I OOOR83105 ' IMPACT ASSESSMENT | WORK GROUP I I ' '•?" I __^_ 1 I I I I I I j I I • FINAL REPORT JANUARY 1983 I 1 I ------- I 1 WORK GROUP I Work Group Co-Chairmen G.E. Bangay, Canada C. Riordan, United States February 1983 IMPACT ASSESSMENT I '§ FINAL REPORT I I i i i I i i I i i i i Submitted to the Coordinating Committee in fulfillment of the requirements of the Memorandum of Intent on Transboundary Air Pollution signed by Canada and the United States. ------- t I I I I I I 1 I I I I I I I 1 I I I TO: Co-Chairmen Canada/United States Coordinating Committee FROM: Co-Chairmen Impact Assessment Work Group I We are pleased to submit our final report completing the Phase III activities of the Impact Assessment Work Group I. Although we have reached agreement on the majority of information and conclusions found in the report, there are a number of instances when Canadians and Americans could not reach agreement. These differences are confined to the aquatic section of the report (Section 3) and has required the preparation of separate summary statements in Section 1. Those portions of the text which represent a lack of Canada/United States consensus are typed in italics. This report completes our activities under the terms of reference contained in the Memorandum of Intent and as such represents the joint efforts by representatives of our two countries to provide information to the negotiators. Sincerely yours, G.E. Bangay Courtney Riordan Canadian Co-Chairman United States Co-Chairman ------- TABLE OF CONTENTS t I I I I ------- 1 1 1 1 1 1 1 1 1 1 1 ^p 1 1 1 1 1 1 1 TABLE OF CONTENTS LIST OF FIGURES LIST OF TABLES PREFACE ACKNOWLEDGEMENTS SECTION 1 SUMMARY 1 . 1 INTRODUCTION 1.2 AQUATIC ECOSYSTEM EFFECTS - Canada AQUATIC ECOSYSTEM EFFECTS - United States 1.3 TERRESTRIAL ECOSYSTEM IMPACTS 1.3.1 Effects on Vegetation 1.3.1.1 Sulphur Dioxide 1.3.1.2 Ozone 1.3.1.3 Acidic Deposition 1.3.2 Effects on Terrestrial Wildlife 1.3.3 Effects on Soil 1.3.4 Sensitivity Assessment 1.4 HUMAN HEALTH AND VISIBILITY 1.4.1 Health 1.4.2 Visibility 1.5 MAN-MADE STRUCTURES 1.6 METHODOLOGIES FOR ESTIMATING ECONOMIC BENEFITS OF CONTROL Page Number XV xxiii xxxi xxxv ii 1-1 1-1 1-7 1-13 1-13 1-13 1-14 1-14 1-14 1-14 1-16 1-16 1-16 1-18 1-18 1-20 ------- TABLE OF CONTENTS (continued) 1.7 NATURAL AND MATERIAL RESOURCE INVENTORY 1.7.1 Introduction 1.7.2 Aquatic - United States 1.7.3 Aquatic - Canada 1.7.4 Agriculture - United States 1.7.5 Agriculture - Canada 1.7.6 Forests - United States 1.7.7 Forests - Canada 1.7.8 Man-Made Materials - United States 1.7.9 Man-Made Materials - Canada 1.8 LIMING 1.8.1 Aquatic Systems 1.8.2 Terrestrial Liming 1.8.3 Drinking Water Supply SECTION 2 INTRODUCTION 2.1 THE EXTENT OF RESOURCES EXPOSED TO ACIDIC DEPOSITION AND POTENTIAL FOR LARGE-SCALE EFFECTS 2.1.1 Methods of Measuring Effects 2.1.2 Hydrologic Cycle 2.2 ATMOSPHERIC INPUT, TRANSPORT AND DEPOSITION OF POLLUTANTS 2.2.1 Emissions of Pollutants to the Atmosphere 2.2.2 Atmospheric Transport of Pollutants 2.2.3 Atmospheric Removal Processes 2.2.4 Alteration of Precipitation Quality ii Page Number 1-21 1-21 1-21 1-22 1-22 1-22 1-22 1-22 1-23 1-23 1-23 1-23 1-24 1-24 2-1 2-2 2-4 2-5 2-5 2-5 2-9 2-13 1 1 1 1 1 1 1 1 1 1 1 1 1 0 1 1 1 1 1 ------- 1 1 I I 1 1 I 1 I 1 1 1 1 1 1 V 1 1 1 TABLE OF CONTENTS (continued) 2.3 REFERENCES SECTION 3 AQUATIC IMPACTS 3 . 1 INTRODUCTION 3.2 ELEMENT FLUXES AND GEOCHEMICAL ALTERATIONS OF WATERSHEDS 3.2.1 Hydrogen Ion (Acid) 3.2.2 Nitrate and Ammonium Ions 3.2.3 Sulphate 3.2.4 Aluminum and Other Metals 3.3 NATURAL ORGANIC ACIDS IN SOFT WATERS 3.4 CATION AND ANION BUDGETS 3.4.1 Element Budgets at Hubbard Brook, New Hampshire 3.4.2 Element Budgets in Canada 3.4.3 Effects of Forest Manipulation or Other Land Use Practices on Watershed Outputs 3.5 AQUATIC ECOSYSTEMS SENSITIVE TO ACIDIC DEPOSITION 3.5.1 Mapping of Watershed Sensitivity for Eastern North America 3.5.1.1 Eastern Canada 3.5.1.2 Eastern United States 3.5.2 Aquatic - Terrestrial Relationships 3.5.3 Geochemical Changes Due to Acidic Precipitation 3.6 ALTERATIONS OF SURFACE WATER QUALITY 3.6.1 Present Chemistry of Aquatic Systems 3.6.1.1 Saskatchewan i Page Number 2-27 3-1 3-1 3-2 3-7 3-7 3-11 3-15 3-17 3-19 3-24 3-27 3-30 3-30 3-35 3-40 3-42 3-45 3-46 3-47 3-59 ------- I TABLE OF CONTENTS (continued) • 1 I I I Page Number 3.6.1.2 Ontario 3-59 3.6.1.3 Quebec 3-65 3.6.1.4 Atlantic Provinces 3-68 3.6.1.5 United States 3-74 3.6.2 Time Trends in Surface Water Chemistry 3-74 3.6.3 Time Trends in Representative Areas 3-76 • 3.6.3.1 Time Trends in Nova Scotia and Newfoundland 3-76 3.6.3.2 Historical Trends in Northern Wisconsin 3-79 3.6.3.3 Historical Trends in New York State 3-83 3.6.3.4 pH Changes in Maine and New England 3-86 3.6.7 pH Declines During Spring Runoff in Ontario and Quebec 3.6.8 pH Depression During Flushing Events 3.7 ALTERATION OF BIOTIC COMPONENTS RECEIVING ACIDIC DEPOSITION 3-100 3.7.1 Effects on Algae 3-104 3.7.2 Effects on Aquatic Macrophytes 3-105 3.7.3 Effects on Zooplankton 3-106 I 1 I I 3.6.3.5 Time Trend in New Jersey 3-88 3.6.4 Paleolimnological Evidence for Recent H Acidification and Metal Deposition 3-90 I 3.6.5 Seasonal and Episodic pH Depression 3-92 • 3.6.6 Seasonal pH Depression in Northern Minnesota 3-92 • and Quebec 3-94 \| Depression During tiusning events M in West Virginia 3-94 • I ,„ I ------- I I I I I I I 1 I 1 I I I I I I I 1 I TABLE OF CONTENTS (continued) 3.8 3.9 3.7.4 Effects on Aquatic Macroinvertebrates 3.7.5 Effects on Bacteria and Fungi 3.7.6 Effects on Amphibians 3.7.7 Effects of Low pH on Fish 3.7.8 Effects of Aluminum and Other Metals on Fish 3.7.9 Accumulation of Metals in Fish 3.7.9.1 Mercury 3.7.9.2 Lead 3.7.9.3 Cadmium 3.7.9.4 Aluminum and Manganese 3.7.10 Effects on Fisheries in Canada and the United States 3.7.10.1 Adirondack Region of New York 3.7.10.2 Ontario 3.7.10.3 Quebec 3.7.10.4 Nova Scotia 3.7.10.5 Scandinavia 3.7.11 Response to Artificial Acidification 3.7.12 Effects of Acidic Deposition on Birds and Mammals CONCERNS FOR IRREVERSIBLE EFFECTS 3.8.1 Loss of Genetically Unique Fish Stocks 3.8.2 Depletion of Acid Neutralizing Capacity 3.8.3 Soil Cation and Nutrient Depletion ATMOSPHERIC SULPHATE LOADS AND THEIR RELATIONSHIP TO AQUATIC ECOSYSTEMS Page Number 3-107 3-108 3-109 3-112 3-115 3-123 3-123 3-123 3-126 3-126 3-126 3-126 3-129 3-132 3-133 3-138 3-139 3-141 3-144 3-144 3-146 3-146 3-146 ------- VI I TABLE OF CONTENTS (continued) " Page M Number 3.9.1 The Relative Significance of Sulphur and • Nitrogen Deposition to Acidification of Surface Waters 3-148 3.9.2 Data and Methods for Associating Deposition • Rates with Aquatic Effects 3-151 3.9.2.1 Empirical Observations 3-153 | Saskatchewan Shield Lakes 3-153 ~ Experimental Lakes Area, Ontario 3-154 * Algoma, Ontario 3-155 M Muskoka - Haliburton, Ontario 3-157 Laurentide Park, Quebec 3-160 | Nova Scotia 3-164 Boundary Waters Canoe Area and • Voyageurs National Park, Minnesota 3-167 Adirondack Mountains of New York 3-171 m The Hubbard Brook Ecosystem, New Hampshire 3-174 Maine and New England 3-177 • Summary of Empirical Observations 3-178 • 3.9.2.2 Short-term or Episodic Effects 3-181 3.9.2.3 Sensitivity Mapping and Extrapolation • •f~^t fi •§- V\ a >• A •*• .-•»«-»£•> s\f IT o c« <- *^ i—1-» f* ** Y^ *\ A f\ *\,~ 1 ft /i ^" I I I I to Other Areas of Eastern Canada 3-184 Terrestrial 3-184 Terrain characteristics of Three Specific Study Areas 3-185 Results of Terrain Extrapolation 3-191 ------- I I I I I I I _ 3.9.3.3 Summary 3-212 vii I I I 1 I I I I I I TABLE OF CONTENTS (continued) Page Number Aquatic 3-192 Possible Magnitude of Effects 3-197 3.9.3 Use of Acidification Models 3-198 3.9.3.1 The "Predictor Nomograph" of Henriksen 3-199 3.9.3.2 Cation Denudation Rate Model (CDR) 3-206 3.9.4 Summary of Empirical Observation and Modelling 3-212 3.10 CRITICAL RESEARCH TOPICS 3-215 3.10.1 Element Fluxes and Geochemical Alterations of Watersheds 3-216 3.10.2 Alterations of Surface Water Quality 3-216 3.10.3 Alteration of Biotic Components 3-217 3.10.4 Irreversible Impacts 3-218 3.10.5 Target Loadings and Model Validation 3-219 • 3.10.5.1 Long-Term Data Collection and Monitoring 3-220 3.11 REFERENCES 3-222 SECTION 4 TERRESTRIAL IMPACTS 4.1 INTRODUCTION 4-1 4.2 EFFECTS ON VEGETATION 4-2 4.2.1 Sulphur Dioxide (S02) 4-2 4.2.1.1 Introduction 4-2 4.2.1.2 Regional Doses of S02 4-3 ------- TABLE OF CONTENTS (continued) 4.2.1.3 S02 Effects to Agricultural Crops 4.2.1.4 S02 Effects to Forest Vegetation 4.2.1.5 S02 Effects to Natural Ecosystems 4.2.2 Ozone (03) 4.2.2.1 03 Effects to Agricultural Crops 4.2.2.2 03 Effects to Forest Vegetation 4.2.3 Acidic Deposition 4.2.3.1 Acidic Deposition Effects to Agricultural Crops 4.2.3.2 Acidic Deposition Effects to Forest Vegetation 4.2.4 Pollutant Combinations 4.2.4.1 S02 - 03 Effects 4.2.4.2 S02 - N02 Effects 4.2.4.3 S02 - 03~Acidic Deposition Effects 4.3 EFFECTS OF ACIDIC DEPOSITION ON TERRESTRIAL WILDLIFE 4.4 EFFECTS ON SOIL 4.4.1 Effects on Soil pH and Acidity 4.4.2 Impact on Mobile Anion Availability and Base Leaching 4.4.3 Influence of Soil Biota and Decomposition/ Mineralization Activities 4.4.4 Influences on Availability of Phosphorus 4.4.5 Effects on Trace Element and Heavy Metal Mobilization and Toxicity viii Page Number 4-3 4-14 4-16 4-25 4-26 4-38 4-39 4-40 4-42 4-46 4-46 4-50 4-50 4-51 4-53 4-54 4-57 4-60 4-61 4-61 1 1 1 1 1 1 1 I 1 1 1 I I 1 1 1 1 1 1 ------- I 1 1 1 1 1 •• I 1 1 ^m 1 1 1 w 1 1 1 1 I 1 1 TABLE OF CONTENTS (continued) 4.5 SENSITIVITY ASSESSMENT 4.5.1 Terrestrial Sensitivity Interpretations 4.5.2 Terrestrial Mapping for Eastern North America 4.5.2.1 Eastern United States 4.5.2.2 Eastern Canada 4.6 RESEARCH NEEDS 4 . 7 CONCLUSIONS 4.8 REFERENCES SECTION 5 HEALTH AND VISIBILITY 5 . 1 HEALTH 5.1.1 Contamination of Edible Fish 5.1.2 Contamination of Drinking Water 5.1.3 Drinking Water From Cisterns 5.1.4 Recreational Activities in Acidified Water 5.1.5 Direct Effects: Inhalation of Key Substances Related to Long Range Transport of Air Pollutants 5.1.6 Sensitive Areas and Populations at Risk - Health 5.1.7 Research Needs 5.2 VISIBILITY 5.2.1 Categories and Extent of Perceived Effects 5.2.2 Evaluation of Visibility 5.2.2.1 Aesthetic Effects 5.2.2.2 Transportation Effects Page Number 4-64 4-67 4-71 4-73 4-78 4-82 4-84 4-86 5-1 5-1 5-6 5-8 5-8 5-10 5-14 5-14 5-15 5-15 5-23 5-23 5-27 ix ------- TABLE OF CONTENTS (continued) 5.3 5.2.3 Mechanisms and Quantitative Relationships 5.2.4 Sensitive Areas and Populations 5.2.5 Data Needs/Research Requirements REFERENCES SECTION 6 EFFECTS ON MAN-MADE STRUCTURES 6.1 INTRODUCTION 6.2 OVERVIEW 6.3 MECHANISMS AND ASSESSMENT OF EFFECTS 6.3.1 Factors Influencing Deposition 6.3.2 Effects of Sulphur Dioxide Pollutant/ Material Interactions 6.3.2.1 Zinc 6.3.2.2 Steels 6.3.2.3 Copper and Copper Alloys 6.3.2.4 Aluminum 6.3.2.5 Paints 6.3.2.6 Elastomers 6.3.2.7 Masonry 6.3.3 Effect of Nitrogen Dioxide and Ozone Pollutant/Material Interactions 6.3.3.1 Metals 6.3.3.2 Masonry 6.3.3.3 Paints 6.3.3.4 Elastomers X Page Number 5-28 5-34 5-34 5-37 6-1 6-2 6-3 6-3 6-6 6-6 6-7 6-9 6-10 6-10 6-11 6-11 6-15 6-15 6-16 6-16 6-16 I 1 1 1 1 1 1 1 1 1 1 1 1 1 I 1 I 1 1 ------- 1 1 • 1 1 mm 1 1 1 1 1 1 1 1 1 1 1 1 1 TABLE OF CONTENTS (continued) 6.3.4 Effect of Ammonia Pollutant /Material Interactions 6.3.5 Effect of Particulate Pollutant /Material Interactions 6.4 IMPLICATIONS OF TRENDS AND EPISODICITY 6.5 DISTRIBUTION OF MATERIALS AT RISK 6.6 DATA NEEDS AND RESEARCH REQUIREMENTS 6.7 METHODOLOGIES 6.8 ASSESSMENT OF ECONOMIC DAMAGE 6.9 REFERENCES SECTION 7 THE FEASIBILITY OF ESTIMATING THE ECONOMIC BENEFITS OF CONTROLLING THE TRANSBOUNDARY MOVEMENT OF AIR POLLUTANTS 7.1 INTRODUCTION 7.1.1 Purpose 7.1.2 Background 7.1.3 Emission-Benefit Relationship 7.1.4 Efficiency and Equity Considerations 7.2 BENEFITS: CONCEPTUAL APPROACHES 7.2.1 Primary Benefits 7.2.1.1 Market Approach 7.2.1.2 Imputed Market Approach 7.2.1.3 Nonmarket Approach 7.2.2 Secondary Benefits Page Number 6-17 6-17 6-18 6-18 6-19 6-20 6-24 6-27 7-1 7-1 7-2 7-4 7-6 7-9 7-10 7-10 7-11 7-12 7-13 xi ------- xii TABLE OF CONTENTS (continued) • Page | Number 7.3 BENEFIT ESTIMATION TECHNIQUES 7-14 I 7.3.1 Aquatic 7-14 7.3.1.1 Recreational Fishery 7-14 I 7.3.1.2 Commercial Fishery 7-19 m 7.3.1.3 Aquatic Ecosystem 7-19 7.3.2 Terrestrial 7-20 I 7.3.2.1 Agriculture 7-20 7.3.2.2 Forestry 7-20 • 7.3.3.3 Ecosystem 7-20 M 7.3.3 Water Supply 7-21 * 7.3.4 Effects on Buildings and Structures 7-21 • 7.3.5 Human Health 7-22 7.3.5.1 Mortality 7-22 | 7.3.5.2 Morbidity 7-23 g 7.3.6 Visibility 7-24 * 7.3.7 Summary 7-26 I 7.4 QUALIFICATIONS, CONCLUSIONS AND RECOMMENDATIONS 7-26 7.4.1 Qualifications 7-26 | 7.4.1.1 Dose-Response Relationship 7-28 7.4.1.2 Inclusion of All Values 7-28 Nothing Feature 7-28 7.4.2 Conclusions and Recommendations 7-30 7.5 REFERENCES 7-32 APPENDIX - REVIEW OF RELEVANT ECONOMIC CONCEPTS 7-34 I 7.4.1.3 Irreversibilities and the All or • I I ------- 1 1 VI 1 1 ^^ 1 1 1 1 1 1 1 1 w 1 1 1 1 1 1 1 TABLE OF CONTENTS (continued) SECTION 8 NATURAL MATERIAL RESOURCES INVENTORY 8.1 INTRODUCTION 8.2 AQUATIC ECOSYSTEM 8.2.1 U.S. Aquatic Resources 8.2.2 Canadian Aquatic Resources 8.3 AGRICULTURAL RESOURCES 8.3.1 U.S. Agricultural Resources 8.3.2 Canadian Agricultural Resources 8.4 FOREST RESOURCES 8.4.1 U.S. Forest Resources 8.4.2 Canadian Forest Resources 8.5 MAN-MADE STRUCTURES 8.5.1 U.S. Historic Inventory 8.5.2 Canadian Historic Inventory 8.6 REFERENCES APPENDIX-TABLES SECTION 9 LIMING 9 . 1 INTRODUCTION 9.2 AQUATIC 9.2.1 Liming as a Mitigative Measure 9.2.2 Liming Programs 9.2.2.1 Sweden 9.2.2.2 Norway 9.2.2.3 United States xi Page Number 8-1 8-6 8-6 8-7 8-12 8-12 8-13 8-22 8-22 8-25 8-25 8-29 8-30 8-34 8-37 9-1 9-1 9-1 9-2 9-2 9-3 9-4 ------- 9.3 9.4 9.5 9.6 TABLE OF CONTENTS (continued) 9.2.2.4 Ontario, Canada 9.2.3 Economic Aspects of Lake Liming 9.2.3.1 Costs in Sweden 9.2.3.2 Costs in Norway 9.2.3.3 Costs in New York State 9.2.3.4 Costs in Canada 9.2.4 Technical Evaluations Necessary in Liming Programs TERRESTRIAL LIMING 9.3.1 The Application of Lime to Agricultural Soils 9.3.2 Economics of Agricultural Liming 9.3.3 Forest Liming 9.3.4 Terrestrial Liming Summary DRINKING WATER SUPPLY COSTS OF CORROSION CONTROL REFERENCES xiv Page Number 9-4 9-5 9-5 9-5 9-5 9-6 9-6 9-6 9-7 9-9 9-9 9-12 9-12 9-12 9-15 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- xv I I I I I 2-4 Precipitation amount-weighted mean annual 2-14 pH in North America. I I I I I LIST OF FIGURES Figure Page Number Number 2-1 Regions of North America containing lakes 2-3 sensitive to acidification by acidic deposition based on bedrock geology. 2-2 Wind patterns for North America based on 2-10 a + b surface stream-lines for January and July. 2-3 Seasonal precipitation patterns for North 2-11 a + b America. 2-5a Precipitation amount-weighted mean H+ 2-15 concentration. 2-5b Precipitation amount-weighted mean H+ 2-16 deposition. 2-6a Precipitation amount-weighted mean SO" 2-19 concentration. I I I I I 2-8a Precipitation amount-weighted mean N03~ 2-23 • concentration. 2-8b Precipitation amount-weighted mean N03~ 2-24 • deposition. 2-9 Percent of normal precipitation in North 2-26 America in 1980. 2-6b Precipitation amount-weighted mean SO" 2-20 deposition. 2-1 a. Precipitation amount-weighted mean NH4+ 2-21 concentration. 2-7b Precipitation amount-weighted mean NH4+ 2-22 deposition. 3-1 Relationship between pH and the relative 3-5 proportions of inorganic carbon species. 3-2 Simplified nitrogen cycle showing chemical 3-8 changes caused by plant and soil processes. 3-3 Simplified sulphur cycle showing chemical 3-10 changes caused by plant and soil processes. 3-4 Aqueous aluminum in equilibrium with gibbsite. 3-13 ------- LIST OF FIGURES (continued) Figure Number 3-5 3-6 3-7 3-8 3-9 3-10 3-11 3-12 3-13 3-14 3-15 3-16 3-17 3-18 Relationship of observed stream concentrations of aluminum to the pH of surface water. Schematic representation of the hydrogen ion cycle . Percent of ionic composition of precipitation for the Hubbard Brook Experimental Forest during 1964 to 1977. Hydrogen ion budget for Hubbard Brook Experimental Forest. Potential of soils and bedrock to reduce the acidity of incoming atmospheric deposition for eastern Canada. Potential of soils and bedrock to reduce the acidity of incoming atmospheric deposition for eastern United States. Total concentration of calcium plus magnesium with respect to alkalinity for lakes in Canada. for [Ca2+ + Mg2+ - alkalinity] vs. lakes in Canada. Hydrographic Regions of Quebec. Sulphate versus [calcium + magnesium - alkalinity] for lakes on the Precambrian Shield in Quebec. Mean and range of sulphate concentrations in Canadian lakes. Mean and range of basin specific yield of excess sulphate compared with atmospheric excess sulphate deposition in precipitation. Areal distribution of sulphate concentrations in Quebec lakes, summer 1980. Relationship between alkalinity and calcium + magnesium for northern Saskatchewan lakes. xvi Page Number 3-14 3-18 3-20 3-23 See map folio See map folio 3-48 3-49 3-51 3-52 3-53 3-56 3-57 3-60 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 Figure Number 3-19 3-20 3-21 3-22 3-23 3-24 3-25 3-26 3-27 3-28 3-29 3-30 3-31 3-32 XVI1 LIST OF FIGURES (continued) Distribution of lakes sampled in Ontario Ministry of the Environment 1981 and 1982 surveys . Mean summer epilimnetic alkalinity, mean summer epilimnetic pH and minimum surface water pH in spring for 16 lakes in Muskoka- Haliburton (1976-1980). Minimum pH values of 57 headwater streams in Muskoka-Haliburton, 1976-80. Calcite saturation indices for 181 lakes in southern Quebec, summer 1980. pH Values for Quebec lakes, summer 1980. values for Quebec lakes, summer 1980. Distribution of calcite saturation index values for New Brunswick, Prince Edward Island and Nova Scotia. Distribution of calcite saturation index values for Newfoundland. Distribution of calcite saturation index values for Labrador. Distribution of surface water alkalinities for the United States. Surface water alkalinity of New England states . Annual changes in median pH and mean discharge-weighted excess SO^" for the St. Mary's and Medway Rivers, Nova Scotia, and the Isle aux Morts and Rocky Rivers, Newfoundland. Geographic distribution of pH levels measured in Adirondack lakes higher than 610 metres elevation, June 24-27, 1975. Frequency distribution of pH and fish population status for 40 high elevation lakes surveyed in the 1930s and again in 1975. Page Number 3-62 3-63 3-64 3-66 3-67 3-70 3-71 3-72 3-73 See map folio 3-75 3-80 3-84 3-85 ------- LIST OF FIGURES (continued) Figure Number 3-33 3-34 3-35 3-36 3-37 3-38 3-39 3-40 3-41 3-42 3-43 3-44 3-45 New Jersey stream pH, 1958-1979, Oyster Creek and McDonalds Branch. Profiles of the lead concentration in four sediment cores from Jerry Lake, Muskoka- Haliburton. Discharge, hydrogen ion load per unit area, pH, and depth of precipitation for each day that: a precipitation event occurred for Harp Lake No. 4. Hydrogen ion content of streams draining Red Chalk Lake watersheds No.3 and No.4 (Muskoka- Haliburton). Mean daily pH for the Shavers Fork River at Bemis, West Virginia and precipitation event pH and accumulation Arborvale, West Virginia. Relative number of taxa of the major taxonomic groups as a function of pH. Generalized response of aquatic organisms to low pH. Age composition of yellow perch (Perca flavescens) captured in Patten Lake, Ontario, pH 4.1. Changes in the age composition of the white sucker (Catostomus commersoni) in George Lake, Ontario. Percent survival of brook trout fry plotted as a function of time in treatment waters at pH level 5.2. Brook trout survival (arcsin transformation) as a function of total aluminum concentration at each pH level. Mercury concentrations in yearling yellow perch vs. epilimnetic pH for selected lakes in Ontario. Age composition of the white sucker population of three lakes in the Muskoka-Haliburton Region of Ontario. xviii Page Number 3-89 3-93 3-97 3-98 3-101 3-102 3-103 3-116 3-117 3-119 3-120 3-124 3-131 1 1 I 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- I I I I I I I I I I I I I I I I I I I xix LIST OF FIGURES (continued) Figure Page Number Number 3-46 Atlantic salmon angling data since 1936 3-134 normalized for comparison between high and low pH rivers. 3-47 Angling records for six Nova Scotia Atlantic 3-136 coast rivers with mean annual pHs 5.0. 3-48 Atlantic salmon rivers of the Maritimes 3-137 divided into 4 pH categories based on significance to salmon reproduction. 3-49 Distribution of alkalinity values for lakes 3-195 in six regions of Ontario. 3-50 Cumulative distribution of alkalinity 3-196 values for lakes in five regions of Ontario. 3-51 Nomograph to predict the pH of lakes given the 3-201 sum of nonmarine calcium and magnesium concentrations (or nonmarine calcium concentration only) and the nonmarine sulphate concentrations in lake water (or the weighted- average hydrogen ion concentration in precipitation). 3-52 The model plot-pH predicted for consideration 3-207 of the sum of cations and sulphate. 3-53 Cation Denudation Rate Model applied to rivers 3-209 of Nova Scotia and Newfoundland. 3-54 Relation of excess sulphate and cation 3-210 concentration for pH 5.3 and 5.8 for basin runoff of 30, 50 and 100 cm/yr. 4-1 Sulphur dioxide emissions in eastern North 4-4 America. 4-2 Geographic distribution of monthly arithmetic 4-5 means for S02. 4-3 Conceptual model of factors involved in air 4-9 pollution effects (dose-response) on vegetation. 4-4 Regression of yield response vs. transformed 4-13 dose for controlled exposures using field chambers. ------- XX LIST OF FIGURES (continued) Figure Page Number Number of mercury in ecosystems. visibilities for North America. a 1950-54 5-17 b 1960-64 5-18 c 1970-74 5-19 d 1976-80 5-20 7-4 Change in demand due to visibility improvement. 7-25 I 4-5 Effects on base cation loss, soil acidification 4-69 and Al^+ solubilization for nonsulphate- adsorbing soils. B 4-6 Effects on base cation loss, soil acidification 4-72 and Al+ solubilization for sulphate- • adsorbing soils. • 4-7 Soil characteristics of eastern United See map ^ States. folio • 5-1 Varying effects of lake pH on the distribution 5-4 I 5-2 Mercury in yearling yellow perch and 5-5 epilimnetic pH relationships. M 5-3 Seasonal and spatial distribution of long- term trends in extinction weighted airport • I I 5-4 Median 1974-76 visibilities (miles) and 5-22 visibility isopleths for suburban/nonurban airports. • 5-5 Visual range as a function of fine mass con- 5-33 centration. • 5-6 Summertime fine particle levels for non- 5-35 urban sites. _ 7-1 Conceptual relationship between emissions and 7-5 economic effects. 7-2 Variation in effects due to different emission- 7-5 • deposition relationships. 7-3 Change in demand due to water quality 7-16 • improvement. I I I ------- xxi I LIST OF FIGURES (continued) • Figure Page Number Number 7-5 Measure of consumer surplus. 7-35 I I I I I I I 7-6 Change in consumer surplus. 7-35 7-7 Change in demand due to visibility improvement. 7-38 7-8 Compensating and equivalent surplus. 7-38 7-9 Producer's surplus. 7-39 7-10 Change in producer's surplus due to change 7-39 in supply. 7-11 Hypothetical change in producer's surplus 7-41 due to reduction in LRTAP deposition. 8-la Annual sulphate deposition regime for 8-3 eastern United States. 8-lb Annual sulphate deposition regime for 8-4 eastern Canada. 8-2 Conceptual scheme for identifying resources 8-5 potentially at risk. 18-3 Forest regions in eastern Canada and acidic 8-26 deposition. I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I xxiii LIST OF TABLES Table Page Numb e r Number 2-1 Current emissions in the U.S. and Canada. 2-6 2-2 Air emmissions from a typical 1000 MW coal 2-7 fired system plant. 2-3 Summary of global sources, annual emmission, 2-8 background concentration, major sinks, and residence time of atmospheric gaseous pollutants. 2-4 Concentrations in bulk deposition and total 2-18 bulk deposition of four ions at four calibrated watershed studies. 2-5 Conversion factors for concentration 2-25 and deposition units. 3-1 The retention of nitrate, ammonium ion and total 3-9 nitrogen by forested watersheds. 3-2 Annual budgets of bulk precipitation inputs 3-22 and stream-water outputs of dissolved substances for undisturbed watersheds of the Hubbard-Brook Experiment Forest. 3-3 Description of watersheds in Canada utilized 3-25 for mass balance studies. 3-4 Net export of major ions for calibrated 3-26 watersheds in Canada. 3-5 Summary of total cation release, hydrogen ion 3-28 production, and the cation release ratio for three manipulated watershed studies. 3-6 Terrestrial factors and associated criteria 3-31 for determining the potential of terrestrial ecosystems to reduce the acidity of atmospheric deposition. 3-7 Terrestrial factors and associated data bases 3-31 utilized for the interpretation of the potential to reduce acidity of atmospheric deposition. 3-8 Terrestrial characteristics of areas having 3-36 high, moderate and low potential to reduce acidity for eastern Canada. ------- LIST OF TABLES (continued) Table Number 3-9 3-10 3-11 3-12 3-13 3-14 3-15 3-16 3-17 3-18 3-19 3-20 3-21 3-22 3-23 Characteristics of map classes for the eastern United States as to the potential to reduce acidity of acidic deposition. Formal names, locations, lake data sources and the laboratories that analyzed data. Regional water chemistry survey results for surface water pH distribution. Summary of the percentage of lakes and streams in each alkalinity class by county or district for Ontario. Some statistics on the ratios of for waters of Quebec. Mean concentrations of ions in the water of four Nova Scotia rivers. Apparent changes in summer pH values in lakes in Nova Scotia and southern New Brunswick during the period 1940-79. pH of streams in Muskoka-Haliburton, Ontario. Monthly discharge hydrogen ion loads and percent of annual total. Spring/summer comparison of average parameter values. Susceptibility of breeding habitat to pH depression for amphibians whose range overlaps areas receiving acidic deposition. Approximate pH at which fish in the LaCloche Mountain Lakes stopped reproduction. Metals residues in yearling yellow perch. Distribution and frequency of occurrence of fish species collected during surveys of Adirondacks Lakes. Summary of biological effects observed. xxiv Page Number 3-43 3-54 3-58 3-61 3-69 3-78 3-82 3-95 3-96 3-99 3-110 3-114 3-125 3-128 3-140 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 Table Number 3-24 3-25 3-26 3-27 3-28 3-29 3-30 3-31 3-32 3-33 4- la 4-lb 4-2 XXV LIST OF TABLES (continued) Avian and mammalian species most likely to be influenced by a reduction in food resources due to acidic deposition. Mean and range of pH values for 21 headwater streams. Periodic pH depressions observed in streams and lakes with different sulphate loadings and corresponding biological effects. Coverage of terrain types in eastern Canada interpreted for their potential to reduce acidity. Summary of terrain types and potential to reduce acidity for all of eastern Canada. Terrain characteristics of watersheds containing the detailed study areas of eastern Canada. Average annual or spring total inflection point alkalinities for nine lakes in the Muskoka-Haliburton watershed study area. Distribution of 141 lake alkalinities, grouped by sensitivity classes, in various terrain types. Calculation of wet sulphate loadings consistent with pH 5.3 or greater in lakes with initial calcium concentrations of 50 yeq/L or greater. Acidification sensitivity of surface waters related to sulphate loading for two pH objectives and three runoff values. Summary of crop effects from S02 exposure in field closed chambers. Summary of crop effects from S02 exposure in field zonal air pollution systems. Sulphur dioxide concentration causing visible injury to various sensitivity grouping of vegetation. Page Number 3-145 3-150 3-183 3-186 3-187 3-190 3-193 3-194 3-205 3-211 4-10 4-11 4-15 ------- LIST OF TABLES (continued) Table Number 4-3 Summary of studies reporting results of SC>2 exposures under laboratory conditions for various tree species. 4-4 Effects of long-term controlled ozone exposures on growth, yield and foliar injury to selected plants. 4-5 The number in 1980 and 1981 that ozone concentrations exceeded the USEPA standard of 0.12 ppm along the U.S./Canada border. 4-6 Summary of growing season: daylight ozone trends in rural locations in southern Ontario, 1976-81. 4-7 Summary of growing season: daylight ozone trends in urban locations in eastern Canada, 1976-80. 4-8 Repesentative tolerance limits to simulated acid precipitation. 4-9 Effects of mixtures of S02 and 03 on plants. 4-10 Acidity related reactions influencing availability of several elements. 4-11 Terrestrial factors and associated criteria limits to assess forest productivity sensitivity. 4-12 The sensitivity of various soil categories to acidic deposition. 4-13 Theoretical sensitivities of terrestrial ecosystems to acidic deposition effects. 4-14 Terrestrial factors and associated data bases utilized for terrain characteristics mapping in eastern Canada and the eastern United States. 4-15 Soil chemical classes and areas dominated by histosols in the eastern United States. 4-16 Terrain characteristics of eastern Canada summarized by soil category. xx vi Page Number 4-17 4-27 4-33 4-34 4-35 4-41 4-47 4-55 4-65 4-66 4-68 4-76 4-77 4-79 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 I 1 1 ------- I I LIST OF TABLES (continued) xxv ii I I I I I I I I I I I I I I I Table Page Number Number 5-1 Canadian and United States drinking water 5-9 guidelines for toxic metals. 5-2 Current health related ambient air quality 5-11 standards. 5-3 Summary of qualitative evidence for 5-24 visibility related values. 5-4 Summary results of iterative bidding 5-26 visibility studies. 6-1 Experimental regression coefficients with 6-8 estimated standard deviations for small zinc and galvanized steel specimens. 6-2 Examples of material loss in one year. 6-12 7-1 Activity categories. 7-7 7-2 Summary of methods. 7-27 7-3 Summary of physical science data needed 7-29 for benefit evaluation. 8-1 Summary of eastern U.S. surface water area. 8-8 8-2 Surface water area with low and moderate 8-9 potential to reduce acidity. 18-3 Summary of surface water area in eastern 8-11 Canada. 8-4 Ranking of U.S. crops by 1978 value of 8-14 production. 8-5 1978 yield of six crops in 38 states by 8-15 deposition regime. 8-6 Select U.S. agricultural crops by state 8-16 receiving greater than 40 kg/ha.yr sulphate deposition. 8-7 Ranking of crops in eastern Canada by 1980 8-18 value of production. 8-8a Value and percentage of total 1980 yield 8-19 of each crop in eastern Canada by deposition regime. ------- LIST OF TABLES (continued) Table Number 8-8b 8-9 8-10 8-11 8-12 8-13 8-14 8-15 1980 yield of major crops in eastern Canada by deposition regime. Value of major crops by province receiving 40 kg/ha. yr sulphate deposition. U.S. hardwood and softwood volume and growth. U.S. forest volume by state receiving greater than 40 kg/ha. yr sulphate deposition. Hardwood, softwood and mixed wood annual growth in eastern Canada by deposition regime. Annual forest growth by province receiving greater than 20 kg/ha. yr sulphate deposition. U.S. historic sites by ambient Canadian historic inventory by province and deposition. APPENDIX TABLES SECTION 8 8-1 8-2 8-3 8-4 3-5 8-6 8-7 U.S. aquatic resources by state and sensitivity category 10-20 kg/ha.yr sulphate deposition. U.S. aquatic resources by state and sensitivity category 20-40 kg/ha.yr sulphate deposition. U.S. aquatic resources by state and acid sensitivity category for greater than 40 kg/ha.yr sulphate deposition. U.S. agriculture resources in areas receiving 10-20 kg/ha.yr of sulphate deposition. Agriculture resources in areas receiving 20-40 kg/ha.yr of sulphate deposition. U.S. agriculture resources in areas receiving more than 40 kg/ha.yr sulphate deposition. U.S. agricultural resources - state totals for six crops. xxviii Page Number 8-20 8-20 8-23 8-24 8-27 8-28 8-31 8-33 8-37 8-38 8-40 8-41 8-42 8-44 8-45 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 Table Number 8-8 8-9 8-10 8-11 8-12 8-13 8-14 8-15 9-1 9-2 XXIX LIST OF TABLES (continued) U.S. forest resources in areas receiving 20-40 kg/ha.yr sulphate deposition - volume. U.S. forest resources in areas receiving 20-40 kg/ha.yr sulphate deposition - growth. U.S. forest resources in areas receiving greater than 40 kg/ha.yr sulphate deposition- volume . U.S. forest resources in areas receiving greater than 40 kg/ha.yr sulphate deposition - growth. 1980 Canadian agricultural production by crop and province for deposition zone 10-20 kg/ha.yr. 1980 Canadian agricultural production by crop and province for deposition zone 20-40 kg/ha.yr. 1980 Canadian agricultural production by crop and province for deposition zone >40 kg/ha.yr. Sulphate deposition for forest resources by province. Cost of corrosion control by lime addition as a function of plant size. Example of cost calculation for feeding lime at a 5 MGD plant. Page Number 8-47 8-49 8-51 8-52 8-53 8-54 8-57 8-58 9-13 9-14 ------- I I I I I I I I I I I I I I I I I I I XXXI PREFACE In August, 1980, the Governments of Canada and the United States signed a Memorandum of Intent concerning transboundary air pollution. This action was taken in response to concern for actual and potential damage resulting from the long-range transport of air pollutants between countries and in recognition of the already serious problem of acidic deposition. Each country has demonstrated concern for the causing of damage to the other's environment by transboundary movement of its pollutants. This concern is rooted in international agreements, such as the 1909 Boundary Waters Treaty, the Great Lakes Water Quality Agreement, and the 1979 E.C.C. Convention on Long-Range Transboundary Air Pollution, all of which both Canada and the United States have signed. The Memorandum noted that both countries hae set a priority on developing a scientific understanding of long-range transport of air pollutants and resulting environmental effects, and on developing and implementing policies and technologies to combat such effects. To achieve the first steps of this overall objective, the memorandum established a plan of action for the period October, 1980 to January, 1982, during which time five documents are to be prepared by the following work groups: 1. Impact assessment 2. Atmospheric modelling of pollutant movements 3A. Strategies development and implementation 3B. Emissions, cost and engineering assessment 4. Legal, institutional arrangements and drafting (preparation of the actual document to be signed). General terms of reference that apply to all work groups were established, together with detailed terms dealing with each work group. General Terms of Reference (as per MOI) 1. The Work Groups shall function under the general direction and policy guidance of a United States/Canada Coordinating Committee co-chaired by the Department of External Affairs and the Department of State. 2. The Work Groups shall provide reports assembling and analyzing information and identifying measures as outlined below, which will provide the basis of proposals for inclusion in a transboundary air pollution agreement. These reports shall be provided by January, 1982, and shall be based on available information. ------- xxxii I I 3. Within 1 month of the establishment of the Work Groups, they • shall submit to the United States/Canada Coordinating Committee • a work plan to accomplish the specific tasks outlined below. Additionally, each Work Group shall submit an interim report by January 15, 1981. 4. During the course of negotiations, and under the general am direction and policy of the Coordinating Committee, the Work • Groups shall assist the Coordinating Committee as required. 5. Nothing in the foregoing shall preclude subsequent alterations • of the tasks of the Work Groups or the establishment of • additional Work Groups, as may be agreed upon by the Governments. • Specific Terms of Reference; Impact Assessment Work Croup The Group will provide information on the current and projected • impacts of air pollutants on sensitive receptor areas, and prepare proposals for the 'Research, Modelling and Monitoring1 elements of an agreement. • In carrying out this work, the Group will do the following. 1. Identify and assess physical and biological consequences • possibly related to transboundary air pollution. 2. Determine the present status of physical and biological • indicators which characterize the ecoloigcal stablity of • each sensitive area identified. 3. Review available data bases to establish historic adverse | environmental impacts more accurately. 4. Determine the current adverse environmental impact within • identified sensitive areas (e.g., annual, seasonal, episodic). _ 5. Determine the release of residues potentially related to • transboundary air pollution, including possible episodic release from snowpack melt in sensitive areas. • 6. Assess the years remaining before significant ecological changes are sustained within identified sensitive areas. _ 7. Propose reductions in the air pollutant deposition * rates (e.g., annual, seasonal, episodic) which would be necessary to protect identified sensitive areas. fl I I I ------- I I I I I I I I I I I I I I I I I I I XXX111 8. Prepare proposals for the "Research, Modelling and Monitoring" elements of an agreement. A time frame was established which called for preparation of the report of Work Group 1 in three phases. The Phase 1 report, an interim working paper dealing primarily with acidic deposition, was completed in February, 1981. The Phase 2 report, which represented a considerable improvement in the information base that was available for the Phase 1 report as well as receiving more thorough peer review, became available in October, 1981. This Phase 3 report, the final report of Work Group 1, contains not only a further refinement and expansion of the data base used in the Phase 2 report, but also a brief treatment of several additional air pollutants. The Phase 3 report has received fairly extensive peer review from governmental, university, and industrial reviewers and together with Phase 3 reports from all other work groups will undergo formal peer review under the auspices of the United States/Canada Coordinating Committee. ------- xxxiv Chairmen: Vice Chairmen: WORK GROUP-1 MEMBERS G.E. Bangay, Environment Canada C. Riordan (1982-83), U.S. Environmental Protection Agency C.I. Harris (1981-82), B.R. Flamm (1980-81), U.S. Department of Agriculture D. Jeffs (1981-83), G. VanVolkenburg (1980-81), Ontario Ministry of the Environment N.R. Glass, U.S. Environmental Protection Agency (1980-82) R.J. Pickering, U.S. Department of the Interior Work Group Structure: Subgroup Aquatic Terrestrial Man-Made Structures Health & Visibility Economic Benefits Canadian Members: Canada - Leader T. Brydges P. Rennie (1981-83) C. Sullivan (1980-81) H. Martin R. Paolini (1981-83) G. Becking (1980-81) A. Castel U.S.A. - Leader R. Wilhour (1981-83) G. Glass (1980-81) J. Corliss (1981-83) C. Harris (1980-81) D. Flinn J. Bachmann R. Luken W. Ayre, New Brunswick Department of Environment G. Beggs, Ontario Ministry of Natural Resources J. Cooley, Fisheries and Oceans Canada F. Elder, Environment Canada K. Fischer, Environment Canada R. Halstead, Agriculture Canada S. Linzon, Ontario Ministry of the Environment L. Metras, Environment Canada (1980-81) H. St. Martin, Environment Quebec H. Sandhu, Alberta Department of Environment W. Shilts, Energy, Mines, and Resources I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I XXXV United States Members: R. Adams, University of Wyoming (1981-83) J. Baker, North Carolina State University (1982-83) J. Bain, U.S. Environmental Protection Agency (1981-83) J. Barse, U.S. Department of Agriculture (1981-83) R. Beadle, U.S. Department of Energy D. Bennett, U.S. Environmental Protection Agency (1981-83) J. Blanchard, U.S. Department of State (1980-82) D. Brakke, University of Western Washington (1982-83) P. Brezonik, University of Minnesota (1982-83) R. Buckman, U.S. Department of Agriculture F. Burmann, U.S. Environmental Protection Agency (1980-81) D. Burmaster, Council on Environmental Quality (1980-81) J. Carter, U.S. Department of the Interior (1981-83) R. Church, North Carolina State University (1982-83) S. Cramer, U.S. Department of the Interior C. Cronan, University of Maine C. Daniel, Council on Environmental Quality (1981-83) M. Davis, U.S. Environmental Protection Agency (1981-82) L. Dochinger, U.S. Department of Agriculture G. Foley, U.S. Environmental Protection Agency J. Fulkerson, U.S. Department of Agriculture (1981-83) W. Heck, U.S. Department of Agriculture M. Heit, U.S. Department of Energy (1981-83) R. Herrmann, U.S. Department of the Interior J. Jacobson, Boyce Thompson Institute D. Johnson, Oak Ridge National Laboratory R. Kane, U.S. Department of Energy R. Livingston, U.S. Environmental Protection Agency (1981-83) H. Marguiles, U.S. Department of Health and Human Services (1980-81) W. McFee, Purdue University J. Miller, National Oceanic and Atmospheric Administration (1980-81) S. Norton, University of Maine (1982-83) B. Ostro, U.S. Environmental Protection Agency (1981-83) R. Phillips, U.S. Department of Energy (1980-82) T. Pierce, U.S. Environmental Protection Agency (1980-81) R. Porter, U.S. Department of State D. Raynal, State University of New York (1981-83) K. Schreiber, U.S. Department of the Interior (1982-83) S. Sherwood, U.S. Department of the Interior (1981-83) D. Shriner, Oak Ridge National Laboratory J. Spence, U.S. Environmental Protection Agency (1981-83) W. Warnick, U.S. Department of Energy (1981-83) S. Wilson, U.S. Environmental Protection Agency (1981-83) T. Wilson, U.S. Department of State (1982-83) T. Williams, U.S. Department of Energy (1982-83) ------- xxxvi Liaison: M. Beaulieu, Department of External Affairs - Canada M. Levine, U.S. Environmental Protection Agency (1982-83) P. Smith, U.S. Department of Agriculture D. Weber, U.S. Environmental Protection Agency (1980-82) I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I xxxvii ACKNOWLEDGEMENTS The Work Group wishes to acknowledge the assistance in the form of technical consultation or peer review of the following individuals and organizations during the several phases of this study. The enthusiasm and dedication of this assistance were fundamental to the successful completion of this report. Inclusion in the following list should not be taken to indicate endorsement of the report by the named individual or organization: F. Adams (Auburn University), A. Anders (University of Wisconsin), N. Baer (New York University), D. Bjonback (Environment Canada), D. Carter (U.S. Department of the Interior), R. Coote (Agriculture Canada), D. Cowell (Environment Canada), E. Cowling (North Carolina State University), T. Crocker (University of Wyoming), S. Dean (Air Products & Chemicals Incorporated), D. Dixon (Wordcom Centres Ltd.), D. Dodge (Ontario Ministry of Natural Resources), J. Donnan (Ontario Ministry of the Environment), C. Driscoll (Syracuse University), H. Eisler (Stelco Incorporated), B. Forster (University of Guelph), N. Foster (Environment Canada), G. Gilbert (Environment Canada), W. Gizyn (Ontario Ministry of the Environment), C. Griffith (Ontario Ministry of the Environment), M. Griffith (Ontario Ministry of the Environment), W. Hart (Environment Canada), A. Harfenist (Environment Canada), R. Harter (University of New Hamphsire), F. Haynie (U.S. Environmental Protection Agency), H. Hirvonen (Environment Canada), B. Hosier (Environment Canada), H. Hultberg (Swedish Water & Air Pollution Research Institute), T. Hutchinson (University of Toronto), M. Hutton (Environment Canada), C. Jackson (University of Georgia), M. Kelly (Tennessee Valley Authority), J. Kelso (Fisheries and Oceans Canada), J. Kerekes (Environment Canada), J. Knetsch (Simon Eraser University), A. Lefohn (ASL & Associates), R. Linthurst (North Carolina State University), R. Livingston (U.S. Environmental Protection Agency), 0. Loucks (Institute of Ecology), A. Lucas (Environment Canada), C. Lucyk (Ontario Ministry of the Environment), J. MacLean (Ontario Ministry of Natural Resources), R. McLean (Domtar Limited), S. Milburn (Environment Canada), H. Miller (U.S. Department of the Interior), K. Mills (Fisheries and Oceans Canada), K. Minns (Fisheries and Oceans Canada), R. Morris (U.S. Department of Energy), I. Morrison (Environment Canada), J. Nicholson (Environment Canada), D. O'Guinn (Northrop Services, Inc.), R. Olson (Oak Ridge National Laboratory), C. Olver (Ontario Ministry of Natural Resources), J. Pagel (Ontario Ministry of the Environment), M. Parker (Wordcom Centres Ltd.), R. Pearson (Ontario Ministry of the Environment), D. Rambo (Northrop Services, Inc.), E. Rhea (Reynolds Metals Company), C. Rubec (Environment Canada), C. Russell (Resources for the Future), R. Saunders (Fisheries and Oceans Canada), D. Schindler (Fisheries and Oceans Canada), C. Schofield (Cornell University), P. Sereda (National Research Council - Canada), K. Shea (U.S. Department of Agriculture), S. Singh (Agriculture Canada), J. Skelly (Pennsylvania State University), J. Smith (Ontario Ministry of the Environment), W. Smithies (Ontario Ministry of the Environment), C. Taylor (University of California - Riverside), M. Thompson (Environment Canada), D. Thornton (University of Minnesota), G. Voigt ------- xxxviii I (Yale Universty), D. Wark (Environment Canada), W. Watt (Fisheries • and Oceans Canada), M. Weaver (Heritage Canada), E. Winkler ™ (University of Notre Dame), N. Yan (Ontario Ministry of the Environment), M. Young (Ontario Ministry of the Environment). H I I I I I I I I I I I I I I I I ------- SECTION I SUMMARY ------- I I I I I I I I I I I I I I I I I I I 1-1 SECTION 1 SUMMARY 1.1 INTRODUCTION Wet and dry deposition of acidic substances and other pollutants are currently being observed over most of eastern North America. The Impact Assessment Work Group was charged with identifying and making an assessment of the key physical and biological consequences possibly related to these transboundary air pollutants. During the Work Group's assessment of these effects it has been necessary to conduct the work along strictly disciplinary lines. Thus the presentation of our findings follows a sectoral approach (i.e., aquatic, terrestrial). While this approach has been useful for organizing and presenting our findings, it has also limited our consideration of the interactions which exist among these sectors. These effects do not occur in isolation. The following sections summarize findings of the Work Group with respect to impacts on aquatic and terrestrial sectors of the biosphere, health and visibility, and man-made structures. There are also summary statements with regard to methodologies for estimates of economic benefits of controls, natural and material resource inventory, and liming. 