forests and surface waters are much more sensitive to acid deposition if harvesting occurs, yet harvest removals in Ontario are poorly quantified. In summary, long-term mass balance studies in south central Ontario demonstrate that soils are acidifying, with Ca losses being a major concern. Critical load models indicate that current S and N deposition exceeds the CL(S + N) for over 40 million hectares of forest land in Ontario with growing evidence linking poor forest health and productivity to exceedance of the CL. Decreasing base cation levels in lakes are predicted, with the decrease being greatest under harvesting scenarios. A decrease in soil base saturation can partly explain why many streams and lakes in the region have not recovered (pH, alkalinity) as expected from reductions in S deposition.

40

Chapter 11: Drought induced pulses of S from Canadian Shield wetlands: ∂

34S and ∂18O in SO42- and DOM

Sherry Schiff

The use of stable isotope ratios to determine the cycling and ultimate fate of elements in the environment has grown exponentially. Mass-dependent fractionation of lighter isotopes such as C, N, S, O and H in particular, has been well-studied. For example, variations in stable isotope ratios of 13C/12C and 15N/14N have been used to examine food sources, trophic structure and energy transfer in lakes while measurements of the fractionation of sulfur (i.e., 34S/32S) have been invaluable in elucidating biogeochemical fluxes and cycling in both terrestrial and aquatic environments.

After microbial SO42- reduction, the residual SO4

2- tends to be ‘heavier’ (i.e., ∂34S more positive) while the products, S2 and organic S, are ‘lighter’ (i.e., ∂34S less positive). Thus, in the equation: ∂

34S o/oo = (34S/32S)sample -1 * 1000 (11.1) (34S/32S)standard where (34S/32S)standard = 0 o/oo, for ∂34S < 0 o/oo there is less 34S whereas for ∂34S > 0 o/oo there is more 34S. We studied the changes in S and O ratios (∂34S and ∂18O) in SO4

2- and DOM in an upland (#47) and wetland (#50) stream, located in the Turkey Lakes watershed, Ontario. In samples collected between 1997 and 2002, ∂34S values in precipitation were ~4 o/oo, whereas ∂34S values in the streams ranged between ~+3 and +9 o/oo for a [SO4

2-] ~5 mg/L but were < o/oo for higher sulfate concentrations (>13 mg/L). There was a general exponential decline in ∂34S values as [SO4

2-]’s increased. Groundwater ∂34S values were very similar to precipitation values at all SO4

2- concentrations except for a few high values (∂34S = +10 to +24 o/oo) in deep groundwater at the wetland site. Figure 11a: Expected trends in ∂18O vs ∂34S in SO4

2-

41

The expected trends in ∂18O vs ∂34S in SO42- are shown in Figure 11a. The results from

∂18O and ∂34S measurements in the wetland stream, prior to and following a drought

event, support the theory that the pulse of sulfate comes from oxidation of S from previously reduced SO4

2- in the wetland. On the other hand, ∂34S values are relatively constant (~+4 to +5 o/oo) in the upland stream, in many groundwater samples and in precipitation, although in some wetland groundwaters, ∂34S reaches values of 15 o/oo. ∂

18O values in precipitation and in both the upland and wetland streams and groundwaters are highly variable (i.e. -2 to +17 o/oo).

Figure 11b: ∂34S values in dissolved organic matter. In the M-H area lakes, ∂34S values measured in dissolved organic matter (DOM) are often similar to those found in precipitation, with values at the wetland site (PC 1-08) tending to be slightly higher (Fig. 11b). In conclusion, ∂34S and ∂18O measurements in SO4

2- can be useful tools for determining sources and/or processes of S cycling. Sulfur is stored in wetlands as SO4

2-, which becomes oxidized to S; any small change in the water table is sufficient to pulse SO4

2-. This storage and release of SO4

2- alters the ∂34S values of stream SO42- and thus it is not

possible to compare ∂34S values in precipitation vs stream waters if wetlands are present.