1.2 AQUATIC ECOSYSTEM EFFECTS - CANADA The potential effects from the deposition of acid and associated ions and compounds (sulphur dioxide, sulphate, nitrate, ammonia, and others) on water quality, and on the aquatic ecosystem, appear to be more fully quantified and understood than for terrestrial ecosystems. Data have been drawn from a number of study areas in eastern North America including Labrador, Newfoundland, Nova Scotia, New Brunswick, the southern part of the Canadian Shield in Quebec, and Ontario. Primary study areas in the U.S. are found in New Hampshire and southern Maine, Adirondack Park in New York, the Boundary Waters Canoe Area of Minnesota, and numerous lakes in north-central Wisconsin. The findings and conclusions of the Work Group with respect to acidification effects are contained in the following statements: Sulphuric acid has been identified as the dominant compound contributing to the long-term surface water acidification process. Nitric acid contributes to the acidity of precipi- tation, but is less important in eastern North America than sulphuric acid in long-term acidification of surface waters. Nitric acid contributes to pH depression of surface waters during periods of snowmelt and heavy rain runoff in some areas. ------- 1-2 Observed Historical Changes Sediments from lakes in Maine, Vermont, and New Hampshire I I Studies of lakes in eastern North America have provided evidence that atmospheric deposition accounts for sulphate levels in excess of those expected from natural processes. In the absence • of effects from mine drainage and industrial waste water, the • symptoms of acidification (e.g., pH depressions of surface waters and loss of fish populations), have been observed only in _ lakes and rivers where the accompanying elevated concentrations I of surface water sulphate (and nitrate in some cases) indicate * atmospheric deposition of these ions. Land use changes, such as fires, logging, and housing developments have taken place in I many areas with sensitive (low alkalinity) surface waters, but • the symptoms of acidification have not been observed unless there is an accompanying increase in surface water sulphate • concentrations. Nitrate concentrations also increase in some I areas, especially during snowmelt. In eastern Canada, the surface waters which have elevated excess • sulphate occur in areas which have high atmospheric deposition ^ of sulphate. All of the surface waters sampled in northeastern North America that have experienced loss of alkalinity also have elevated excess sulphate concentrations. In areas with less acidic deposition, loss of alkalinity in surface waters has not been observed. In Quebec, sulphate concentrations in surface • waters decrease towards the east and north in parallel with • deposition patterns. Sulphate concentrations are equal to or greater than the bicarbonate concentration in lakes in the _ southwest part of the Province. This indicates that the surface I water chemistry has been altered by atmospheric sulphur • deposition. I I I indicate increased atmospheric acidic deposition has affected I terrestrial and aquatic ecosystems as measured by changes in metal concentrations and diatom populations. It has been inferred from the sediment record that the rate of acidification • of aquatic ecosystems has increased since the late 1800s as • measured by declines in metals (zinc, copper, iron, calcium, magnesium and manganese) in the sediments. Conditions of low pH maintain metals in the water column, where they can be flushed out of the system before being deposited in the sediments. Diatom data are less complete, but they also indicate a M statistically significant pH decline since the early 1900s. I In this report numerous historical chemistry records have been examined for waters not influenced by local urban or industrial • discharges. Reviews have been conducted for 2 rivers in B Newfoundland and 6 in Nova Scotia; 7 lakes in Nova Scotia and 3 in New Brunswick; 40 lakes in Adirondack Park, New York; 250 • lakes in New England; 2 streams in New Jersey Pine Barrens; and | 275 lakes in Wisconsin. Historical records which are available ------- I I I I I I I I I I I I I I I I I I I 1-3 from areas of soils and bedrock with a low potential to reduce acidity exposed to acidic deposition, show an increase in sulphate and corresponding decrease in alkalinity and pH. Areas of similar lithology and land use practices, but not receiving significant acidic deposition do not show similar losses of alkalinity. Lakes in the Adirondack Mountain range have some of the lowest alkalinity values and are located in watersheds with a low potential to reduce acidity. They are located in the eastern U.S. in a zone receiving high acidic deposition (26-40 kg/ha.yr of sulphate in precipitation 1978-81). Historical data on fish and pH are available for 40 high elevation Adirondack lakes. In the 1930s, only 8% of these lakes had pH less than 5.0; 10% had no fish whereas in the 1970s, 48% had pH less than 5.0 and 52% had no fish. In some cases, entire fish communities consisting of brook trout, lake trout, white sucker, brown trout, and several cyprinid species apparently have been eliminated over the 40-year period. The New York Department of Environmental Conservation has concluded that at least 180 former brook trout ponds are acidic and no longer support brook trout. The relative contribution of natural and anthropogenic sources to acidification of these lakes is not known. In New England, deposition of wet sulphate has been measured to be 17-40 kg/ha.yr. A study of 95 lakes for which there are historical pH data from the 1930s to the 1960s has indicated that 36% either had the same pH or higher while 64% now have lower pHs. For 56 lakes, a comparison of historical alkalin- ities to modern values indicated that 30% of the lakes had increased and 70% had decreased in alkalinity. Over the period of record, measured alkalinity values have decreased by an average of 100 yeg/L. The lakes were small to medium size oligotrophic to mesotrophic with moderately to very transparent water, low to moderate concentrations of humic solute, low alkalinity and conductance and with moderately disturbed to pristine watersheds. For four rivers in Nova Scotia data from 1980-81 showed a decrease in bicarbonate, ah increase in sulphate and hydrogen ion concentrations when compared to 1954-55 data. Short-Term pH Depressions While the rate of change of water quality of lakes (i.e., the time required for a lake to become acidified) is one of the least well-defined aspects of the acidification process, there is evidence that current acid loadings are damaging to fish populations and other biota due to short-term pH depressions following snowmelt and storm runoff. Both sulphate and nitrate are associated with short-term changes in water chemistry but in the majority of surveyed cases sulphate appears to be the larger contributor to the total acidity. ------- 1-4 I I Short-term pH depressions, and elevated concentrations of metals, particularly aluminum, have been observed during periods of high infiltration or runoff. Metal accumulation in surface • waters (Al, Mn, Fe, Zn, Cd, Cu, Pb, and Ni), first noted in | streams and lakes of Scandinavia, also has been reported from such places as Hubbard Brook, the Adirondacks, and the Great _ Smoky Mountains of the U.S., and the southern Precambrian Shield I area of Ontario, Canada. Artificial acidification of a lake in ™ the Experimental Lakes Area of Ontario has also shown rapid mobilization of metals from lake sediments to the water column. I Data for 57 headwater streams in Muskoka-Haliburton show that 65% experience minimum pH values less than 5.5 and 26% have m minimum pH values less than 4.5. Some inlet streams were • observed to have pH values below 4.0 during spring snowmelt. Data from intensive studies of 16 lakes in the Muskoka- I Haliburton area of Ontario currently receiving about 23-29 • kg/ha.yr sulphate in precipitation have shown that lakes which have summer alkalinity values up to about 40 yeg/L, experience • pH depressions to values below about 5.5 during snowmelt. In | Ontario and OueJbec there are about 1.5 million lakes on the Precambrian Shield. In Ontario, of the 2,260 lakes sampled on im the Precambrian Shield, 19% have alkalinities below 40 yeg/L. I In the Shield area of Quebec, a 1981 survey of 162 lakes indicated 37% were extremely sensitive to acidification (CSI greater than 5.0), while 15% had summer pH values less than 5.0 I (alkalinity less than 0). I A very large number of surface waters are being affected by acidic deposition, even though the total number of lakes and rivers in eastern North America which are known to have been acidified (alkalinity less than 0) by atmospheric acidic « deposition is a relatively small percentage of the total aquatic I resource. ^ Biological Effects • Detailed studies of watersheds have been carried out in sensitive regions of North America and Scandinavia under a range • of sulphate deposition rates. The results of the studies | conducted in North America are described below. Observed changes in aquatic life have been both correlated with • measured changes in the pH of water and compared for waters of * different pH values. Differences have been documented in species composition and dominance and size of plankton B communities in lakes of varying pH. Study results show that the | number of species is lower in low pH lakes compared to lakes of higher pH. These alterations may have important implications m for organisms higher in the food chain. Individual lakes often • experience several symptoms of acidification at the same time. For example, in Ontario, Plastic Lake inlet streams have low pH I ------- I I I I I I I I I I I I I I I I I I I 1-5 and high aluminum concentrations during spring runoff and extensive growth of filamentous green algae, and fish kills have been observed in Plastic Lake. For those regions currently receiving loadings of sulphate in precipitation of less than 17 kg/ha.yr (Wisconsin, Minnesota and northwestern Ontario), there have been no observed detrimental chemical or biological effects. For regions currently receiving between 20 and 30 kg/ha.yr sulphate in precipitation there is evidence of chemical alteration and acidification. In Nova Scotia rivers which currently have pH less than 5 there have been salmon population reductions as documented by 40 years of catch records. Fish stocks have remained viable in adjacent rivers with pH values presently greater than 5. Water chemistry records (1954-55 to 1980-81) have indicated a decline in pH to values presently less than 5 for other rivers in the same area. In Maine there is evidence of pH declines over time and loss of alkalinity from surface waters. In Muskoka-Haliburton there is historical evidence of loss of alkalinity for one study lake and there is documentation of pH depressions in all study lakes and streams with low alkalinity. Fish kills were observed in the shore zone of a study lake during spring melt. In the Algoma region there are elevated sulphate and aluminum levels in some headwater lakes. For regions currently experiencing loading greater than 30 kg/ha.yr there are documented long-term chemical and/or biological effects and short-term chemical effects in sensitive (low alkalinity) surface waters. In the Adirondack Mountains of New York, comparison of data from the 1930s with recent surveys has shown that some more lakes have been acidified. Fish populations have been lost from 180 lakes. Elevated aluminum concentrations in surface waters have been associated with low pH and survival of stocked trout is reduced by the aluminum. In the Hubbard Brook study area in New Hampshire where the influx of chemicals is limited principally to precipitation and dry deposition there are pH depressions in streams during snowmelt of 1 to 2 units. Elevated levels of aluminum were observed in headwater streams. Many species of frogs, toads and salamanders breed in temporary pools formed by the mixture of spring rains and snowmelt. Such pools are subjected to pH depression. Embryonic deformities and mortalities in the yellow spotted salamander which breeds in temporary meltwater pools have been observed in New York State where the acidity of the meltwater pools was 1.5 pH units lower than that of nearby permanent ponds. Population densities of ------- 1-6 the bullfrog and woodfrog were reduced in acidic streams and ponds in Ontario* Target Loadings Sulphate in precipitation has been used as a surrogate for total acid loading. Sulphate in precipitation can be reliably I I I A lake acidification experiment in northwestern Ontario clearly shows that alterations to aquatic food chains begin at pH values slightly below 6.0. The remarkable agreement between these « whole-lake experiments and observational studies in Scandinavia I and eastern North America provides strong evidence that the observed declines in fisheries are caused by acidification and not by other ecological stresses. I Extent of Effects The terrestrial mapping analysis for eastern Canada supported by • surface water chemistry has demonstrated that the watersheds of sensitive (low alkalinity) aquatic ecosystems where effects _ have teen observed have a low potential to reduce acidity and • are representative, in terms of soil and geological characteristics, of much larger areas of eastern Canada. Similarly, using related but different criteria, maps have been • developed which characterize considerable areas of the northeastern United States as having low potential to reduce m acidity. Therefore, there is reason to expect that there are • sensitive surface waters in these other areas which would experience similar effects if subjected to deposition rates _ comparable to those in the study areas. However, quantification • of the number of lakes and rivers susceptible to acidification ~ in both countries will require validation of the terrestrial mapping methodologies and increased information on the chemistry I of lakes and streams. I The present empirical evidence covers a broad spectrum of m physical and climatological conditions across northeastern North I America and therefore provides a reasonable basis on which to make judgements on potential loading effect relationships. However the data do have some deficiencies. More data on • historical trends of deposition and associated chemical and • biological characteristics would improve our understanding of long-term rates and effects of acidification. In addition, a • better understanding of all the mechanisms involved in the | acidification process will enhance our ability to estimate loading/response relationships precisely. Therefore any • estimates of loading/response relationships should be • strengthened in the light of new scientific information as it becomes available. I I I I ------- I I I I I I I I I I I I I I I I I I I 1-7 measured. Jt is recognized that dry deposition of sulphate and sulphur dioxide, and the wet and dry deposition of nitrogen oxides, nitric acid, particulate nitrate and ammonia, as well as other compounds also contribute to acidic deposition. Based on documented effects, wet and dry deposition of sulphur compounds dominate in long-term acidification. Sulphur deposition also predominates in the majority of cases surveyed involving short-term pH depressions and associated effects. Insufficient data are available to relate nitrate deposition to short-term water quality effects. Therefore, we are unable to determine a nitrate dose-response relationship. The models, which are based on theory, that have been considered, permit a quantification of the target loadings in terms of geochemical basin sensitivity. Although these models require further validation, the derived loading estimates are generally supportive of the empirical observations for the study areas discussed above. Based on the results of the empirical studies, interpretation of long-term water quality data, studies of sediment cores and models that have been reviewed, we conclude that acidic deposition has caused long-term and short-term acidification of sensitive (low alkalinity) surface waters in Canada and the U.S. The Work Group concludes on the basis of our understanding of the acidification process that reductions from present levels of total sulphur deposition in some areas would reduce further damage to sensitive (low alkalinity) surface waters and would lead to eventual recovery of those waters that have already been altered chemically or biologically. Loss of genetic stock would not be reversible. The Canadian members of the Work Group propose that present deposition of sulphate in precipitation be reduced to less than 20 kg/ha.yr in order to protect all but the most sensitive aquatic ecosystems in Canada. In those areas where there is a high potential to reduce acidity and surface alkalinity is generally greater than 200 \ieq/L, the Canadian members recognize that a higher loading rate is acceptable. As loading reductions take place and additional information is gathered on precipitation, surface water chemistry and watershed response, it may be possible to refine regional loading requirements. 1.2 AQUATIC ECOSYSTEM EFFECTS - UNITED STATES Acidic deposition has been reported in the literature as a cause of both long-term and short-term episodic depressions in pH and loss in alkalinity in some lakes and streams in the U.S. and Canada. ------- 1-8 I I Elevated concentrations of toxic elements, such as aluminum, and biological effects including losses in fish populations have been reported to accompany some of these pH depressions. In most of the • reported cases, clear relationships were not established between | acidic deposition and observed effects. Conclusions are based on an understanding of the acidification process although mechanisms which ^ control this process are often not completely understood. • The following summary statements are observations reported to be occurring in areas receiving acidic deposition. I Both sulphuric and nitric acid contribute to the acidity of precipitation. It appears, however, that sulphuric acid IB contributes more to long-term acidification of surface waters I than does nitric acid. Nitric acid can contribute to pH depression of surface waters during periods of snowmelt and _ heavy rain runoff in some areas. Studies of lakes in eastern • North America indicate that atmospheric deposition accounts for ™ sulphate levels in some waters in excess of those expected from natural processes. Lake study areas are located in Labrador, I Newfoundland, Nova Scotia, New Brunswick, the southern part of | the Canadian Shield in Quebec, and in eight regions of Ontario. Primary study areas in the U.S. are found in New Hampshire and • southern Maine, Adirondack Park in New York, the Boundary Waters • Canoe Area of Minnesota, and numerous lakes in north-central Wisconsin. There is evidence of long-term reductions of pH and alkalinity • and other water quality changes for some low alkalinity surface waters. The rate of change of pH and alkalinity in lakes is one H of the least well defined aspects of the acidification process. | However, there is evidence of short-term pH depressions in some waters following high runoff from snowmelt and storm activity. M Both sulphate and nitrate are associated with short-term changes I in water chemistry but, in the majority of surveyed cases, sulphate appears to be the larger contributor to total acidity. Short-term pH depressions and elevated concentrations of metals, • particularly aluminum, iron, zinc, and manganese have been observed during periods of high runoff. Metal mobilization from • some watersheds, first noted in streams and lakes of | Scandinavia, also has been reported from such places as Hubbard Brook, the Adirondacks, and the Great Smokey Mountains of the • U.S., and Sudbury, Muskoka, and Plastic Lake in Ontario, Canada. • Artificial acidification of a lake in the Experimental Lakes Area of Ontario has shown mobilization of metals from lake sediments to the water column. I Sediments from lakes in Maine, Vermont, and New Hampshire suggest increased acidity in aquatic ecosystems. It has been • inferred from declines in metals (zinc, copper, iron, calcium, | magnesium and manganese) in the sediments that the acidity of the water increased since the late 1800s. Low pH maintains _ I ------- I I I I I I I I I I I I I I I I I I I 1-9 metals in the water column, where they can be flushed out of the system before being deposited in the sediments. Diatom data are less complete, but they also indicate a pH decline since the early 1900s. There are few historical records of chemistry of low alkalinity waters not influenced by local urban or industrial discharges (i.e., 6 rivers in Nova Scotia; 7 lakes in Nova Scotia and 3 in New Brunswick; 40 lakes in Adirondack Park, New York; 250 lakes in New England; 2 streams in the New Jersey Pine Barrens; 270 lakes in Wisconsin). The above locations are exposed to various levels of acidic deposition. Some surface waters in these areas have shown a decrease in alkalinity and/or pH. In Wisconsin, however, most lakes surveyed had increased in alkalinity and pH. The total number of lakes and rivers in eastern North America that are thought to have been acidified by acidic deposition is a very small percentage of the total aquatic resource. In the absence of effects from mine drainage and industrial waste water, the symptoms of acidification (e.g., long-term pH declines and/or short-term pH depressions of surface waters with loss of fish populations) have been observed only in clearwater lakes and streams with accompanying elevated concentrations of sulphate and/or nitrate. Natural acidification processes do occur but their effects appear greatest in coloured surface waters. Land use changes, such as fires, logging, and housing developments, have taken place in many areas with low alkalinity surface waters. However, the symptoms of acidification have not been observed in clearwater lakes and streams except in areas receiving high levels of acidic deposition. Lakes in the Adirondack Mountain range exhibit some of the lowest alkalinity values found in the eastern United States and are located in a zone presently receiving high acidic deposition (30-40 kg/ha.yr of sulphate in precipitation). In this area, 52% of the 214 high elevation lakes sampled in 1975 had pH values less than 5.0. Seven percent had pH values between 5.0 and 6.0. The New York Department of Environmental Conservation has concluded that at least 180 former brook trout ponds are acidic and no longer support brook trout. The factors causing these population extinctions have not been demonstrated. New England currently receives wet sulphate deposition loadings of 17-40 kg/ha.yr. A study of 95 relatively small low alkalinity lakes in New England for which historical data were available showed that 64% had decreased in pH. However, accompanying historical deposition data are not available. A comparison of present alkalinity values with historical values for 56 lakes indicated that 70% had decreased in alkalinity. Two other studies have indicated pH declines in some lakes surveyed in Maine. The relative contributions of natural and anthropogenic sources to acidification of these lakes is not known. ------- 1-10 I I Data from intensive studies of 17 lakes in the Muskoka-Haliburton area of Ontario currently receiving about 20-30 kg/ha.yr sulphate in precipitation have shown that some lakes with summer alkalinity • values up to about 40 yeg/L experience pH depressions to values below • 5.5 during snowmelt. One inlet stream was observed to have pH values as low as 4.1 during spring snowmelt. Other inlet streams had pH _ depressions but pH did not drop as low. Of 2,624 lakes surveyed in • Ontario, 50% had alkalinity of less than 200 yeg/L, a value that may be regarded as the upper limit for potential effects of acidic deposition; 13% of the lakes sampled in the province had alkalinities I below 40 yeg/L. While these lakes may be representative of the areas m sampled, they may not be representative of lakes located elsewhere in the Shield. In another survey of 199 lakes of the Precambrian Shield 4H of Quebec 7.5% had alkalinity of approximately 50 yeg/L or less. M There are about 1.5 million lakes on the Precambrian Shield in the provinces of Ontario and Quebec; but it is not possible at present _ to extrapolate results of the surveys to the total population of • lakes. ™ Observed changes in aquatic life have both been correlated with I measured changes in the pH of water and inferred by comparisons of | waters of different pH values. Differences have been documented in species composition and dominance and size of plankton communities in mt lakes of varying pH. Study results show that the number of species • is lower in low pH lakes compared to lakes of higher pH. These differences may have important implications for organisms higher in the food chain, but studies to date have not been done that might • establish this connection. • Many species of frogs, toads and salamanders breed in temporary pools Ij formed by the mixture of spring rains and snowmelt and subject to pH | depression. Embryonic deformities and mortalities in the yellow spotted salamander, which breeds in temporary meltwater pools, have « been observed in New York State where the acidity of the meltwater I pools was 1.5 pH units lower than that of nearby permanent ponds. Population densities of the bullfrog and woodfrog were lower in acidic streams and ponds than in those of higher pH sampled in • Ontario. These data are very limited and therefore the extent of the ™ problem is unknown. Atlantic salmon populations have disappeared from nine rivers in Nova | Scotia but remain in rivers in the same area having higher pH due to greater alkalinity. Decreases in alkalinity and the pH of water over « time have been observed in some low pH rivers in Nova Scotia. • However, historical chemical data do not exist for the period of major decline in angling success nor do they exist for rivers in which fish declined. • Detailed studies of watersheds and clusters of lakes have been carried out in regions of North America and Scandinavia containing low alkalinity lakes and streams under a range of sulphate deposition I I I ------- I I I I I I I I I I I I I I I I I I I 1-11 rates. The results of those studies conducted in North America are summarized below. There have been no reported chemical or biological effects for regions currently receiving loadings of sulphate in precipitation at rates less than about 20 kg/ha.yr. Evidence of chemical change exists for some waters in regions currently estimated or measured to be receiving between about 20-30 kg/ha.yr sulphate in precipitation. In Nova Scotia rivers, 40 years of historical records document reductions in angling success for Atlantic salmon in nine rivers of low pH. Records over later periods for other nearby rivers document decreases in alkalinity and pH. In Maine there is evidence of pH declines over time and loss of alkalinity from some surface waters. In Muskoka-Haliburton historical evidence documents loss of alkalinity for one lake and pH depressions in a number of lakes and streams. Fish confined to the inlet of one lake died during spring melt. In the Algoma region there are elevated sulphate and aluminum levels in some headwater lakes. Long-term chemical and/or biological effects and short-term chemical effects have been observed in some low alkalinity surface waters experiencing loadings greater than about 30 kg/ha.yr. In Quebec, sulphate concentrations in surface waters decrease towards the east and north in parallel with the deposition pattern of sulphate. Sulphate concentrations are equal to or greater than the bicarbonate concentration in some lakes in the southwest part of the province. In the Adirondack Mountains of New York comparison of data from the 1930s with recent surveys has shown that more lakes are now in low pH categories. The relative contribution of natural and anthropogenic sources to acidification of these lakes is not known. The New York Department of Environmental Conservation has concluded that at least 180 former brook trout ponds are acidic and no longer support brook trout, although a direct association with acidic deposition has not been established. In the Hubbard Brook study area in New Hampshire there are pH depressions in some streams during snowmelt of 1 to 2 units. In the watershed studies summarized above, sulphate in precipitation was used as a surrogate for total acid loading. Sulphate in precipitation can be reliably measured. It is recognized that dry deposition of sulphate and sulphur dioxide, and the wet and dry deposition of nitrogen oxides, nitric acid, particulate nitrate and ammonia, as well as other compounds, also contribute to acidic deposition. The use of a single substance as a surrogate for acidic loadings adds unknown error owing to site-to-site variability in: (1) composition of deposition, and (2) ability of watersheds to neutralize incoming acidity. Wet and dry deposition of sulphur compounds appeared to predominate in long-term acidification. ------- 1-12 I I Insufficient data are available to related nitrate deposition to short-term water quality effects. Therefore, we are unable to develop nitrate loading/response relationships. • The terrestrial mapping analysis for eastern Canada has demonstrated that the watersheds in which some surface waters have been observed « to experience effects are representative, in terms of soil and • geological characteristics, of larger areas of eastern Canada. The level of variability within terrain classes is not known. An alkalinity map of the U.S. shows the location of regions where the m mean alkalinity of most of the sampled surface waters is less than 200 yeg/L. There is reason to believe that some of these low • alkalinity surface waters could experience effects similar to those <|0 noted in detailed study sites receiving similar total acidic deposition loadings. However, quantification of the number of lakes M and rivers in both countries susceptible to acidification at specific • loading rates would require validation of mapping methodologies and increased information on loading rates and the chemistry of lakes and streams. The present empirical evidence covers a broad spectrum of • physical and climatological conditions across northeastern North V America and therefore provides a basis on which to make only qualitative judgements regarding relationships between acidic loading m rates and effects. • Based on the results of the empirical studies, interpretation of _ long-term water quality data and studies of sediment cores that have • been reviewed, we conclude that acidic deposition has caused long- ™ and short-term acidification of some low alkalinity surface waters in Canada and the U.S. Based on our understanding of the acidification H process the Work Group concludes that reductions from present levels | of total sulphur deposition would reduce further chemical and biological alterations to low alkalinity surface waters currently « experiencing effects and would lead to eventual recovery of those m waters that have been altered by deposition. The U.S. members conclude that reductions in pH, loss of alkalinity, • and associated biological changes have occurred in areas receiving * acidic deposition, but cause and effects relationships have often not been clearly established. The relative contributions of acidic M inputs from the atmosphere, land use changes, and natural terrestrial | processes are not known. The key terrestrial processes which provide acidity to the aquatic systems and/or ameliorate atmospheric acidic M inputs are neither known or quantified. The key chemical and • biological processes which interact in aquatic ecosystems to determine the chemical environment are not known or quantified. Based on this status of the scientific knowledge, the U.S. (fork Group • concludes that it is not now possible to derive quantitative B loading/effects relationships. I I I ------- I I I I I I I I I I 1 I I 1 I 1 I I I 1-13 1.3 TERRESTRIAL ECOSYSTEM IMPACTS The effects of transboundary air pollution on terrestrial ecosystems have been reviewed on the basis of direct effects on vegetation, effects on soils, and effects on wildlife. 1.3.1 Effects on Vegetation Three main pollutants are of concern with regard to vegetation effects. These pollutants are sulphur dioxide, ozone, and acidic deposition. Ozone and acidic deposition occur at concentrations above background levels at long distances from emission sources. Sulphur dioxide is more of a concern to vegetation in proximity to point sources of emissions than at long distances, where dispersion effects can reduce atmospheric levels to those of background. 1.3.1.1 Sulphur Dioxide Near point sources, the adverse effects of sulphur dioxide on vegetation can be both visible and subtle (without development of visible foliar injury). Visible effects can be associated with both doses of high concentrations of sulphur dioxide over short periods of time and low concentrations over extended periods. However, in a few specific cases, atmospheric sulphur dioxide deposition may have beneficial effects on agricultural vegetation grown on borderline or sulphur deficient soils. Visible effects of sulphur dioxide have occurred on pine forests in Canada subjected to average growing season concentrations of sulphur dioxide of 0.017 ppm. Visible injury to the perennial foliage of coniferous trees results in premature needle drop, reduced radial and volume growth and early death of trees. Reduced growth and yield of crops without the development of visible injury have also been found in certain field experiments. Annual doses of sulphur dioxide of 0.02 ppm have been associated with habitat modifications in grasslands and the elimination of certain sensitive species of lichens near point sources. Lichens may be markedly affected by sulphur dioxide and are considered as bioaccumu- lators of very low level sulphur dioxide exposures. Direct effects including visible injury, effects on reproductive capacity and species mortality have been encountered in the field at concentra- tions of sulphur dioxide as low as 0.006 - 0.03 ppm annual average. Despite such documented evidence of instances of direct effects, obviously not all, but probably most exposures to sulphur dioxide on a regional scale are below levels producing phytotoxic reactions. However, long-term, low-dose studies have demonstrated direct effects on lichen communities and indirect effects on several plant species. ------- 1-14 1.3.1.2 Ozone 1.3.2 Effects on Terrestrial Wildlife Soils vary widely with respect to their properties, support different vegetation communities, are subjected to different cultural I I I Ozone is the most important long-range transported pollutant with respect to vegetation effects. Air masses carry ozone and its precursors over long distances and can affect crops and forests in rural areas remote from sources. As a specific example, ozone • related crop injuries in southern Ontario have been reported • associated with high ozone levels in air masses moving across Lake Erie. In the U.S., experimentally derived crop yield losses ranging from 2 to 56% (crop dependent) were equated with seasonal 7 hr/day • mean ozone concentrations of 0.06 - 0.07 ppm. Yield losses in the 9 various crops were as follows: kidney bean 2%, soybean 10%, peanut 14-17%, and lettuce 53-56%. Although direct effects of ozone have • been documented on forest growth, an estimate of loss is difficult to • calculate because of the limitations stated in the main report. 