42

Chapter 12: Acidity, Climate Change, UV, DOC and Photochemistry in Lakes

Lewis Molot

There are several reasons for studying DOC in lakes: 1) DOC regulates visible light penetration which affects epilimnetic (upper layer) mixing depth, photosynthesis and heat budgets, 2) chromophoric DOC acts as a sunscreen and sharply attenuates UV intensity, 3) DOC transports Fe and P (essential for metabolism) to and within lakes, 4) DOC binds metals and organic contaminants with varying affinities and can reduce toxicity, 5) DOC adds acidity and 6) DOC provides microbial substrate. This chapter will focus on how interactions (i.e., climate change, UV and acidity) influence DOC concentrations in lakes and thus affect UV exposure to organisms. It begins with a review of results from the Muskoka-Haliburton (M-H) mass balance study and then broadens to include experimental results. Using the basic mass balance approach, and assuming steady state and 1st order kinetics, the concentration of DOC in a lake is: [DOC] = Inputs (12.1)

q + ν

where q is the areal water discharge (m/yr) and ν is the mass transfer or loss coefficient which describes sediment and atmospheric losses (m/yr). Thus, the [DOC] in lakes is dependent on inputs, which are affected by runoff and % wetlands and thus by climate; q,

which is controlled by the lake shape and runoff, and ν, which is affected by UV radiation, lake acidity and respiration.

In seven M-H lakes, previous calculations ascertained that there was a net annual transfer of carbon to the sediments and the atmosphere. In acidified lakes (alkalinity < 20 µeq/L)

there was an accelerated loss to the atmosphere (νatmosphere = 2.8 ± 0.8 m/yr) and a lower

accumulation rate in the sediments (νsediment = 0.9 ± 0.1 m/yr) whereas for alkalinity > 20 µeq/L, sediment losses exceeded atmospheric losses (1.9 + 0.5 vs 1.0 + 0.9 m/yr, respectively). In laboratory experiments, the loss of DOC in stream waters incubated under solar radiation in UV-transparent containers, is described by the equation: d[DOC]/dt = -K [DOC] (12.2) where K is normalized for variation in broadband UV intensity, Iuv, such that: auv = K / Iuv (12.3) The relationship between auv and alkalinity is best described as a 2nd order decay constant for M-H stream waters (Figure 12a).

43

Figure 12a: Relationship between auv and alkalinity in Dorest stream waters.

The photo-Fenton pathway explains the increased DOC decay rate at low pH/alkalinity. Photo-reduction of organically bound ferric Fe+3 in near surface waters (via ligand-metal charge transfer) is followed by dissociation: Fe3+LDOC + UV => •LDOC + Fe2+ (12.4) where LDOC and •LDOC are organic ligand and organic ligand radicals, respectively. At pH > 6, Fe+2 is rapidly oxidized to Fe+3 by O2 but at pH << 6, oxidation by O2 is slow, allowing the production of hydroxyl radicals: Fe2+ + H2O2 � Fe+3 + OH* (12.5) Hydroxyl radical oxidation of DOC leading to loss of DOC and loss of colour (absorbance in the visible region) explains the increased clarity of acidic lake waters and is supported by experimental evidence. In one experiment, photo-bleaching of Dickie 5 stream water at pH 4.3 was significantly inhibited with the addition of 50 mM KI (a hydroxyl radical quencher). Similarly, in 1 mM NaF addition experiments, DOC loss was inhibited (i.e. K decreased) at pH 4 and 4.9, and at pH 4.3 there was increased fluorescence. Fluoride binds to Fe, forming a non-reactive iron-fluoride compound. Binding prevents hydroxyl radical formation and also frees chelation sites on DOC which in turn increases fluoresence from these sites. After 4 days of irradiation however, 1/3 of the Fe binding sites are lost and amorphous particulate Fe hydroxides form. Therefore, as predicted by the photo-Fenton pathway (eq’n 12.4 and 12.5), Fe accelerates loss of DOC at pH < 6. The loss of some binding sites leads to loss of some Fe to the

0

2

4

6

8

10

12

14

16

-50 0 50 100 150 200 250

Initial Alkalinity, ueq/L

auv

44

sediments but the relatively UV-insensitive DOC that remains in the lake limits the loss of Fe by complexation (i.e., free Fe+3 is not very soluble in the presence of oxygen). This

was evident in the 7 M-H lakes, where the mass transfer coefficient for Fe, νFe (m/yr), was found to decrease as [DOC] increased (Figure 12b).

Figure 12b: Mass transfer coefficient for Fe to the sediments, νFe, vs [DOC] (mean 1980-1988) Two approaches were then used to assess whether or not DOC provides adequate protection from UVB, especially in clear, acidic waters. First, the distribution of UVB-sensitive Daphnia sp was examined in 258 systems in Ontario and eastern North America. Daphnia were found to be present in clear, acidic systems with relatively high exposure rates, implying that the loss of similarly sensitive species in acidic systems is not likely due to increased direct UVB exposure unless waters are very shallow and very clear or missing species are considerably more sensitive to UVB. Second a comparison between the maximum depth (Zmax) to depth of 1% penetration of UV at 320 nm (Z320,1%) in approximately 1000 systems across Canada was made to determine the proportion of ‘optically clear’ systems (i.e. Zmax < Z320,1%) and the extent to which systems that are not clear may become so (‘at risk’) should they lose 50% of their DOM (‘at risk’). The proportion of ‘optically clear’ systems was found to be low (< 6%) across Canada with the exception of three ecozones (e.g., alpine and arctic tundra) where between 13% and 20% were optically clear. The proportion of systems ‘at risk’ is 0% in most regions and 5-9% in five regions from four ecozones. These results suggest that DOM levels are adequate to prevent large-scale, loss of sensitive species from direct exposure to elevated UVB in most regions of Canada.

4

6

8

10

12

14

16

18

0 1 2 3 4 5 6

D OC , mg / L

45

Chapter 13: Summary of OPG/NSERC Chair Results Peter Dillon

Although estimated emissions of SO2 by OPG increased by 40%, from 1993-96 (avg. 89,500 tonnes/yr; Fig. 1a) to 1998-2002 (avg. 149,300 tonnes/yr; Fig. 1a and 2c), deposition of S in south-central Ontario did not increase measurably as a result of these changes in S emissions (Fig. 3a). Similarly, we measured no increase in N (Fig. 3a) or Hg deposition, although in the latter case, the data available for comparison are relatively few. Long-term chemical recovery of lakes lagged behind decreases in S deposition; recovery remains at about half (measured by decrease in sulphate; Fig. 3b) to almost non-existent (measured by increase in pH; Fig. 5a) compared with the anticipated changes, and biological recovery has not occurred to any significant degree. The recovery in lakes is counteracted by drought cycles within the lakes’ watersheds that result from regional and global-scale climate phenomena (see Chapters 3, 5 and 6). The phenomena (the Southern Oscillation and the North Atlantic Oscillation) are affecting eastern North America, including our study area, very strongly (Table 3a) and are increasing in frequency with elevated levels of greenhouse gases believed to be the primary reason. Three models were used to assess the status of Ontario lakes with respect to their critical loads of acidifying substances. The Steady State Water Chemistry (SSWC) Model was used to estimate the critical load of acidity for approximately 1500 lakes in five counties/regions of Ontario and the exceedance (the amount by which deposition exceeds the critical load) calculated (Chapter 8). Based on current (1995-99) S deposition, 13% of the lakes have critical loads that are exceeded, with up to 22% in one county (Parry Sound). The critical load of acidity that will protect 95% of the lakes in this region is 34 meq/m2/yr, which is about 20% below the current deposition level. This method gives a non-conservative estimate of potential damage, i.e. minimum damage, and an upper estimate of the allowable or critical load, i.e. a best-case scenario. However, because forest harvesting results in the loss of base cations (Ca, Mg, Na, K) from the watershed, i.e. of acid neutralizing components, this represents another mechanism of ecosystem acidification. The number of lakes where the critical load of acidity is exceeded increased from 13% with no harvesting to 23% with wood only, 56% with stem only and 72% with whole-tree harvesting (Chapter 10). Since much of this region is mixed hardwood forest and is subject to periodic selective harvesting, the second case (56%) is most realistic. The First-order Acidity Balance (FAB) model allows the simultaneous calculation of the critical loads of both S and N (Chapter 8). Based on current S and N deposition, the critical load is exceeded in ~60% of the lakes (with no forest harvesting) in Ontario (Fig. 8c), vs. 13% based on the SSWC model. Critical load exceedance can be reduced by decreasing the S deposition, the N deposition or a combination of the two (Fig. 8d). Reducing the S deposition further would be more effective than reducing the N deposition; a 50% decrease in S deposition would lower the portion of lakes with exceedances to about 12%, while a 50% decrease in N deposition would lower this number to only 32%.