1.3.1.3 Acidic Deposition Acidic deposition in the form of simulated rain has been demonstrated • to induce a variety of direct and indirect effects on plants grown •! under greenhouse or semicontrolled conditions. Foliar injury, growth reductions, and growth stimulations have been found under these m growing conditions following treatment with simulated acidic precipi- I tation. However, visible foliar injury has not been documented in the field for vegetation exposed to ambient levels of acidic — precipitation. The potential effects of acidic deposition on forest • growth have been difficult to assess because of the complicating - influence of other environmental and climatic factors. To date, there have been too few studies to establish a clear relationship on • the interactions of acidic deposition/sulphur dioxide/ozone to reach H a definitive conclusion on effects. I Direct effects of acidic deposition on terrestrial wildlife have not • been reported and are not considered likely. Nevertheless, in some • instances, indirect effects have been suggested through three possible mechanisms: (• 1) contamination by heavy metals mobilized by acidity; 2) reduction in nutritional value of browse or food source; • and I 3) loss of browse species or impairment of habitats. I 1.3.3 Effects on Soil • I I ------- I I t I I I I I I I I I I 1 I I I 1 I 1-15 practices, are situated in different climatic zones, and are exposed to a broad spectrum of acid loadings. The following effects of acidic deposition probably occur and in some cases are supported by observation, although the number of field situations where investi- gators have been able to attribute acidity to precipitation or to compare present with former soil pH value is small. On soils derived from calcareous parent materials, the effects of acidic deposition will lead to only insignificant increases in lime requirement, except in situations near strong point emitters. Heavy metal deposition from these same point source emitters may also cause soil toxicities. On acid soils, the absence of clear effects upon tree growth from radial-increment measurements covering several decades suggests there will be no short-term effects attributable to acidic deposition. From the few field situations where earlier investigations permit a comparison over a reasonable time-frame, there is evidence that less acutely acid soils increase in acidity and lose bases at a faster than normal weathering rate. For acutely acid soils, pH may show only minor changes, while over the same period moderate to appreciably larger amounts of soil aluminum are mobilized. These depend upon whether the forest cover is deciduous (e.g., beech) or coniferous (e.g., spruce). From one comprehensive field investigation, it has been suggested that the additional amounts of aluminum brought into solution kill feeding roots and permit the invasion of fungi causing tree "dieback", but it is not known whether this phenomenon would occur on other sites and soils. What appears well established from a variety of hydrological, limnological and catchment studies is that acidic deposition can lead to the release of additional amounts of soluble aluminum, thus disturbing previous aluminum/calcium ratios in soils, sediments and streamwaters. An eventual reduction in base status and fertility is suggested. The sulphate component of acidic deposition appears to be adsorbed by soils containing active aluminum and iron oxides, but where these are absent or present in limited amounts, sulphate functions as a balancing anion, leading to the leaching loss of bases and other cations. The fate of the nitrate component depends upon wet precipitation/ snowmelt characteristics. Nitrate, reaching the surface organic horizons of acid forest soils is held there for assimilation by tree roots during the growing season. There are, however, forested catch- ments in the northeast where nitrate is passed to water bodies. The lack of appropriate experimental approaches from which the effects of acidic deposition on soil might be assessed and safe deposition ceilings estimated, has caused scientists to exploit ------- 1-16 1.3.4 Sensitivity Assessment 1.4 HUMAN HEALTH AND VISIBILITY 1.4.1 Health 1 I indirect or special situations. These include working near strong point sources, studying soils treated with acidifying fertilizers, and designing lysimetric experiments incorporating simulated acid ft rains. From such approaches, a variety of soil effects have been I demonstrated, usually of an undesirable nature, but at the present time the problem remains of quantifying the dose-response reactions ._ in the field situations. • I 1 1 Regions which may be sensitive to acidic deposition have one or more components (i.e., forests, aquatic life, soil, or water) susceptible JB to degradation under the influence of acidic deposition. Relative m sensitivity of these components is reflected in the rate at which an ecosystem component degrades under a particular acidic deposition loading. Different underlying criteria have to be used to represent • sensitivity for the different ecosystem components, such as rate of ™ tree growth, characterization of the soil-base status, or water alkalinity. Because so little is known about the acidic deposition dose-response relationships, the underlying criteria are often imprecise. Therefore, relative sensitivity can only be approximately represented or mapped, and then perhaps for only a few species, • ecosystems or theoretical effects. • Attention is focused on the sensitivity of soils and bedrock because results from studies which address vegetation and ecosystem effects • are limited and not well understood at this time. In the approach ™ used, the emphasis has been to map a combination of potentially important soil attributes as a best available indicator of relative sensitivity. Soil attributes incorporated include texture, depth to carbonate, pH and cation exchange capacity, as well as glacial and bedrock features. Incompleteness of survey data for certain _ important properties (e.g., sulphate adsorption capacity, internal • proton production, and the role of dry deposition) precludes their use in identifying detailed sensitivities of land or aquatic resources. As far as possible, the eastern parts of the United • States and Canada are mapped using a similar conceptual framework • which indicates the general extent of areas of different possible sensitivities to the effects of acidic deposition. The significance • of these categories will increase as more effects are documented. <• I I Available information gives little cause for concern over direct health effects from acidic deposition. The potential indirect health mt effects associated with transboundary air pollution discussed are: • (1) contamination of the food chain with metallic substances, I I ------- I I I I I I I I I I i I I 1 I I I I I 1-17 especially mercury; (2) leaching of watersheds and corrosion of storage and distribution systems, leading to elevated levels of toxic metals; and (3) health implications of recreational activities in impacted waters. The principal conclusions of the report are as follows: Acidification of lakes is a concern because it may be related to increased mercury contamination of the food chain, thus increasing the health risks associated with high levels of consumption of contaminated organisms. A correlation exists between low pH in lakes and higher mercury concentrations in some species of fish, although the mechanism for this accumu- lation is not presently known. In addition to the effects produced by acidic deposition, the increased input of anthropo- genic sources (air or water effluents) of mercury and other heavy metals may further complicate the issue and lead to health problems when affected fish are consumed by humans in large amounts. Acidic deposition may liberate metals in some groundwaters, surface drinking water supply systems and cisterns. However, groundwater may also be acidic due to increased partial pressure of C02 at depths of a few metres or more. This should not be confused with acidity due to atmospheric deposition. Elevated metal concentrations in acidified drinking water supplies have been found. Lead levels in tap water from cisterns were much higher than those found in the source water; about 16% of the households sampled in one western Pennsylvanian county had tap water levels in excess of the United States drinking water standards. Surface drinking water supplies which are not treated (i.e., small communities or individual water supplies) are susceptible. No adverse health effects resulting from consumption of such water have been reported. Concern has been expressed that recreational activities in acidified waters, such as swimming, may prove to cause eye irritation. To date, no compelling evidence has been forthcoming that indicates that humans are being adversely affected by these waters in their current state. With respect to the direct inhalation of transported air pollutants for which standards exist (i.e., particulate matter, photochemical oxidants, sulphur oxides, and nitrogen oxides), no adverse human health effects are anticipated, providing the ambient air quality standards are not exceeded (see Table 5-2). However, in regions where transboundary air pollution contributes to the violation of the standard, health related problems cannot be ruled out. Although some concern has been expressed over the effects of acid sulphates on mortality/morbidity, the available data appear insufficient to single out this species as the sole pollutant of ------- 1-18 concern in the sulphur-particulate complex. As with the gaseous pollutants, the long-range transport of particulate matter should only be viewed as a concern when violation of the ambient air quality standards occur. I 1 I 1.4.2 Visibility I Effects of transboundary air pollution on visibility are related to fine particle air quality and only indirectly to acidic deposition. • The major precursors of acid deposition that can significantly affect • visibility are sulphuric acid and various ammonium sulphate aerosols. These form a large fraction of the fine particle loadings that • dominate visibility impairment from anthropogenic sources. Available V data do not suggest that nitrates (predominantly in the vapour phase) play a significant role in impairment of visibility, but visible _ brown plumes from NC>2 have been reported at a distance of 100 km • from a few isolated point sources. ^ From available information on background and incremental fine fl particle loadings and relative humidity, estimates of visibility • impacts (reduction in visual range and contrast, discolouration from haze or plumes) can be made. Analysis of airport data indicate a • substantial decline in regional summertime visibility in the eastern • U.S. and portions of southern and eastern Canada between 1950 and 1975, with stable or small improving trends since that time. These — changes may be associated with changes in the level and distribution • patterns of sulphur oxide emissions. ™ Areas such as those found in western North America, are the most • sensitive to visibility degradation. Usually, good visibility is f| valued most highly in natural settings such as parks and wilderness areas. Any area, however, with normal viewing distances of a mile or M more may be affected by episodic regional haze carrying acid • precursor substances. Studies of the value of visibility and public perception indicate that the public cares about visibility and is willing to pay for maintaining or improving it. Accurate economic • assessments are not, however, available for eastern North America. • 1.5 MAN-MADE STRUCTURES '| Certain airborne chemicals can accelerate deterioration of materials. « There is evidence that materials in urban areas of Europe and North I America have suffered and are suffering from exposure to these ' pollutants. Materials at risk include statuary and structures of cultural value as well as commonly used construction materials. In • the present discussion, exterior surfaces are the focus of interest. m It is reasonable to assume that acidic deposition due to long-range • transport and transformation of air pollutants contributes somewhat • to material effects. Current understanding of material decay I I ------- I I I I I I J I I I I I I I I I I I I 1-19 processes leads to the tentative conclusion that local sources of corrosive pollution mask the effects resulting from long-range transport of acidic deposition. The principal findings of the Work Group are: The majority of sensitive materials tend to be located in urban/suburban areas. However, materials at risk cannot be assumed to be proportional to population density. Relationships between concentration of corrosive gases and damage are better documented than relationships between acidic precipitation or particulates and deterioration. The main groups of materials which are damaged by outdoor air pollutants are: metals, coatings and masonry. The pollutants are delivered to the surfaces in wet and dry form. It is generally accepted that 862 is the primary species causing damage to materials. The importance of nitrogen compounds is closely related to its particular species and may increase with the predicted increases in NOX emissions relative to S02 emissions. Chemical degradation processes include deterioration of calcareous building materials by the removal of calcium carbonate through conversion to calcium sulphate and the removal of protective corrosion products on metals, particularly zinc and copper. Mechanical deterioration of masonry occurs when calcium sulphate enters the porous material and causes internal rupturing due to the pressure of crystallization or hydration. Regional field studies, chamber tests and atmospheric corrosion sites have indicated the nature and extent of accelerated corrosion associated with metal-pollutant interactions. Dose-response relations have been determined for SC>2 and low-carbon steel and zinc. In some areas of eastern North America, urban centres have experienced extensive and significant deterioration of zinc coverings. Common materials of construction at risk include, limestone, carbon steel and galvanized steel sheet. Carbon steels must be coated in order to provide useful service life and, thus the coating becomes the material at risk. Dose-response relations have been determined for sulphur dioxide and ozone for some paints and coatings. In some urban centres, ozone can have a significant impact on the durability of elastomers. ------- 1-20 I I For porous materials such as masonry, the long-term accumulation of pollutants is a major concern especially for deterioration associated with sulphate. • Materials at risk and some active corrosion agents have been identified in numerous field and laboratory tests. Confidence in dose-response relationships is weakened in some cases because • of incomplete monitoring of air quality and meteorological ™ parameters in field tests. I 1.6 METHODOLOGIES FOR ESTIMATING ECONOMIC BENEFITS OF CONTROL Traditionally, the decision-making process has required an fl appreciation of the costs and benefits associated with following a prescribed set of actions. Basic to this process has been the transformation of the implications of these actions, (i.e., • converting changes in crop yield and fish catches, into comparable w units of measurement). Monetary units are widely accepted as providing comparable weighting units for individual variables. In 4 order to provide the Canada/United States Coordinating Committee with f guidance in this important area, the Work Group has undertaken a review of the methodologies available for assessing the economic « benefits of controlling long-range transport of air pollution. • The following are the conclusions of the Work Group: A number of methodologies have been reviewed but presently the • basic conclusion of this effort is that application of available approaches for conducting a benefit/cost analysis must either • omit real but intangible benefits or include a wide uncertainty B range. Despite these real limitations, these methodologies can provide a useful estimate of benefits for some sectors. — There are several techniques which can be applied to determine * the primary economic benefits associated with a particular receptor category recognizing that option and legacy values are II not captured. However, the lack of data on dose-response • relationships limits the application of most of these techniques at this time. For some sectors, differences in producers' • income may provide benefit estimates even in the absence of • explicit dose-response data. The value of the secondary benefits can be estimated for • specific economic sectors and regions, to derive a partial * estimate of the impacts in various geographical areas. It is evident that more economic research is required. Economic '| techniques have yet to be rigorously tested in some sectors, such as historical value, and are limited in their treatment of • option and legacy values, and in dealing with the issues of • property rights. I I ------- I I I I I f I I I I I I I I I I I I I 1-21 The initial design of future research efforts to document the effects of acidic deposition should reflect the data require- ments for an economic benefit estimate. Interdisciplinary cooperation at the design stage is the best way to ensure results which are amenable to economic analysis. 1.7 NATURAL AND MATERIAL RESOURCE INVENTORY 1.7.1 Introduction A natural and material resource inventory is a necessary component of an assessment of the benefits of emission reductions. Consequently, the Work Group attempted to compile an inventory for aquatic, terrestrial and man-made resources. In all cases, the sectoral inventories are incomplete and sometimes lacking in sufficient detail. For example, not only does the aquatic inventory not include an accurate accounting of lakes and streams with their associated alkalinity, but it also does not include a consideration of the population size and diversity of aquatic organisms depending on the maintenance of a stable aquatic environ- ment. Similarly the terrestrial inventory has been limited to only a consideration of hardwoods and softwoods because a comprehensive inventory at the species level is presently lacking. The inventory has been established on the basis of sulphate depo- sition regimes coincident with the location of terrestrial features such as soils and bedrock which have a limited capacity to reduce the impact of acidic deposition on aquatic regimes. In no cases were there sufficient data to indicate which particular resources are being damaged by acidic deposition. Thus, this inventory is a categorization of resources potentially at risk, rather than a list of resources now adversely affected by acidic deposition. The completion of this inventory has served to underline the considerable weakness which exists in our ability to adequately quantify the extent of the resource at risk. 1.7.2 Aquatic - United States Approximately 36,000 km^ of the eastern U.S. surface water area (25%) is located in areas of low and moderate potential to reduce acidity (high and moderate sensitivity) and of deposition greater than 20 kg/ha.yr sulphate in precipitation. Only 24% are located in areas with a high potential to reduce acidity (low sensitivity) and of deposition greater than 20 kg/ha.yr sulphate in precipitation. The actual surface water area would be more restricted if data had been available on surface water chemistry (i.e., alkalinity). Additional refinements on the inventory should include data on this ------- 1-22 variable as well as more accurate measurements of surface water area. 1.7.3 Aquatic - Canada 1.7.4 Agriculture - United States 1.7.5 Agriculture - Canada 1.7.7 Forests - Canada I 1 I Approximately 52,000 km2 of surface water area is estimated to be • at risk in areas with deposition exceeding 20 kg/ha.yr. Of this 9 total, about 54% (28,000 km2) is located in areas with a low potential to reduce acidity (high sensitivity). The inventory could • be improved by better data availability on actual surface areas of | waters and kilometres of rivers and streams. Moreover, actual data on aquatic alkalinity and aquatic biota will be required to define ^ more accurately the extent of the resource at risk. • I Major crops in the eastern U.S. (corn, soybeans, hay, wheat, tobacco and potatoes) are grown under varying sulphate deposition regimes. •• Soybeans and tobacco are the only ones, however, with approximately • 20% of their yield grown under sulphate deposition greater than 40 kg/ha.yr. For the other crops, less than 10% of their total yield is _ grown under sulphate deposition greater than 40 kg/ha.yr. • I Many of Canada's most valuable crops are grown in areas of high deposition. These include both grains and vegetables. Importantly, • for 6 of the 12 crop types included in the inventory, more than 50% • of their individual total yields is grown in areas where sulphate deposition exceeds 40 kg/ha.yr. Only 4% or less of each crop is — grown in areas experiencing annual deposition levels of 10-20 • kg/ha.yr sulphate in precipitation. ™ 1.7.6 Forests - United States I The annual forest growth in those states east of the 100° meridian in m 1977 was 476 million m3. Approximately 10% of this combined • hardwood and softwood growth occurs under sulphate deposition regimes greater than 40 kg/ha.yr. Over 75% of the growth occurs under sulphate deposition regimes between 20-40 kg/ha.yr. • I Canadian forest growth occurs in a slightly different pattern than in the U.S. Of the total annual yield of 150 million m3, about 10% of • the hardwood growth is located in the highest deposition area, but • only 1% of the softwood growth and 8% of the mixed growth. I I ------- I I I I I I I I I I I I I I I I I I I 1-23 Approximately 64% of the hardwood and 70% of the mixed growth occurs in the area of moderate deposition, but only 28% of the softwood growth. 1.7.8 Man-Made Materials - United States There is no adequate U.S. inventory of renewable or cultural resources. Past efforts to create an inventory of renewable resources have combined per capita material estimates and census data on population distribution. These per capita estimates have been shown to be very site specific and are not an adequate basis for creating a national inventory. The only inventory prepared by the Work Group is one on historic resources exposed to various levels of ambient sulphur dioxide. 1.7.9 Man-Made Materials - Canada As in the case in the U.S., Canada has no adequate inventory of renewable materials or cultural resources. The historic resources inventory includes historical landmarks, buildings and monuments and parks. The inventory presented here indicates the numbers of each of these which are located in 2 categories of deposition: greater than 40 kg/ha.yr and under 40 kg/ha.yr. Geographically, these resources are located in the area around Quebec City, one of the earliest towns in Canada, and in southwestern Ontario (Windsor-Sarnia). 1.8 LIMING Mitigation of the effects of acidic deposition by adding neutralizing agents to the receptors has been an obvious action to be considered. Limestone is most often used although other chemicals have been tried. The term "liming" has often been used to describe such treatments and in this section will be used to describe artificial neutralization experiments regardless of the chemical or chemicals actually used. Extensive work has been carried out on aquatic systems affected by acidic deposition. However, the application of lime products to aquatic resources will not address the potential for damage to forests or buildings and structures. 1.8.1 Aquatic Systems Liming will not eliminate all problems associated with acidification of surface waters but may be necessary on a limited basis as a means of temporarily mitigating the loss of important aquatic ecosystem components. However, it cannot be used in all situations. ------- 1-24 1.8.2 Terrestrial Liming 1.8.3 Drinking Water Supply Liming techniques have been effectively applied to the treatment of low pH municipal supplies. The per capita costs range from $0.18 to$0.57. I Further, its long-term viability and impact on fish populations needs | additional study. The following observations support this overall conclusion: • Liming can only treat certain aquatic situations, mostly lakes, and must be repeated periodically. It is not practical to fl locate and treat small temporary meltwater pools because of m their large number and widespread occurrences. These pools, however, are an important habitat for amphibians and dependent M wildlife. The technology for reliably treating high discharge I rivers (such as the salmon rivers of the eastern North American coast) is not available. _ Swedish experimental liming programs report some success in * being able to promote the growth and reproduction of fish populations. However, all results to date are from experiments • which have been run for five years or less. The long-term V effectiveness of liming to protect aquatic ecosystems is not known. As a result of liming acidic waters, aluminum poisoning « of salmon and rainbow trout has been encountered. • No experimental data on liming are available for surface waters containing some of the important sport fish species in North • America, such as muskellunge, walleye and bass. " Anthropogenic acidic deposition will alter the original uniqueness of "wilderness" aquatic environments. The additions of neutralizing agents will further modify the character of these ecosystems and will not preserve the "wilderness" nature • of these waters. • I I The liming of forest lands to neutralize potential acidic deposition effects on terrestrial ecosystems has serious limitations. These • include evidence that liming would not prevent direct foliar injury; | that under certain conditions lime additions can disrupt important soil biological relationships and adversely affect forests; and that « the area coverage required would tend to be so large as to be • economically prohibitive. I I I I ------- SECTION 2 INTRODUCTION 1 I I I ------- I I I I I I I i I I I I I I I I I I I 2-1 SECTION 2 INTRODUCTION World attention was drawn to the problem of transboundary air pollutants and their deposition on surface waters in 1972, when Sweden and Norway reported the so-called "acid rain" phenomenon. From these Scandinavian studies, scientists in many other nations became increasingly aware that, because atmospheric dilution does not eliminate waste, there may be effects on "receiving" aquatic ecosystems, caused by the transport and deposition of pollutants. Since 1972, the acidic deposition phenomenon has become recognized in North America, as detailed in many articles by both Canadian and U.S. scientists. 2.1 THE EXTENT OF RESOURCES EXPOSED TO ACIDIC DEPOSITION AND POTENTIAL FOR LARGE-SCALE EFFECTS Acidic deposition is currently being observed over most of eastern North America. The effects on watersheds and aquatic resources are most strongly expressed in the areas where elevated inputs of acid combine with low natural acid neutralizing capacity (ANC) of soils and water to reduce the pH of surface water, leading to effects on aquatic ecosystems. Over most of this area, acidic deposition, sulphate particulates, and oxidants occur together. In addition, there are local exposures occurring to sulphur dioxide, nitrogen oxides and fluorides, with biological uptake and subsequent cycling of these compounds. Although acidic components of acidic deposition remain the focus of this report, due to their important impacts on aquatic/terrestrial ecosystems and on human health and man-made structures, the effects attributable to oxidants are also considered. Hydrogen ion concentration (acid, H+) is a critical factor controlling the rate of most chemical reactions. Processes such as solubilization, corrosion, and mobilization of minerals and metals are accelerated by increasing the acid concentrations in soils and water. Soil weathering and nutrient balances are altered by changes in the acidity of soilwater. Household water supplies from shallow wells, or acidic surface waters, in turn, can be modified by the further mobilization of metals from lead and copper pipes. The hydrogen ion load (mass per area per time) affects the extent of chemical reactions in soils and other materials whereas the concentration affects the rates of reactions. For example, the total amounts of chemical constituents leached from soils annually is more closely related to the annual load on hydrogen ion than to the concentration in any precipitation event. The load is also used in the following ways: (1) in comparison with loads of acid-forming ions to determine the influence of acid neutralizing materials in ------- 2-2 I atmospheric deposition; (2) to simulate the acidity of streams B receiving snowmelt runoff from an accumulated snowpack; and (3) to determine the acidity of lakes which average the inflow from a number _ of runoff events. fl The effect of acidic deposition on watersheds is quite different from one region to another due to differences in climate, soils and B geology. Generally speaking, ecosystems seen as sensitive to acidic B deposition are characterized as having thin soils, low in exchange- able bases and cation exchange capacity, overlying granitic bedrock • (noncalcareous). Figure 2-1 provides a small scale overview of areas B seen as sensitive based on bedrock geology. Efforts are now underway and preliminary results are presented in Sections 3.5 and 4.5 of ^ larger scale mapping of sensitive areas. In the United States, the B four most susceptible regions are the Northeast, the Appalachian ™ Mountains, the Minnesota-Wisconsin-Michigan highlands, and the western mountain areas of Colorado, Oregon, Washington, Idaho and I California. In Canada, sensitive regions include parts of the B Atlantic Provinces and portions of the Precambrian Shield areas of Ontario and Quebec. Other areas may be considered sensitive based on -M soil characteristics or other variables. B What is known of the complexity and geographic range of acidic deposition ecosystem interaction and long distance transport of air B pollutants pose a significant dilemma for federal, state and B provincial regulatory agencies. Aquatic life is apparently being damaged by regional air emissions, but air quality standards were not • designed to protect water quality. Nevertheless, important resources B over a large part of the continent appear to be at risk and new multinational control approaches may be required. ^ 2.1.1 Methods of Measuring Effects Lakes, rivers, and watersheds act as "collectors" of atmospheric B pollution. Therefore, one research approach has been to study lakes and watersheds as large-scale "calibrated" collectors since the • surface environment experiences a total loading that is an | integration of all deposition processes. This approach has led to establishing "calibrated watersheds" as monitoring sites which are « combinations of streams, lakes, and plant communities under intensive B measurement. In these watersheds, hydrologic weirs are set up in streams entering and leaving small study lakes or settling pools. The flows of water and dissolved substances are measured upon B entering and leaving the lake, and these data are combined with B measures of atmospheric inputs and water loss by evaporation, to calculate "substance budgets". The difference between the inputs • measured by the budgets and inputs measured from wet deposition £ monitoring can provide a preliminary estimate of dry and gaseous deposition. . Detailed sampling of biota within such a watershed, together with ™ chemical data, allow an assessment of the chemical and biological B I ------- I I I I I I I I I I I I I I I I I I I 2-3 Location of Ten Calibrated Watersheds: 1. Experimental Lakes 2. Boundary Waters Canoe Area 3. Northern Highlands, Wl 4. Saulte Ste. Marie 5. Dorset 6. Sagamore Lake 7. Hubbard Brook 8. Laurentide 9. Kejimkujik Park 10. Coweeta Figure 2-1. Regions of North America containing lakes that may be sensitive to acidification by acidic deposition, based on bedrock geology, showing where calibrated watershed studies on sensitive areas are in progress (modified from Galloway and Cowling 1978). ------- 2-4 2.1.2 Hydrologic Cycle I I effects that air pollution can have on the system. Calibrated watersheds have been established at a number of locations, such as Kenora, Sault Ste. Marie, and Dorset in Ontario; Laurentide Park, • Quebec; Kejimkujik Park, Nova Scotia; Hubbard Brook, New Hampshire; f Coweeta, North Carolina; and Sagamore Lake, New York (Figure 2-1). I Although the hydrologic cycle seems to be well-known, questions do fl emerge as to the potential for high evaporation/precipitation ratios V to concentrate sulphate, and the availability of water for many of the acid-forming reactions, as well as for the wet deposition and M soil flux processes. For example, in low-humidity regions, or during • drought periods, long-distance gaseous transport of SC>2 may provide a greater fraction of the deposition than in wet regions. Similarly, conditions of low rainfall and high evaporation, or seasonal droughts • will alter the soil solution flux processes and associated reactions. w In regions where annual precipitation is less than potential annual evaporation, movement of dissolved ions is upward (calcification). M This movement of bases would tend to neutralize acidic deposition | falling onto soil surfaces. Indeed, in regions where potential evapotranspiration approaches total rainfall, flushing of H+ or _ 50^2- becomes limited to short-season processes or those that • occur only every few years. Because of the evidence that in many poorly-buffered northern soils, I the sulphate ion is a relatively conservative substance (Harvey • et al. 1981), high rates of evaporation can leave the precipitation sulphate concentrated in the soil solution (and lake water) by a «| factor controlled by the evaporative losses. The equations for lake • sulphate concentration developed by Henriksen (1980), show this factor plus dry deposition to be 1.9 for central Norway. Regions of proportionately high evaporative losses are expected to have higher • observed sulphate concentrations in lake water than are predicted by ™ the Henriksen equations for a given atmospheric loading rate (Glass and Brydges 1981). These processes vary with precipitation and • temperature patterns between regions, from one watershed to the | next, and from areas having strong topography. Thus, local processes governing the hydrologic balance need to be • considered as a part of the surface water acidification process. Knowledge of the periodicity of atmospheric cycles, and of the geographic patterns of these transport processes and precipitation is • essential to understanding what happens over long periods to • sensitive aquatic systems, as well as when and where it will happen. I I I I ------- I I I I I 1 I I I I I I I I I I I I I 2-5 2.2 ATMOSPHERIC INPUT, TRANSPORT AND DEPOSITION OF POLLUTANTS 2.2.1 Emissions of Pollutants to the Atmosphere Extensive research attributes most of the acidity in rainfall in eastern North America and elsewhere to the presence of sulphuric and nitric acids. These acids are formed by a complex series of chemical and physical processes during and subsequent to the burning of fossil fuels, ore smelting, and petroleum refining. Vehicular transportation, construction, agriculture, municipal incineration uses of home wood burning stoves and natural processes also contribute to the atmospheric burden. Other substances are also emitted to the atmosphere during these processes. Prevailing weather conditions in eastern North America foster the large-scale movement of pollutants within and between Canada and the United States, so that the movements of pollutants are regional issues. Current emissions in the United States and Canada have been estimated by Work Group 3B (Table 2-1). Substantial increases in these emissions are expected if consumption of fossil fuels continues to increase. Estimated emission rates for other constituents from the burning of coal are presented in Table 2-2. A recent U.S. National Academy of Sciences report (NAS 1978) further estimated that from 25 to 30% of the present day atmospheric mercury burden is due to man-made emissions. Much study is presently being devoted to the characterization of emissions from both natural and anthropogenic sources. Table 2-3 presents a comparison of these sources for several gases. Rasmussen et al. (1975) estimated that greater than 90% of the global anthropogenic S02 is emitted from the Northern Hemisphere. It is evident that the natural sources of many gases far exceed the man-made sources on a global basis. However, because such natural gases are usually well distributed throughout the atmosphere, their concentration, known as the background concentration, is extremely low. Anthropogenic sources of many pollutants are centered near urban complexes and, therefore, their local pollutant concentrations are higher and may pose major threats to the urban environment. This spatial concentration of pollutant emission sources causes many atmospheric constituents to exceed their natural levels several fold. 2.2.2 Atmospheric Transport of Pollutants The fate of a pollutant once emitted into the atmosphere depends on several factors, some meteorological and some a function of the pollutants themselves. It is important to have information about these factors since sensitive receptor areas are often located at considerable distances from the pollutant source regions. ------- 2-6 TABLE 2-1. CURRENT EMISSIONS IN THE U.S. AND CANADA (106 Tons) U.S.A. (1980 Estimated) CANADA 1979a TOTAL N0x sox N0x sox N0x sox Utilities 6.2 19.5 0.3 0.8 6.5 20.3 Industrial Boilers/ 7.1 7.3 0.6 1.1 7.7 8.4 Process Heaters/ Residential/ Commercial Nonferrous 0.0 2.0 0.0 2.2 0.0 4.2 Smelters Transportation 9.0 .9 1.1 0.1 10.1 1.0 Other - - 0.2 1.1 0.2 1.1 TOTAL 22.3 29.7 2.2 5.3 24.5 35.0 a Inco, Sudbury at 1980 emission rate. From: Canada/United States Work Group 3A Interim Report "Strategies Development & Implementation" Feb. 1981, Ottawa, Ont. 1 1 1 1 I 1 1 1 • " 1 1 1 1 1 1 1 1 1 ------- I I I I I I I I I I I I I I I I I I I 2-7 TABLE 2-2. AIR EMISSIONS FROM A TYPICAL 1000 MW COAL-FIRED STEAM PLANT3 Const Ituent Ash Total carton Total su 1 phur Water Total nitrogen Al Ca Cl Fe K Mg Na SI Tl Organic C Fluoranthane Benzo(gnl Jperylene Benzo(a)pyrene Benzo(a)pyrene Pyrene Perylee Phenanthrene Corenene Ag Au As B Be Br Cd Co Cr Cu F Ga Ga Hg LI Mn Mo Nl P Pb Ra Rb Sb Se Sn Sr Ta Te Th Tl U V W Zn Zr Mean concentration 11.4 70.3 3.3 9. 1.3 1.3 0.77 0.14 0.9 0.16 0.05 0.05 2.49 0.07 - - - - - - - - - 0.1 0.001 14. 102. 1.6 15.4 2.5 9.6 13.8 15.2 61. 3.1 6.6 0.2 9. 49.4 7.5 21.1 71.1 34.8 40. 1.3 2.1 4.8 34. 0.16 50. 3.1 680. 5. 32.7 3. 272. 72.5 In coalb (*> (*> (?) <*) <*) <*> «) (I) <*> (*) (%) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) (ppm) Annual air emission (kg/yr) 2.5 x 106 1010 (as C02) 107 (as CO) 65 x 106 450 x 106 106 (as NO*) 0.24 x 106 0.14 x 106 7 x 10^ (mostly vapor) 0.31 x 106 0.062 x 106 0.014 x 106 0.02 x 106 0.54 x 106 0.034 x 106 5,000. 35. 14. 13. 7. 13. 6. 3. 0.6 31. 0.3 3,500. 3,100. 80 70,000. (mostly vapor) 680. 640. 1,700. 915. 15,000. (40* as vapor) 172. 1,600. 1,000. 365. 1,500. 940. 1,300. 2,700. 1 1,000. 0.1 (Cl) 1,200. 360. 335. (20$vapor) 1,200. 1,100. 6.5 2,600. 96. 33,000. 250. 3,400. 90. 37,000. 2,200. a Plant has electrostatic preclpltator efficiency of 99.5<, no scrubbers, and consumes 5 x 10 tons of coal per year. b Western, midwest and eastern coal mean of 101 samples. From: A.W. Andren personal communication; Bauer et al. 1982a, 1982b; EPRI 1980; Klein et al. 1975; NAS 1977; ORNL 1977. ------- 2-8 8 CO 51 9. — — 10 s-° L. V, 3 1 4- (A UJ U C O Q. O 4- c (0 8 * L. 4- g 3 to O (0 o c 1 £ 4- C n j= +- "o 0. A 0) ID i_ in * in ? o tj = = WQO 2 t ^ Z « * S 8 !!* * •*- o o g1 s § ? _ -o ^ _ ? S Q. — O > S o 5 8 to to IB to T T •— in 0* o> o. 2 o — X X CO CM O 1 "o CTi — 0 X X S £ (8 10 I « o S >• to •Q fl (O "o £ 5 E 5 4- 4- (A .C (A 34-3 .0 L £) ^ — O Ul *•» J_ CM S £ Oxidatio o o o X o o o X - o "o m (A 8 8 £ — O 5 S is ,. in c O -1 ex 3 — 4- Chemlca Sewage t/> x" o in » 1 O 6 (A t = £ « ut — 4- <= S 5 . % 8 Si § | if | £. CL "O 0. w « 1 i o o X o CT* in 9 i 8 % 3 I" O m 9 I o CM •z. in o iN £ Oxidatio • X 0* ON O X 0* O X rO | 4- 10 C — — o c S O (A 4- "0 O S i z 4- -O — O c «J .C 10 — o. m v> o c .0. 4- VI 3 1 o o z in ± ? Ol c 1 reactions; nitrate; Scave 3 £ E c Photochei Oxidatio * CM 1 CM i c ~ ° •2 § «i •c ° ID .+ 8 i CD in | Combust i" r- Q c h S02; Oxidatio anglng 4- > — +. 4- a tr ~c °. ON O X o X s 5 a o o> o o .* 9 N — 9 U- 4- i i C 4- = S 1 * J> « 1 1 oxidation (0 c Photochei - ON 0 X ON O X r- c y> U) 8 k n u — « s1$ 54 "o 4^ "O 3 C 5 £ X 4- (9 VI o 5 4- E < O 0) > (A t £ 1 9 >• n i f-» (D £ ~ £ f- **" "~ * n — *D » a c 4- CM « 9 P-- 4- f^ CO C?» 9 >• W O> Ol 4- * — — O 9 « 0 ~~ 01 E ~~ S£ « | C « A ^ JO ^ -O O 0 r-- o o* 0* *~ in *~ o f- (A (ft O\ c tn oo *. — c P-- — O to 4- -o c r- w •0 O) — 4- c r- -a 9 S §2 § ? +. I/) (A — — (_ 3 C 9 t_ -0 ffl e — £ a> TS * tn .0 -a "b o 9 « O O O 2S. 4- Q; CK 01 to to to 10 A .O J3 .O ^ § U. I I I I I I I I I I I I I I I I I I ------- I I I I I I I I 1 I I I I I I I I I I 2-9 Several types of meteorological factors influence long distance transport. The prevailing wind regime over much of the eastern U.S. and Canada is one of westerly winds. This pattern is complicated by seasonal trends, in that there is a southerly component in the summer and a northerly component in the winter (Figures 2-2 a and b). "Long-range transport" is facilitated by tall stacks, high wind, and a stable lower atmosphere (i.e., where the temperature increases with altitude). Absence of precipitation also increases the distance of transport. Figures 2-3 a and b show isopleths of precipitation for the North American continent. The numbers on the contours represent the average number of centimetres of water falling on the land during two periods, warm and cool, over 12 months. The amount of precipitation in any particular locality usually varies from year to year, but over a long period its average is fairly constant. The precipitation patterns shown in Figures 2-3 a and b partially govern the removal processes of pollutants from the atmosphere. The properties of the pollutants also will determine their ultimate fate in the atmosphere. Junge (1977) has argued that atmospheric constituents may be put into three categories, each describing the fate of a set of compounds: (1) accumulative gases; (2) gases determined by chemical or physio-chemical equilibria with the earth's surface; and (3) gases and particulate matter (aerosols) determined by steady-state conditions of their cycles. The third category comprises most trace gases and particulate matter and is the prime concern of this report. The atmospheric concen- tration of these constituents is determined by dynamic processes between sources and sinks. The average atmospheric residence or turnover time varies between constituents. A large percentage of compounds with relatively short residence times (a few days) are deposited within tens to hundreds of kilometres from the point of emission. Compounds with longer residence times may travel thousands of kilometres. Galloway and Whelpdale (1970) estimate, for example, that some two-thirds of sulphur emissions in eastern North America are deposited there, the remainder being transported out over the Atlantic Ocean. Table 2-3 presents typical residence times for other selected parameters. 2.2.3 Atmospheric Removal Processes Substances transported through the atmosphere are removed via wet and dry processes. There are presently a number of deposition models, both empirical and theoretical, which may be used in delineating pollutant deposition patterns. The suitability of these models depends on the time and space scales of the transport processes under consideration and the complexity of the chemicals of interest. The transport process models require knowledge of the chemical and physical characteristics of the airsheds involved, for example, reaction rates under ambient conditions; the concentration in the ------- 2-10 Figure 2-2. Wind patterns for North America based on surface stream-lines for (a) January and (b) July (Bryson and Hare 1974). I I I I I I I I I I I I I I I I I I I ------- I I I I I 1 I I I I I I I I I I I I I 2-11 Total Precipitation (cm) Apr-Oct 1979 so Total Precipitation (cm) Nov 1979-Mar 1980 33 Figure 2-3. Seasonal precipitation for North America patterns, total precipitation as water depth (cm), shown for (a) "Summer" April - October 1979, and (b) "Winter" November 1979 - March 1980. Data reporting sites (A) are from NADP and CANSAP precipitation monitoring networks (Glass and Brydges 1982). ------- 2-12 I 1 solid, vapour, and liquid states; size distributions, morphology, and sorptive characteristics of aerosols; and the vertical, aerial, and temporal variability of these parameters. • Particles and unreactive gases in air may be removed by rainout (in-cloud processes) and washout (below-cloud removal). The wet flux • of these substances is a function of their concentration in • precipitation and the amount of precipitation. The particle washout ratio has been completed by several authors for different chemicals ^ (Slinn et al. 1978). Slinn et al. (1978) have also summarized data • on enhanced solubility coefficients for reactive gases. These gases ™ include SC^, where the dissolution, hydrolysis and oxidation to sulphuric acid are considered. fl| Accurate and direct measurements of dry deposition, both for aerosols and gases, are not possible at present (Hicks and Williams 1980). • The mass transfer is especially difficult to estimate for trace • chemicals because long sampling times are required (often greater than 24 hours) and meteorological conditions may change drastically during such a sampling interval. Dry flux estimates will undoubtedly • change in the future as deposition measurement techniques and models • improve. At the present time, it seems that the best experimental strategy is to collect accurate data for atmospheric constituents • with the best possible time resolution, at an appropriate reference £, height, and with as much meteorological information as possible. Several approaches are available for indirectly calculating mass • transfer of aerosols to the earth's surface. The most popular approach has been to use the relation by Chamberlain (1966): F = VDCZ (1) I where F = flux, VQ = deposition velocity, and Cz = pollutant • concentration at a certain reference height. Deposition velocity • data, determined by wind tunnel experiments for several particle diameters, roughness lengths, and friction velocities, have been furnished by Sehmel and Sutter (1974), Cawse (1974) and Holler and • Schumann (1970). The data, which have been summarized by Gatz • (1974), represent time-averaged deposition velocities for a variety of meteorological conditions and thus do not necessarily give • realistic values for aerosol depositions to water. Sievering et al. ^jf (1979) has used the profile method for estimating fluxes across the air/water interface. Hicks and Williams (1980) have proposed a new « spray capture model, indicating that very little (if any) transport • is possible during calm conditions. Slinn (1980) has proposed a more * sophisticated resistance model, where aerosol growth in the surface layer is included. Sehmel and Hodgson (1974) have presented a model fl based on dimensionless integral mass transfer resistances. Surface V integral resistances were evaluated with deposition velocities of monodispersed aerosols determined in wind tunnel experiments. • I I ------- I I I I I I I I I I I I I I I I I I I 2-13 Similar models are also available for gaseous deposition to various surfaces. Garland (1980), Gramat (1980) and Liss and Slater (1974) have devised models based on resistance of transfer to various surfaces, such as grass, snow, water, and forest canopies. These models usually include an aerodynamic, stagnant film, and stomatal (for vegetation) resistance. The same caveats are necessary on these models as are applied to dry particle deposition. The relative importance of each process (i.e., dry vs. wet deposition), is still being evaluated. Although the results are not fully conclusive, modelling and mean balance studies indicate wet and dry deposition of sulphur compounds are of equal importance in north Europe and North America (Fowler 1980; Haines et al. 1981). Dry deposition seems to be of lesser importance in remote areas. Harvey et al. (1981) conclude that dry deposition is relatively more important than wet deposition in areas like the Ohio Valley, whereas the opposite is true in remote Canadian Shield lakes. 