46

The dynamic model, MAGIC, was used to evaluate the potential effects of the proposed S emission reduction scenario based on Canada’s post-2000 Acid Rain Strategy and the proposed Clear Skies Legislation in the United States (Chapter 9). Reductions of 50% in S deposition by 2010 from 2000 baselines will result in lake water concentrations in 2050 recovering to levels predicted for the early 1900s. However, soils in the region will continue to acidify with Ca losses dominating depletion of the exchangeable pool. Long term mass balance measurements for Ca, Mg, K, S and N for forested watersheds clearly showed that large losses of Ca from the plant-available soil pool have occurred during this study period (Chapter 10). These losses (up to 40 % of the Ca present in soil in 1983; Fig. 10a) have resulted in a decrease in soil pH (soil acidification). Losses of Ca from soil will lead to further decreases in Ca levels in lakes, which will be particularly evident in harvested regions. Calcium concentrations in lakes have declined by up to 30% over the past 20 years (Fig. 7b), and predicted decreases in lake Ca levels are between ~20 – 70% of current levels, depending on the harvest intensity. Low Ca concentrations may, in the future, be limiting to biological recovery in lakes. In the future, the discontinuity between the increase in OPG’s S emissions and the lack of change in S deposition in south-central Ontario needs to be verified. The contribution of OPG to S (and N) deposition in Ontario also should be ascertained, largely through modeling efforts linked to field observations. Exceedances of critical loads are greatly affected by the assumptions that all of the models used world-wide make about nitrogen dynamics in catchments. Many of these assumptions may be wrong. We need to learn the long-term fate of N in forests and what factors influence release of N to aquatic systems. New approaches (e.g. isotope analysis) are now available that will aid in this. Our major study site now has no net biomass growth, thus mimicking an “old growth” forest. This makes it an ideal site for determination of how much carbon is accumulated in such systems and directly pertains to the fate of greenhouse gases, particularly CO2 and N2O. We need to determine whether trees can access Ca from sources other than the exchangeable pool. If there are other sources, then the critical load models need to be modified, and systems in Ontario can tolerate higher levels of acid deposition than believed previously. Although emission inventories are imprecise, OPG is almost certainly the biggest source of Hg in the province. The contribution of OPG to the Hg pool currently cycling in the Ontario environment is unknown; this is very relevant should remediation of ecosystems be considered. This can be addressed, at least in part, through isotope studies. The acceptable level of Hg deposition is also unknown at this time. In fact, we do not know if reduced Hg emissions will have any effect on Hg levels in fish or other biota. We believe that the model framework used in MAGIC in suitable for Hg, and can be used to estimate the relationships between Hg deposition and Hg in biota, e.g. fish. Other important questions which need to be addressed include:

• how long will lake recovery be delayed?

47

• will existing and/or planned future S and N emission targets be adequate to protect watersheds, including lakes, streams, forests and soils, or will additional decreases in S emissions be required?

• how widespread is soil acidification, and will it lead to loss of forests in Ontario? • how important are links between Hg cycling, acidification and climate change? • what will be the impact of soil acidification on trace metal cycling? • how long will it take before biological recovery will occur, and by how long will

it follow chemical recovery? • are there useful remediation methods that can be applied?

48

Alphabetical List of Attendees

Julian Aherne Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: 748-1011 ext. 5351 Ronald Bell Senior Air Policy Advisor Regional Air Issues Section Air Policy and Climate Change Branch Ontario Ministry of the Environment [email protected] Phone: (416) 314-4933 FAX: (416) 314-2979 Heather Broadbent Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: 748-1011 ext. 7348 Krista Chomicki University of Waterloo Department of Earth Sciences - CEIT 200 University Ave W Waterloo, ON N2L 3G1 [email protected] Phone: (519) 888-4567 ext. 7284 Peter Dillon Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: (705) 748-1011 ext. 7536 FAX: (705) 748-1569 Cathy Eimers Trent University

49

Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: 748-1011 ext. 7451 Hayla Evans Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: (705) 748-1011 ext. 7370 FAX: (705) 748-1569 Joe Findeis Ontario Ministry of the Environment Dorset Environmental Science Centre 1026 Bellwood Acres Rd P.O. Box 39 Dorset, ON P0A 1E0 [email protected] Phone: (705) 766-2030 FAX: (705) 766-2254 Martyn Futter Ontario Ministry of the Environment Dorset Environmental Science Centre 1026 Bellwood Acres Rd P.O. Box 39 Dorset, ON P0A 1E0 [email protected] Phone: (705) 766-1291 FAX: (705) 766-2254 Roland Hall University of Waterloo Department of Biology Waterloo, ON N2L 3G1 [email protected] Phone: (519) 888-4567 ext. 2450 FAX: (519) 746-0614 Andy Hoffer Advisor, Corporate Environment Ontario Power Generation