2.2.4 Alteration of Precipitation Quality The seasonal quantity and quality of precipitation are important for determining the potential for acidic deposition impacts on the environment. Acid pollutants accumulating in the snowpack have a higher potential for causing deleterious effects on organisms and habitats in areas with higher amounts of snowfall than in areas with lower amounts of snow accumulation. This is due to the rapid flushing of accumulated acid during snowmelt. Large storms, on the other hand, tend not to have as low a pH for the entire rainfall as do light rains. Thus, the distribution of precipitation during the year, the temporal behaviour of rainfall, and the location of pollution sources within rainfall pathways are linked to the potential for damage to the aquatic ecosystems. In addition, many areas in the east with the greatest annual precipitation have the least buffering capacity in soils and waterways. Distilled water in equilibrium with atmospheric carbon dioxide has a pH value of about 5.6. Results of CANSAP + NADP monitoring presented in Figure 2-4 show large areas of North America which are receiving precipitation with a pH less than 5.6. This results in elevated concentration and deposition of acids to the surface as shown in Figures 2-5 a and b. All precipitation contains a wide variety of chemical constituents from sources such as sea spray, dust particles and the natural cycling of carbon, nitrogen and sulphur. The discharge of wastes to the atmosphere increases the amounts of compounds containing elements such as nitrogen, carbon and sulphur, and adds to the variety of compounds, such as PCBs, PAHs and heavy metals, which are found in rainfall. The four ions usually of most importance to rainfall acidity are: hydrogen (H+), ammonium (NH^+), nitrate (N03~) and sulphate (S0^^~). Other ions (e.g., calcium) may be important ------- 2-14 V ,1980 pH Precipitation amount - weighted mean annual pH in North America for the calendar year 1980. Figure 2-4. Legend Canada United States •CANSAP »NADP QAPN DMAP3S ------- I I 2-15 Legend Canada United States •CANSAP «NADP • APN "MAP3S »OME 1980 CHH Figure 2-5a. Precipitation amount - weighted mean hydrogen ion concentration in 1980 ( pmoles per litre). ------- 2-16 0.01 kg/ha =1 m mole/m2 Precipitation amount- weighted mean hydrogen ion deposition for 1980 (m moles per square metre). Figure 2-5b. Legend Canada United States •CANSAP «NADP OAPN QMAP3S *OME I ------- I I I I I I I I I I I I I I I I I I I 2-17 under some conditions. Some of the nitrogen and sulphur-containing pollutants are oxidized to nitric and sulphuric acids, so that the acid content of precipitation is mainly a secondary result of the primary emissions. Table 2-4 lists the concentrations of these four major ions in bulk precipitation and total bulk deposition for various sites in North America. Precipitation at pH 5.6 has a hydrogen ion content of about 2.5 ueq/L (microequivalents/litre). It is evident that the most westerly study area, the Experimental Lakes Area, has an acid concentration of about 4 times this value, while Dorset and Hubbard Brook are about 30 times this value. Sulphate is the dominant anion in terms of eq/L (equivalents/litre). In the wet precipitation at Kejimkujik National Park, Nova Scotia, the most easterly study area, the pH is about 4.6, while sulphate is the second highest anion, surpassed by chloride (41 yeq/L), which is a reflection of the strong maritime influence on the precipitation in Nova Scotia. Figures 2-6, 2-7 and 2-8 illustrate the concentration and deposition patterns of sulphate, ammonium and nitrate ions, respectively. Both sulphate and nitrate ion concentrations are highest in the east with high values also recorded in southern Alberta and Saskatchewan. Table 2-5 defines the conversion factors for ion deposition and concentration. The percent of normal precipitation for 1980 is shown in Figure 2-9. While most areas received at least 75% of the normal precipitation, others received up to twice as much precipitation as normal. ------- 2-18 TABLE 2-4. CONCENTRATIONS IN BULK DEPOSITION AND TOTAL BULK DEPOSITION OF FOUR IONS AT FOUR CALIBRATED WATERSHED STUDIES (concentrations in bulk deposition yeq/L) Muskoka- Hubbard Kejimkujik3 Sagamore ELAb Haliburtonc Brookd Park6 Lakef H+ NH4+-N N03~-N so42~ Deposition H+ NH4+-N N03~-N so4 - 11 70-90 21 34-36 18.5 36-41 30 77-89 in meq/m^.yr from 10 55-58 22-28 25-34 20.7 62-64 72-74 12.2 23.7 60.3 bulk deposition 96 16 30.6 79 24 4.4 12.4 33(28.5) 34 80-95 6 20-26 17 37-50 46(40) 81-95 a Wet deposition only, ( ) indicating excess sulphate. b Schindler et al. 1976 c Scheider et al. 1979 d Likens et al. 1977 6 Kerekes 1980 Johannes and Altwicker 1980 NOTE: differences in values for areas, compared to the isoplot figures are due to year to year variations. I I I I I I I I I I I I I I I I I I I ------- I I 2-19 Legend Canada United States •CANSAP «NADP • APN "MAP3S • OME 1980 Figure 2-6a. Precipitation amount - weighted mean sulphate ion concentration for 1980 ( ymoles per litre). ------- 2-20 • 3.4 • 7.8 • 51 • 10 • 16 »3.8 .46 |»49 • 48 * 5° • 45 • 22 \ \ 27 • 0.961 kg/ha =1 m mole/m2 • 33 20 1980 D SO4= Precipitation amount - weighted mean sulphate ion deposition for 1980 (m moles per square metre). Figure 2-6b. Legend Canada United States •CANSAP «NADP QAPN DMAP3S *OME I I I I ------- 2-21 Legend Canada United States •CANSAP »NADP • APN BMAP3S »OME 1980 CNH4+ Figure 2-7a. Precipitation amount - weighted mean ammonium ion concentration for 1980 ( ymoles per litre). ------- 2-22 \J c 0 \ f f? 0.18 kg/ha =1 m mole/m2 Precipitation amount - weighted mean ammonium ion deposition for 1980 (m moles per square metre). ure 2-7b. Legend Canada United S • CANSAP »NADP nAPN QMAP3S *OME ------- 2-23 Legend Canada United States •CANSAP »NADP • APN "MAP3S »OME 1980 Figure 2-8a. Precipitation amount - weighted mean nitrate ion concentration for 1980 ( ymoles per litre). ------- 2-24 0.62 kg/ha = 1 m mole/m2 Precipitation amount - weighted mean nitrate ion deposition for 1980 (m moles per square metre). Figure 2-8b. • Legend Canada United States •CANSAP »NADP QAPN nMAP3S 4OME » I ------- I I I I I I I I I I I I 1 I I I I I I 2-25 TABLE 2-5. CONVERSION FACTORS FOR CONCENTRATION AND DEPOSITION UNITS ION Example for CONCENTRATION mg/L PER pmole/L DEPOSITION kg/ha PER mmole/m2 H+ NH+ Na+ Ca2+ Mg2+ so42~ N03 ci- 0.0010 0.0180 0.0230 0.0401 0.0243 0.0961 0.0620 0.0355 0.010 0.180 0.230 0.401 0.243 0.961 0.620 0.355 2- 0.0961 mg/L equal 1 ymole/L f\ 0.961 kg/ha equal 1 miaole/m ------- 2-26 I I A 1980 % normal precip Figure 2-9. Percent of normal precipitation in North America in 1980. I I I ------- I I I I I I I I I I I I I I I I I I I 2-27 2.3 REFERENCES Andren, A. Personal communication. Water Chemistry Department, University of Wisconsin, Madison, WI. Bauer, C.; Vern Mark, K.; Price, B.; and Andren, A.W. 1982a. Organic vapour emissions from a coal-fired steam plant. Final Report, U.S. Environmental Protection Service, University of Wisconsin, Madison, WI. (in preparation) Bauer, C.; Andren, A.W.; and Knaebe, M. 1982b. Chemical composition of particulate emissions from a coal-fired steam plant; Variation as a function of time and size. Final Report, U.S. Environmental Protection Service, University of Wisconsin, Madison, WI. (in preparation) Bryson, R.A., and Hare, F.K. 1974. In World Surveys of Climatology, ed. H. Landsberg. Vol. VII. The Climates of North America. Elsevier Press. 432 pp. Cawse, P.A. 1974. A survey of atmospheric trace elements in the United Kingdom. A.E.R.E. Hawwell Report No. R-7669, HMSO, London. Chamberlain, A.C. 1966. Transport of gases to and from grass-like surfaces. Proc. R. Soc. Lond. A296:45-70. Electric Power Research Institute (ERPI). 1980. Inventory of organic emissions from fossil fuel combustion for power generation. ERPI Report EA-1394, Palo Alto, CA. Fowler, D. 1980. Removal of sulphur and nitrogen compounds from the atmosphere in rain and by dry deposition. In Proc. Int. Conf. Ecological Impact of Acid Precipitation, eds. D. Drablos and A. Tollan, pp. 22-32. SNSF-Project, Sandefjord, Norway, 1980. Galloway, J.N., and Cowling, E.B. 1978. The effects of precipitation on aquatic and terrestrial ecosystems - a proposed precipitation chemistry network. J. Air Pollut. Control Assoc. 28:229-235. Galloway, J.N., and Whelpdale, D.M. 1980. An atmospheric sulfur budget for eastern North America. Atmos. Environ. 14:409-417. Garland, J.A. 1980. Dry deposition of gaseous pollutants. In Proc. WMO Symp on Long Range Transport of Pollutants and its Relation to General Circulation Including Stratospheric/Tropospheric Exchange Processes. WMO (Geneva), 538:95-103. Glass, G.E., and Brydges, T. 1982. Problem complexity in predicting impacts from altered precipitation chemistry. In Proc. Int. Symp. Acidic Precipitation and Fishery Impact in Northeastern North America. American Fisheries Society, Ithaca, NY., 1981. (in press) ------- 2-28 I I I Goldberg, E.D., and Bertine, K. 1971. Fossil fuel combustion and the major geochemical cycle. Science 173:233-235. Gramat, L. 1980. I Gramat, L.; Rodhe, H.; and Hallberg, R.L., 1976. Global sulfur cycle. In Nitrogen, phosphorus and sulfur global cycle, eds. • V.H. Svensson and R. Soderlund, pp. 89-134. Scope Report 7, • Ecol. Bull. (Stockholm) 22. Raines et al. 1981. Harvey, H.H.; Pierce, R.C.; Dillon, P.J.; Kramer, J.P.; and • Whelpdale, D.M. 1981. Acidification in the Canadian aquatic I environment; scientific criterion for an assessment of the effects of acidic deposition on aquatic ecosystems. Nat. Res. Council Canada Report No. 18475, Ottawa, Ont. 369 pp. • Henriksen, A. 1981. Acidification of freshwaters - a large-scale titration. In Proc. Int. Conf. Ecological Impact of Acid • Precipitation, eds. D. Drablos and A. Tollan, pp. 68-74. SNSF - | Project, Sandefjord, Norway, 1980. Hicks, B.B., and Williams, R.M. 1980. Transfer and deposition of I particles to water surfaces. In ORNL Life Sciences Symposium * Series. (in press) Johannes, A.H., and Altwicker, E.R. 1980. Atmospheric inputs to • three Adirondack lake watersheds. In Proc. Int. Conf. Ecological Impact of Acid Precipitation, eds. D. Drablos and A. • Tollan, pp. 256-257. SNSF - Project, Sandefjord, Norway, 1980. | Junge, C.E. 1972. The cycle of atmospheric gases - natural and _ man-made. Q. J. R. Meteorol. Soc. 98:711-729. • 1974. Residence and variability of tropospheric trace gases. Tellus 26:477-488. • 1977. Basic considerations about trace constitutents in the atmosphere as related to the fate of global pollutants. In M Fate of pollutants in the air and water environments, Part I, • ed. I.H. Suffet. New York: J. Wiley and Sons. Kellogg, W.W.; Cadle, R.D.; Allen, E.R.; Lazrus, A.L.; and I Kartell, E.A. 1972. The sulfur cycle. Science 174:587-596. • Kerekes, J.J. 1980. Preliminary characterization of three lake basins sensitive to acid precipitation in Nova Scotia, Canada. In Proc. Int. Conf. Ecological Impact of Acid Precipitation, eds. D. Drablos and A. Tollan, pp. 232-233. SNSF - Project, Sandefjord, Norway, 1980. I I I I ------- I I I I I I I I I I I I I I I I I I I 2-29 Klein, D.H.; Andren, A.W.; Carter, J.A.; Emery, J.F.; Feldman, C.; Fulkerson, W.; Lyon, W.S.; Ogle, J.; Palmy, Y.; Van Hook, R.I.; and Bolton, N. 1975. Pathways of thirty-seven trace-elements through a coal-fired powerplant. Environ. Sci. Technol. 9:973-979. Likens, G.E.; Bormann, F.H.; Pierce, R.S.; Eaton, J.S.; and Johnson, N.M. 1977. Biogeochemistry of a forested ecosystem. New York: Springer-Verlag. 146 pp. Liss, P.S., and Slater, P.G. 1974. Mechanism and rate of gas transfer across the air-sea interface. In Atmosphere-surface exchange of particulate and gaseous pollutants, pp. 345-368. ERDA Symposium No. 38, Richland, WA. Liu, S.C. 1978. Possible non-urban environmental effects due to carbon monoxide and nitrogen oxide emissions. In Man's impact on the troposphere, eds. Levine and Schryer, pp. 65-80. NASA Ref. Publ. 1022. Moller, U. , and Schumann, G.J. 1970. Mechanisms of transport from the atmosphere to the earth's surface. Geophys. Res. 75:3013-3019. National Academy of Sciences (NAS). 1977. Nitrogen oxides: Medical and biological effects of environmental pollutants. National Academy of Sciences, Washington, DC.333 pp. . 1978. An assessment of mercury of the environment. National Academy of Sciences, National Research Council, Washington, DC. 185 pp. Oak Ridge National Laboratory (ORNL). 1977. Environmental, health, and control aspects of coal-conversion: an information overview. In ORNL-EIS-94, Volume 1, eds. H.M. Braunstein, E.B. Copenhaver, and H.A. Pfuder. Oak Ridge National Laboratory, Oak Ridge, TN. Rasmussen, T.M.; Taheri, M.; and Kabel, R.L. 1975. Global emissions and natural processes for removal of gaseous pollutants. Water, Air, Soil Pollut. 4:33-64. Robinson, G., and Robbins, R.C. 1970. Gaseous nitrogen compound pollutants from urban and natural sources. J. Air Pollut. Control Assoc. 70:305-306. Rodhe, H. 1978. Budgets and turnover time of atmospheric sulfur compounds. Atmos. Environ. 12:671-680. Scheider, W.A.; Snyder, W.R.; and Clark, B. 1979. Deposition of nutrients and ions by precipitation in south-central Ontario. Water, Air, Soil Pollut. 12:171-185. ------- 2-30 I I Schindler, D.W.; Newbury, R.W.; Beaty, K.G.; and Campbell, P. 1976. Natural water and chemical budgets for a small Precambrian lake basin in central Canada. J. Fish. Res. Board Can. • 33:2526-2543. | Sehmel, G.A., and Hodgson, W.H., 1974. Predicted dry deposition « velocities. In Atmosphere-surface exchange of particulate and • gaseous pollutants, pp. 339-419. ERDA Symposium No. 38, Richland, WA. Sehmel, G.A., and Sutter, S.L. 1974. Particle deposition rates on a • water surface as a function of particle diameter and air velocity. J. Rech. Atmos. 8:911-920. • Sievering, H.; Dave, M.; Dolske, D.A.; Hughes, R.L., and McCoy, P. 1979. An experimental study of lake loading by aerosol « transport and dry deposition in the southern Lake Michigan I Basin.EPA-905/4-79-016, EPA Progress Report, U.S.™ Environmental Protection Agency, Governor's State University, Park Forest, IL. • Slinn, W.G.N. 1980. Precipitation scavenging. In Meteorology and power production. U.S. Department of Energy, Washington, DC. • Slinn, W.G.N.; Basse, L.; Hicks, B.B.; Hogan, A.W.; Lai, D.; Liss, P.S.; Munich, K.O.; Sehmel, G.A.; and Vittori, 0. 1978. Some aspects of the transfer of atmospheric trace constituents • past the air/sea interface. Atmos. Environ. 12:2055-2087. I Soderlund, R. , and Svensson, V.H. 1976. The global nitrogen cycle. In Nitrogen, phosphorus and sulfur global cycle, eds. V.H. Svensson and R. Soderlund, pp. 23-74. Scope Report 7, Ecol. Bull. (Stockholm) 22. _ Spedding, D.J., 1972. Sulphur dioxide adsorption by seawater. Atmos. Environ. 6:583-586. I Stewert, R.W.; Hameed, S.; and Pinto, J. 1978. The natural and perturbed troposphere. In Man's impact on the troposphere, eds. Levine and Schryer, pp. 27'-74. NASA Ref. Publ. 1022. • Sze, N.D. 1977. Anthropogenic CO emissions: Implications for the atmospheric CO-OH-CH4 cycle. Science 195:673-675. _ I I I I ------- SECTION 3 AQUATIC EFFECTS ------- I I I I I I I I I I I I I I I I I I I 3-1 SECTION 3 AQUATIC EFFECTS 3.1 INTRODUCTION This assessment is structured to address three major questions concerning aquatic effects of acidic and pollutant deposition in North America: 1. What is the nature and extent of the chemical alteration of the hydrologic cycle due to pollutant deposition? 2. What is the nature and extent of biotic alteration in aquatic ecosystems as a result of acid-induced chemical alterations? 3. What is the geographical distribution and acid-loading tolerance of watersheds of various sensitivities? Several approaches were used to evaluate these questions. Firstly, emphasis was placed on identifying and substantiating historical (long-term) changes in aquatic systems possibly related to long-range transport of acidifying substances. This evaluation has required some consideration of the complexity of hydrologic systems, as well as of the complexity and the extent of aquatic resources that are at risk. Included are detailed documentations of affected aquatic environments, both chemical and biotic components, and definition of time trends for observed changes. Secondly, consideration was given to the significance of the episodic nature of atmospheric pollutant loading and flushing processes, such as snowmelt, as well as the seasonal character of the receiving environments and biota, such as periods of fish spawning. Thus, these sections relate pollutant loading levels to the observed extremes in chemical conditions and biological effects. Finally, this section focuses on the aquatic ecosystems and biota that are sensitive to acidic deposition. It was, therefore, necessary to define an acid-loading tolerance, to identify regions sensitive to acid inputs, to identify aquatic resources at risk from higher acid-loading levels, and to discuss recovery possibilities for aquatic systems showing apparent damage. 3.2 ELEMENT FLUXES AND GEOCHEMICAL ALTERATIONS OF WATERSHEDS For a complete understanding of the effects of acidic deposition on aquatic ecosystems, it is necessary to examine the fate of ions deposited from the atmosphere, directly on aquatic systems and indirectly through deposition on watersheds. In the latter case deposition may result in geochemical alterations of watersheds. These geochemical alterations must be considered before a complete ------- 3-2 I I understanding of chemical inputs to and changes in aquatic ecosystems can be achieved. 3.2.1 Hydrogen Ion (Acid) Hydrogen ions (acid) (H+) drive most chemical weathering reactions. | They are supplied from both external and internal sources. The major external source is acid supplied by atmospheric deposition • (meteorological input). Internal sources are biological and chemical • processes occurring within the watershed. Carbon dioxide (C02) in the atmosphere represents a large, but • I occasionally rate-limited, reservoir of carbonic acid Carbon dioxide contributes to both internal and external sources of hydrogen ions. The major process of chemical weathering is the • exchange of protons (H+) for cations (Ca2+, Mg2+, Na+, and K+) . The | proton source for the weathering process is derived from the external supply (precipitation) and from internal biochemical generation. In _ a typical calcium carbonate-or silicate-bearing soil or rock, this I normal weathering process gives rise to waters having calcium and bicarbonate as the major ionic constituents. (See standard texts on limnology; e.g., Hutchinson [1957] or Wetzel [1975].) • The hydrogen ion cycle within soils is quite complex and not well understood. At the Hubbard Brook Watershed, New Hampshire, the • average external net annual input of hydrogen ion equivalents I observed over the 1963-74 decade was 86.5 +_ 3.3 meq/m^.yr (milli-equivalents/ square metre. year) (Likens et al. 1977b). If this _ were the only source of H+ ions at Hubbard Brook and the ecosystem • were in a steady state, one might expect this hydrogen ion input to ™ be balanced by hydrogen ion exports plus the net rate at which ionic Ca, Mg, K, Na, and Al are leached from the soil. In fact, there are I more of these cations removed from the ecosystem each year than there • are external hydrogen ions to replace them. The difference is statistically significant, and implies the yield of internally • generated H+ and/or an underestimate of dry deposition and/or the I influence of ammonium and nitrate ions on the charge balance. Internal sources of H+ at Hubbard Brook were identified as : _ (1) nitrogen compounds, particularly NH4+; (2) reduced carbon • oxidized in the soil; (3) organic acids, such as citric, tartaric, • tannic, and oxalic acids, produced by biological activity within the soil; (4) oxidation of small amounts of sulphide minerals in the H bedrock; and (5) the uptake of cations (e.g., K+, Ca^+) by the || forest vegetation and the forest floor. Currently, in eastern North America, the amounts of hydrogen ion • being deposited generally are in the range of 50-100 meq/m^.yr for areas receiving the highest acidic deposition. To neutralize this acid input, a base equivalent of 25-50 kg/ha. yr (kilograms/ • hectare. year) of calcium carbonate would be required. Carbonate • soils can neutralize this amount of acid for an indefinite time with I I ------- I I I I I I I I I I I I I I I I I I I 3-3 only a small percentage increase in total runoff of calcium and magnesium salts which is a small loss compared to the total stored in the watershed. However, in areas underlain by rocks resistant to weathering and with shallow noncalcareous soils, such as much of the Precambrian Shield region, the amount of salts and alkaline materials normally leached are on the order of 10-100 meq/m^.yr. External hydrogen ion loadings to these areas are of the same order of magnitude as this leaching rate. When hydrogen ion inputs exceed the levels of available Ca and Mg, other less available metals are leached. For example, some of the acid results in leaching of such cations as aluminum, iron, zinc and manganese. In some cases, hydrogen ion inputs exceed the ability of the soils to fix hydrogen ions and excess hydrogen ions are exported to surface waters. In most parts of the Precambrian Shield, current levels of hydrogen ions from rainfall are neutralized within the soils of the watersheds during most of the year. Retention (neutralization) of hydrogen ions deposited in bulk deposition has been measured at 88, 94 and 98% on an annual basis, at the Experimental Lakes Area (Ontario), Hubbard Brook (New Hampshire) and Muskoka-Haliburton (Ontario), respectively (Schindler et al. 1976; Likens et al. 1977b; Scheider et al. 1979c). On the other hand, hydrogen ions deposited in snow tend to be stripped from snow crystals early in the spring snowmelt process, and much of the total annual H+ export from a watershed occurs during a brief period in the spring. This large volume of water, coupled with less opportunity for infiltration and interaction with the soil, has resulted in some cases in "shock level" concentrations of acid exported to streams and surface waters of lakes (Schofield 1981). Hultberg (1977) reported on such shock level pH declines in Swedish lakes and rivers and demonstrated that in some cases these pH declines were associated with fish kills. The total ionic strength of surface waters is determined largely by the hydrological and geochemical properties of the catchment basin. "Soft" waters, of low ionic strength, occur within basins having chemically resistant and very little readily-exchangeable material, often associated with igneous bedrock or its soil derivatives. "Hard" waters of higher ionic strength are derived from basins having greater amounts of carbonate lithology (see Section 3.5). The amount of cations exported from a basin thus becomes a parameter which, under similar hydrologic and acid-loading conditions, characterizes the basin in an integrated sense. The chemical composition of receiving waters is dependent on the types of weathering reactions within the surrounding watershed. If the weathering has been the result of reactions with C02 and carbonic acid, the major ionic constituents in surface waters will be bicarbonate and calcium. When strong acids such as ^SO^ are introduced (for example, as acidic deposition) into a bicarbonate-weathering system, the generation of bicarbonate alkalinity may be altered (see Section 3.3). Instead of weathering resulting primarily from reactions with carbonic acid and yielding bicarbonate ions as a major end product: ------- 3-4 CaC03 + H2C03 —> Ca + 2HC03 (1) or (K-feldspar) 30g + 2H2C03 + 12H20 —» 2K + 2HC03 + 6H4Si04 (2) m3Si3°10(°H)2 I *•' i OTT/irv f 1 \ ^1 I I the reaction of sulphuric acid with limestone or other rocks yields I sulphate as a major anion: I CaC03 + H2S04 —*• Ca2+ + SO^2" + H20 + C02 (3) • Alternatively, the reaction can be considered as a progressive titration of bicarbonate alkalinity. _ Ca2+ + 2HC03 + H2S04 —> Ca2+ + S042~ + H20 + 2C02 (4) • Alkalinity of waters is a measure of the reserve acid neutralizing • capacity (ANC) that remains to be titrated to any chosen pH level. | Dissolved carbonate species (HC03 and C03^+), if present in sufficient concentrations, react together as a buffering system, M tending to retard or limit changes in pH. (See Figure 3-1 [Wetzel • 1975] for the relationship of the inorganic carbonate species to pH.) For a monoprotic acid [HA]: alkalinity [ANC] = [A~] + [OH~] - [H+] (5) • For a diprotic acid [H2A]: • alkalinity [ANC] = [HA"] + 2[A2~] + [OH~] - [H+], (6) where the acids are HA and ^A, respectively (Stumm and Morgan • 1970). The major source of buffering in freshwaters is the carbonate system. Therefore for surface waters: [ANC] = [HC03~] + 2[C032~] + [OH"] + [B~] - [H+] (7) I where £B~ is the sum of all titrable bases (Lerman 1978). Thus, • the loss of bicarbonate during the ^804 titration represents a | decrease in the buffering capacity of the water (lower alkalinity). As a result of the above reactions, 804^- replaces HC03~ in the ionic • balance of outflow waters until the titration endpoint is reached, ™ that is, when all the HC03~ has been consumed (Kramer 1981). The HC03~ remaining at any stage above the titration endpoint largely I determines the pH or alkalinity of the waters, although organic I I I I ------- I I I I I I I I I I I I I I I I I I I 3-5 8 9 10 11 12 13 Figure 3-1. Relationship between pH and the relative proportions of inorganic carbon species of CC>2 (H2CC>3), HCO^, and C0o~ in solution (from Wetzel 1975). ------- 3-6 I I materials may provide some additional buffering at lower alkalinities (see Section 3.3). Beyond the titration endpoint, large concentrations of hydrogen ion will be present and other buffers such • as aluminum or humic materials may also become important in the J control of pH of the waters (see Section 3.2.4). Thus, the concentration of cations in surface waters for any given alkalinity _ reflects a basin's ability to produce cations and may be used as an I index of its capacity to neutralize acidic deposition added to the ™ basin. Henriksen (1980) and Thompson (1982) have used this assumption and • the necessity of ionic charge balance to estimate surface water sensitivity or the ability of a basin to respond to an external • stress of acidification. Hesslein (1979) has applied similar I assumptions and used alkalinity to estimate acid loadings which would be required to produce acidification. Thus, if arbitrary "loading" _ or acidification stress levels are specified, alkalinity can provide • a quantified measure of the sensitivity of a basin to further • acidification of waters. For example, HCO^ concentrations of 100 to 200 yeq/L, have been identified as approximate levels below H which a basin may be considered to be sensitive to acidification | (Altshuller and McBean 1979; Glass and Loucks 1980). When the flow rates through a basin are specified, the alkalinity provides a flux • or basin yield of reserve ANC. If significant loss of alkalinity has • not occurred this equates to the Ca^+ or cation flux used in the Cation Denudation Rate (CDR) model of Thompson (1982). Alkalinity (concentration or flux) or CDR therefore provide techniques to I estimate quantitatively the capacity of a drainage basin to withstand • acid loading (see Sections 3.9.2 and 3.9.3). Acidification of nonorganic surface waters by external sources of | H may thus be a combination of two processes: (1) a retardation of the development of alkalinity in the watershed (Kahl et al. 1982), • and (2) a titration (Henriksen 1979) of surface water alkalinity. I The Calcite Saturation Index (CSI), was defined by Conroy et al. (1974) as the undersaturation of waters with respect to CaC03. As I modified by Kramer (1981): • CSI = log K - log [Ca2+] -log [HCOZ] -pH • where log K = 2.582 - 0.024t | t - temperature (°C) and [ ] are concentrations. The CSI allows for assessment of pH and alkalinity on a single • logarithmic scale. Saturation with respect to calcium carbonate gives a value of zero with degree of undersaturation on an increasing positive scale. Kramer (1976) considered values greater than CSI =3 • to indicate waters sensitive to acidification. To date, a • quantitative relationship between acidification potential and CSI units has not been developed. I I I ------- I I I I I I I I I I I I I I I I I I I 3-7 3.2.2 Nitrate and Ammonium Ions Atmospheric deposition of nitrate is only about one-third to one-half as great on an equivalent basis as the sulphate deposition in eastern North America, but in some areas of the western United States nitrate may represent up to 60% of the annual acid fractions in rainfall (Lewis and Grant 1979; Liljestrand and Morgan 1978). Nitrogen deposition can result in either acidification or neutralization of surface waters depending on the ionic form. Nitrogen, as nitrate ions (N03~), can be incorporated directly by vegetation resulting in the release of hydroxyl ions (OH~) into the environment (Figure 3-2). The hydroxyl ions neutralize hydrogen ions and raise the pH of the soil and water. Natural decomposition of nitrogenous plant material releases hydrogen ions, but net annual accumulation of plant tissue dominates in most ecosystems (Bormann and Likens 1979). Hence, net production of neutralizing capacity from nitrate addition is often dominant, especially during warm periods of the seasons (see data from Harvey et al. 1981; Brewer and Goldman 1976). This is particularly significant where forest harvest rather than decomposition removes plant materials, because the neutralizing portion of the cycle is left in the system and a portion of the acidification source (decomposition) is removed. Ammonium salts and sulphate particulates are present in both dry and wet deposition. Ammonium is a source of hydrogen ions (Figure 3-2) when the nitrogen is utilized by plants. This release of hydrogen ions can be a significant source of acidification in soils and surface waters. Nitrogen is usually in short supply in terrestrial habitats, and is readily incorporated and retained by ecosystems (Reuss 1976) (Table 3-1). Nitric acid and ammonium salts are stored in snowpack and released as acid components to streams and lakes during spring snowmelt and may, therefore, be partially responsible for the documented episodic increase in acidity in aquatic ecosystems. During the growing season, however, both terrestrial and aquatic vegetation use most of the deposited nitrate and ammonium ions, except for periods of heavy rainfall. Because nitrate ions often occur at higher concentrations in precipitation than do ammonium ions, there is often a net production of alkalinity. 3.2.3 Sulphate Sulphur, like nitrogen, is an essential plant nutrient and the incorporation of sulphate into vegetation releases hydroxyl ions (Figure 3-3). As opposed to nitrogen, sulphur in soil is usually in adequate supply for plant growth. Additions of sulphur may not be entirely incorporated into living tissue. Sulphate ions can also be absorbed by soils and reduced by bacterial action. This reaction consumes acid and raises the pH of the soil-water environment. ------- 3-8 PLANT NH« r R-C —I NH/ «- NO, F SOIL OR i ? organic nitrogen r WATER T* R r R -k NH H + i H* * 3^ ^ kj| 2H + -J > OH 2H + j * .--^ , k Mr i«~ R organic nitrogen Figure 3-2. Simplified nitrogen cycle showing chemical changes caused by plant and soil processes (from Reuss 1976). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-9 TABLE 3-1. THE RETENTION OF NITRATE, AMMONIUM ION AND TOTAL NITROGEN BY FORESTED WATERSHEDS IN SEVEN CALIBRATED WATERSHED STUDIES % Retention in the watershed on an annual basis Substance ELAa Muskoka- Hubbard Kejimkuiik Sagamore Woods Panther Haliburtonb Brookc Park3 Lake6 Lake6 Lake6 Ontario Ontario New Nova Scotia New York New York New York Hampshire N03~ - 75 15 99 43 70 15 95 89 98 90 90 90 Total Nitrogen 81-90 a Schindler et al. 1976. b Scheider et al. 1979c. c Likens et al. 1977b. d Kerekes 1980. e Galloway et al. 1980 (figures estimated from published bar graphs). ------- 3-10 — c 1 SOIL S 2+ P^ *> ^^\ A ^2H * / 2OH~^ H 2H + aerobic ^^^ N ^- k o ^ r VJ 2H + 0 ^ ± 2« 4 anaerobic 2- f% P4 2- /^ °4 Figure 3-3. Simplified sulphur cycle showing chemical changes caused by plant and soil processes (modified from Reuss 1976). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-11 Sulphide (S^ ) can subsequently be oxidized back to sulphate, resulting in the production of hydrogen ions. In spite of these possible reactions in granites and related rock types, much of the SC>2 and S0^~ deposited in acidic deposition is not retained. Sulphate is leached out of soils and is often the anion balancing the presence of H+ and other cations in surface and shallow ground waters. The amount of 80^2- in runoff from the Shield areas is very close to the amount deposited in precipitation. At the Experimental Lakes Area (ELA) in Ontario, Schindler et al. (1976) found the atmospheric 80^" input measured in bulk precipitation and the 804^" export in the runoff were in balance. Likens et al. (1977b) found 67% of the total input in runoff at Hubbard Brook, New Hampshire. Kerekes (1980) reported that outputs of sulphate were about 80% of the annual inputs for the Lower Mersey River system in Nova Scotia. In the Adirondack Mountains of New York, Galloway et al. (1980b) observed that sulphate inputs and outputs were in balance for two lake/watershed systems, while for a third watershed some accumulation of sulphur may be occurring within the terrestrial system. In some cases, the SO^" in surface waters is greater than the total input measured in precipitation and the difference may be due to sulphur inputs in dry deposition (see Section 3.6.1). Although little sulphate is retained in granitic watersheds, in certain kinds of soils, such as are common in the southeastern U.S., a large portion of sulphate inputs may be retained in the soil by soil adsorption processes (Johnson et al. 1980). This will have the very important effect of retarding the movement of cations, including H+, from the soil to aquatic systems. (See Section 4.4.2 for further discussion.) 3.2.4 Aluminum and Other Metals Surveys of waters in regions affected by acidic deposition indicate elevated levels of aluminum (Al), cadmium (Cd), copper (Cu), lead (Pb), manganese (Mn), nickel (Ni) and/or zinc (Zn) in many acidic lakes and streams (Aimer et al. 1978; Beamish 1974; Conroy et al. 1976; Henriksen and Wright 1978; Schofield 1976b). These increased concentrations of metals may result from either increased atmospheric loading (associated with or independent of acidic deposition) or increased metal solubility caused by increasing surface water acidity. Elevated concentrations of Cd, Cu, Pb, and Ni are probably derived from increased atmospheric deposition. For these metals, deposition and concentrations significantly above background levels occur principally in lakes and streams in relatively close proximity to pollutant sources (e.g., Sudbury region of Ontario; Conroy et al. 1976). Although increased atmospheric loadings of these metals may occur in conjunction with acidic deposition, acidic deposition and acidification of surface waters are not direct causative factors. On the other hand, increased concentrations of Al, Mn, and Zn can occur without increased atmospheric metal loadings. For example, addition ------- 3-12 I I of acid to limnocorrals in the Experimental Lake Area, Ontario, produced substantial increases in lake water concentrations of Al, Mn, Zn, and Fe at pH levels 6 and 5 (Schindler 1980). Elevated • concentrations of these metals result from an increase in solubility | at lower pH levels (Stumm and Morgan 1970) and their mobilization from the surrounding watershed and lake and stream sediments H (Galloway et al. 1980a). Elevated concentrations of Al, Mn, and Zn • in acidic waters are for the most part, a direct consequence of atmospheric deposition and acidification. cycling of metals have focused on aluminum. One of the effects of soil acidification is the mobilization of aluminum. The solubility • of this metal is pH dependent, with a minimum solubility at about • pH 6 (May et al. 1979; Stumm and Morgan 1970) (Figure 3-4). Several reports have documented elevated aluminum concentrations in acidic surface waters (Figure 3-5) (Cronan and Schofield 1979; Dickson 1978; • Driscoll et al. 1980; Richard 1982; Wright and Gjessing 1976; Wright « et al. 1980), and in effluent from lysimeters in soils treated with acid solutions (Abrahamsen et al. 1977; Dickson 1978). While • aluminum ordinarily is leached from the upper soil horizon of podsol | soils by carbonic acid, tannic and humic acids, and organic chelation, it is usually deposited in lower horizons. Under the • influence of strong acids in precipitation, however, the aluminum may • be mobilized in the upper (slightly acid) soil horizons and transported by saturated flow through the surface layers into lakes and streams (Cronan and Schofield 1979; Herrmann and Baron 1980). • Elevated aluminum concentrations in streams have been shown to occur • during the spring melt of the snowpack, when large quantities of ff1" ions are released into the saturated surface layers (Driscoll 1980b; M Seip et al. 1980). | The mechanism supplying Al^+ to soil water, and therefore to • shallow interflow water, is the dissolution of aluminum minerals or • exchange reactions on soil organic matter. Norton (1976) and Reuss (1976) suggest the following as an explanation of weathering reactions for aluminum minerals: • A1(OH)3 + H A1(OH)2+ + H+^ A1(OH)2+ + H20 - A1(OH)2+ + H+^A13+ + H20 " These reactions are likely to occur in watersheds where there are no • carbonates to consume H+. In such instances, the reactions above • become the primary buffering mechanism (N.M. Johnson 1979; Kramer 1976). The pH at which this buffering occurs is around 4.5-5.0 • (Johannessen 1980). Henriksen (1980) has shown that lakes with pH | 4.6-4.8 have a higher pH than expected from a theoretical titration I I ------- I I I I I I I I I I I I I I I I I I I 3-13 4 - 5 - 6 - 7 - 8 - 9 PH Figure 3-4. Aqueous aluminum in equilibrium with gibbsite (after May et al. 1979). ------- 3-14 1000 r 500 O) 2 200 - o c o O E 3 100 - 50 - 20 - 10 4.0 5.0 6.0 PH 7.0 8.0 Figure 3-5. Relationship of observed stream concentrations of aluminum to the pH of surface water (modified from Wright and Gjessing 1976). I I I I 1 I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-15 curved based only on bicarbonate buffering, and that the extra buffering can be explained by the presence of aluminum. In aquatic systems, aluminum forms a variety of complexes with water and its constituents, including hydroxide, fluoride, silicate, organic matter, and sulphate (Driscoll et al. 1980). In surface waters of the Adirondack Region of New York, Driscoll (1980b) found aluminum-organic complexes were the predominant monomeric aluminum form (average = 44%). Concentration increased linearly with total organic carbon content. Aluminum-fluoride complexes were the most abundant inorganic form (average = 29% of the total monomeric Al), with concentrations increasing with decreasing pH, although their formation was generally limited by fluoride concentration. 3.3 NATURAL ORGANIC ACIDS IN SOFT WATERS Surface and ground waters can have low pH values or become acidified as a result of natural processes including: 1) natural chemical weathering of pyrite and other sulphide-rich rocks (Herrmann and Baron 1980; Huckabee et al. 1975); 2) net oxidation of reduced organic material due to aerobic biological decay (Likens et al. 1969); 3) oxidation of reduced inorganic material following a lowering of water tables, lake levels, with subsequent exposure to oxygen (Urquhart and Gore 1973); 4) strong cation exchange, especially by Sphagnum sp., with subsequent release of H+ (Clymo 1967); and 5) production of organic acids which are dissociated in the pH range 3 to 6 (Oliver and Slawych 1982). Natural acidification due to chemical weathering (process 1) is usually identifiable because of local geologic conditions (e.g., bedrock geology and Fe-rich secondary soil and sediment mineralization). Processes 2 and 3 are not steady-state phenomena and can generally be related to mechanical disturbances in the watershed or meteorological changes. Process 4, common in humid temperate or sub-arctic climates, is normally distinguishable by analysis of the local hydrology, vegetational studies, and the presence of coloured (humic) waters (related to process 5). A major portion of the dissolved organic carbon in natural waters is organic acids, especially humic and fulvic acids. These acids are produced (process 5) by microbial degradation of plant and animal matter. They are poorly characterized in terms of chemical and physical properties but serve two important functions. They display ------- 3-16 I I acidic properties and contribute significantly to acidity in some organic-rich waters. Secondly, these organic compounds chelate various metals that: (1) increase total metal solubility, and • (2) may decrease the concentration of biologically available metals • (Reuter and Perdue 1977). The relationship between colour (Platinum units) and dissolved • organic carbon (DOC) has been evaluated by several workers (e.g., Juday and Birge 1933) and the relationship between DOC and organic acid has been evaluated empirically by Thurman and Malcom (1981). 4 The extent of dissociation of the acid can be estimated by methods • developed by Oliver and Slawych (1982). Thus, the organic anion concentration can be estimated with a knowledge of DOC and pH, both • commonly made measurements. Alternatively, the organic anion • concentration can also be estimated based on a complete chemical analysis (cations and anions) using an ion balance approach. Many ' igs and organic rich soils have undiluted water pH values in • the r^nge of 3.5 to 4.5 due to high concentrations of DOC and associated acidity. The high IT1" concentration is not totally If balanced by S042~, N03~ Cl~, or HC03~ and the PH is dearly j| determined largely by organic acid production and cation exchange (Clymo 1967). M The synoptic surveys of acidic clearwater lakes (Dickson 1980; Haines 1981b; Haines and Akielaszek 1982; Norton et al. 1981a; Wright and Henriksen 1978; among others) have concentrated on lakes that have • relatively low or no water colour, and therefore having low DOC, low 9 organic acid content, and low organic anion concentrations. Ion balances are achieved largely using only H"1", major cations and sulphate for lakes with pHs below about 5.5 where HC03~ becomes relatively unimportant. Natural soil processes in well-drained terrain may produce • considerable acidity due to soil respiration (which raises dissolved C02 and carbonic acid concentrations) and biological breakdown of organic material to produce organic acids and chelators. Water I percolating through the soil profile may commonly have pH levels ™ lowered to near 4.0 in the organic horizons. As these solutions descend further, acidity is consumed by inorganic reactions including mineral weathering and desorption of cations. Additionally, organic compounds precipitate with increasing pH and/or oxidize to C02 and H20. The result is that acidic soil solutions commonly have their • pH raised from about 4.0 to 5.5-6.5 within a few vertical meters of • travel (Cronan 1982). Should these solutions emerge as surface water, the pH would be elevated, "nonacidic", and HC03~ would be a major charge balancing anion, along with sulphate. However, if • soils are shallow and unreactive, solutions may reach streams prior • to effective neutralization (A.H. Johnson 1979) and prior to the development of the maximum allowable HC03~ alkalinity. The • addition of excess acidity (as ^804 or (Nfy^ SO^) to soil waters | decreases the pH of soil solutions further (even for soils with pH I I I I ------- I I I I I I I I I I I I I I I I I I I 3-17 values originally around 4). As a result, given the same flow path, a smaller proportion of the acidity will be consumed. Therefore, surface water emerges with a lower pH, lower alkalinity, and possibly elevated concentrations of cations due to accelerated cationic leaching (Abrahamsen 1980). A number of processes may ameliorate the impact of increased acid loading, the most important of which are net uptake of NC>3~ by plants (Reuss 1976) and SO^" adsorption by soils (Johnson et al. 1980). Humic materials have recently been shown to have low buffering capacity even when present in high concentrations (Wilson 1979). The weak buffering capacity they do exhibit is between pH 4 and 5 (Driscoll 1980a; Wilson 1979), the pH region in which the endpoint of alkalinity titrations occurs. In systems with low alkalinity, the presence of humics can lead to a significant underestimation of alkalinity when the usual acidimetric determination method is used (Driscoll 1980a). In addition, these substances can influence the bioavailability of acid-leached cations such as Al, Mn, Fe and Zn by acting as chelators. 3.4 CATION AND ANION BUDGETS "Calibrated lakes and watersheds, that is, natural catchments for which the input and output rates of substances can be measured, are an established research tool in environmental studies. For example, the development of strategies for the management of eutrophication of lakes by phosphorus control was based largely on mass balance studies and models (Dillon and Rigler 1975; Oglesby 1977a, 1977b; Reckhow 1979; Vollenweider 1975). "Common reasons for the use of this approach include: (a) the relative importance of different inputs of a pollutant can be assessed and abatement planned accordingly; (b) mass balances can be used with mathematical models to predict the chemical concentrations of compounds in the receiving body, either the stream draining the calibrated watershed, or the calibrated lake itself; (c) the quantitative accounting of the flow of substances in the watershed or lake may provide information concerning the processes and mechanisms occurring there." (Dillon et al. 1982) Ionic balances of watersheds have been used as a means of quantifying net basin chemical fluxes (Figure 3-6). This approach is being used to evaluate the effects of acidic deposition on element budgets. Several studies have been underway since the early 1960s. One of the earliest studies and the longest continuous record (1963-present) is ------- 3-18 | 1 1 Allochthonous Sources of Hydrogen Ion A 1 precipitation J>°|| • "^x^r^j dry deposition 't^ drainage water £M *i^r % 1 |^ >-«. _ ^*! ^YV^, Hydrogen Ion Sources Hydrogen Ion Sinks oxidation reduction cation uptake anion uptake pyrite oxidation oxide weathering NH * uptake Stream Exports 1 H* HCO3", OH - ligands, organic Figure 3-6. Schematic representation of the hydrogen (Driscoll 1980a). anions 1 1 1 1 _ 1 1 1 1 1 1 1 ion cycle 1 1 1 ------- I I I I t I I I I I I I I I I I I I I 3-19 from Hubbard Brook, New Hampshire, summarized by Likens et al. (1977a). 3.4.1 Element Budgets at Hubbard Brook, New Hampshire The ionic composition of bulk precipitation at the Hubbard Brook ecosystem is essentially characterized by acids, such as H2SC>4 and HNC>3. In contrast, water leaving the system is characterized mainly by neutral salts, composed of Ca^+, Mg2+ and Na+ balanced in solution by SO^" and, to a lesser extent, by chloride, nitrate, and bicarbonate species. The chemical and biological reactions of hydrogen ion, nitrate, ammonium, and sulphate are very important in driving displacement and weathering reactions. Observed trends and annual ion budgets for 11 years at Hubbard Brook demonstrate the influence of atmospheric inputs on surface water quality. High rates of H+, N03~, and S0^~ inputs were observed throughout the period. The average annual weighted pH of precipitation from 1964-65 through 1973-74 ranged between 4.03 and 4.21. The lowest value recorded for a storm at Hubbard Brook was pH 3.0 and the highest was 5.95. During the period 1969-1974 (the latter being the last year of the 1977 summary), no weekly precipitation average exceeded a pH of 5.0. Fluctuations in hydrogen ion deposition can be explained in large part by the fluctuation in total precipitation. Concentrations for SQ^~ and Nfy"1" varied from year to year, but showed no statistically significant time-trends for the period. In contrast, annual weighted N03~ concentrations were about 2.3-fold greater in 1971-74 than they were in 1955-56 (Likens et al. 1977b). From 1964 to 1970, there was a general downward trend in the percentage sulphate contribution to the total anion equivalents (Figure 3-7). During the period 1970-77, the rate of decline decreased or perhaps the trend even reversed. The proportion of nitrate to the total anion equivalents has increased throughout the period. Two conclusions were drawn: (1) nitric acid was of increasing importance in precipitation at Hubbard Brook (Likens et al. 1976), and (2) the average annual change in nitrate was somewhat smaller after 1970, apparently due to slower increases in nitrate concentration relative to sulphate in precipitation. The proportion of hydrogen ion to the total cations increased throughout the period even though the total equivalent concentration of cations decreased (Likens et al. 1980). The Hubbard Brook study site is an isolated headwater catchment. As a result, the influx of chemicals is limited principally to precipitation and dry deposition, and the outflow to drainage waters. Theoretically, differences between annual input and output for a given chemical indicate whether that constituent is being accumulated within the ecosystem, is being lost from the system, or is simply passing through the system. Likens et al. (1977b) were, therefore, ------- 3-20 80 70 « 60 c Si "5 .2 50 3 a UJ « 40 30 S 20 10 i I l ( i i i 1964-65 66-67 68-69 70-71 72-73 74-75 76-77 Year Figure 3-7. Percent of ionic composition of precipitation for the Hubbard Brook Experimental Forest during 1964 to 1977, £M+ is sum of all cations (Likens et al. 1980). I I I I 1 1 I I I I I I I I 1 I I I I ------- I I I I I I I I 1 I I I t I I I I I t 3-21 able to estimate with reasonable accuracy the mean annual budgets for most of the major ions (Table 3-2). Over the long term, there was considerable variation. However, calcium, magnesium, potassium, sodium, sulphate, aluminum, and dissolved silica budgets indicated net annual losses. Net annual gains of ammonium, hydrogen ion, and phosphate occurred in these undisturbed, accreting watershed ecosystems. Nitrate and chloride budgets indicated a net accumulation in all but 3 of the 11 years of study. Overall, during 1963-74 there was an annual net loss of total dissolved inorganic substances from the experimental watersheds amounting to 74.7 kg/ha. yr. The average net output of dissolved inorganic substances minus dissolved silica (1963-1974) was 38.4 kg/ha. yr. The smallest annual net loss of dissolved inorganic substances (27.8 kg/ha, or 7.0 kg/ha for total material minus dissolved silica) occurred during 1964-65, the driest year of the study. The largest net losses of dissolved inorganic substances occurred during the wettest year, 1973-74 (139.7 kg/ha). Likens et al. (1977b) also noted the complexity of computation of the long-term cationic denudation rate in the Hubbard Brook ecosystem because of the need to consider accumulations in living and dead biomass. The net accretion of biomass should be viewed as a long-term sink for some of the nutrients supplied from the weathering reactions. The total amount of cations sequestered by this means is 72.2 meq/m^.yr. They concluded that: (1) cations stored within the biomass must be included in calculations of contemporary weathering ; ( 2) the rate of storage is a consequence of the current state of forest succession and changes with time; and (3) the existence of the forest and its state of development must be included in geological estimates of weathering. If this appraisal of the biological system at Hubbard Brook is correct, the flux of cationic nutrients being diverted into biomass accretion (72.2 meq/m^.yr) must be added to that actually removed from the system in the form of dissolved load (126.7 meq/m^.yr) and particulate organic matter (1.0 meq/m^.yr). Therefore, the best estimation of cationic denudation (net loss from ecosystem plus long-term storage within the system) at Hubbard Brook is about 200 These long-term estimates of cationic denudation at Hubbard Brook allow estimation of the relative importance of external and internal sources of H+ ions. The external supply rate is 100 meq/m^.yr and, by difference, the internal source becomes 100 meq/m^.yr. This suggests that under prevailing biological and chemical conditions (perhaps altered by changes in atmospheric precipitation), external and internal generation of H"1" ions play nearly equal roles in driving the weathering reactions at Hubbard Brook (Figure 3-8). ------- TABLE 3-2. ANNUAL BUDGETS OF BULK PRECIPITATION INPUTS AND STREAM-WATER OUTPUTS OF DISSOLVED SUBSTANCES FOR UNDISTURBED WATERSHEDS W THE HUBBARD BROOK EXPERIMENT FOREST (Likens et al. 1977b) Substance (kg/ha) CALCIUM 1 nput Output Net MAGNESIUM Input Output Net ALUM 1 NUM Input Output Net AMMON 1 UM Input Output Net HYDROGEN 1 nput Output Net SULPHATE Input Output Net NITRATE Input Output Net BICARBONATEd Input Output Net 1963 1964 1965 1966 1967 1968 to to to to to to 1964 1965 1966 1967 1968 1969 3.0 2.8 2.7 2.7 2.8 1.6 12.8 6.3 11.5 12.3 14.2 13.8 -9.8 -3.5 -8.8 -9.6 -11.4 -12.2 0.7 1.1 0.7 0.5 0.7 0.3 2.5 1.8 2.9 3.1 3.7 3.3 -1.8 -0.7 -2.2 -2.6 -3.0 -3.0 a a a a a a 1.6c 1.2 1.7 1.9 2.1 2.2 -1.6 -1.2 -1.7 -1.9 -2.1 -2.2 2.6° 2.1 2.6 2.4 3.2 3.1 0.27C 0.27 0.92 0.45 0.24 0.16 2.3 1.8 1.7 2.0 3.0 2.9 0.85C 0.76 0.85 1.05 0.96 0.85 0.08C 0.06C 0.05 0.07 0.06 0.09 0.77 0.70 0.80 0.98 0.90 0.76 33.7b 30.0 41.6 42.0 46.7 31.2 42. 7b 30.8 47.8 52.5 58.5 53.3 -9.0 -0.8 -6.2 -10.5 -11.8 -22.1 12.8° 6.7 17.4 19.9 22.3 15.3 6.7° 5.6 6.5 6.6 12.7 12.2 6.1 1.1 10.9 13.3 9.6 3.1 a a a a a a 6.2b 4.6b 6.2 9.4 9.6 7.0 -6.2 -4.6 -6.2 -9.4 -9.6 -7.0 1969 1970 1971 to to to 1970 1971 1972 2 16 -14 0 3 -3 a 2 -2 2 0 2 0 0 0 29 48 -18 14 29 -14 a .3 1. .7 13. .4 -12. .5 0. .5 3. .0 -2. a .2 1. .2 -1. .7 3. .51 0. .2 3. .93 1. .09 0. .84 1. .3 34. .1 51. .8 -16. .9 21. .6 24. .7 -3. a 6.0 7. -6 .0 -7. 5 1 9 12 4 -11 5 0 1 2 6 -2 a 8C 1 8 -1 9 2 23 0 7 2 18 0 14 0 04 0 6 33 1 46 5 -13 6 21 9 18 3 2 a 1b 6 1 -6 .2 .4 .2 .4 .8 .4 .7C .7 .8 .05 .8 .97 .13 .84 .0 .8 .8 .4 .7 .7 .6b .6 1972 to 1973 1.2 15.6 -14.4 0.5 3.3 -2.8 a 2.3C -2.3 2.5 0.18 2.3 1.08 0.16 0.92 43.4 64.0 -20.6 26.3 19.2 7.1 a 9.0b -9.0 1973 to 1974 2 21 -19 0 4 -4 a 3 -3 3 0 3 1 0 0 52 84 -31 30 34 -3 a 12 -12 .0 .7 .7 .4 .6 .2 .2C .2 .7 .42 .3 .14 .20 .94 .8 .7 .9 .9 .8 .9 .5b .5 3-22 ITHI N 1 1 1 Total Annual 1963-1974 mean • kg/ha kg/ha • 23.8 151.2 -127.4 6.3 34.6 -28.3 a 21.9 -21.9 31.6 3.7 27.9 10.62 1.13 9.49 418.3 580.3 -162.0 209.5 177.5 32.0 a 84.2 -84.2 2. 13. -11. 0. 3. -2. a 2. -2. 2. 0. 2. 0. 0. 0. 38. 52. -14. 19. 16. 2. a 7. -7. a Not measured, but trace quantities. b Calculated value based on weighted average concentration and on amount of precipitation or streamflow during the c Calculated from weighted concentration for 1964-1966 Based on annual concentration of 0.50 mg/1 (Juang and d Watershed 4 only. during years when chemical measurements — ^ 1 7 • 5 1 6 1 • 1 I 0 1 9 34 - 1 96 | 10 • 86 1 0 8 m 81 0 1 i • 9 1 £• 1 1 were made * specific year. times precipitation Johnson 1967). for 1963-1964. 1 1 ------- 1 1 1 1 1 1 1 1 1 1 1 1 t 3-2: HUBBARD BROOK HYDROGEN ION BUDGET ^ Allochthonous Precipitation Dry Deposition ^^^^f, JnPUtS + 362 "M^ Ir 'Hi — n Weathering Reactions Net Accumulation in Forest Floor and Ca -1055 Mg - 288 Na - 252 K - 182 S + 25 Al - 211 Fe - 78 P + 83 TOTAL -1957 Stream pH (H*) Stream Alkalinity (HCOp Discrepancy in Charge Balance Forest Biomass Ca + 475 Mg + 74 Na + 7 K + 156 S - 125 Fe + 103 HQ-- 47 NH4++ 144 P - 90 + 697 Stream Exports -100 ' +126 + 26 _ (organic anions, hydroxide ligands) 1 • 1 1 SUMMARY Hydrogen Ion Sources Hydrogen Ion Sinks Budget Discrepancy Figure 3-8. Hydrogen ion budget (meq/m^ 1 + 2541 - 2428 + 113 .yr) for Hubbard Brook Experimental Forest (Driscoll and Likens in press). ------- 3-24 3.4.2 Element Budgets in Canada I I The calibrated watershed technique for measuring rates of movement of • elements has also been used in Canada and results have been • summarized by Harvey et al. (1981): "Input-output budgets (mass balances) for major ions are being • measured at a number of locations in Canada as described in Table 3-7 [Table 3-3 this report]. In all cases, mass balance _ measurements have excluded possible contributions via subsurface • flow, although the evidence available for these lakes suggests that these contributions are negligible (Schultz 1951). Net exports of Ca2+, Mg2+ and K1" are shown in Table 3-8 It [Table 3-4 this report] for Canadian watersheds, along with 9 input of H"1" by precipitation. Output of HC03~ and input- output data for S042~, NH^+ and NC>3~ are also included, where • reported. No Canadian information on inputs and outputs of • aluminum was found. Nicolson (1977) reported only output of major ions from 12 watersheds in the Experimental Lakes Area, _ northwestern Ontario; input by precipitation to the nearby • Rawson Lake watershed (Schindler et al. 1976) was used to ™ calculate a net export for these 12 watersheds." With one exception, Clear Lake, all of the watersheds studied had a net output of the major cations (Ca2+ + Mg2+ + Na+ + K+). The net M export of Ca2+ + Mg2+ dominated the ion budgets particularly • in the watersheds in British Columbia which contain some calcareous till. The study sites in British Columbia received a larger amount of precipitation (260 to 450 cm/yr) a factor which may increase the • export of cations. Potassium export is low in all cases reflecting • the biological demand for this element in the watersheds. In all cases there is a net accumulation of NH4+ + N03~ in the • watersheds. | The export of cations from ELA and Rawson Lake watersheds on the ^ Precambrian shield in Western Ontario was about 30-40% of the export V from the Hay Lake watersheds in the Muskoka-Haliburton area. The ' corresponding H"1" inputs were 5-10 times greater at Harp Lake, also in Muskoka-Haliburton. I More sulphate was exported from the watersheds than entered via wet or bulk deposition. In some cases (Rawson Lake and Jamieson Creek) • the differences may be within experimental error, and in some cases | the input may be underestimated due to dry deposition and canopy effects. « I I I ------- 3-25 ^ co ON * in 4- 0 ^. 0 L. 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C 0 O JO — > •• o o T3 0 00 JL <5 «^ c o •H L. 3 JO JO 3 CL C 3 ^_ >* ^- ^ U o 0 jC -o c ID •t ID in *. € JO •l 1 £ in . in 0 0 CL c ID CO E 4- in 0 b csi CO o r- m CD (D CL ID X •t T3 CD f~ (/) — -C 4- 'i* 10 — ^_ «— 4- O O N "8 CL 3 O _ ID JC in ^- ID 0 C CO In ~O CO 1 co c O in ON ON •o 0 jc in t_ 0 ID CO C ,— o s> o o JC CL 1_ in o 4- E — ID in 4- O CO a. E T3 0 C TO ID 3 in u, 4- L. C CD — 8- . — C — 0 (D JO I/) 0 C JC L. 4- 5 » I I ------- 3-26 oo ON • ID 4- 0 >*, 0 L. ID X „. Q i 0 Z CO Q LU S cc i- 3 a i— i o § CO Z o g —3 i Li- ID 1- ct o ^ Ul I— LU Z • 1 to LU m i- x >• •4- CSI O E +• CT 3 0 0. E C «-» CO in ID i CN CO 1 IO o 3- z in ID i to Q 1 r~* \^ >• CN E Q- 0 •-' t 0 Q. 0 4- 0 ' i i z in ID Tt X z E ^ •^ 4. 10 z ^. CSI 01 CN ID O •o 0 £1 in 0 +- ID S to 1 ^ • ^ to " o o CN IO » ^> 1 ^ f^ 1 in • in f °°. in in • 00 .— CN ^ VO ON VO CSI fc •^ 0 • 0 in O C 0 O > — 3 4- O ID 0 C C L. (D ID > 0 1 ON ^- ^ oo CSI CN 1 CN O 00 CN to "* ON in ^. • ^> in vO . — • o CD fc ^ 0 0 O c O 0 E 10 o i ^_ in 0 co co to i co °i o CSI co ^]- • to o CO 0 *^- CSI ID VO vO •h _J LU * in TJ 0 J2 in L. 0 4- ID 0 0 * H- r++ » LO in • o «— .— ^. • vO ^" o ^ r- O L. ID | C O c i_ 0 4- tn 0 3 O T vO ON ON o to ^ * — to -o 0 to •i- • L. CN o to Q. 1 0 Z ID CSI •t- • 0 0 i— to 1 I** vO • • CSI M- tO ON *3" tO • • O CN O VO ON vO — O CSI tO CSJ l~- — CN *- in ^_ T3 O 0 1 O -C ^ ID • — in in j^ i_ l_ *D O ID 0O 0 -* •(- +- — .c in c ID L. in 3 O 5 ID L- S •I- 0 0 c 4- •- c ^ O ID 0 O ID 3 -* -f- _l C -D ID L. I- 3 _l 3 c 0 in .0 O +- Q- — in in L. i_ — 20 3 ID ID ID 3 O X X CC LL. • $X) 3 in •o 0 L! 4- ID 0 ID c O in ID OL o >4_ C O (D .— Q. O 0 CL C +- 3 O. C 0) c in 3 4- Q. ^_ 3 0 • ^_ +- 0 3 C CL 3 0 O -f- in E in — O 4- L. in O LU ID -O I I I I 1 1 I 1 I I I I I I 1 1 I I I ------- I I I I I I I I I I I I ff I I 1 I i I 3-27 "The mechanism for cation export is apparent in some cases. At Carnation Creek, the output of HCO^" is substantial, suggesting that carbonic acid is the principal weathering agent responsible for cation export. At Jamieson Creek, the output of cations greatly exceeds the output of HCC>3~ and supply of IT1", while at the Haney watersheds, the opposite situation is observed. In view of these contradictory observations, the mechanisms for cation export in this area is uncertain. At the Experimental Lakes Area in northwestern Ontario (including the Rawson Lake studies), the release of cations probably is a result of carbonic acid weathering, although the H+ loading of 7-10 meq/m .yr in precipitation may account for 20% of the net cation export. On the other hand, in southern Ontario where the cation export is ~3 times greater than in northwestern Ontario, 50% or more of the cation yield probably results from input of H* in strong acid form. Although the evidence is circumstantial, it appears likely that the increased H+ input of southern Ontario has resulted in a two- to four-fold increase in net output of cations." 3.4.3 Effects of Forest Manipulation or Other Land Use Practices on Watershed Outputs Land use practices within watersheds have been suggested as an influence on acidification (Henderson et al. 1980; Likens et al. 1978; Rosenqvist et al. 1980). Henderson et al. (1980) have summarized results from watersheds at Hubbard Brook (New Hampshire), Fernow (West Virginia), and Coweeta (North Carolina) which were experimentally manipulated through a series of forest cutting practices (Table 3-5). The work was designed to estimate changes in streamflow concentrations of cations, particularly the potential effects of H+ concentrations. At Hubbard Brook, after felling of all vegetation and herbicide treatments for three successive years, nitrogen discharges into stream flow increased by 245.9 kg/ha.yr. Export of dissolved Ca^+ and K+ increased by 65.2 and 28.7 kg/ha.yr respectively compared to a control watershed (Bormann et al. 1974). Increased acidity from biomass decomposition amounted to 69.9 x 10^ iJeq/ha.yr of H+. This additional acidity is presumed to have been a major contributor to the accelerated loss of cations from the soil, shown in Table 3-5. Strip cutting of one-third of the vegetation at a second Hubbard Brook watershed produced significantly less effect on soil leaching rates. Organic matter decomposition was about 5% of that of the total vegetation removal (500 kg/ha.yr versus 10,500 kg/ha.yr). Subsequently, internal H+ production was also less, as was resultant cation leaching than in the deforestation experiment discussed above (Likens et al. 1977a). Commercial clear-cutting at the Fernow watershed generated fewer H+ equivalents possibly because only economic biomass was removed, ------- 3-28 TABLE 3-5. SUMMARY OF TOTAL CATION RELEASE, HYDROGEN ION PRODUCTION, AND THE CATION RELEASE RATIO FOR THREE MANIPULATED WATERSHED STUDIES (Henderson et al. 1980) H+ produced (eq/ha) Hubbard Brook, New Hampshire Deforested 69,960 Strip-cut 8,400 Fernow, West Virginia Clear-cut 960 Fertilization 55,710 Coweeta, North Carolina Clear-cut and 360 cable-logged Total cation release (eq/ha) 6,850 390 170 2,420 50 Cation release ff*" produced (eq/eq H+) 0.10 0.05 0.18 0.04 0.14 I I I I t I I I I I I I f I I 1 I I I ------- I I I t f I I f I 1 I I 1 I I 1 I I I 3-29 reducing overall decomposition rates and resultant H"1" formation (Henderson et al. 1980). The Coweeta clear-cut and cable logging experiments resulted in even less production of H+ ions. When the Fernow watershed was fertilized with 260 kg/ha of urea, a 10-fold increase in stripping of calcium ions occurred, plus a 6-fold increase in magnesium, a 50% increase in potassium leaching, and a 3.6-fold increase in sodium ion denudation (Henderson et al. 1980.) The possibility of changes in land use causing acidification of surface waters, rather than atmospheric inputs of acid, has been explored in great detail by two recent studies in Norway. Seip (1980) concluded that, while land use changes probably have contributed to the acidification process in some areas, "there is no reason to doubt that the increase in the deposition of acidifying components has played an important role in the acidification of freshwater." Drablos et al. (1980) also reviewed land use changes in relation to lost fish populations in lakes and could find no relationship between the two. The greatest number of lakes from which fish populations have been lost occurred in areas without farming activity. Although it is well documented that land use changes affect the quality of runoff, including pH, these reports conclude that the large scale acidification of lakes in Scandinavia is apparently not due to land use changes. In Canada, all of the surface waters which have elevated excess sulphate occur only in areas which have high atmospheric deposition of sulphate (Figure 2-6b). Land use changes, such as logging, have taken place in many areas, including those areas which do not have excess sulphate in surface waters (see Section 3.6.1). All of the surface waters sampled in Northeastern North America that have experienced loss of alkalinity also have elevated excess sulphate concentrations. In areas with less acidic deposition, loss of alkalinity in surface waters has not been observed. These observations indicate that loss of alkalinity from surface waters is associated with increased sulphates resulting from atmospheric deposition rather than land use changes. Wright et al. (1980) summarized their observations as follows: "Acidified lakes often barren of fish are found in southern Norway, southern Sweden, southwestern Scotland, the Adirondack Mountains, New York, and southeastern Ontario. These areas have in common granitic or other highly siliceous types of bedrock, soft- and poorly-buffered surface waters and markedly acidic precipitation (average pH below 4.5)." A recent USGS report (Peters et al. 1981) provided a 14-year data analysis of precipitation in New York state (nine stations) and stream chemistry. "Statistical analyses of chemical data from several streams throughout New York yielded little evidence of temporal trends resulting from acid precipitation, except in the Adirondack mountains, where the soils lack significant buffering capacity. In most areas of the state, chemical contributions from ------- 3-30 3.5.1 Mapping of Watershed Sensitivity for Eastern North America I I urbanization and farming, as well as the neutralizing effect of carbonate soils, conceal whatever effects acid precipitation may have on chemical quality of streams." (Peters et al. 1981) m In summary, the experiments concerning different forest and vegetation-removal practices showed wide variation in the short-term — (less than five years) patterns of H* produced and cation releases. • A survey of European information on land use changes found no ™ evidence that land use changes had an important role in acidification of water or impact on fish populations. Therefore, we conclude that ft although land use changes can affect the quality of runoff and loss • of alkalinity in surface waters, land use changes do not appear to have a major impact on alkalinity nor pH changes in surface waters, with the exception of some waters affected by mine drainage. I 3.5 AQUATIC ECOSYSTEMS SENSITIVE TO ACIDIC DEPOSITION I The roles of soils, bedrock and vegetation in regulating surface water chemistry must be considered when assessing the sensitivity of II aquatic ecosystems to acidic deposition. The geochemical properties | of a watershed provide the primary controls determining surface water alkalinity. Sensitivity evaluations can be based on parameters such mt as lake and stream alkalinity or calcite saturation index. These • parameters do not necessarily reflect the long term capacity of watersheds to buffer or neutralize acidic deposition. Ideally, aquatic and terrestrial data should be evaluated in combination. • Unfortunately, the present data base is not sufficient to do so for ™ all of eastern North America. Data on terrestrial systems (especially soils and bedrock) are more readily available. • Therefore, terrestrial data have been used to identify areas likely | to contain potentially sensitive aquatic ecosystems for all of eastern North America. The mapping of such areas is based on an M estimation of the capacity or potential of the terrestrial system • within an area to reduce the acidity of incoming atmospheric deposition. To identify aquatic regimes already acidified and those most susceptible, in terms of present levels of acidic deposition, it ft will be necessary to compare terrestrial-based mapping with aquatic • chemistry data and regional deposition maps of SO^" in precipitation (Section 3.9). I Cowell et al. (1981) considered a number of characteristics of terrestrial environments to be essential for the assessment of aquatic sensitivity (Table 3-6). Important factors include soil il chemistry, soil depth, drainage, landform relief, vegetation type and W bedrock geology. Each of these factors plays a significant role in ameliorating the effects of acidic deposition. It is important to m evalute as many factors as practical in order to derive an overall • assessment for any area. Single factor assessments can be I I ------- I I 3-31 II cc 0 i— f ^ i- CO 1 • _i cc tec LU 1— u_ 1 ^Hi _ 00 (— O> s - It- • £ « LU 4- X , ||1 E L. LU CO 1- 4- BS-S a: -Q o ® LL. •— M- ^ •— ICC Q l! Q 1- ILU - f— tn < 0 0 & 01 o 1* 5 O LU Z X C/) cc i §£ fe _i >- < f- 10^ O in O LU < cc CC LU LU X 1- 1- 1 ^" 1 10 LU 1-1 m * z o :POSITI UJ Q O CC LU Q- - c E s. .2 S io" H- t — — o. i- co O L. CO 4- 0 4- c E *«. in 4- V- O m ID o x> ID o o • Oin O— ID E — *3- • • — V X* ^t inin —> TO ID o co TO X >• O 4- — ID IO O O. X X Xt Xi O (DC ID O o- Q. cc •— x: ID 10 LL- •— V V V IDID — a. TO O \ •. l/lin CT — U CM 4- f>-«» CO 3OCC — CO >- >• .. 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(- o (DO— — — CM'* x: x (0 xi o o O LLI ^-^ ^* ^* in ^ *•* «-» o 01 E 0 2 8- in Q. 0) CM ID 4- \/ i. in CO E 3 O 0 4= O - t 1 — 8 > X a. — .4^ CD (D — O) — ID CD -c c ce a. ID E (D t- L. Q Q O M- — — -o O O ID in in _j ^^ o CM C CD 1 in SO •F 0 0 «4- in 'c ID 0 a. u in CD in L. (D a. in in O — mixed discontinu (20 to 60% x»x ^^ o in in 3 3 O O 3 3 C T3 — O C CO O XI 0 1_ CD 0 O C C .2 2. 4- 4- (0 ID 4- 4- CO CO TO TO CO CO in a) in 4- TO (DC c c co — O) >. u in ID CD m 4- « jO — O C 4- CO ID — 4- 1- N ID TO 4- C L. 5 .- § L. 4- CT 8 c 5 § £ t C CT) O « in . . 2 t V) in xi o CO CO XI — TO — C ID ID C (D O CD > _ j; £: ^_ L. m in « z ID T3 m CT m , c JC ffl X> in (D U ID jt in o 4- CO O 1-0) 4- O • £ ID L. Ul O C (D — • O O JC *• ID XI — O ID C L. C (D E O 8(D S ID 4- 0 >~ L. (fl C — CD 4- -0 O O U — 3 C > TO 3 £ •o .-8 xi * in -o c in A: m in ID 4- O X) -* c o L. o > (D L. (D O CD — 4- L. 4- (D O C TO — > — — CO C E — 4- 4- •— o 3 in m ID L. — CT ID ^ C (D O CO — O O CO X) O O XI XI o U L. « — in (D (D CO .c 3 (D O 4- c a. o 4- c ID O L. CO ID O c 4- Q L. C c O in E (D O X) CO ID O X) £ 1- E 4- — U 4- 8— m ID ID — — E o y s — "ID — ID — ID — L. — L. CO L. CO 4- CD 4- ID 4- ID S ID S E .* 4- O TO C O C CD U — L. XI >- ID CD — a. co (D — — C — I I ------- 3-32 I I misleading, especially in areas where soil mineralogy differs from underlying bedrock lithology such as glacially-derived soils and old, • deeply weathered soils over limestone (such as in United States, • south of the glacial limit). In both the U.S. and Canadian mapping, soils and bedrock are the • primary factors assessed. It is assumed that resultant lake or V stream water chemistry will reflect the combined interaction of the varying soil and bedrock characteristics on acidic deposition. J| Surface and shallow groundwater flow conditions are assumed to best m characterize the surface water regime. Groundwater residence times and deep groundwater circulation were not considered. ^ Certain types of vegetation, especially broad-leaved deciduous trees, ™ are capable of reducing acidity of intercepted precipitation (Fairfax and Lepp 1975, 1976). The nature of chemical modifications by 9 vegetation species, however, is not yet fully understood. Therefore, • the effect of vegetation type and cover on aquatic system sensitivity has not been included in this analysis. •m Lucas and Cowell (1982) have mapped the potential of soils and bedrock to reduce acidity of atmospheric deposition across eastern — Canada (east of Manitoba). A similar study has been carried out by I Olson et al. (1982) for the eastern U.S. The mapping was coordinated ™ in order to produce a comparable basis for evaluation. Although the conceptual framework is similar, data availability and quality varied B considerably both between and within countries. 0 The maps of eastern Canada and the eastern United States presented • here combine bedrock, soil and certain other factors (Table 3-7) in • order to interpret the potential ability of terrestrial ecosystems to reduce acidity. A low potential implies that acidic deposition could reach aquatic systems with little neutralization. Many of the low • potential ecosystems are naturally acid and may contribute a high • capability to acidify incoming precipitation because of organic acids, especially in areas receiving low inputs of mineral acids. High potential areas would generally be capable of reducing acidity such that impacts to aquatic systems would be minimal. Specific factors used in the mapping for both Canada and the United • States are listed in Table 3-7. The assessment of relative potential to reduce acidity may be inferred from Table 3-6. The methodologies for combining and weighting the variables shown in Table 3-7 are • discussed below. ™ The map of eastern Canada (Figure 3-9 in map folio) is presented at • the compilation scale of 1:1,000,000. The U.S. map (Figure 3-10 in J| map folio) is shown at 1:5,000,000. I I I I ------- I I I I I t I I I 1 I I I I I I I I I 3-33 TABLE 3-7. TERRESTRIAL FACTORS AND ASSOCIATED DATA BASES UTILIZED FOR THE INTERPRETATION OF THE POTENTIAL TO REDUCE ACIDITY OF ATMOSPHERIC DEPOSITION (After Lucas and Cowell 1982; Olson et al. 1982) TERRESTRIAL FACTORS/SURROGATES DATA SOURCES3 EASTERN CANADA 1) Soil Chemistry Surrogates: i) Texture (sand, loam or clay) - Quebec, the Maritimes and Newfoundland/Labrador, northern Ontario ii) Depth to Carbonate (high, low or no lime) - Ontario iii) Glacial Landforms - northwestern Ontario iv) Organic Soils (^50% of mapping unit) 2) Soil Depth - shallow (25 cm to 1 m) - deep (>1 m) 3) Bedrock Geology - type - % exposed (<25 cm deep) Ecodistrict Data Base (Environment Canada 1981a, b, c) Ontario Land Inventory (MNR 1977) Pala and Boissonneau 1979 Ecodistricts (Environment Canada 1981a, b, c) and Ontario Land Inventory (MNR 1977) Ecodistricts (Environment Canada 1981a, b, c) and Ontario Land Inventory (MNR 1977) Shilts et al. 1981 Ecodistricts (Environment Canada 1981a, b, c) and Ontario Land Inventory (MNR 1977) ------- 3-34 TABLE 3-7. CONTINUED TERRESTRIAL FACTORS/SURROGATES EASTERN UNITED STATES 1) Soil Chemistry i) mean soil order pH (in distilled water) ii) S04~ adsorption (assumed for Ultisols only) 2) Elevation - landform - 2000 ft a.s.l 3) Bedrock Geology - type 4) Land Use - urban areas - cultivated (managed) soils DATA SOURCES3 Soil Map (USGS 1970) Johnson et al. 1980 Hammond's Landform Map (USGS 1970) Topographic Map (USGS 1970) Hendrey et al. 1980; Norton 1982 1977 National Resource Inventory (USDA 1978) 1978 Census of Agriculture (USDC 1979) a All U.S. data sources listed have been compiled within the Geoecology Data Base (Olson et al. 1980). I I I I I I I I I I I I I I I 1 1 I I ------- I I I I I I I I I I I I I I I I I I I 3-35 3.5.1.1 Eastern Canada The map of Eastern Canada (Figure 3-9) was prepared from several data sources (Table 3-7). Quebec, the Maritimes and Newfoundland- Labrador were interpreted using the Ecodistrict Data Base (Environment Canada 1981 a, b, c) described in Cowell et al. (1981) and the bedrock sensitivity evaluation of Shilts et al. (1981). In northeastern Ontario, north of 50°N latitude, Ecodistricts were used in combination with the Ontario bedrock geology maps (Ontario Ministry of Natural Resources, Maps 2198 and 2200). In northwestern Ontario, the recent physiographic mapping of Pala and Boissonneau (1979), and bedrock geology mapping (Ontario Ministry of Natural Resources, Maps 2199 and 2201) provided the basis for interpretations north of 52°N. Interpretations for the area to the south are based on Shilts et al. (1981) and the Ontario Land Inventory (OLI) (OMNR 1977; also described in Richards et al. 1979). The OLI was originally generated at 1:250,000. As much information as possible has been retained in the 1:1,000,000 scale mapping presented here. This partially accounts for the apparent variation in map detail. A cautionary point regarding the use of the Ecodistrict data base in Quebec, the Maritimes and Newfoundland-Labrador must be emphasized. The Ecodistrict delineations are based on a series of biophysical factors including geology and soils. However, the units are not based solely on these two factors. In an attempt to isolate the critical geological factor in sensitivity assessment, the bedrock sensitivity evaluation compiled by Shilts et al. (1981) was superimposed directly on the Ecodistrict map south of 52°N latitude. As no similar map is available for soils for eastern Canada and, with the premise that the ecodistricts delineate major soil characteristics the Ecodistrict base is assumed as the soil base for the combined map. Because of this assumption, for the resultant subdivisions of ecodistricts, the soils data represent the dominant characteristics as described for the original ecodistrict and not the more site specific combined units. However, the Shilts et al. (1981) interpretation was used primarily to improve the resolution in areas where carbonate predominated or where soils were thin and discontinuous and the bedrock sensitivity was most important in the overall evaluation. In assessing map units, each factor in Table 3-7 is assigned a high, moderate or low potential to reduce acidity of atmospheric deposition independently (except percent bedrock exposure). Dominant factors were then combined and weighted in order to derive an overall rating for the map units. Subdominant characteristics were not considered. Specific combinations of the factors mapped as high, low or moderate potential for reducing acidity are identified in Table 3-8. This table shows 74 classes (of which 65 actually occur) which have been grouped into high, low and moderate potentials to reduce acidity. In addition there are 10 classes representing terrain dominated by organic deposits for which no specific interpretation has been made. ------- 3-36 I TABLE 3-8. TERRESTRIAL CHARACTERISTICS OF AREAS HAVING HIGH, MODERATE AND LOW POTENTIAL TO REDUCE ACIDITY FOR EASTERN CANADA (after Lucas and Cowell 1982) I TERRAIN DESCRIPTION Polygon Soil Classification Depth HIGH POTENTIAL Hla TO REDUCE ACIDITY Hlb Hie H1d Hie Hlf H1g Hlh Hli H1J Hlk H2a H2b H3a H3b H3c MODERATE POTENTIAL Mia TO REDUCE ACIDITY Mlb Mic M1d Mle Mlf Mlg Mlh Mil Mlj M1k Mil Mlm Mln Mlo Mlp M1q Mir Mis Mlt Mlu Mlv deep deep deep deep deep deep shal low shal low shal low shal low bare shal low shal low deep deep deep deep deep deep deep deep deep deep deep deep deep deep deep shal low shal low shal low shal low shal low shal low shal low shal low bare bare Soil Bedrock Texture Lithology clay loam sand clay loam sand clay loam sand clay, loam or sand clay clay clay clay clay clay clay loam loam sand sand clay clay loam loam sand sand clay clay loam loam sand sand clay, loam or sand clay, loam or sand Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 1 Type 2 Type 3 Type 2 Type 3 Type 4 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 Type 2 Type 3 % Bedrock Outcropping 0-49 0-49 0-49 50-99 50-99 50-99 0-49 0-49 0-49 50-99 100 0-49 0-49 0-49 0-49 0-49 50-74 50-74 50-74 50-74 50-74 50-74 75-99 75-99 75-99 75-99 75-99 75-99 50-74 50-74 50-74 50-74 50-74 50-74 75-99 75-99 100 100 • MAP AREA • Km 78,890 65,960 8,105 N/A 1,004 109 7,989 5,305 8,959 1,405 N/A 4,317 5,467 20,567 65,470 101,420 7,104 N/A 83 5,543 N/A 9,615 N/A N/A 2,325 489 N/A N/A N/A 1,038 3,114 740 48 14,345 374 2,218 415 14 % of Eastern Canada 2.51 2.10 0.26 N/A 0.03 <0.01 0.25 0.17 0.29 0.04 N/A 0.14 0.17 0.66 2.09 3.23 0.23 N/A <0.01 0.18 N/A 0.31 N/A N/A 0.07 0.02 N/A N/A N/A 0.03 0.10 0.02 <0.01 0.46 0.01 0.07 0.01 <0.01 _! • 1 1 1 | 1 1 1 ------- I I 3-37 • TABLE 3-8. CONTINUED 1 MODERATE POTENTIAL ITO REDUCE ACIDITY 1 ff • LOW POTENTIAL TO REDUCE ACIDITY 1 1 1 * 1 ORGANIC TERRAIN 1 1 1 TERRAIN DESCRIPTION Polygon Classification M2a M2b M3 M4a M4b M5 M6a M6b M7a M7b M7c Lla Lib L1c Lid Lie L2a L2b L2c L2d L3 L4a L4b L4c Ola Olb Olc Old Soil Depth3 deep shal low shal low shal low shal low shal low shal low shal low deep deep deep deep deep deep shal low bare deep deep shal low shal low shal low deep deep deep Soil Texture clay clay loam sand sand clay loam loam loam loam loam clay loam sand clay, loam or sand loam sand loam sand sand sand sand sand organics organics organics organics Bedrock Lithologyc Type 4 Type 4 Type 4 Type 2 Type 3 Type 4 Type 2 Type 3 Type 2 Type 3 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 4 Type 2 Type 3 Type 4 Type 1 Type 2 Type 3 Type 4 % Bedrock Outcropping 50-74 50-74 0-49 0-49 0-49 0-49 0-49 0-49 0-49 0-49 0-49 75-99 75-99 75-99 75-99 100 50-74 50-74 50-74 50-74 0-49 0-49 0-49 0-49 0-50 0-50 0-50 0-50 MAP AREA km 82 982 60,388 15,234 202,167 13,776 14,155 51,297 35,546 104,406 64,804 3,322 6,538 80,800 10,405 47,460 2,150 11,905 156,862 527,190 15,595 161,509 676,252 205,748 40,426 52,949 142,200 % of Eastern Canada < 0.01 0.03 1.93 0.49 6.44 0.44 0.45 1.64 1.13 3.33 2.07 0.11 0.21 2.58 0.33 1.51 0.07 0.38 5.00 16.80 0.50 5.15 21.55 6.56 1.29 1.70 4.53 ------- 3-38 TABLE 3-8. CONTINUED I 1 I TERRAIN DESCRIPTION MAP AREA I Polygon Soil Soil Bedrock Classification Depth9 Texture Lithology0 ORGANIC TERRAIN6 02a 02b 02c 02d 03a 03b 03c 03d organics organics organics organics organics organics organics organics Type 1 Type 2 Type 3 Type 4 Type 1 Type 2 Type 3 Type 4 % Bedrock Outcropping 51-74 51-74 51-74 51-74 75-99 75-99 75-99 75-99 km 34 N/A 207 377 48 N/A 55 327 % of Eastern Canada • < 0.01 N/A < 0.01 0.01 0.01 N/A < 0.01 0.01 4 a Soil depth is defined as follows: deep - >1 m average soil thickness shallow - 25 cm - 1m average soil thickness bare - <25 cm average soil thickness b Soil texture is used to interpret soil sensitivity for most of eastern Canada. In Ontario where depth to carbonate information is available, the following corresponding classes were used: clay - high or very high lime loam - moderate and low lime sand - low base or no lime c Bedrock sensitivity classes were defined by Shilts et al . (1981) on the basis of lithology. Specifically: Type 1 - Limestone, marble, dolomite Type 2 - Carbonate-rich siliceous sedimentary: shale, limestone; non- calcareous siliceous with carbonate interbeds: shale, si dolomite; quartzose sandstone with carbonates Type 3 - Ultramafic rocks, serpentine, noncal careous siliceous sedimenta rocks: black shale, slate, chert; gabbro, anorthosite: gabbro diorite; basaltic and associated sedimentary: mafic volcanic rocks . Type 4 - Granite, gneiss, quartzose sandstone, syenitic and associated a I kal ic rocks. ^ Average bedrock outcropping within each map unit is shown as a percent of map unit. e Organic materials