50

700 University Avenue (H19E2) Toronto, ON M5G 1X6 [email protected] Phone: (416) 592-6399 Bill Keller Ontario Ministry of the Environment Co-operative Freshwater Ecology Unit 1222 Ramsey Lake Road Sudbury, ON P3E 2C5 [email protected] Phone: (705) 671-3858 Dolly Kothawala Ontario Ministry of the Environment Dorset Environmental Science Centre 1026 Bellwood Acres Rd P.O. Box 39 Dorset, ON P0A 1E0 [email protected] Dean Jeffries Research Scientist, Aquatic Ecosystem Impacts Research Branch National Water Research Institute Environment Canada Canada Centre for Inland Waters 867 Lakeshore Road, P.O. Box 5050 Burlington, ON L7R 4A6 [email protected] Phone: (905) 336-4969 FAX: (905) 336-6430 Gerry McKenna Environmental Advisor Nanticoke GS Ontario Power Generation [email protected] Phone: (519) 587-2201 ext. 3703 David McLaughlin Supervisor, Phytotoxicology Investigations Unit (ESSD Liaison - Port Colborne Project) Biomonitoring Section Environmental Monitoring and Reporting Branch Ontario Ministry of the Environment

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125 Resources Rd, North Wing, 2nd Floor Toronto, ON M9P 3V6 [email protected] Phone: (416) 314-6266 Business Cell: (416) 723-0420 FAX: (416) 314-6270 Chris Metcalfe Dean of Research and Graduate Studies Trent University 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: (705) 748-1011 ext. 1272 FAX: (705) 748-1587 Brad Mills Ontario Ministry of the Environment Dorset Environmental Science Centre 1026 Bellwood Acres Rd P.O. Box 39 Dorset, ON P0A 1E0 [email protected] Phone: (705) 766-2440 FAX: (705) 766-2254 Lewis Molot York University Faculty of Environmental Studies 4700 Keele Street Toronto, ON M3J 1P3 [email protected] Tel: (416) 736-2100 ext. 22613 FAX: (416) 736-5679 Jessica Mueller Department of Geography and Environmental Studies Wilfrid Laurier University 75 University Ave. W. Waterloo, ON N2L 3C5 [email protected] Andrew Paterson Ontario Ministry of the Environment Dorset Environmental Science Centre P.O. Box 39

52

Dorset, ON P0A 1E0 [email protected] Phone: (705) 766-2951 Wolfgang Scheider Manager - Biomonitoring Section Environmental Monitoring and Reporting Branch Ontario Ministry of the Environment 125 Resources Rd, North Wing Toronto, ON M9P 3V6 [email protected] Phone: (416) 327-6535 Sherry Schiff University of Waterloo Waterloo Centre for Groundwater Research 200 University Avenue West, Room 222A Waterloo, ON N2L 3G1 [email protected] Phone: (519) 888-4567 ext. 2473 Kathryn Sentence University of Waterloo Waterloo Centre for Groundwater Research 200 University Avenue West, Room 222A Waterloo, ON N2L 3G1 [email protected] Keith Somers Dorset Site Manager - Biomonitoring Dorset Environmental Science Centre Ontario Ministry of the Environment 1026 Bellwood Acres Rd P.O. Box 39 Dorset, ON P0A 1E0 [email protected] Phone: (705) 766-2408 FAX: (705) 766-2254 Leonard Surges Manager, Sustainable Development and Product Policy Noranda Inc. / Falconbridge Limited Queen's Quay Terminal 207 Queen's Quay West, Suite 800 Toronto, ON M5J 1A7

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Temporary E-Mail: [email protected] [email protected] Phone: (416) 982-6900 FAX: (416) 982-3543 Shaun Watmough Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: (705) 748-1011 ext. 1647 Jianjun Yang Trent University Environmental and Resource Studies 1600 West Bank Drive Peterborough, ON K9J 7B8 [email protected] Phone: 748-1011 ext. 7348

opg/nserc research chair watershed biogeochemistry 2002 … · · 2005-11-20opg/nserc research...

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