are the dominant soil constituent wherever organics are indicated. I i i i i i i i i i ------- I I 1 I 1 I I I I I I I I I I I I I I 3-39 These are areas of high natural acidity which contribute organic acids to enclosing watersheds. Percent bedrock exposure and soil depth were key parameters for weighting the relative contribution between bedrock type and soil chemistry. Generally, the emphasis was on the bedrock capacity to reduce acidity where bedrock exposure was greater than 50% of an area. If soils were deep (greater than 1 m) and occurred in more than 50% of the map unit, then soil chemistry was emphasized. Soil and bedrock potentials to reduce acidity in map units having combinations of shallow soils and bedrock exposures less than 75% were either averaged or assigned the highest potential. Soil chemistry was interpreted using texture and depth to carbonate because of the nature of the data bases. However, in comparing these with the smaller-scale Soil Map of Canada (Clayton et al. 1977) there appears to be a good correlation with soil order. Hence, sand or no lime soils are dominantly acid Podzols (or "Rockland") and clay or high lime soils are dominantly Luvisols and Gleysols. Loam or low lime soils tend to be more varied including Podzolic, Luvisolic and Brunisolic orders. Map units identified as having a high potential to neutralize acidic deposition are predominantly areas underlain by carbonate bedrock (HI) or areas dominated by deep clay or high lime soils (H3). These each cover approximately 6% (Table 3-8) of the map area represented in Figure 3-9. The former assumes at least some interaction between carbonate-rich bedrock and precipitation prior to entering the aquatic regime. This is probably valid for most of eastern Canada where limestones have either been exposed or buried under carbonate- rich tills by the latest glaciation. In the Hudson Bay Lowland (northernmost Ontario and part of northwestern Quebec) organic deposits blanket the carbonate-rich substrate. Although large streams, rivers and lakes in this region intersect mineral soil, smaller peatland lakes, ponds and streams have naturally soft waters which developed as peat material accumulated over the carbonates. All such organic terrains are considered separately for this reason. The two dominant combinations of soil and bedrock identified as having a moderate potential to reduce acidity are shallow loam or low lime soils overlying bedrock of moderate (M6) and deep loam or low lime soils (M7). Each of these occupy approximately 7% of eastern Canada. The distribution of all moderate classes is highly variable across eastern Canada. All five combinations identified as having a low potential are recorded in eastern Canada. In Ontario and Newfoundland-Labrador the dominant class is deep sand (L4). Shallow sands (L3) are frequently found in the more northerly regions, notably in Quebec and Ontario. These two classes are predominately acid Podzols. Areas of high bedrock exposure (L2) are common to shore zones of lakes and northern areas. ------- 3-40 1 I Major areas of organic soils overlying noncalcareous bedrock (05b, 05c and 05d) are identified in western New Brunswick, southern Labrador and Newfoundland and in the west central portion of Quebec. • Large areas of peatland are identified adjacent to the Hudson Bay | Lowland in northwestern Ontario. Throughout central and western Ontario are numerous small pockets of organic soils. These areas M are, to varying degrees, undergoing natural organic acidification and V hence already contribute low pH, low bicarbonate waters to enclosed watersheds. It is not clear to what degree peatlands and organic groundwater are affected by, or in turn modify, incoming • anthropogenic mineral acids. • I 3.5.1.2 Eastern United States The potential for terrestrial systems to reduce acidity of atmos- _ pheric deposition was determined by combining information on soil M chemistry, bedrock geology, terrain characteristics, and land use (Table 3-7). A map covering the eastern 37 states (Figure 3-10) was produced at Oak Ridge National Laboratory to characterize the • relative potential for areas to reduce acidity of acidic deposition m prior to being transferred to aquatic systems. The analysis utilized available national resource inventories and was interpreted according • to the current understanding of mechanisms of transport and • alteration of acid inputs in terrestrial systems (Seip 1980). County-level data from the Geoecology Data Base (Olson et al. 1980) H were used in the analysis to provide a regional perspective. As more • detailed data or new studies are completed, the resolution or inter- ™ pretation of the map may need to be revised. Initially, counties that were predominantly ( > 50%) urban or • agricultural were excluded from the analysis. Management and land use practices (liming, fertilizing, etc.) in these areas would tend A to dominate modifications resulting from acidic deposition. The 1977 H National Resource Inventory (USDA 1978) was used to define land in urban built-up areas and transportation corridors. The 1978 Census _ of Agriculture (USDC 1979) provided data on cropland. This resulted • in 1,648 of the 2,660 counties in the east being included in the ™ analysis. They contained predominantly forest, range, or pasture. Rapid surface runoff of precipitation or snowmelt can preclude 9 significant interaction with soils or bedrock. Steep areas with greater than 160 m of relief and elevation greater than 600 m based • on Hammond's landform map (USGS 1970) and a general topographic map • (USGS 1970) were identified as areas in which topography dominated the movement of rainfall to streams and lakes. Counties covered by 50% or more of soil types with a surface pH of ^ less than 5.0 were assigned a low potential to reduce the acidity of incoming precipitation. However, it should be noted that these soils • are naturally acid and could contribute natural acidity to aquatic | ecosystems. Criteria for natural acid generation in the soil are I I ------- 1 I I f I I I I I I I I I I I I I I I 3-41 lacking and thus it is not known how significantly acidic deposition adds to the natural acidification in areas with low soil pH (see Section 3.5.2). Thus, the interpretation of all areas having a low potential to reduce acidity as being highly sensitive ignores natural acidity contributions. Chemical and physical soil characteristics employed in the analysis represent average values for the A horizon (upper 20-25 cm) for the 82 great soil groups occurring in the eastern United States. These values were obtained from published literature (Klopatek et al. 1980). The great soil groups were combined to estimate values for the 195 soil mapping units identified (USGS 1970) in the east. Although the exact proportions of great soil groups within map units are not readily available, the dominant great soil group was given a weighting factor of 0.66 to calculate average map unit values. Proportions of soil mapping units within counties were estimated from the 1:7,500,000 scale soil map of the United States (USGS 1970). Sulphate adsorption capacity of soils provides additional neutralization of acidic water infiltrating the soil. Sulphate adsorption prevents H+ transport and can increase soil cation exchange capacity (see Section 4.5 on soil sensitivity mapping). Ultisols generally have high sulphate adsorption capacity, although few studies have determined the current status of adsorption capacity in existing Ultisols, such as occur extensively in the southern United States (Johnson et al. 1980). Counties containing 50% or more Ultisols were identified on Figure 3-10 as having high potential to reduce acidity. Bedrock influence was based on the occurrence of type 1 (low to no ability to neutralize acidic inputs) and type 2 (medium to low ability to neutralize acidic inputs) bedrock as defined and mapped by Hendrey et al. (1980). Counties having 50% or more area in type 1 and 2 were defined as having low potential to reduce the acidity of acidic deposition which comes in contact with bedrock. These are also designated sensitive. The remaining counties were generally dominated by type 4 (greater ability to neutralize acid inputs) with a high potential to neutralize acid water coming in contact with bedrock. Such areas are often called insensitive to acid rain. The influence of these factors on the ability of the watersheds to neutralize acid inputs was evaluated on a county by county basis. Although counties are generally uniform in size in the eastern United States, some of the larger counties occur along the Canada-United States border in Maine and Minnesota. For each factor, 50% or greater of land surface area was used as dominance criterion to classify counties. Therefore, significant areas can exist within counties that differ from the final designated classification. Thus, Figure 3-10 displays the broad regional patterns but evaluation of an individual county requires more detailed analysis to determine the extent and coincidence of the various factors within that county. The analysis identified the dominant factor(s) in each county that ------- 3-42 3.5.2 Aquatic - Terrestrial Relationships 1 I would determine if a county had relatively low, moderate or high potential to reduce the acidity of acidic deposition. The moderate class would both be between the extremes in reducing acidity and also B may be more variable, that is, within a moderate county, there may be I both areas of low and high of sensitivities. Seven classes were used to describe the combinations that occurred • (Table 3-9) with the agricultural/urban areas shown as blank on the map (Figure 3-10). The factors in each class and the assignment of _ low, moderate or high potential are defined in Table 3-8. Classes 3, • 4 and 6 are combinations of soils and bedrock having opposite ™ potentials for reducing acidity. In these areas, the soil depth and other terrain characteristics (such as glaciation or soil pans) will tt determine whether soil or bedrock properties dominate. Class 3 • consists of low pH soils overlaying bedrock with high ability to neutralize acid. In the south these soils are generally very thick • and bedrock would be an insignificant factor. However, in northern • glaciated areas, the thin, porous soils would probably result in a high potential to reduce acidity of precipitation through the — interaction with the bedrock. • I The maps shown in Figures 3-9 and 3-10 identify areas of low, moderate or high potential to ameliorate the impact of acidic • deposition on aquatic regimes. Map units having the lowest capacity I to reduce the acidity of atmospheric deposition should not be interpreted as representing the total land area with acidified lakes and streams in eastern North America. These are the areas where • acidification would theoretically be most pronounced provided the • input of anthropogenic acids add significantly to natural acid production or, in the case of bedrock dominated systems, exceeded the acid neutralizing capacity of the strata. In order to determine which aquatic ecosystems are already acidified, detailed water chemistry data are necessary (Section 3.6). However, as noted tm earlier, soil and bedrock information provides the best indication of • the long term capacity of watersheds to buffer acidic deposition. I I Many questions remain as to how the dilute acid in precipitation can • be transported through terrestrial systems without being dominated by ™ organic/soil buffering mechanisms. Current knowledge of terrestrial- aquatic transport fail to account mechanistically for the large changes in surface water pH attributed to acidic deposition. However, lake and stream acidification effects are observed in water- sheds where soils are present (Section 3.9). These changes are « through mechanisms not now fully understood. Such mechanisms • probably relate to rapid drainage through soil macropores as represented by root channels, voids surrounding coarse fragments (such as common in glacially-deposited soils) and other routes. Thus • at this stage sensitivity criteria are hypothetical (e.g., soil • texture). The maps presented in this section should not be I I ------- I I I I I I I I I I I I I I I I I I I 3-43 TABLE 3-9. CHARACTERISTICS OF MAP CLASSES FOR THE EASTERN UNITED STATES AS TO THE POTENTIAL TO REDUCE ACIDITY OF ACIDIC DEPOSITION (Olson et al. 1982) Class Potential to No. of Reduce Acidity Counties Characteristics Low Low Low-High Moderate High Moderate High 72 89 114 291 326 241 515 Steep slopes, high relief, high elevation Low soil pH, sensitive bedrock Low soil pH, nonsensitive bedrock Low soil pH, sensitive bedrock, sulphate adsorption Low soil pH, nonsensitive bedrock, sulphate adsorption High soil pH, sensitive bedrock High soil pH, nonsensitive bedrock ------- 3-44 I I considered strictly as sensitivity maps. They are objective representations of soil and bedrock characteristics as provided by available data bases. Map units are identified by their "potentials" B to reduce acidity based on one interpretation of the criteria. The II criteria are plainly visible if anyone should desire some other interpretation or sensitivity assessment. The application of these • maps for surface water sensitivity interpretation can, at present, • only be tested using emprically based surface water acidification data (Section 3.9). • Low pH soils (eastern U.S.) and acid podzolic soils (eastern Canada) • are representative of much of the area identified in Figures 3-9 and 3-10 as having the lowest potential to reduce the acidity of tt rainfall. According to Wiklander (1973/74) as reported in Seip | (1980) as soil pH decreases below 5.0 there is an increasing probability that streams and lakes within a watershed will receive H acid and aluminum associated with anion inputs from the terrestrial • systems due to increased acidic sulphate deposition. Because these soils generally have low quantities of basic cations (e.g., Ca^+, ^ Mg^+) a significant portion of the increased cation concentration • required to balance the increased sulphate input must be H+ and Al • (Johnson and Olson in press). Because these soils are naturally acid, they are also the ones most likely to contribute natural B acidity to surface waters (Rosenqvist 1978). Such soils do indeed 0 have the lowest potential to reduce acidity of rainfall, but they also have the potential to acidify incoming rainfall in areas where • low acidic deposition occurs (Johnson 1981; Johnson and Cole 1977). • Thus, the question of acidic deposition effect is one of quantity, that is, to what extent does acidic or sulphate deposition via the mobile anion mechanism described by Seip (1980) contribute to the • natural acidity of waters from such soils? • Aquatic ecosystems most sensitive to acidification and Al mobility, • therefore, are those areas identified as having a limited ability to J| reduce acidity. They may also receive significant inputs of anthro- pogenic acids. What constitutes a "significant input" can only be • determined at present by monitoring surface water chemistry in areas • undergoing acidification. It should then be possible to extrapolate these results to other areas by comparing watershed characteristics such as bedrock, soils, proportion of open water to watershed area • and vegetation. This would need to be carried out at a more detailed W level than the present mapping. However, the maps in Figures 3-9 and 3-10, in combination with acidic deposition maps, illustrate the distribution of the areas within which efforts need to be concentrated in eastern North America. I Groundwaters and surface waters which cross areas of differing • capacity to reduce acidity would reflect the chemistry of the most reactive bedrock or soils upstream from any sample point (Hendrey et al. 1980). Thus, both local and regional hydrological and hydro- • geological conditions need to be assessed when comparing measured • aquatic chemistry with potential of areas to reduce acidity. I I ------- I I I I I I I I I I I I I I I I I I I 3-45 It is also important to consider the topographic position of streams and lakes within watersheds. Generally, the most susceptible aquatic resources are those in the headwater portion of watersheds, or in small enclosed watersheds. During spring snowmelt runoff reaches surface waters with little or no contact with soils or bedrock, resulting in episodic pH declines. This effect may occur even in watersheds that are not very sensitive to long-term acidification. 3.5«3 Geochemical Changes Due to Acidic Precipitation Nearly all precipitation is processed terrestrially before becoming surface water. Thus, changes in soil chemistry might be expected as a result of atmospheric deposition of acids and metals. Many geochemical changes in soils are difficult to measure directly, due to large reserves of elements, and due to complex ecology. However, changes in outputs of soil and groundwaters from stressed systems may be manifested as changes in surface water chemistry. In dilute waters such as found in the Adirondacks and New England, any change should be readily observed, if historical data cover the time interval of change. Acidification rate (or lack of) is in part a function of relative maturity of the water in question. Low order streams and small headwater ponds will reveal acidification effects before major streams, rivers and lakes (Raines 1981b). Johnson and Reynolds (1977) examined the chemistry of headwater streams in New Hampshire and Vermont. These streams ranged from pH 5.0 to 7.8. Streams situated in sensitive bedrock such as granite or quartz monzonite ranged from pH 5.0 to 6.8. Total dissolved solids were generally very low (12-30 ppm), lower than TDS for many waters in areas that are not experiencing acidic precipitation. Similar results are reported for Hubbard Brook (Likens et al. 1977a). The implication is that cation denudation in New England is relatively low, in spite of acidic precipitation (Johnson et al. 1972, 1981). Studies by Schofield (1982) and Johnson et al. (1972) concur that it is not possible to conclude that increases in weathering (or increases in dissolved load) have occurred in areas receiving acidic deposition. While major cations are generally low in sensitive terrain, trace metals such as Zn, Mn, Al and Cu have been shown to be elevated in acidified systems (Norton et al. 1981b; Schofield 1982). This increase is a function of the solubility relationships for the metals, as well as atmospheric inputs of heavy metals (Galloway et al. 1980a). Continued inputs of acids may alter soil pH regimes, and result in mobilization of metals into ground- and surface-waters (Burns et al. 1981; Johnston et al. 1981; Kahl and Norton 1982). Aluminum mobilization may neutralize acids, as suggested by N.M. Johnson (1979), but once mobilization has occurred, Al species may buffer acidic waters at low pH (about pH 4.9), much like the carbonate buffer system, but at a lower pH. Once acidified, recovery ------- 3-46 of waters with high concentrations of Al may be hindered by these Al hydrolysis reactions. 3.6 ALTERATIONS OF SURFACE WATER QUALITY I I Atmospheric inputs of Pb and Zn can be demonstrated in organic soils | in New England. In Massachusetts, Siccama et al. (1980) report that Pb is accumulating in the forest floor at a rate of 30 mg/m .yr. • No increase was reported for Zn, but Zn is much more mobile than Pb. I Benninger et al. (1975) estimate that the retention time for Pb at Hubbard Brook Experimental Forest (HBEF) is nearly 5,000 years: thus Pb is largely being retained terrestrially. Other studies have • reported concentrations of Pb and Zn above background in the • northeast U.S. (Kahl and Norton 1982; Lazrus et al. 1970; Reiners et al. 1975; Schlesinger and Reiners 1974). Hanson et al. (1982) found tt a gradient of Pb in sub-alpine litter, suggesting that Pb was more | concentrated in litter (and therefore in precipitation) in southwest New England, than in northeast New England or the Gaspe" Peninsula. « Concentrations of these metals may be hundreds of times those found • in underlying inorganic soils (Kahl and Norton 1982). The study by Johnson and Reynolds (1977) did not reveal any markedly • acidic streams in New Hampshire or Vermont (low pH = 5.0). Burns • et al. (1981) also report nonacidic headwater streams in New Hampshire (mean pH 6.1). However, they conclude that significant • acidification has occurred since the 1930s, based on historical jg colorimetric data. They report that their colorimetry data agreed with their pH meter data. Regardless of the validity of time trend « comparison, they also conclude that alkalinity to total base cation • ratios are 0.2-0.5 in New England. This indicates that some chemical weathering is occurring through the reaction with strong acids, rather than by carbonic acid, with implications for surface B alkalinities in the future. Similar findings have been presented by m Cronan et al. (1978) for New Hampshire sub-alpine soils, where sulphuric acid, instead of carbonic or organic acids, is supplying • most of the hydrogen ion for weathering reactions. The net result is £ a replacement of HC03~ by SOz^ as the dominant anion inwaters of the Adirondacks and New England. This replacement is complete in very — acidic waters, and may be used as an estimate of acidification when • only partial replacement by sulphate has occurred (Henriksen 1979). ™ Schofield (1982) reports an apparent, although not significant, increase in 80^2" in Adirondack lakes during the past 15 years, B although the validity of the comparison between methods is unknown. V I The chemistry of surface water is an integrative measure of precipitation inputs and watershed influences. Altered precipitation • inputs may cause biogeochemical changes and account for differences " in regional water chemistry. The available data are discussed below as three topics: the present chemistry of aquatic systems; evidence H I I ------- I I I I I I I I 1 I I I I I I I I I I 3-47 of time-trends in water quality measures; and the patterns of seasonal or episodic variations in water quality. 3.6.1 Present Chemistry of Aquatic Systems A decade of results in Scandinavia indicates that as lakes are subjected to acidic precipitation cations are mobilized and some of the bicarbonate ions are replaced by sulphate. As a result, the normal relationship between the dominant cations, calcium and magnesium, and alkalinity is altered. Although these lakes are not necessarily acidic, the alkalinity will be less than predicted, from the sum of calcium plus magnesium (Henriksen 1980). The report by Harvey et al. (1981) gives a similar evaluation of lake data for North America. Their description is as follows: "Comparable data for lakes on the Canadian Shield are shown in Figure 4-3 [Figure 3-11 this report]. Lakes that can be considered unaffected by acidic deposition include those in the Northwest Territories, probably Labrador and Newfoundland, and northern Manitoba and Saskatchewan. These lakes have close to a 1:1 relationship between [Ca2+ + Mg2+] and [HCC^"], as do several lakes in calcareous pockets in the Killarney area of Ontario. [Ca2+ + Mg2+] may be overestimated for several of the Newfoundland and Labrador lakes, because the concentrations are not corrected for sea salt contributions. Many of the other lakes, however, have a HC03~ deficiency relative to Ca2+ plus Mg2+; most of the Killarney lakes, including all the La Cloche Mountain lakes ... almost all of the lakes within a 100 km radius of Sudbury, Ontario, and all of the Muskoka-Haliburton lakes and Nova Scotia-New Brunswick lakes (although the latter data were also not corrected for sea salt contributions), have HC03~ deficiencies. The distance below the [solid] line in Figure 3-11 may be an indication of the extent of acidification; lakes found below zero alkalinity [the hatched line] are considered to be acidified. If the other major source of Ca2+ and Mg2+ in lake water is weathering by strong acids in precipitation, and if most of this acid is associated with sulphate, then there should be a good relationship between [S042~] and [Ca2+ + Mg2+ - HC03~] on an equivalent basis. This relationship provides an estimate of the Ca2+ and Mg2+ not derived from carbonic acid weathering. Data for Canadian Shield lakes are shown in Figure 4-4 [Figure 3-12 this report]; the agreement between [S0^2~] and [Ca2+ + Mg2+ - HC03~] for most lakes is very good although it can be argued that this may be expected, based on the principle of charge balance and implies no cause-effect relationship. All of the lakes in Nova Scotia and New Brunswick have excess Ca2+ and Mg2+, suggesting that Ca2+ and Mg2+ may be supplied in part by sea salt ------- 3-48 400 300 200 § 100 D • « • MANITOBA, SASKATCHEWAN A LA CLOCHE MT LAKES • SUDBURY AREA x NWT, LABRADOR, NFLD O MARITIMES O MUSKOKA-HALIBURTON, ONT ® ELA ® MALI BURTON © OTTAWA R. dromoge © UPPER GREAT LKS. dromog. I I I I I I I I -100 100 200 300 Co2%Mg2*} 400 500 600 Figure 3-11. Total concentration of calcium plus magnesium with respect to alkalinity for lakes in Canada. Average concentrations for groups of lakes are shown as letters. Individual lake data are shown as symbols. Maritime lakes are not corrected for seasalt contribution. Solid line represents theoretical relationship for lakes unaffected by acidic deposition (see text for explanation and data sources). The scatter of data is due partly to the various techniques used to measure alkalinity (modified from Harvey et al. 1981). I I I I I I I I I ------- 1 1 1 1 | 600 • 500 400 1 \m 1 i™ 200 • 100 1 1 1 H Figur 1 1 1 3-49 / • . : •/. ; •••'/ • ^^ .* ; I? ' • MANITOBA, SASKATCHEWAN I !• »'H-V '• • A LA CLOCHE MT LAKES * '•/•:' . • A • .. j& ' :.vr ' ' SUOBURY AREA .^5® "•". •' * N.Wt, LABRADOR, NFLO. ah ''Xv." "'" • D MAR1TIMES **"y&' >^*'* '• * " ° MUSKOKA'HAI-I8URTON. ONT ~ ''/^ ' ' ' ' • ® ELA „ / * * ® HALIBURTON s O ~~* . / 9 OO @ OTTAWA R drainagt / <6 QcD D B / ®< • ® UPPE" GREAT LKS drainage "• / / » x / 1 1 1 1 1 1 1 1 1 1 1 1 100 200 300 400 500 600 [Co2** Mgz*-ALKALINITYj(/ieq/L) e 3-12. [Ca2+ + Mg2+ - alkalinity] vs. [S042~] for lakes in Canada. Solid line represents theoretical relationship for [Ca2+ + Mg2+] not derived from carbonic acid weathering reactions (modified from Harvey et al . 1981). ------- 3-50 I I The fact that most lakes or groups of lakes have reasonably close correspondence between [Ca2+ + Mg2+ - HC03~] and [S042~] supports the hypothesis that the Ca2+ and Mg2+ • content of lakes is not derived from carbonic acid weathering, • but is related to the input of sulphate. Because the other major strong acid anion, NOg", is a nutrient in lakes and • streams, (i.e., is non-conservative), it is not possible to | incorporate it into a more complete relationship." In a study of the most recent available data for lakes in Quebec, H Bob§e et al. (1982) have shown that a similar relationship between * [Ca2+ + Mg2+] - [alkalinity] and [SO^2"] exists on the southern slope of the Canadian Shield in Quebec. Figure 3-13 shows the • different hydrographic regions sampled and Figure 3-14 shows the • sulphate versus [Ca] + [Mg] - alkalinity relationship for six of the regions. The highest sulphate concentrations and greatest alkalinity fl| deficiencies were observed in the southwest part of the province I (Region 04). The concentrations of sulphate and the alkalinity deficits decrease to the north and east. In Region 10, sulphate — concentrations average about 30 yeq/L and the alkalinity values were • equal to or slightly greater than the calcium plus magnesium values ™ indicating no alkalinity deficit. This supports the hypothesis of atmospherically deposited sulphur being a major influence on lake B chemistry. • Current data on pH, alkalinity, sulphate, and other chemical M variables are available for surface waters in a wide variety of I climatic, geological and biological conditions in eastern North America. These data give an idea of the current chemistry of aquatic _ systems, but do not necessarily indicate how or when that status was • achieved, or whether it is currently changing. Some insight into • these questions is given by comparisons of water quality data from areas with quite different rates of acidic deposition. This approach I was used by Thompson and Button (1982) for lakes in Canada from ELA • (Experimental Lakes Area, Kenora, Ontario) eastward to Labrador and Newfoundland (Figure 3-15). The formal names of the lake regions, m data sources , and the range of latitude and longitude including the • sampled lakes are shown on Table 3-10. Concentrations of sulphate, and of excess sulphate near the coast, were multiplied by the basin runoff of water, obtained from hydrographic records of the basins or • were approximated from hydrographic charts (Fisheries and Environment ™ Canada 1978) to determine the sulphate flux or specific yield of sulphate for each basin in units of mass per unit area per year. • Lakes used for this comparison exclude those where geological sources | and direct industrial or municipal discharges of sulphur to the water body were obvious. Therefore, Thompson and Hutton (1982) assumed « that the primary source of the basin sulphate yield was atmospheric. • As a test of this assumption they compared the values of sulphate flux with the estimated atmospheric loading of sulphate in I I ------- I I I I I I I I I I I I I I I I I I I 3-51 HUDSON BAY GULF OF ST LAWRENCE 03 UNITED STATES ONTARIO Figure 3-13. Hydrographic Regions of Quebec (Bobge et al. 1982). ------- 3-52 | 1 30-i 24 - 18- •4 O — 1 £. 06- o- 30- Q 24- •^ IT 0 18- 12- T"1 ' <% 06- O CO 30- 24- 18- 12- 06- 0- , REGION 04 ! R = 0.51 A\ ! N = 71 /T " j • /w.^.r. 1 '. ..------- I I I I I I I I I I I I I I I I I I I 3-53 0) CO eo to to o CO »- CM CO CM CM CM CM CO to CM o in CO CM 6 o CM LABRADOR IS. OF NFLD HALIFAX NEW BRUNSWICK LAFLAMME MAURICIE N. OF OTTAWA ALGONQUIN HALIBURTON SUDBURY ALGOMA THUNDER BAY QUETICO ELA D O O O O O o o *> o w o w o in ra eo CM CM T- »- (I/barf) uojjBjjuaouoo a)Bi|d|ns rt__ lean (*) and range of SO^ concentrations and excess concentrations in anadian lakes on an approximately west to east axis. The numbers to the 1 he means are the number of samples; where there are two numbers, the lower umber of lakes (Thompson and Button 1982). m bo ------- 3-54 Latitude Longitude 0 4J td U 0 3 U> cj 3 O en B to CO 4J 01 en id 4J & CU 3 m -a o CO ON 1 m CO o CO ON m o ON -a- 1 m CO o ON 00 01 Q. •H E 1 Institute, \ rea, Kenora Freshwater al. 1976 < 4-1 CO 01 Sx cd co J -H a •-H CO id cu u pa B 01 « S 0 iH iH M M 0) CO X B U O 8 o CM ON 1 0 O o O ON b o o ON 1 b o o 00 4-1 01 Jnviron >ratory MH W X iJ JJ rH •s s o 14 00 4-1 01 CO QJ -H lie Lacs Ontario Hit Unit Lakes Thunder Ba) S4J B 0) ac des ssessm J > 00 z$ CO •H B B IH d 01 £ 4J CO s §1 -- X CO 4-1 4J M 0 CO o ON CO 1 b o o ON 00 b o ON b -a- 0 CO 4J 01 Snviron >ratory M-l W (D Oj *£ I-J rH (^ nj O B O 14 00 4J Q) CO tf •H »-" rt es, T. Marshall, Ontario Mil Natural Resources Thunder 833 50 S& H u < CO iH >N E CO iH m £ IH O CU J eu 4J id M i-H O 3 CO -en to •O - to OJ E 0 CO -H U > k4 District, Environment s, National Forestry S« §cu •H O M H MH < CU CO • cu o 3 . o >t iH 01 IH M a £§ eu 3 4J •H 4-1 CO B H O M Resea u CU 4J cd 3 o 0 0 CM CO 1 o CO o O 00 o CM o 1 b CO o NO 4J B eu 01 rH B id E TJ O X > B -a to 4J eu CO MH IH °s 14 4J 4J CO CO •H 10 ri n\ e Trout Lakes, Ontario Mil istry of Laboratori< Ji B td IH CO 4-1 O 01 CJ -H CJ iH IH r4 h cd 3 4J 4J O CO B CO °°.'S >N B~rH M E td 3 3 M ,0 CJ 3 *O 4-1 3 • cd en ^ Z O o ON 1 O O 0 CO ps. O CM o m 1 b o B 01 01 iH e id E -O 0 X 6 T3 cd X 4J 4J CU CO MH IH 14 4-1 4-1 fd CO •H CO ^- di Fisheries Ontario Mil s, D. Loftus, Laboratori< m oi •a M B to CO , 4J co •H B •H O •H M id 4J § m m o CO 1 in -a- o b o o NO -a- 1 m CM o m 4J B 01 eu -^ B *0 B TJ 0 X M 01 •H Pd E -0 CO 0) 4-1 01 CO MH t-t IH 4J 4J 10 -H CO IH n\ Assessment Unit Ontario Mil ario Ministry Laboratori< 4-1 CO CO E 01 01 O CJ •H M M - 3 CU CO O X r* CO CO CJ CU •H -H pd rH B • CO •H HH IH 3 3 O1 - 4-1 B 10 id O CU Z 00.* rH CO MH < rJ 0 o o NO 1 CO o CO o o NO 1 b -a- 0 in 4-1 E 1 01 wironn M rH •8 01 01 4J en iH |H th of Ottawa, Quebec, Mil I'Environnement Ste. Foy 0 oi 0 T3 Jg J!8 01 CO s-g E£ •H CJ CO CU CU ,^N M 01 CO 3 rJ & O 0 CO 1 O m 0 CM -a in o -a 1 m -a- o vO 4J •H rH CO 01 : Canada, Wat E 01 I •H I <0 i— | -S rH CO E O •ri 4-1 td id E •H CO 01 3 rH •H 3 01 3 00 , Environment Branch, Lot n X 0 3 -H 01 00 i-l 01 B P£ U at a 01 iH * o id •H -0 M td 3 B £3 b 0 0 b CO ON O i-H id cu : Canada , Wat igueuil ieux, Environment Quebec Region Branch, Lot B cu . I-H CO •a • cd Pd E id -CJ cu af lamm onment rJ M •H CJ > CO B O o NO 1 CO o NO CM o m •a- 1 0 o in rH to B 01 S to T3 to E S to O CJ 4J O w Brunswick, Environment Branch, Mot Ol Z E U 01 ^ X 0 4J CO 3 0> 0 -H CO ^ 5§ CO CO IH 01 Ol id oi 1-J PH CM CM o -3- NO 1 CM -a o CO NO b o o m 1 S 0 -a iH id cu S id •o id B CO CJ E i o u •H > B W •H id iH X Halifa near CO 0) 3 a 0 o £° x" o B cd B 0 •H 00 S CJ •H 4J E to rH «" •o to E to onment IH •H w 0 1 o CO in o o m 1 o sr o rH id cu S CO C3 E E id O CJ 4J O he island of Environment 1981a Branch, Moi 4J C iH o id i-H 10 O 01 ater 1 undlan 3 o T3 >W S z m o NO 1 O CM 0 s b CO o in 1 in o CM rH CO 01 : Canada , Wat icton r 1981b Environment Branch, Mot •H id rH CJ CO cu to t-H IH 0 T3 to IH X 3 I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-55 precipitation for the different areas studied. The estimated atmospheric excess sulphate deposition rates in precipitation (flux per unit area per year) in units comparable to the basin specific yield as well as the range of estimated deposition for 1977-80 were obtained from measurements or interpolation of measurements from the CANSAP precipitation network (Barrie and Sirois 1982). These values are shown on Figure 3-15 for direct comparison to the basin-specific yield of surface waters. Also shown on Figure 3-16 are dry deposition of sulphate calculated from measurements of sulphur oxides in air at four Air Pollution Network (APN) stations (ELA, Long Point on Lake Erie, Chalk River, Ontario and Kejimkujik National Park, Nova Scotia (Barrie 1982). The agreement between the estimated deposition and basin specific yield of sulphate is generally good but shows greater yield than deposition in the areas of highest yield (i.e., the region between Thunder Bay, Ontario and Halifax, Nova Scotia). This deficiency of sulphate measured in precipitation as compared to basin yield of sulphate may, at least in part, be due to dry deposition of sulphate and sulphur dioxide. The dry deposition would be greater in regions nearer to or downwind from industrial sources. There may also be some release of sulphate previously stored in the basin. Contributions from geologic or other sources cannot be entirely dismissed in all cases. However, in these areas the evidence is strong that the atmospheric deposition of sulphate is the primary source of the basin yield of sulphate. In the province of Quebec there exists a strong south to north gradient of lake sulphate concentrations as illustrated in Figure 3-17 and Figure 3-14. The data were obtained from 256 lakes in the province. The highest observed concentrations are in the southwest portion of the Province, and reach 180 yeq/L. The concentrations decrease gradually toward the north and to the east to values around 30 yeq/L. More than 80% of the lakes have sulphate concentrations higher than 60 yeq/L (Bobee et al. 1982), equivalent to the upper background level for lakes on the Precambrian Shield (Harvey et al. 1981). Haines and Akielaszek (1982) reviewed available data on surface water pH distributions in sensitive regions. They found that the regions receiving precipitation of lower pH had higher percentages of low pH lakes (Table 3-11). A number of other studies have documented the present status of surface water resources in regions of Canada and the United States. They are summarized in the following sections. ------- -*—i co «- IN a -i O a) co >i O M <4-l "O §4J CO Cfl M 3 <4-l t-H M ft -H O cfl 00 4-1 H eX -H CO CJ 0) CO M CO CX cu o a X! -H 0> S ^ O O 4-1 J3 CO CO 4J O O 0) co S 4-1 H 0) i-l •< M co 3 O co a • cfl 0) x^ 01 *s e ^J ^^ C^ cu ^H oo T3 C cu ^-1 O 4J 0) -H Cfl •H 4J 0 CO 4J O O CO •H a CU "4-1 CD •H T3 <4-l CJ O CU 01 PL, 4-1 01 co ct) bO a co CO •H O M •rl CO pi CU C cfl O 4-1 CO cfl "H -O ^ cfl c a o 6 •rH O 4J CJ •H CO 14-1 O O ex 0) CO •O X Cfl bO Ol CO CJ CM *-l -H 00 !-i CT> *^ CU '"^ cfl a c CO O a o 4J Cfl 0 4J £ m 5 s cfl re vO f-H I CO cu !-i 3 bO •H ft. o> •H 3 Ol "O M CJ ,C M H H C CO ctj O PQ O -H •^> • 4J Ol ^ -H J^ 4-1 CM CO M Cfl 00 O O ,*"i ^^ Ci* 5 CX —i CU 0) 3 0) C CO -H >» M M O ctf 4-1 C! Cfl O CO CU *£ o •H CO -H & CJ O 4-1 i-H cu a co cfl M CU 4-1 43 CX"O CO O I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-57 ------- 3-58 TABLE 3-11. REGIONAL WATER CHEMISTRY SURVEY RESULTS FOR SURFACE WATER pH DISTRIBUTION (Raines and Akielaszek 1982) LOCATION NUMBEI OF LAKI AND STRI Areas New England West Sweden West Sweden South Sweden South Norway South Norway Denmark Scotland Nova Scotia Quebec Central Ontario La Cloche Mountains, Ontario Sudbury, Ontario Adirondack Mountains , New York I PERCENT IN pH RANGE JO 5AMS <5 5-6 where Precipitation 226 314 15 51 155 719 14 72 21 25 26 152 150 849 Areas where North Norway Northwest Wisconsin North Minnesota 77 265 85 8 36 27 2 18 64 29 26 52 12 8 28 13 25 Precipitation 0 0 0 21 21 47 20 38 33 57 36 24 40 58 34 15 30 >6 Averages 71 43 27 78 44 3 14 38 24 48 34 38 72 45 Averages > 13 6 0 87 94 100 REFERENCE pH 4.6 Haines and Akielaszek 1982 Aimer et al. 1974 Dickson 1975 Malmer 1975 Wright et al. 1977 Wright and Snekvik 1978 Rebsdorf 1980 Wright et al. 1980 Watt et al. 1979 Jones et al. 1980 Scheider et al. 1979a Beamish and Harvey 1972 Conroy et al. 1976 Pfeiffer and Festa 1980 pH 4.6 Wright and Gjessing 1976 Lillie and Mason 1980 Glass and Loucks 1980 I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-59 3.6.1.1 Saskatchewan Measurements of total alkalinity, calcium, magnesium and pH were analyzed for some 300 lakes in Saskatchewan's Precambrian Shield and fringe Shield regions (Liaw and Atton 1981). Concentrations of alkalinity in these lakes varied from 10 to 1740 yeq/L. Forty-four percent of the lakes surveyed had alkalinities of 200 yeq/L or less. Measurements of lakewater pH, ranged from 5.6 to 8.2, and indicate that, at present, Saskatchewan's Shield lakes are circumneutral. Lakes with pH values between 6.5 and 7.5 accounted for nearly 80% of all lakes investigated. Concentrations of calcium ranged from 7 to 630 yeq/L. About 54% of the lakes surveyed had calcium of 80 neq/L or less, while 25% had between 80 and 160 yeq/L. Concentrations of magnesium varied from 0 to 130 yeq/L. Approximately 65% of the lakes measured had magnesium concentrations of 24 yeq/L or less, whereas 32% had between 24 and 36 yeq/L. Concentrations of calcium plus magnesium showed a one-to-one relationship to alkalinity (Figure 3-18). This relationship is expected in areas where alkalinity production is by bicarbonate weathering in the absence of strong acids. 3.6.1.2 Ontario Alkalinity data for 2,624 lakes in Ontario are shown in Table 3-12 (OME 1982). The categories from 1 to 5 indicate decreasing sensitivity to acid deposition. The 48% of the lakes in categories 2 and 3 had some measurable alkalinity less than 200 yeq/L and may be regarded as sensitive to acidic deposition. The spacial distribution of the lakes sampled is shown in Figure 3-19. In Precambrian areas, up to 90% of the lakes are less than 200 yeq/L. Five percent of the lakes had alkalinity values less than zero, i.e., acidified, these lakes are located mainly in the Manitoulin and Sudbury areas which have been subjected to deposition from smelting operations in Sudbury. Scheider et al. (1981) indicate that acidic deposition to the area is substantial. Chan et al. (1980) concluded that much of the deposition is due to long-range transport of acid. The influence of long-range transport is relative to the historic local emissions, with respect to acidifying lakes, cannot be determined. Data from 16 intensively studied lakes in Muskoka-Haliburton are plotted in Figure 3-20. The average epilimnetic summer pH and the lowest spring pH observed in the surface waters are plotted against the mean summer alkalinity values. The data cover 4 years (1976- 1980) . Lakes with alkalinity of less than 40 yeq/L experienced pH depressions in surface waters to values less than 5.5. A few streams showed pH values below 4.0 in some cases (Figure 3-21). At Algoma, spring pH values of about 5.0 occur in the surface waters of study lakes with alkalinities less than 40 yeq/L (Scheider 1983). ------- 3-60 1500- ^ >. 1000- 500- y = -18.204 + 0.915x r=0.970 n = 281 500 1000 1500 Calcium+ Magnesium (jjeq/L) Figure 3-18. Relationship between alkalinity and calcium + magnesium for northern Saskatchewan lakes. Broken lines indicate 95% confidence limits of predicted values (Liaw 1982). I I I I I I I I I I I I I I I I I I I ------- I 3-61 I I in V) ro 0 j>- c 1 _ * O3c ro^ xo(D^«)3 EOi_(04- ja~ci-L.inO4-— Ov)j= OOjCJC C >» — 4- O O IDT) C — T3 .* — 4- C7i3Ouo^;_ ISZOCLOQ: ID in o oo o • r» • CM CM • 4- Q • -a O c O 3 O 18 ID 1. 4- l- 5 ID Q. to i- •— mtoo\mr^-'— infNoNf^^ in O\ O CM •— o CM •— to to — OfOOtO«— OOOmvO^fO OI^O — inoOCMtCM^-O — — •— •— *~ • •— «ot~ • •— in«-vo • — CM CM — CM CM in • to •Ttav • •cor— in • • •— VO — CM CM CM • IO • IO fO • • to to ^ • • CM — • • • • • O 4- 4-4- O O in in in •o Q • a Q o> a i. • o tn >• u> O 3»a>o « o — ro c o O-D>OO am — • 1. UJ — O4- Ero» oo o:s c >- u ro — o CjjOO OOQL-O^UO i-o>-i.OE3T3inO — SX>-M ro 1- O 3 (D D) l_ — Q. U_ ID I. jz O 4- **- O) O O ID O 1. ro t * — o L. — 4- JC in x 58 c c 3 3 O O o o ID JQ ------- 3-62 I Figure 3-19. Distribution of lakes sampled in Ontario Ministry of the Environment 1981 and 1982 surveys. Numbers indicate number of lakes sampled within each grid cell (1° by .5°) and letters identify county or district as per Table 3-12. I I I I I I ------- 1 1 1 1 -]> £ ^H if. V 1 1 1 *A 1* • ** • 1 * • Iile ilc • * 5p-=F • *A w 1 1 *A V 1 Iq 10 o iq q 10 c Is- ^ T I Hd 1 ' 1 1 o plO T" - " U) -co - " o T~ • 10 -o T- _ lo . . 10 N- - o (O " . • _ _IO " - o ~co . _w - ~ o 3 i 3-63 o CO vD C ON cd *™( r-N ^ * C x« / O 4-> a ^ pt ^j 0 -H •H iH 4-1 Cfl 0) EC § 03 •H ^ rH O •H ^ Pu CO ^ M O" 1 -H (U 1 3 3 to O* to a) C (fl 0) iH >> <1J *^"^ rt ^O • ^~ »s 03 O M Jv i-J 0 , bo C -rH -H .t CM r= -H o. Q. -H w H i n) 111 J------- o c 0> 30 25H 20H 15H 0) DC 3-64 3.5 4.0 4.5 5^0 5!5 6.0 6.5 7.0 Minimum pH Observed in the Streams Figure 3-21. Minimum pH values of 57 headwater streams in Muskoka-Haliburton, 1976-80 (Scheider 1983). ------- I I I I I I I I I I I I I I I I I I I 3-65 3.6.1.3 Quebec Vast areas of the Province of Quebec are composed of non-calcareous lithology, and glacial transport of materials has not provided calcareous tills to modify the local soil structures. Only in the marine sediments of the St. Lawrence Valley and calcareous lithology of the Gaspe" region and a few other areas are sufficient buffering materials found. A few local occurrences of more adequately buffered waters are found in the Gatineau Lakes, Lake St. Jean, Lake Mistassini, and the Harricana River in northwest Quebec. It is estimated that the Province contains greater than one million lakes, the vast majority of which have not been surveyed. Present surveys have been limited to the southern, more accessible areas. Examination of the alkalinity and CSI values of the surface waters of the surveyed area gives an indication of sensitivity of waters of Quebec to acidification. A distribution of calcite saturation index (Conroy et al. 1974) of 181 lakes surveyed during the summer of 1980 (Bobge et al. 1982) is illustrated in Figure 3-22. Values lower than 3 are not very sensitive (15% of the lakes surveyed have a value less than 3.5). Values between 3 and 5 are potentially sensitive (48% of the surveyed lakes have a value between 3.5 and 5.5). Values higher than 5 are extremely sensitive to acidification (37% of lakes of the shield have a value higher than 5.5). The distribution of CSI of surveyed lakes in Quebec, is illustrated in Figure 3-22. It is evident that nearly all waters, other than the St. Lawrence Valley and the Gaspe" Regions (Regions 01, 02 and 03, as per Figure 3-13), have CSI equal to or greater than 3 and are, therefore, sensitive to acidification. Surveys of the lakes of Laurentide and La Mauricie Parks appear to indicate a greater sensitivity than do lakes in the surrounding regions. This may be an actual indication of local differences in terrain geochemistry, but is believed to result from over-estimation of alkalinity or pH in the older measurements. The actual sensitivity of lakes in Quebec may, therefore, be even greater than indicated by the older surveys (Ahern and Leclerc 1981; Jones et al. 1980). Bobe'e et al. (1982) have shown that 19 of 20 of the lakes sampled on the Precambrian Shield in Quebec south of 50° latitude have alkalinities less than 200 yeq/L, and thus are considered to be sensitive to acidic deposition. For the same region, summer values of pH were below 5.0 for 15% of these lakes, and below 5.5 for 41% (Figure 3-23). The pH frequency distribution has two modes: one between 5.0 and 5.5 and one from 6.0 to 6.5. In a lake with only carbonate species to buffer the water, a pH value of 5.5 indicates that the lake can have rather large pH fluctuations. "The importance of the sulphate anion in the lakes of the Canadian Shield (in Quebec) appears clear upon examination of the relationship between bicarbonate and sulphate. ------- 0 o m o 00 3-66 0 U3 o o 0 o 0 00 o 00 M 01 I c M O en en 0) oo M o 14-1 en CU o •H C CM O 00 •H ------- I I I I I I I I I I I I I I I I I I I 0 (XI LO 0 ID 3-67 o o o en o 10 o h- o to CN 00 o P3 O 00 CO Jl o 0) I CO (!) S3 Pu CO CM CO 3 bO •H ------- 3-68 3.6.1.4 Atlantic Provinces I I According to Dickson (1975), this ratio should greatly exceed one in lakes not influenced by (atmospheric) sulphates. In the southern portion of the Shield region • (in Quebec), the sulphate ion dominates in these lakes. | For the entire area of the Shield (in Quebec), 84% of the lakes have a HC03~/S042~" ratio less than one." • [Translated from Bob€e et al. 1982] (See Table 3-13 and Figure I 3-24.) " I I Except for isolated regions of calcareous lithology, mostly in northern and eastern New Brunswick, the northern peninsula of Newfoundland and all of Prince Edward Island, the Atlantic provinces have soils and bedrock that provide limited acid neutralizing _ capacity. Much of the area is of very complex geology which has been • indicated in the sensitivity maps (see Section 3.5). However, summaries of the surface water chemistry by Clair et al. (1982), Wiltshire and Machell (1981), and Thompson et al. (1980) have shown • that large portions of these waters are very dilute and poorly • buffered. Clair et al. (1982) have summarized the Atlantic Provinces water chemistry in terms of the Calcite Saturation Index (Kramer • 1976). CSI maps for New Brunswick, P.E.I., Nova Scotia, Island of | Newfoundland and Labrador are illustrated in Figures 3-25, 3-26 and 3-27. If CSI of 3 or greater is taken as an index of highly _ sensitive waters, it is evident that large portions of the surface I waters are sensitive to acidification. * The loss of alkalinity and consequent decline in pH of some lakes and I rivers of Nova Scotia have been well documented (see Thompson et al. m [1980], Watt et al. [1979], Wiltshire and Machell [1981]). Wiltshire and Machell (1981) applied Henriksen's (1979) comparative • relationship to data from 16 lakes in Nova Scotia and suggested that | acidification (loss of alkalinity) of 40 to 50 yeq/L has occurred over the past two decades to 1979, which is consistent with measured _ pH declines. Watt et al. (1979) have shown similar pH declines for • lakes near Halifax but attribute this decline to sulphate deposition ™ from local sources. Some pH increases have also been identified in rivers of southwestern Newfoundland, (Thompson et al. 1980). • Thompson and Button (1982) concluded that lower levels of sulphate • deposition over Newfoundland and Labrador have apparently resulted in only moderate alkalinity replacement. • Bogs are a common feature of the Atlantic provinces and waters often carry significant organic contents. Although there is a need to more clearly define the role of these natural acids in determining the • acidity of waters and subsequent influences in metal availability, ^ ionic balances by Thompson (1982) for a number of these waters suggest that the major acidity is due to inorganic ions. • I I ------- I I I I I I I I I I I I I I I I I I I 3-69 TABLE 3-13. SOME STATISTICS ON THE RATIOS OF HC03/S042~ FOR WATERS OF QUEBEC DERIVED FROM LEGENDRE ET AL. (1980), BY HYDROGRAPHIC REGION (see Figure 3-13). VALUES OF THE RATIO 0.6 0.6 - 1.2 1.2 - 1.8 1.8 TOTAL LAKES NORTH OF THE ST. LAWRENCE RIVER WEST EAST HYDROGRAPHIC REGION 04 05 06 07 N %N %N %N % 24 21.5 5 8.6 3 50.0 9 64.3 41 36.6 15 25.9 1 16.7 3 21.4 19 17.0 14 24.1 1 16.7 2 14.3 28 25.0 24 41.4 1 16.6 0 112 100.0 58 100.0 6 100.0 14 100.0 VALUES OF THE RATIO 0.6 0.6 - 1.2 1.2 - 1.8 1.8 TOTAL LAKES SOUTH OF THE ST. LAWRENCE RIVER WEST EAST HYDROGRAPHIC REGION 03 02 01 N % N % N % 21 95.4 13 76.5 5 100.0 1 4.6 3 17.6 - - 0 1 5.9 22 100.0 17 100.0 5 100.0 ------- 3-70 CNJ 00 ON CD VOJ ,C O M O oo to 03 0) O 0) CO o; iH 0) I CNJ I CO o CN I CO 3 oo •H I I I I I 1 I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-71 Ill o _J o u_ o z o 1- m DC H- co Q X LJLJ Q Z Z O H- QC P^ <(- CO CO HI D < > § T3 W 0) O C •H O •H C !H CQ 01 00 CO 01 4-1 3 0) iH Q U > -H « X iH CU U T3 C E •H O U C M-l O •H 13 3 -H CO O co E x-^ QJ 4J CO O 4.) i-l O nj o O C/3 4-1 CO O > o •H 13 •u C 3 cfl fi •H 13 ^ C •w CO CO ^H •H CO Q M CN I 3 bo •H ------- 3-72 DISTRIBUTION OF CALCITE SATURATION INDEX VALUES Figure 3-26. Distribution of calcite saturation index values for Newfoundland (modified from Clair et al. 1982). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-73 DISTRIBUTION OF CALCITE SATURATION INDEX VALUES Figure 3-27. Distribution of calcite saturation index values for Labrador (modified from Glair et al. 1982). ------- 3-74 3.6.1.5 United States I I I The national map of total alkalinity of surface waters by Omernik and • Powers (1982) illustrates general patterns of surface water sensi- ™ tivity to acidic deposition on the conterminous United States (Figure 3-28 in map folio). A large number of regions in the U.S. have mean annual alkalinity values below 200 ueq/L in surface waters. In the portion of the country where continental glaciation resulted in high densities of natural lakes, these low alkalinity areas M comprise: (1) much of New Hampshire and central and southern Maine; • (2) the Adirondacks in northeast New York; and (3) the northeastern tip of Minnesota and a portion of northcentral Wisconsin. An analysis of 300 headwater lakes and streams in six northern New V England sites shows that alkalinity values of less than 200 yeq/L • cover most of the regions examined, with widespread values <20 yeq/L as shown in Figure 3-29 (Raines 1981b; Raines and Akielaszek 1982). • In the West, streams and lakes with average alkalinity values below | 200 yeq/L are generally found in the higher mountainous areas, particularly the Cascade Range of Oregon and Washington and the ^ Sierra Nevadas in California. I Elsewhere in the United States, sensitive surface waters are primarily streams, small lakes and general purpose reservoirs. For I these waters there are several areas of low alkalinity values: (1) a • discontinuous region extending from southeastern New York to western Pennsylvania and central West Virginia; (2) eastern North Carolina; • (3) central South Carolina and southeastern Georgia; (4) an area | centered on the southwestern end of the Blue Ridge Mountains; (5) a band extending from southeastern Louisiana to northeastern Florida; _ (6) southeastern Texas and westcentral Louisiana; and (7) smaller I areas in southern New Jersey, northwest Alabama, southern Arkansas * and northern Louisiana, and the Quachita Mountains across the Oklahoma/Arkansas border. 1 As indicated by the cautionary note on the face of the alkalinity map, the "map is intended to provide a synoptic illustration of the • regional patterns of surface water alkalinity in the United States. • As such, it affords a qualitative graphic overview of the sensitivity of surface waters to acidification. The map should not be used for _ making quantitative assessments of the extent of alkalinity or I sensitivity" (Omernik and Powers 1982). ' 3.6.2 Time Trends in Surface Water Chemistry | Questions of past and potential future changes in surface water • acidification are best answered by detailed analysis of available I long-term data. Such studies also give an indication of natural trends or an anthropogenic effect. Care must be taken in any historical studies, however, to account for differences between older • methods of measurement and current methods. Precautions have been " taken in the following analyses to correct for methodological I I ------- I I I I I I I I I I I I I I I I I I I 3-75 <20/jeq/L 20-200 ^eq/L >200 jueq/|_ states ------- 3-76 differences, but these corrections add uncertainty to some of the comparisons. 3.6.3.1 Time Trends in Nova Scotia and Newfoundland I I I 3.6.3 Time Trends in Representative Areas Historic water quality data exist for several areas in eastern North • America. The data must be verified based on sampling and analytical methodologies before comparisons with modern measurements can be made. Acidification represents a loss of alkalinity, but measure- • ments of alkalinity are sparse, and therefore, most time trend | comparisons have been restricted to pH changes. The following section will review time trends in pH measurements and alkalinities • where available. • I Several rivers in Nova Scotia and Newfoundland were sampled and analyzed by Thomas (1960) during the period 1954-55 in Newfoundland flj using carefully described analytical methods. Several of the same | sample locations were continued under the Environment Canada, National Water Quality monitoring program in 1965. This monitoring w on a monthly basis was continued through 1974 after which most • stations were reverted to seasonal sampling. Monthly sampling was reinstated for some stations in 1979. The methodologies are described and data are archived in the Environment Canada National I Water Quality Data Archive, NAQUADAT. pH has been determined • potentiometrically throughout the period of record. Laboratory samples from the early data of Thomas were stored in soft glass • sample bottles which may have caused a small increase in the | laboratory measured value of pH. This factor would be unimportant for the field measured pH values. Sulphate was measured by a • BaCl2 precipitation gravemetric method prior to 1954. Later • determinations were by the colourimetric procedure which was automated in 1973. This data set has been employed in several studies to analyze trends » and changes that may have occurred in the chemistry of the waters during the period of record. Thompson et al. (1980) examined the pH • records for three rivers of Nova Scotia and three rivers of | Newfoundland. The pH values were accepted as comparable and, while statistical analysis was not undertaken, there appears to have been a • decrease in the discharge-weighted mean pH of the Tusket, Medway and I St. Mary's Rivers of Nova Scotia during the period 1966 through 1974. The one year of record 1954-55 indicates a discharge-weighted mean pH greater than the following years of record. The Isle aux Morts, • Garnish, and Rocky Rivers of Newfoundland indicate minimum discharge- • weighted pH in 1972 or 1973 in the period of record 1966-1980 or 1981. Values have increased since that period. • I I ------- I I I I I I I I I I I I I I I I I I I 3-77 Glair and Whitfield (1983) have subjected portions of the record of several of these rivers ta statistical analysis, where the record was continuous and uniform. They classified the rivers by the CSI sensitivity index (greater than -3, insensitive and less than -3 sensitive). Periods of record were limited to 1965-66 through 1978-79 and some records terminated in 1973-74. They have reported decreasing trends for pH of the Medway, Isle aux Morts, and Rocky Rivers and stationary records for the Piper's Hole, Lepreau and Mersey. Among the insensitive rivers, all were either stationary or increasing in pH records. Glair and Whitfield (1983) also analyzed the trend of sulphate but did not apply a correction for seasalt contribution. While some trends were reported, they are likely to be strongly influenced by seasalt in this marine coastal environment and cannot be interpreted as trends in excess sulphate. Watt et al. (1983) has compared the major ion concentrations (corrected for seasalt) for several Nova Scotia rivers for the 1954-55 through 1980-81 period using the same data set. He has found smaller values of bicarbonate, greater values of sulphate and greater hydrogen ion concentrations, all at greater than 1% significance level for the Roseway, Medway, Mersey, and La Have Rivers as shown in Table 3-14 (from Watt et al. 1983). While caution is required in interpreting SO^- trends due to interference of organic anions in measurement procedures, the bicarbonate and hydrogen ion concentration trends are not subject to such caution. Other major ions did not show significant changes. Farmer et al. (1980), using the same data base, has compared major ion concentrations (unweighted) for 1954-55 with 1978-79 for the Mersey River. The pH has decreased from 5.8 (range 5.4 to 6.6) in 1954-55 to 5.2 (range 4.9-5.4) in 1978-79 while sulphate has increased from 1.6 mg/L (range 0.1 to 3.0 mg/L) in 1954-55 to 3.3 mg/L (range 1.0 to 5.0 mg/L) in 1978-79. Other rivers of Southwest Nova Scotia examined by Farmer et al. (1980) include the Tusket, Clyde, Roseway, Jordan, and Medway. Decreased pH values were observed in all these rivers. pH of the La Have River has changed little over the same time period "... reflecting deposits of sandstone in this area." Farmer et al. (1980) have stressed that the Nova Scotia rivers having the greatest pH change and the lowest pHs in 1978-79 were also the most highly coloured. Thus the contribution of humic and/or fulvic acids to the total acidity may be significant. Extensive direct measurements of the organic anion concentration have not been reported. Preliminary measurements by Oliver and Slawych (1982) of samples from the West, Medway and Mersey Rivers indicated an organic acidity of 95, 94 and 53 yeq/L respectively as compared to an estimated precipitation acidity of 29 yeq/L. Thompson (1982) has observed that for the Roseway, Mersey and Medway rivers while "their pH's have been thought to be dominated by naturally occurring organic acids, their low pH's can be explained quite well on the basis of simple inorganic chemistry." More direct measurements of the organic anion concentrations are needed to define the relative contributions to these waters of very low total ionic strength. ------- 3-78 Ion Concentrations LeHave 1954-55 0.072 0.040 0.019 0.008 0.001 0.070 0.017 0.007 1980-81 0.036 0.081 0.024 0.006 0.001 0.069 0.030 0.017 Average difference -0.036 +0.039 +0.010 -0.004 +0.009 -0.012 +0.006 +0.009 Significance level <0.001 <0.001 N.S. N.S. <0.001 N.S. N.S. <0.001 I I TABLE 3-14. MEAN CONCENTRATIONS (meq/L) OF IONS IN THE WATER OF FOUR NOVA SCOTIA RIVERS IN 1954-55 AND 1980-81. AVERAGE DIFFERENCES _ WERE CALCULATED AS 1980-81 CONCENTRATION MINUS 1954-55. THE I SIGNIFICANCE LEVELS (EXCEPT H+) ARE FROM THREE-WAY VARIANCE • ANALYSES. THE S042~, Na+, K+, Ca+, AND Mg2+ IONS HAVE BEEN CORRECTED FOR SEASALT INFLUENCE (Watt et al. 1983) I I River Years HCO^ So£~* Na+ K+ H+ Ca2+ Mg2+ A13+ • Roseway 1954-55 0.049 0.033 0.002 0.016 0.014 0.060 0.004 0.014 I 1980-81 0.007 0.089 0.031 0.007 0.040 0.028 0.010 0.027 Medway 1954-55 0.055 0.031 0.016 0.006 0.001 0.047 0.016 0.007 1980-81 0.013 0.059 0.018 0.005 0.005 0.045 0.017 0.014 I Mersey 1954-55 0.044 0.023 0.014 0.008 0.002 0.045 0.010 0.009 • 1980-81 0.022 0.053 0.017 0.005 0.006 0.031 0.012 0.016 I I I I I * Caution is required in interpreting SO^ trends due to the interference of organic anions in measurement procedures. • I I I ------- I I I I I I I I I I I I I I I I I I I 3-79 The time records of the median pH and excess sulphate discharge for two rivers of Nova Scotia (Medway and St. Mary's) and for two rivers of Newfoundland (Rocky and Isle aux Morts) are shown in Figure 3-30. The pH values are the median value for the n observations of the calendar year. Maximum and minimum pH observed are also shown. Excess sulphate (seasalt corrected) discharge calculated from the n observations of sulphate concentration and the measured run off (calculated as indicated on the figure) are also shown. Figure 3-30 illustrates the difficulty in making statistical analysis of the observations. Record breaks are present and unevenly spaced observations render statistical analysis to establish trends over time impossible. Thus Clair and Whitfield (1982) could treat only portions of the record. Figure 3-30 may be examined to illustrate the temporal variability of pH and excess sulphate discharge which is not revealed by the statistical trend analysis or the changes between two time periods as reported by Watt et al. (1983) or Farmer et al. (1980). The dominant feature is the minimum pH that occurs in the 1971-73 period. Excess sulphate discharge reaches a maximum for the Newfoundland rivers during the same period. Wiltshire and Machell (1981) have reported on a re-survey of 16 lakes in Nova Scotia and New Brunswick, which had historical data going back to the 1930s in some cases. Eleven of the lakes are remote from local sources in Halifax and Saint John. The data for 10 remote lakes with the most reliable historical information are summarized in Table 3-15. Between 1950 and 1979 data indicate that there has been a decline in pH in all cases, most notably since the 1950s. All but one of the 10 lakes, however, still had a pH> 5.5 in 1979. Calculated alkalinity changes (Table 3-15) show declines ranging from 5.5 to 55 P eq/L. 3.6.3.2 Historical Trends in Northern Wisconsin Juday et al. (1935) described the pH-C02 relationships of lakes in northeastern Wisconsin. Between the period 1925-41, measurements were made of pH, alkalinity and conductivity in 518 lakes. Historic pH was measured colorimetrically between 1925 and 1932; from 1932 to 1941 a quinhydrone electrode was used. Two groups have remeasured pH, alkalinity and conductivity in separate subsets of the 518 lakes (Bowser et al. 1982; Eilers et al. 1982). The 53 lakes sampled by Bowser et al. (1982) ranged from alkaline, mesotrophic lakes to brown-water bogs to clear-water, oligotrophic lakes. Multiple historical measurements were available and modern-day sampling was adjusted to a similar seasonal sampling period. All three lake types showed increases in pH, alkalinity, and conductivity over the 50-year period. Bowser et al. (1982) attributed their results to: (1) short duration of acidification of precipitation in Wisconsin; (2) changes in vegetation and shoreline land use; and (3) the importance of groundwater to many of the lakes. Eilers et al. (1982) sampled 275 of the lakes surveyed by Birge and Juday. They selected 180 for analysis and comparison with earlier measurements. They also found ------- 3-80 8.8- X a -100 i 1 1 1 i 1 1 1 1 5 10 12 14 24 12 12 12 12 4 —• Excess SO ST. MARY'S RIVER 5 14 1 1 1 1 13 12 8 12 11 4 4 4 3 4 5 18 1984/88 98 86 87 68 69 70 71 72 73 74 75 76 77 78 79 80 81 S • S ? CO < *p J 71 ? a o °- S. 5 -100 6.6- I a 9 9111 3 4 ISLE AUX MORTS RIVER —I 1 1 1 1 1 1 1 1 1 1 - • 10 12 13 12 12 13 11 11 10 12 10 10 ROCKY RIVER 66 66 67 68 69 70 71 72 73 74 75 76 77 78 79 p «' o 3- <» o » £1 80 81 Year Figure 3-30. Annual changes in median pH and mean discharge- weighted excess SO^" for the St. Mary's and Medway Rivers, Nova Scotia, and the Isle aux Morts and Rocky Rivers, Newfoundland. Data are from the sources indicated. Upper and lower numbers shown represent range of values, (•) is median and n is the number of samples for each year. I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-81 Figure 3-30. CONTINUED Data for Figure 3-30 were calculated according to following format: 1. Excess SO^2~ = Total S042~ - 0.14C1" (both in mg/L) 2. Mean discharge-weighted excess S042~ = ^(Excess S042~ times Sample Date Discharge) Z Sample Date Discharge 3. Runoff = Mean Annual Discharge (m-Vyr) divided by drainage area (m2) = m/yr = m-Vm2.yr. 4. Runoff for Water Years 1954/07 - 1955/06 and 1955/07 - 1956/06 were calculated from mean monthly discharges. For the Mersey the long-term runoff times simple mean SO^2" concentration was used. 5. Excess SO^2" (meq/m.yr) = Mean discharge-weighted excess SC>42~ times Runoff. 6. Chemical data are from NAQUADAT. Discharge data are from various reports of the Water Survey of Canada, Ottawa. ------- 3-82 1 • 1 TABLE 3-15. APPARENT CHANGES IN SUMMER pH VALUES IN LAKES IN NOVA SCOTIA AND SOUTHERN NEW BRUNSWICK DURING THE PERIOD 1940-79 (Wiltshire pH pH change ca 1940a 1950sb 1979C pre-1950s post-1950s Boarsback N.S. 4.7 4.7 4.4 .0 -.3 Jesse N.S. 6.5 6.5 5.8 .0 -.7 Lily N.S. 6.5 5.8 -.7 Kerr N.B. 6.8 6.6 6.0 -.2 -.6 Creasey N.B. 6.7 6.7 6.0 .0 -.7 Tedford N.S. 6.3 6.6 6.3 +.3 -.3 Sutherland N.S. 7.0 6.3 -.7 Gibson N.B. 7.0 6.7 6.4 -.3 -.3 Black Brook N.S. 6.8 6.4 -.4 Copper N.S. 7.3 7.0 -.3 a Data from Smith 1937a, 1937b, 1948, 1952, 1961. k Data from Hayes and Anthony 1958. c Wiltshire and Machell 1981. ^ Calculated by Liljestrand pers. comm. PCO~ assumed as 10~^ mean) with bicarbonate as major buffer. and Machell 1981) Calculated alkalinity change in yeq/L pre-1950 post-1950 0 -20 0 -14 -14 -12 -16 0 -21 5.5 -5.5 -41 -25 -18 -20 -55 •5 atm. (a global 1 1 1 1 I 1 1 I 1 1 • * 1 1 1 1 ------- I I I I I I I I I I I I I I I I I I I 3-83 that most lakes had increased in pH, alkalinity and conductivity. The lowest pH values were found in lakes having no inlet or outlet and no contact with groundwater (Eilers et al. 1982; Schnoor et al. 1982). 3.6.3.3 Historical Trends in New York State Peters et al. (1981) analyzed precipitation data and stream water chemistry data from a nine-station monitoring network in New York State. The data covered the years 1965-78. Sulphate concentration in precipitation decreased 1-4%/yr, while N(>3~ increased by 4-13% each year. An increase in the total amount of precipitation over the period resulted in an increase in total acid loading. Variable neutralization of hydrogen ion, perhaps by particles in dry depo- sition, was suggested because the observed trends in hydrogen ion concentration do not correlate well with those for sulphate or nitrate. In most areas of New York, urbanization, farming and carbonate soils have masked any effects of increased acid loading. For the East Branch of the Sacandaga River in the Adirondacks, nitrate increased 0.004 meq/L.yr, while sulphate decreased 0.0041 meq/L.yr. Sulphate concentrations exceed bicarbonate for the stream indicating little interaction with soils or ground water. Consequently, with the increases in acid loadings in precipitation over the period, alkalinity has decreased 0.083 meq/L.yr (Peters et al. 1981). In a survey designed to identify acidic lakes in the Adirondacks, Schofield (1976c) sampled 214 high altitude lakes (Figure 3-31). A complete set of chemical analyses was obtained, but no internal checks can be made on the data, because sulphate was determined by difference. The pH range of sampled waters was 4.3-7.4. Fifty-two percent were listed as pH <5.0; 7%, pH 5.5-6.0. pH measurements were made in the laboratory following aeration of the sample. Increased elevation and low pH of ponds and lakes were positively correlated. The combined influences of heavier precipitation at higher elevations, the smaller surface area and watershed size which characterizes most headwater ponds, the prevalance of granitic bedrock and shallow soil deposits in the higher elevations, and the direct impingement of acidic cloud water are all possible factors. In addition, N.M. Johnson (1979) showed that neutralization occurs as contact time with the substrate increases, such as occurs as water flows downhill into progressively larger streams. For a subset of 40 of these 214 high elevation lakes, historical data on pH are available from the 1930s (Schofield 1976c; Figure 3-32). These early pH values represent colorimetric measurements. Several authors (Norton et al. 1981a; Pfeiffer and Festa 1980; Schofield 1982) have examined the agreement between the two pH methods. Although certainly not exact, qualitative comparisons appear appropriate. For this subset of 40 lakes, in 1975, 19 lakes had pH ------- 3-84 (Ouake V placid 0 S 10 15 20 Km. Figure 3-31. Geographic distribution of pH levels measured in Adirondack lakes higher than 610 metres elevation, June 24-27, 1975 (Schofield 1976c). I I I I I I I 1 I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-85 20 10 W o> x. ------- 3-86 I I below 5.0 and all of these lakes had no fish. In the 1930s, only 3 lakes had values below 5.0 and a total of only 4 lakes had no fish (Schofield 1976c). • In a larger survey that included Schofield's 1975 sites (Schofield 1976c), a 1980 report by the New York Department of Environmental — Conservation (NYDEC) (Pfieffer and Festa 1980) reported that 264 of • 849 (25%) lakes sampled in the Adirondacks had a pH <5.0. The report * linked this acidity to fish losses in these lakes. Since publication of this report, however, both the NYDEC (1982) and others (Schofield • 1982) have recognized that many of the pH values reported were too 9 low (due to problems with the pH meter). Pfeiffer and Festa (1980) also presented a comparison of colorimetric pH measurements for the • 1970s and 1930s for a set of 138 Adirondack lakes. In general, • historic pH readings were higher than the comparable current measurements. 3.6.3.4 pH Changes in Maine and New England Several synoptic studies have been done for New England surface | waters. Davis et al. (1978) studied 1936 pH readings taken from 1,368 Maine lakes during the period 1937-74 in an effort to see if m they could find pH decreases associated with the acidic precipitation • of that area (4.4------- I I I I I I I I I I I I I I I I I I I 3-87 lakes and then mean slopes from 1937 to 1974. The mean slopes were added to obtain a total H"1" concentration change for the entire period. Given a starting pH of 6.89 (mean of 123 values 1937-42), the final (1974) pH would be 5.79, an increase in acidity of 12.6 times. Using a t-test, the authors also found that the mean annual increase in H+ concentration based on the mean slopes for each year was significantly different from zero change with p <0.0001. The authors noted, however, that this procedure more strongly weights data pairs with long time separations, thus possibly invalidating the use of a t-test. The second procedure Davis et al. (1978) used was to average the 376 single slope values. This gave a mean of 0.115 peq/yr H+ concen- tration change. By t-test, this mean is significantly different from zero p <0.1, but not at p <0.05. If a disproportionately greater decrease in pH occurred in the 1950s (as the authors hypothesized), this procedure would give greater weighting to the more frequent data pairs beginning about that time and would thus overestimate total change (Davis et al. 1978). Procedure III the authors used was to weight each data pair (H+ concentration) slope linearly in inverse proportion to the time interval between each reading. These weighted slopes were then averaged for each year that they applied. Using an initial pH of 6.89 in 1937, the authors noted that pH decreases by 1950 to only 6.83. By 1961, however, the pH has decreased to 5.91, so that 73% of the increase in acidity has occurred in this latter time period. The authors believed that this 73% increase in acidity was actually an underestimate for this time period. Davis et al. (1978) also discussed some alkalinity data they had for 44 of the 258 lakes cited above. These data were from the period 1939-71, a total of 96 values and 52 pairs. No information was given on the analytical method(s) used to determine alkalinity. Applying their Procedure I to those data, they obtained a decrease of about 6.34 ppm (as CaCC^; from 11.82 to 5.48 ppm, typically; corresponding to a decrease of 127 yeq/L from 236 to 109 yeq/L) over the period. This was much less than expected from pH changes from the same period and observed relationships between pH and alkalinity. The authors noted that "the discrepancy may be due in large part to the inadequate sampling and great variance of the alkalinity data, including the fact that 67% of the pairs had their initial member in 1960 or later" (Davis et al. 1978). The authors concluded from their study that between the years 1937 and 1974 H* concentration in Maine lakes increased about 1.0 peq and pH decreased from about 6.85 to 5.95. Further, nearly three-quarters of this change occurred in the 1950s. "This is the first demonstra- tion of a pH decrease due to acidic precipitation on a large region of lowland lakes in the United States" (Davis et al. 1978). ------- 3-88 I I Norton et al. (1981a) measured pH in 94 New England lakes (82 in Maine, 8 in New Hampshire, 4 in Vermont) for which historical pH existed from the period 1939-46. Eleven (12%) of these lakes had • pH <5.0 in 1978-80. The lakes sampled were small, oligotrophic- • mesotrophic, and located in forested areas on noncalcareous bedrock. The recent sampling (1978-80) was done during July-October but not on • the same monthly dates as the historic sampling. These samples were • collected at 1 m depths, and the lakes were stratified at the time of sampling. The pH values of the recent samples were measured in the field with • (1) a portable pH meter with combination electrode, and (2) a Hellige color comparitor. Except for three spurious cases of low pH lakes, U the authors found that "reasonable agreement exists for these two jj methods, especially at higher pHs" (Norton et al. 1981a). The authors presented their results in plots of: (1) old colori- • metric pH vs. recent colorimetric pH, and (2) recent colorimetric pH vs recent electrometric pH. They concluded that their study "confirms the results of Davis et al. (1978) regarding an overall • decrease in the pH of Maine lakes" (Norton et al. 1981a). ™ Norvell and Frink (1975) found that the pH and alkalinity in :fl sensitive (alkalinity <200 peq/L) lakes in Connecticut had not | changed significantly from 1937 to 1973. Haines (1981a) reports a number of Connecticut rivers as being "sensitive", due to alkalin- M ities <200 ueq/L, but pH in these waters is 6.4-7.1 except in smaller • lower order streams. In Maine, the pH of major rivers is greater than 6.5, with the lowest values in eastern Maine (the area with the highest precipitation pH in New England) (Haines 198la). • Haines and Akielaszek (1982) sampled 95 lakes for which there were historical pH data from the 1930s to the 1960s. Of these, 36% either • had the same or higher pH while 64% were lower. For 56 lakes there | were also fixed end point titrations of alkalinity. A comparison of historical alkalinities to modern values indicated that 30% of the ^ lakes had increased and 70% had decreased in alkalinity. The • historical alkalinity values averaged 166 yeq/L and recent samples 68 peq/L. I 3.6.3.5 Time Trend in New Jersey A.H. Johnson (1979) described a 17-year decline in pH of headwater | streams in the New Jersey Pine Barrens which drain relatively undisturbed watersheds (Figure 3-33). The trend is statistically _ significant and has amounted to approximately 0.4 pH units over the • period. In the sandy soils of this region, relatively little " neutralization of acid inputs occurs by ion exchange or mineral weathering as precipitation moves through the soil. The low level of H neutralization is evidenced by the low pH of shallow groundwater, | averaging 4.3 for 78 samples in 1978 through 1979. The great I I ------- I I I I I I I I I I I I I I I I I I I 3-89 6 °- 6 5 OYSTER CREEK O O o MCDONALDS BRANCH O 1960 1970 1980 Figure 3-33. New Jersey stream pH, 1958-1979, Oyster Creek and McDonalds Branch. Closed circles represent samples in which anion and cation equivalents balanced, and calculated and measured specific conductances were equal. Open circles are samples for which the chemical analyses were incomplete, or for which discrepancies in anion and cation and conductivity balances could not be attributed to errors in pH. The closed triangle represents the average pH determined in a branch of Oyster Creek in a 1963 study. Open triangles are monthly means of pH data collected weekly from May 1978 to January 1979 during a University of Pennsylvania trace metal study (A.H. Johnson 1979). ------- 3-90 variability in pH values of streams in 1978-79 is thought to be due to storm events. 3.6.4 Paleolimnological Evidence for Recent Acidification and Metal Deposition I I Some precipitation pH data suggest a trend toward lower pH values in southern New Jersey (A.H. Johnson 1979). Precipitation samples, collected at several sites in the Mullica and Cedar Creek basins in _ 1970 through 1972, had an average pH of 4.4. Samples collected near • Oyster Creek for seven months in 1972 had an average pH of 4.25. * From May 1978 to April 1979, the average pH of weekly precipitation samples at McDonalds Branch was 3.9. • I To supplement the sparse information from long term records on water _ quality in eastern North America alternative techniques have been • developed to define time trends related to surface water acidifi- * cation. Paleolimnological analyses of lake sediments have traditionally been • used to reconstruct many aspects of the evolution of lake/drainage basin ecosystems including terrestrial and aquatic vegetational M succession (Bradstreet et al. 1975), fire history (Patterson 1977), • trophic status (Davis and Norton 1978; Stockner and Benson 1967) and even the occurrence of blight or disease (Bradstreet and Davis 1975). ^ Long-term changes in meteorology, morphology of the lake basins, soil I development, land use, or surface water chemistry can be partially * determined from the sediment record. I Both chemical and biological records are left in the sediment record. By studying modern biota in relation to modern water quality for many lakes, changes in ecosystems which have taken place over time can be • reconstructed for water quality parameters, for example, pH (Davis • and Berge 1980; Davis et al. 1980, 1982; Renberg and Hellberg 1982). Normally, measurable changes in natural acidity are on the order of ^ centuries and are accompanied by changes in other sediment character- • istics. Land use changes such as logging followed by reforestation • can bring about perturbations in pH of surface waters (both increases and decreases) with a general return to equilibrium in 10 to perhaps • as much as 50 years (Likens et al. 1978; Pierce et al. 1972). | Davis and Berge (1980) and Davis et al. (1980, 1982) have demon- M strated, using fossil diatom data, pH declines in the last 30-70 • years from pH values of greater than 5.5 to values less than 5 in lakes in Norway with relatively undisturbed drainage basins. The present pH of these lakes is too low to be explained by the concen- fl trations of naturally occurring organic acids. Sulphate is the • dominant anion and is apparently atmospheric in origin (Wright and Henriksen 1978). • I I ------- I I I I I I I I I I I I I I I I I I I 3-91 Norton et al. (1981b), Davis et al. (1982), Evans and Dillon (1982), Dickson (1980), and others have demonstrated that heavy metal (especially Pb, Zn, and Cu) deposition rates started increasing over 100 years ago in eastern North America and Scandinavia indicating polluted air masses existed in the late 1800s. Cores from Swedish Lapland (Davis et al. 1982) do not show these increases. There, the pH of precipitation is approximately 5.0. By inference, precipitation in eastern North America was probably somewhat acidified by the late 1800s. No change in the biology, defined from the sediment cores is observable until at least the early 1900s. Thus biological effects appear to lag behind definable chemical changes (Brakke et al. 1982; Davis et al. 1982). As the acidity of precipitation increases, leaching of Zn, Cu, Ca, Mg and Mn from organic matter and soils of the terrestrial ecosystem also increase. At near neutral surface water pH (greater than pH 5.5), Zn and Cu from the terrestrial leaching processes are accumulated in the sediments. However, as surface waters become more acidic, Zn and Cu from the watershed remain in the water column to be exported from the lake. As a result, acidification of surface waters will decrease sedimentation of Zn and Cu. Lead, on the other hand, accumulates in sediments independently of pH (Davis et al. 1982). Calcium, Mg and Mn also decline in lake sediments as acidity increases for two reasons. Firstly, as a result of acidic deposition falling on the watershed, terrestrial detritus becomes depleted of Ca, Mg, and Mn prior to entry into the aquatic ecosystem and incorporation into the sediments. Secondly, acidified waters prevent these metals from being resorbed to sediment organic matter and, like Zn and Cu, are exported from the lake system (Norton et al. 1981b). Experiments on lake sediment microcosms (Kahl et al. 1982) indicate that if lake water pH is increased, the sediments absorb Ca, Mg, Mn, and Zn from the water column at a rate which would enrich the sediments. The observations and information from field studies suggest that acidification strips cations from terrestrial detritus and prevents them from resorbing onto the detritus (Dickson 1980; Kahl et al. 1982; Norton et al. 1980). Norton et al. (1980) showed that the chemistry of soil organic matter in New England, New Brunswick, and Quebec is partly controlled by atmospheric deposition of acids and metals. They found decreased Zn, Mn, Ca, and Mg in regions receiving high H+ loading. Cores from acidic clear water lakes in New England (pH less than 5.5) with undisturbed drainage basins (5 of the 30 lake samples taken over at least the last 50 years) show declines in sediment concentrations of Zn, Ca, Mg, and Mn starting as early as about 1880 suggesting increased leaching of sediment detritus prior to entry into the lakes (Davis et al. 1982; Kahl et al. 1982) or reduced sedimentation rate. All acidified lakes in Norway and New England with pH less than 5.0 have shown declines in Zn and Cu in recent sediment. Lead, on the other hand, is not released from surficial sediments unless the pH is less than 3.0 (Davis et al. 1982). ------- 3-92 I I Measurement of atmospheric loadings via both wet and dry deposition techniques is plagued by a variety of uncertainties (Section 2.2.3). The lake sediments integrate materials deposited directly on the lake flj surface from the atmosphere with elements leached from the terres- | trial watershed. Increased mobilization of metals from watersheds during hydrologic events when pH is depressed has been discussed • (Section 3.2.4). The calibrated watershed approach to measuring • deposition is ineffective for trace metals because metal levels found in lake and stream water are often below analytical detection limits usually employed. Trace metals are rapidly removed from the water • column in most lakes (residence times are typically of the order of ™ days), and stored in the sediment. Calculation of metal loadings to lakes may sometimes be derived from information collected from the M lake's sediments. Profiles of lead concentration in four sediment | cores from Jerry Lake, Ontario are shown in Figure 3-34. Dillon and Evans (1982) demonstrated that input of lead from • anthropogenic sources to eight lakes in southern Ontario resulted only from atmospheric deposition directly on the lakes' surfaces; that is, deposition on the lakes' watersheds was effectively retained • in the watersheds. The whole-lake lead burdens estimate the total • atmospheric deposition of lead during the period when anthropogenic emissions have existed. Regional anthropogenic lead burdens measured • for Muskoka-Haliburton, Ontario (Dillon and Evans 1982) and a remote || northern site, Schefferville, Quebec (Rigler 1981) are 680 (range 610-770) mg/m^ and 37 (range 31-59) mg/m^, respectively. • 3.6.5 Seasonal and Episodic pH Depression I Although survey data, both current and historical, can be used to document long-term trends in a synoptic sense, the samples usually represent one or a few measurements at any one location and are often collected during the summer. This limited sampling period provides no record of pH and other chemical changes which take place in relation to seasonal cycles or major weather events. Individual pH _ values during the summer do not reflect these cyclic and episodic • aspects of the loading/episodic response relationship. If short-term * changes in water chemistry coincide with sensitive periods in the life cycles of fish (e.g., spawning and hatching), significant I mortality and reduced reproduction can occur. The following data m describe recent results on episodic pH declines; the extent to which these phenomena occurred in the past is not known. •• 3.6.6 Seasonal pH Depression in Northern Minnesota _ I Siegel (1981) reported on the effect of snowmelt on Filson Creek and ™ Omaday Lake in northeastern Minnesota. He found concentrations of sulphate increased in Filson Creek and Omaday Lake during snowmelt B from less than 2 to 12 mg/L in 1977 and from less than 2 to 4 mg/L in | 1979. During snowmelt, pH decreased from 6.6 to 5.5 in 1979. I I ------- I I I I I I I I I I I I I I I I I I I 300 250 200 T3 D) *«, O) c O 0) O c O O .a a. 150 100 50 Jerry Lake Sediment Cores 1979 3-93 LAKE DEPTH AT CORE SITE 11.0 m 20.2 m 17.3 m 32.3 m 0.0 Figure 3-34. 0.8 1.6 CDWA (g/cm2) Profiles of the lead concentration in four sediment cores from Jerry Lake, Muskoka-Haliburton. Depth within core is expressed as cumulative dry weight per unit area (CDWA) (modified from Dillon and Evans 1981). 2.4 ------- 3-94 3.6.8 pH Depression During Flushing Events in West Virginia I I Alkalinity and concentrations of total calcium, magnesium, and sodium in the creek during snowmelt reflect the simple dilution of stream- flow with more dilute precipitation. Depression of pH values to less • than about 5.7 indicate that base flow (pH~6.5) has been diluted • with meltwater that contains some mineral acids. I 3.6.7 pH Declines During Spring Runoff in Ontario and Quebec Detailed surface water chemistry studies have been conducted in lakes • near Muskoka-Haliburton, Ontario, on the Precambrian Shield. Jeffries et al. (1979) compared pH values of a series of small streams in the study area, before and during spring runoff. The pH • declines of the lake outflows demonstrated that the top portions of I the entire lakes were acidified. The lowest stream pH values observed, 4.1 to 5.1 (Table 3-16), were within a range capable of • causing damage to some aquatic organisms, particularly fish (see • Section 3.7.7 for discussion). As much as 77% of the measured annual acid export of the streams occurred in April (Table 3-17). A typical m hydrograph and pH response for one of the streams during the snowmelt • period is illustrated in Figure 3-35 and Table 3-17. ™ I The pH. of streams was depressed for periods of as little as a few hours during times of heavy runoff, during the summer months (Figure 3-36; Scheider et al. 1979b). Heavy fall rains also cause depressed pH in runoff for days at a time. Jeffries et al. (1979) * observed as much as 26% of the total annual hydrogen ion runoff, from • small watersheds, in October. As a control for eastern work, ELA is probably one of the best that • can be obtained at temperate latitudes. Mean annual pH of bulk • precipitation ranged from 4.9 to 5.2 over the past 10 years, calculated on a volume-weighted hydrogen ion basis. No directional • trend was observed in pH values. There is a pH depression in the | area related to spring snowmelt, of about 0.2 to 0.5 pH units in inflow streams and about 0.2 to 0.3 in lake outlets. Minimum values M observed in spring runoff have been as low as 4.5 but are generally • above 5.0 (data from ongoing studies at the Freshwater Institute, Fisheries and Oceans, Winnipeg, Canada). A comparison of the water chemistry of 70 lakes in Quebec, sampled at W the spring isotherm and at a summer stratification of 1980, has been performed by Bob£e et al. (1982). [This information is summarized in • their table 5.2, Table 3-18 in this report.] The authors observed | that the mean values of conductivity, alkalinity and pH were generally lower during the spring than during the summer while the _ ratio [804^"]/[HC03~] was much higher in the spring. • I Seasonally low pH and regular patterns of pH declines have been documented for the Little Black Fork and Shavers Fork Rivers by the • I ------- I I I I I I I I I I I I I I I I I I I 3-95 TABLE 3-16. pH OF STREAMS IN MUSKOKA-HALIBURTON, ONTARIO, CANADA: STREAM pH IS GIVEN PRIOR TO SPRING RUNOFF (MID-MARCH 1978) AND AT MAXIMUM RUNOFF (MID-APRIL 1978) (Jeffries et al. 1979) PH Watershed Harp Lake Dickie Lake Chub Lake Red Chalk Lake Maple Lake Lake Simcoe Lake of Bays Stream 3 3A 5 6 6A Outflow 5 6 11 Outflow 1 2 Outflow 1 2 3 4 Outflow Maple Creek Black River (at Vankoughnet) Oxtongue River Mid-March 6.1 6.0 5.9 6.2 5.4 6.3 4.6 4.6 4.9 5.6 5.8 5.2 5.5 6.1 4.5 6.0 6.2 6.1 6.2 6.3 6.3 Mid-April 5.1 5.6 4.8 5.3 5.0 5.0 4.3 4.4 4.1 4.9 5.1 4.7 4.8 5.6 4.3 5.5 5.5 5.9 5.8 5.9 6.1 ------- 3-96 Pj pd o S3 H S3 Pd g 00 < ON H -« O H ;M 5! H r-» S3 r- W ON U -H Qd W W OH S3 co pi Q En S3 • O O M S3 S3 W W S O << O iJ O W ^ >H H ON W taj 1^- O O\ « H rt W O o H O -H J S3 H ffi MH S3 g I CO (a j 0) ^ •o cfl a s-s •H o- W o o o O CD O 00 00 •—i r^ CM m o o o m ON oo CN O ON sosom oooocomcN i-toovor--. r-- cd oo ON o so r^ I-H co 4- >rH ON in O 00 sO ON T^« r—4 CO in fcd 3 o cr* CM o o m so o i—i ONin^som o C W CM * >*O O-, B^S cu i-Hoo r~»mO>-H o m •—' m tn g cN^HOooONinr^ COOOOOCNON co p o ooooomso coi-Hcocoo —i ^H m CM -H i-H r^ oo i--. r~- ON ON •—I ^H (1) >, • 4J • • • .... Cd CrHbOP-i4J>O CJ3)HM>, W 333------- I I I I I I I I I I I I I I I I I I I 3-97 HARP LAKE No. 4 5 10 15 May Figure 3-35. Discharge (upper line), hydrogen ion load per unit area (middle line), pH (lower line), and depth of precipitation for each day that a precipitation event occurred for Harp Lake No. 4. Daily H"1" load to the respective lakes can be calculated by multiplying by the watershed area: lake area (A------- 3-98 30 20 10 n RED CHALK No. 4 - ^^\ _l 1 1 1 1 1 1 L. __L. ±....1111 II 1 II II 70 r 60 50 40 30 20 10 RED CHALK No. 3 1900 2100 2300 0100 0300 0500 0700 0900 1100 1300 1500 July 12 / 77 Rain Event Time (hr) July 13 177 pH 4.06 ~2cm Figure 3-36. Hydrogen ion content of streams draining Red Chalk Lake watersheds No.3 and No.4 (Muskoka-Haliburton, Ontario) showing effects of a 2 cm rainfall (pH 4.06) between 11:00 p.m. July 12, 1977 and 3:00 a.m. July 13, 1977 (Scheider et al. 1979b). I I I I I I I I I I I I I I I I I I I ------- 3-99 2 O H O /-s W CM K OO ON CJ -H M HI • PH rH § rt O 4-> O Ol Dd 3 oj SrS o frj m cj •< S 2 § !H ps! MH O v^ fc x^, Q 4-J H co H 01 3? 3 4J cj j ^ ^J '^s C u <1) *o c.o 3 W 4-* £3 W ^ > M W Pi 0 W J3 H EH W W ^ J IZ PH [&3 m o •— i •^^ > u •H c •tH t-H CO r* rH < ON , 4-> •r( > •H 4-1 O 3 13 n § P. O * 00 ON -•v^ i-H 4-1 -H 1 r-. rH 4J CN 4-1 C 0) T3 * 3 rH 4J CN W •**. 00 ON -O rH Q) rH •H CO cx * o * 00 rH ^ m 01 ON > CM o> rH &-S UO * •-H QJ CM 42 •"^ ON 4-1 OO -H jj CO 0) Ol o C Ol rJ Ol >4H <4H •H T3 ^-N 4J W c PH CO Cfl 01 O •4-1 ^ TH O CO <4H hJ -H DO d 6 'H DO *~s O -H CO M rH d 3 01 O O rQ r3 rH B O 3 U 2 * ------- 3-100 3.7 ALTERATION OF BIOTIC COMPONENTS IN AQUATIC SYSTEMS RECEIVING ACIDIC DEPOSITION I I U.S. Forest Service in Monongahela National Forest (Dunshie 1979). Due to the sandstone geology of the watershed, the tributaries and the river are poorly buffered and subject to rapid changes in water • quality. The lowest pH values in both streams (Little Black Fork is • a control area, with no logging or coal mining) normally occurred during the winter and early spring, apparently because of snowpack IB melting. The highest pH occurred during low stream flow periods in jf the summer and fall. Even though summer and autumn are the periods of highest precipitation inputs (see below), more extensive contact H between soils and precipitation may have lead to greater neutrali- • zation at these times than during either winter or spring flushing events. The effect of rainfall on river pH is more apparent when individual «• events are examined. A graphic presentation pairing daily river and precipitation events with pH during summer periods is shown in • Figure 3-37. During the growing season, a storm event with a j§ subsequent increase in discharge can significantly lower river pH below the natural nonstorm daily variation. The magnitude of this • downward shift is dependent upon rainfall characteristics (pH, • amount, intensity, and area distribution) and antecedent soil moisture. Downward shifts in river pH ranging from 0.6 to 0.9 pH units, occurred on July 11 and 26, and on August 15 and 25, 1977. On • three of these days, at least 3.3 cm of rainfall fell within a • 48-hour period; pH of the rainfall for these dates ranged from 3.7 to 4.2. • Nearly 13 years of pH data have been collected at the Bowden Fish Hatchery river intake on the Shavers Fork River, showing lower pH . values during winter and spring compared to summer conditions. This • is important for aquatic organisms and has been measured in other poorly buffered streams. This pH trend occurred in streams and tributaries independent of watershed disturbance by mining (Dunshie • 1979). I I Many changes in biota have been linked to acidification of surface • waters. In some controlled whole lake and laboratory experiments a * causal relationship with decreased pH has been established. In the majority of cases, the observed changes in biota have simply been fl correlated with observed changes in pH and other parameters, but | causality has not been established. For many biological communities, acidification has been accompanied by decreases in species diversity • and changes in species dominance. Acidification may also be • accompanied by species extinctions, or decreases in overall community standing stocks. This topic has been reviewed recently by Raines (1981c). Generalized summaries of responses of aquatic organisms to • low pH are given in Figures 3-38 and 3-39 (Eilers and Berg 1982), and ™ are presented here as a simplified overview of the complex I I ------- I I I I I I I I I I I I I I I I I I I 3-101 A Rainfall pH I Rainfall Accumulation 10 15 20 25 30 5 10 15 7/77 Figure 3-37. Mean daily pH for the Shavers Fork River at Bemis, West Virginia and precipitation event pH and accumulation at Arborvale, West Virginia (Dunshie 1979). ------- 3-102 Algae Insects Molluscs Sponges Leeches Zooplankton Fish Frogs 75 50 25 75- 50 25 75 50- 25 75 50 25 75 50 25 75 50 25 75 50 25 PH Figure 3-38. Relative number of taxa of the major taxonomic groups as a function of pH (Eilers and Berg 1982). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-103 c 0 O L_ CD a. 100 75 - 50 - 25 - 0 Major Aquatic Community Impacts yMi i n*lit*i i I f- 1 I 3 snails5 zooplankton I insects I I phytoplankton CO •^ CO 1 Sprules (1975) - Ontario 2 Beamish (1976) - Ontario 3 Bell (1971) - Laboratory TL50 4 Yan and Stokes (1978) - - Ontario PH 5 0kland (1969) - Scandinavia 6 Wright et al. (1976) - Norway 7 Kwiatkowski and Roff (1976) - Ontario 8 Snekvik (1974) - Norway Figure 3-39. Generalized response of aquatic organisms to low pH (Eilers and Berg 1982). ------- 3-104 3.7.1 Effects on Algae I I interactions described below. It is important to note that the data were derived from literature surveys of the relationship between the distribution of groups of organisms versus lake and stream pH values. • The quantitative description of these relationships may not reflect • the response of individual taxa. Definitive experiments are required to demonstrate whether such • changes are directly attributable to increases in hydrogen ion • concentration or whether they are attributable to secondary ecosystem interactions, such as elevation of trace metal levels or disruptions B of normal food chains. In spite of incomplete understanding of the | actual mechanisms underlying observed changes accompanying pH declines, it appears that acidification of surface waters brings • about major quantitative and qualitative changes in structure and • function of aquatic ecosystems. Disruption of the normal food chains may occur long before the lakes have been acidified in a chemical sense. • I The free-floating (planktonic) and attached (benthic and epiphytic) algae are the major primary producers in most aquatic ecosystems and « directly or indirectly provide most of the food for zooplankton and • ultimately for fish. Evidence gathered mainly from synoptic surveys in Scandinavia, Canada and the United States has indicated that the species diversity of benthic and planktonic algal communities is less • in acidified lakes. Yan and Stokes (1976) observed only nine species • of phytoplankton in a single sample from Lumsden Lake (pH 4.4; Beamish and Harvey 1972), in the La Cloche Mountains in Ontario, but • observed over 50 species in each of two nearby nonacidic lakes, | having pH over 6.0. Diversity indices for phytoplankton populations in the La Cloche Mountain lakes are much less in lakes with pH values • below 5.6 (Kwiatkowski and Roff 1976). In Scandinavian lakes numbers • of phytoplankton species are also much less in lakes with pH values below 5.5 (Aimer et al. 1978; Leivestad et al. 1976). Some long-term functional adaptations to certain acidic environments • may occur. Raddum et al. (1980) have suggested that such a mechanism explains the observation that a group of relatively recently • acidified clearwater lakes in Norway have less diverse phytoplankton Jj assemblages than naturally acidic, humic lakes. Additionally, the bioavailability and toxicity of trace metals may be lower in the « brownwater acidic lakes because metals may be complexed with humic • materials. Although species diversity of phytoplankton generally decreases with • increasing acidity, biomass (Yan 1979) and productivity (Aimer • et al. 1978; Schindler 1980) are often not reduced by acidification. However, if phosphorus (the nutrient that normally limits phyto- • plankton productivity in soft-water lakes) is immobilized to some • degree in acidic lakes because of complexation with aluminum and I I ------- I I I I I I I I I I I I I I I I I I I 3-105 humic material (Aimer et al. 1978), this would result in reduced primary productivity. To date, data from lakes in Scandinavia and eastern Canada indicate no significant correlations between pH and phytoplankton biomass or productivity (Harvey et al. 1981). Phytoplankton communities of nonacidic oligotrophic lakes in eastern Canada are typically dominated by chrysophytes (Schindler and Holmgren 1971) or by diatoms (Duthie and Ostrofsky 1974). In contrast, strongly acidic lakes are generally dominated by dino- flagellates. In Sweden, the dinoflagellates, formed 85% of the biomass in lakes of pH 4.6-5.5 (Dickson et al. 1975). Of 14 lakes in central Ontario, dinoflagellates formed between 30 and 70% of the phytoplankton biomass in 4 lakes having pH 4.2-4.8, but only 2-30% of the biomass in 10 lakes with pH levels of 5.8-6.8 (Yan 1979). In certain poorly buffered lakes, some of the phytoplankton species may interfere with recreational use of the lakes. For example, in five lakes in Ontario and New Hampshire with pH 5.5-6.2, obnoxious odours developed during the summers of 1978, 1979, and 1980 (Nicholls et al. 1981). The odours have been shown to be caused by the growth of the planktonic Chrysochromulina breviturrita. This species was first discovered in 1976, but it is now known to inhabit more than 40 lakes in Ontario, most of which are acidic (Nicholls et al. 1981). The "invasion", and associated odour production, by this organism is apparently a recent phenomenon. Although the relationship between lake acidification and the proliferation of this species has not been proven, data collected thus far indicate that dominance of this species to an extent causing the serious odour production, is restricted to acidic lakes. Acidified lakes and streams are often characterized by increased growth of benthic filamentous algae. In Sweden, Ontario and Quebec, unusually dense and extensive masses of filamentous algae (mainly Mougeotia, Zygogonium and Zygnema sp.) proliferate in the littoral zones of many lakes with pH values of 4.5-5.5 (Blomme 1982; Grahn et al. 1974; Hendrey et al. 1976; Hultberg and Grahn 1975; Schindler 1980; Stokes 1981). These filamentous algal growths are associated either with macrophytes or other substrates or exist as floating "clouds" near the lake bottom. The accumulations of algae may reduce light availability to macrophytes, change microclimates for benthic macroinvertebrates and restrict fish feeding and spawning. Some depreciation of shoreline recreational values and activities, especially swimming, may result from this growth of algae. 3.7.2 Effects on Aquatic Macrophytes Information on the effects of acidification on macrophyte communities of soft-water lakes is still incomplete. Scandinavian investigators have suggested that when lake water pH declines, typical macrophyte dominants are replaced by very dense beds of Sphagnum (Grahn et al. 1974; Hendrey et al. 1976; Hultberg and Grahn 1975). The loss of ------- 3-106 3.7.3 Effects on Zooplankton I I some macrophyte species and the correlative increase in Sphagnum abundance may be indirectly related to depressed pH, through changes in inorganic carbon availability (Raven 1970; Steemann-Nielsen 1944, I 1946). In Scandinavia, the decline of macrophyte species and the • concurrent Sphagnum invasion begins as pH falls to about 6.0, and proceeds rapidly when pH falls below 5.0. In Lake Golden in New York • (pH 4.9), Sphagnum is abundant (Hendrey and Vertucci 1980), and in • Beaverskin Lake in Kejimkujik National Park in Nova Scotia, a clear lake of pH 5.3, Kerekes (1981) has reported extensive Sphagnum _ growth. In Ontario lakes, some species of Sphagnum have been I identified (Harvey et al. 1981), but accumulations as dense as those ™ recorded in Scandinavia have not been observed. I Sphagnum moss coverage of littoral zones creates a unique habitat that is considered unsuitable for some species of benthic inverte- brates or for use as fish spawning and nursery ground (Hultberg and • Grahn 1975). It may reduce the appeal of freshwater systems for • certain recreational activities. Through the release of hydrogen ions and polyuronic acids, Sphagnum could acidify their immediate _ surroundings should they accumulate (Clymo 1963; Crum 1976). • I I Four major groups of animals contribute to zooplankton communities: protozoans, rotifers, crustaceans and insects. Zooplankton are an m important food for many species of fish, particularly for younger • individuals. Thus, they are an essential component of the aquatic food chain, transferring energy and materials from the primary producers (algae) to consumers, including fish and man. Acidifi- • cation apparently results in reduced zooplankton biomasses, as both • the numbers and average size of community members are reduced (Yan and Strus 1980). As a result, food availability to higher trophic levels may be decreased. Acidification of lakes is accompanied by changes in the occurrence, « abundance and seasonal succession of species, and in the diversity of • crustacean (and other) zooplankton. It is often assumed that the direct cause of these changes is differences in tolerance among zooplankton species to increased H+ concentration. However, • acidification also increases the transparency of lakes, increases the 9 concentration of potential toxicants such as Cd^+ (Aimer et al. 1978) which is toxic to zooplankton at less than 1 Pg/L (Marshall and • Mellinger 1980), and produces quantitative and qualitative changes in | zooplankton predator and prey species (Harvey et al. 1981). Hence, the immediate causes for the changes in zooplankton communities that _ do occur, while linked to increased acidity, may be quite complex. • The most important components of zooplankton communities are usually the rotifers and crustaceans. Of these, the crustaceans usually form I 90% of the biomass (Pederson et al. 1976), while rotifers, because I they have shorter generation times, may be responsible for 50% of the I I ------- I I I I I I I I I I I I I I I I I I I 3-107 zooplankton productivity (Makarawicz and Likens 1979). Available studies on the effects of acidification on rotifer populations are contradictory; both smaller (Roff and Kwiatkowski 1977) and larger (Malley et al. 1982; Yan and Miller 1981) standing stocks have been observed in acidic lakes. Studies from very acidic lakes (Smith and Frey 1971) and the Smoking Hills Lakes of the Northwest Territories (Havas 1980) indicate, however, that some species of rotifers can survive when all crustacean zooplankton have been eliminated or could not survive at the low pH conditions. The diversity of zooplankton communities has been reported in several studies to be greatly reduced by acidification (Raddum et al. 1980; Sprules 1975). Whereas nonacidic lakes typically contain approxi- mately ten species of planktonic Crustacea in mid-summer collections, Sprules (1975) observed that the number of species and species diversity of acidic lakes in the La Cloche Mountains in Ontario was drastically reduced. In several cases only a single species, Diaptomus minutus, remained. The diversity of littoral cladocerans has also declined with acidification (Brakke et al. 1982). The decrease in number of species and diversity is apparently related to low pH and not to changes in aquatic macrophytes (Kenlan et al. 1982). Sediment core studies in New England and in Norway suggest that changes in littoral cladoceran assemblages occurred simultaneously with calculated dates of pH declines based on diatom analyses (Brakke et al. 1982; Davis et al. 1982). Some predacious zooplankton, for example cyclopoid copepods (Raddum et al. 1980) and Epischura lacustris (Malley et al. 1982), are very sensitive to acidification, and are often absent from acidic lakes. Densities of other predators, such as some species of Chaoborus (Eriksson et al. 1980a) and Heterocope saliens (Raddum et al. 1980), apparently increase. The significance of these changes in predator populations to zooplankton community structure is not yet understood although it may be important (Eriksson et al. 1980a). 3.7.4 Effects on Aquatic Macroinvertebrates Numerous aquatic macroinvertebrates are known to be affected by low pH conditions. In some cases an entire phylum appears to be affected, while in other situations susceptibility is species- specific. Evidence indicates that molluscs, in general, are highly susceptible to reduced pH (J. 0kland 1980; Raddum 1980; Wiederholm and Eriksson 1977), often being restricted to habitats with pH greater than 5.8-6.0. Similarly, all species of oligochaetes studied thus far have been found at lower densities in acid waters (Wiederholm and Eriksson 1977). Sensitivity to low pH has been inferred from field investigations for certain Arachnids, Crustaceans and Insects. Arachnids were only ------- 3-108 I I briefly mentioned by Grahn and his co-workers (1974); acarinids were absent in waters with pH values below 4.6. No macro-crustaceans were found below pH 4.6 (Grahn et al. 1974). Gammarus lacustris was • absent from waters with pH below 6.0 (J. j&kland 1969), while the | crayfish, Astacus astacus was rare in lakes where the summer pH value was less than 6.0 (Svardson 1974). Orders of Insecta exhibit a wide • range of sensitivities to pH. While the numbers of species of • Ephemeroptera and Plecoptera appear to be positively correlated with pH, larvae of Chironimidae (Diptera), Hemiptera and Megaloptera are often abundant in acid lakes (Aimer et al. 1978). Hutchinson et al. • (1978) reported an example of extreme tolerance by larvae of red • chironomids, Chironomus riparius, to waters of pH 2.2 in the Northwest Territories. I Although the field studies mentioned above provide evidence of the effects of acidification on certain species, the pH of a natural & system has rarely been altered experimentally, and the impacts on • invertebrates noted. The documented effects of decreased pH include the disappearance of Mysis relicta in Lake 223, an experimentally acidified lake in the Experimental Lakes Area (Malley et al. 1982), • elimination or reduction of Ephemeroptera populations in a stream in • the Hubbard Brook Experimental Forest in New Hampshire (Fiance 1978; Hall et al. 1980), and decreased emergence of some species of • Plecoptera, Trichoptera and Diptera in the same stream (Hall et al. • 1980). Those species with acid-sensitive life stages (such as emergence in insects) which can coincide with low pH snowmelt, or ^ other events, such as low pH flushing, may be especially sensitive to • acid deposition. ™ In considering the distribution of the above species in relation to B waters of varying pH no causative relationship between hydrogen ion • concentration and the observed changes has been determined as yet. Other factors vary with pH, including concentrations and availability • of nutrients, bicarbonate, and various metals. From the results • available, however, it appears that molluscs (perhaps because of their requirement for calcium) and moulting crustaceans (perhaps _ because of their large demand for calcium at the time of moult) are • the macroinvertebrates most sensitive to low pH levels. It is still • unclear why certain groups of aquatic insects are more sensitive than others. H 3.7.5 Effects on Bacteria and Fungi • The decomposition rate of fixed carbon, both allochthonous and autochthonous organic matter, is largely determined by microbial processes in the water column and in the surface layers of sediment. • Several studies have demonstrated that rates of decomposition of • organic matter are decreased at low pH values. In a laboratory study, for example, Bick and Drews (1973) demonstrated that as pH was lowered, the number of bacteria and protozoans decreased, populations of fungi increased, and the rates of decomposition and nitrification I I I ------- I I I I I I I I I I I I I I I I I I I 3-109 were reduced. Traaen and Laake (1980) measured decomposition rates of homogenized birch litter and glucose/glutamate mixtures. When the pH was decreased from 7.0 to 3.5, litter decomposition dropped to 30% of control levels, and a shift from bacterial to fungal dominance was observed. Traaen (1980) further observed that rates of weight loss of decomposing birch leaves and aspen sticks after one year in the laboratory or one to two years in field situations were significantly lower at pH levels less than 5.0. Reductions in numbers of heterotrophic bacteria have been observed previously in aquatic habitats acidified by acid mine drainage (Guthrie et al. 1978; Thompson and Wilson 1975; Tuttle et al. 1968, 1969). Caution must be exercised, however, in extrapolating results from such studies to situations where the source of protons is atmospheric because the pH is often much lower in acid mine drainage lakes, and the concentration of dissolved substances, including metals, much higher. Rao et al. (1982) studied the effects of acidic precipitation on bacterial populations of the Turkey Lakes, Ontario and Kejimkujik, Nova Scotia. They observed reduced numbers of nitrifying bacteria and sulphur cycle bacteria in low pH lakes and streams. Bacterial activity as measured by oxygen consumption rate and biodegradation or organic material was 50% less and 30-40% less respectively in acid-stressed environments compared to nonacid-stressed areas. Microbial transformations of sulphur and nitrogen species may influence lake acidity and alkalinity (Brewer and Goldman 1976). Schindler (1980) showed that increases in SO^- concentrations stimulated sulphate-reducing bacteria in lakes that develop anoxic hypolimnia. The reduction of SO^" yields OH~ thereby increasing akalinity. Stimulation of SO/^- reduction has been used with success to reclaim acid mine drainage waters. Sulphate-reducing bacteria require anoxic conditions, and are stimulated by large quantities of organic matter (i.e., they prefer conditions typical of eutrophic lakes). However, acidified lakes are not eutrophic and many have oxygenated hypolimnia. 3.7.6 Effects on Amphibians Many species of frogs, toads, and salamanders breed in temporary pools. These pools are formed by a mixture of snowmelt water and spring rains and may have low pH values during the spring. Because of the vulnerability of this habitat to pH depressions, amphibian populations are expected to be one of the earliest forms of wildlife to be affected by the acidification of fresh waters. Temporary pools used as breeding sites by Jefferson's (Ambystoma jeffersonianum) and yellow-spotted salamanders (A. maculatum) in New York were found to have pH values 1.5 units lower than nearby permanent ponds (Pough and Wilson 1977). The amphibian species of eastern Canada considered most susceptible to the effects of acid deposition because of their breeding habitat are listed in Table 3-19. ------- 3-110 TABLE 3-19. SUSCEPTIBILITY OF BREEDING HABITAT TO pH DEPRESSION FOR THOSE AMPHIBIANS IN NORTHEASTERN NORTH AMERICA WHOSE RANGE OVERLAPS AREAS RECEIVING ACIDIC DEPOSITION (modified from Clark and Fischer 1981) Potential for acidification of egg-laying habitat Habitat Species high meItwater pools moderate permanent ponds low streams lakes bogs logs and stumps Ambystoma maculatum - Yellow-spotted salamander Ambystoma laterale - Blue-spotted salamander Ambystoma tremblayi - Tremblays salamander Bufo americanus - American toad Pseudacris triseriata - Chorus frog Rana sylvatica - Wood frog Rana pipiens - Northern leopard frog Hyla crucifer - Northern spring peeper Hyla versicolor - Gray tree frog Necturus maculosus - Mudpuppy Notophthalmus viridescens - Red-spotted newt Bufo americanus - American toad Hyla versicolor - Gray tree frog Pseudacris triseriata - Chorus frog Rana catesbeiana - Bullfrog Rana clamitans - Green frog Rana pipiens - Northern leopard frog Rana septentrionalis - Mink frog Eurycea bislineata - Northern two-lined salamander Necturus maculosus - Mudpuppy Rana catesbeiana - Bullfrog Hemidactylium scutatum - Four-toed salamander Plethedon cinereus - Red-backed salamander I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-111 Detrimental effects of acidity on adult amphibians have been shown in a number of field surveys. In England, Cooke and Frazer (1976) reported that no adult newts were caught from ponds of pH less than 3.8. The natterjack toad (Bufo calamita) was not found in ponds below pH 5 (Beebee and Griffin 1977) in England. The common toad (Bufo bufo) did not occur where pH was less than 4.2, and the smooth newt (Triturus vulgaris) occurred only rarely in ponds at pH values less than 6.0. Hagstrom (1977) observed that the common toad and common frog (Rana temporaria) disappeared when pH levels reached 4.0-4.5. In New Hampshire, when a section of Hubbard Brook was artificially acidified to mean pH 4.0, salamanders disappeared from the study area (Hall and Likens 1980). Pough (1976) noted heavy embryonic mortalities and deformities in the yellow-spotted salamanders which breed in temporary meltwater ponds with pH less than 6.0. In central Ontario, Clark and Euler (1981) reported that the numbers of egg masses of yellow-spotted salamanders and male calling densities (an estimate of population size) of spring peepers (Hyla crucifer) were positively correlated with pH. This latter species often breeds in stream inflows and outflows or along the littoral zone of lakes, habitats also subjected to particularly heavy acid loads as a result of snow melt (Clark and Euler 1981). Bullfrog (Rana catesbeiana) and wood frog (Rana sylvatica) densities were also reduced in acidic streams and ponds (Clark and Euler 1981). Strijbosch (1979) reported a negative correlation between pH and percentages of dead and moulded egg masses of frogs and toads in the Netherlands. Laboratory experiments have demonstrated that reductions in pH are both directly and indirectly responsible for mortalities and deformities found during amphibian embryonic development. Gosner and Black (1957) studied the sensitivity of 11 species of frogs and toads to conditions of depressed pH and found that the embryos were more sensitive than adults. Frogs may undergo iono-regulatory failure due to acidic conditions (Fromm 1981) similar to that reported for fish (Leivestad and Muniz 1976; McWilliams and Potts 1978; Muniz and Leivestad 1980; Packer and Dunson 1970). In the case of the cricket frog (Acris gryllus) and northern spring peeper, an exposure of embryos to water in the vicinity of pH 4.0 for a few hours resulted in greater than 85% mortality. Beebee and Griffin (1977) noted abnormalities in natterjack toad spawn exposed to low pH, and Noble (1979) observed delayed development and embryonic mortality in the leopard frog (Rana pipiens) at pH less than 4.75. The leopard frog may be more sensitive to low pH than the latter study indicates. Schlichter (1981) found decreased sperm motililty at pH values less than 6.5 and the percentage of eggs which formed healthy embryos decreased below pH values of 6.3. A similar study using the common frog reported that sperm motility was reduced to 50% of maximum at pH values of 6.4-6.7 and to 0% at pH values less than 6.0 (Gellhorn 1927, cited in Schlichter 1981). ------- 3-112 I I Cook (1978) found no significant correlation between pond pH and percent embryonic mortality in either the yellow-spotted salamander or Jefferson's salamander studied in six ponds with mean pH values of I 5.3-5.6. In contrast Pough (1976) found heavy embryonic mortalities I and deformities for both species in waters with pH values less than 6.0. Egg transplant studies suggest that yellow-spotted salamander • eggs from acidic ponds are more tolerant to acidity than eggs from • neutral ponds (Nielsen et al. 1977). While Hagstrom (1977) reported the elimination of the common toad at pH values of 4.0-4.5, Cooke _ (cited in Beebee and Griffin 1977) found this species in waters of • pH 4.2 and noted that tadpoles were able to tolerate this hydrogen • ion level. It is likely that other factors influenced by the acidity of the | water may cause detrimental effects upon amphibian development. For example, Huckabee et al. (1975) suggest that the combined effects of m low pH and increased concentrations of aluminum, manganese and zinc • may be the cause of the high mortality of shovel-nosed salamander (Leurognathus marmoratus) larvae in Great Smoky Mountain National P ark";• Frogs, toads, and salamanders are important components of both aquatic and terrestrial ecosystems. Orser and Shure (1972) reported • that amphibians are among the top carnivores in temporary ponds and | small streams, and are important predators of aquatic insects. In turn, they serve as a high protein food source for other wildlife M (Burton and Likens 1975b). Many birds and mammals depend heavily on • these species for food (Burton and Likens 1975a; Cecil and Just 1979; DeBenedictis 1974). 3.7.7 Effects of Low pH on Fish The purpose of this section is to review briefly how fish respond to | low pH conditions. This will be done on the basis of documented changes in fish population related to acidification, other field _ evidence and laboratory substantiation. For more comprehensive • treatments of this subject, the reader is referred to reviews by ™ Fromm (1980), Haines (1981c) and Spry et al. (1981). In addition, there is extensive literature available on laboratory studies (see • Doudoroff and Katz 1950; EIFAC 1969), that were designed to elucidate • mechanisms of pH toxicity. These laboratory results are reviewed, as they are useful in explaining field observations and suggesting new • directions for field studies. • Results from laboratory experiments demonstrate how overall water ^ quality (i.e., hardness, ionic strength) can affect pH toxicity. For • example, as ionic strength and water hardness increase, the short- ™ term sensitivity of fish to waters with pH values of 4 is decreased (reviewed in Spry et al. 1981). The ameliorative effects of high B Ca^+ and ionic strength appear most beneficial to early larval stages • at intermediate pH values (^5). This is consistent with field I I ------- I I I I I I I I I I I I I I I I I I I 3-113 observations that fish communities disappear from more dilute waters at higher pH levels than they do from lakes with higher concentra- tions of salts (Leivestad and Muniz 1976). In addition to hardness and ionic strength, survival of fish in water of low pH is influenced by the type of acid present (Packer and Dunson 1972; Swartz et al. 1978), temperature (Kwain 1975; Robinson et al. 1976), the level of dissolved carbon dioxide in the water (Neville 1979), and by the presence of metals (Baker and Schofield 1980; Swartz et al. 1978). Salts are lost from plasma and body tissue of fishes exposed to low pH conditions. Leivestad et al. (1976) found that Na+ and Cl~ in blood plasma and K+ in muscle tissue declined in brown trout at low pH levels. Increases in the concentration of Ca2+ enabled the trout to regulate better ionic balances (Leivestad et al. 1980). Recent studies by Saunders et al. (1982, in press) have shed light on possible mechanisms affecting survival, growth, and the smelting process in Atlantic salmon. Under low pH laboratory conditions it was found that parr-smolt transformation was impaired, and ATPase activity was lowered, resulting in a decreased salinity tolerance of smolts. Salmon raised under low pH regimes (i.e., pH 4.2-4.7) were found to have significantly lower plasma Na+ and Cl~ levels, which was indicative of an impaired osmoregulatory ability in fresh water. Field evidence suggests that the susceptibility to low pH appears to be species-specific. From his studies of La Cloche Mountain lakes, Beamish (1976) estimated the pH at which reproduction ceased in 11 species of fishes (Table 3-20). As well as interspecific differences in sensitivity, variability in sensitivity has also been observed among different strains of the same species (Robinson et al. 1976; Swartz et al. 1978). However, it is likely that the acidification of lakes and rivers in North America is proceeding too rapidly to enable genetic selection for acidic tolerant strains to occur naturally (Schofield 1976b). Results of laboratory and field studies have demonstrated that some species of fish are particularly sensitive to low pH levels in certain reproductive stages (reviewed by Spry et al. 1981). Low pH can inhibit gonadal development (Ruby et al. 1977, 1978), reduce egg production (Craig and Baksi 1977; Mount 1973) affect egg and sperm viability (EIFAC 1969; Menendez 1976) and inhibit spawning (Craig and Baksi 1977; Menendez 1976). Embryonic development may also be affected by low pH (Swartz et al. 1978; Trojnar 1977) and low environmental pH can affect egg internal pH (Daye and Garside 1980). Generally, fry appear less resistant to low pH than eggs (Spry et al. 1981), and therefore fry may be particularly vulnerable to low pH conditions associated with spring melt and storm events. Hulsman and Powles (1981) conducted experiments on walleye eggs. The eggs were incubated in situ in a series of small streams in the La Cloche area of Ontario. The various sites ranged in pH from 4.60 to 6.72. Hatching success was significantly reduced in the clear dilute streams with pH values less than 5.40. ------- 3-114 TABLE 3-20. APPROXIMATE pH AT WHICH FISH IN THE LACLOCHE MOUNTAIN LAKES STOPPED REPRODUCTION (Beamish 1976) 6.0 to 5.5 5.5 to 5.2 5.2 to 4.7 4.7 to 4.5 Species Smallmouth bass Micropterus dolomieui Walleye Stizostedion vitreum Burbot Lota lota Lake Trout Salvelinus namaycush Troutperch Percopsis omiscomaycus Brown bullhead Ictalurus nebulosus White sucker Catostomus commersoni Rock bass Ambloplites rupestris Lake herring Coregonus artedii Yellow perch Perca flaveseens Lake chub Couesius plumbeus Family Centrarchidae Percidae Gadidae Salmonidae Percopsidae Ictaluridae Catostomidae Centrarchidae Salmonidae Percidae Cyprinidae I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-115 One mechanism which appears to contribute to species extinction in acidified systems is the failure of recruitment of year classes. In a study of 38 La Cloche lakes, Ryan and Harvey (1980) reported evidence of recruitment failure in yellow perch (Perca flavescens) populations in the two lakes of lowest pH values: Patten Lake (pH 4.1) and Terry Lake (pH 4.3). The age group composition of yellow perch in Patten Lake is illustrated in Figure 3-40. Ryan and Harvey (1981) also found evidence of reduced and missing year classes of young fish in five populations of rock bass (Ambloplites rupestris) in acid-stressed La Cloche lakes. The absence of older individuals in populations of fish in some acid-stressed lakes has also been reported (Harvey 1980; Ryan and Harvey 1980). This effect is illustrated by the changes in age composition of white suckers in George Lake, Ontario from 1967 to 1979 (Figure 3-41), Rosseland et al. (1980) also reported the absence of post-spawning age perch and brown trout in three lakes within the Tovdal River System, Norway. In the field, there have been several reports of fish kills appar- ently related to the low pH of rivers and lakes. In Scandinavia, for example, Jensen and Snekvik (1972) reported mass mortality of Atlantic salmon (Salmo salar), and Leivestad and Muniz (1976) reported a brown trout (Salmo trutta) kill. Both fish kills have been correlated with reduced water pH, although Al was not measured in either case. In North America, Harvey (1979) reported mortalities of several species, primarily pumpkinseeds (Lepomis gibbosus) in Plastic Lake, Ontario, during spring snowmelt runoff and pH depression. Surface water pH was 5.5, while the pH of the major inlet stream was 3.8. During the spring of 1981, some in situ bioassays were conducted in Plastic Lake (Harvey 1981). Rainbow trout, Salmo gairdneri, were placed in cages at four locations in Plastic Lake and at four locations in the control, Beech Lake. Three nonmetal cages of 35 fish were situated at each location. No mortality occurred at any of the cage sites in the control lake (pH 6.09-7.34, alkalinity 132-390 ;ieq/L). In Plastic Lake, however, mortality ranged from 12% at the lake outlet site (pH 5.0-5.85) to 100% at the inlet site (pH 4.03-4.09). Although aluminum concentrations were not measured at the time of the 1979 fish kill and aluminum data for 1981 is not yet available, total aluminum concentrations in Plastic Lake during the 1979 and 1980 ice-free season varied between 9 and 30 ug/L in the lake, and between 240 and 490 Pg/L in the major inlet. 3.7.8 Effects of Aluminum and Other Metals on Fish Concentrations of metals can be elevated in acid-stressed lakes (Beamish 1974a; Raines 1981c; Scheider et al. 1979b) because of increased atmospheric deposition, increased mobilization from the sediments and/or mobilization from the watershed (see Section 3.2.4). ------- 3-116 35 30 25 x: 0) b 20 "o 0> 1 15 z 10 5 0 PATTEN LAKE - - cw i i i 012345 Age in • 1 1 1 1 1 678 Years Figure 3-40. Age composition of yellow perch (Perca captured in Patten Lake, Ontario, pH 4 Harvey 1980). 1 Q 1 1 10 1 1 1 1 1 1 1 1 1 1 1 1 1 f lavescens) .1 (Ryan and • 1 1 ------- I I I I I I I I I I I I I I I I I I I 3-117 w il E 3 •z. 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 Age in Years Figure 3-41. Changes in the age composition of the white sucker (Catostomus commersoni) in George Lake, Ontario (Harvey et al. 1981). ------- 3-118 I I I I One of the most important consequences for fishes of watershed acidification is the mobilization of aluminum from the watershed to the aquatic environment (Cronan and Schofield 1979). Elevated levels of aluminum in waters have been shown to have serious effects on fish within the pH range normally considered not harmful to aquatic biota (Baker and Schofield 1980). _ Spry et al. (1981) give a simplified description of the complex chemistry of aqueous aluminum. The solubility of aluminum is minimal at pH 5.6-6.0, increasing as pH increases or decreases outside this • range (Figure 3-42). At pH greater than 5.5, soluble aluminum is • mostly anionic; at pH less than 5.5 it exists increasingly as a cation. The solubility of aluminum is apparently regulated by some • form of aluminum trihydroxide solid, Al(OH)3(s), which is minimally || soluble at pH values of 5.6-6.0 (Driscoll 1980b). Fewer hydroxyl ligands at lower pH allow the aluminum to become cationic, eventually • becoming Al^+ at pH values less than 4.5 to 5.0. Cationic aluminum • is able to form complexes with a number of ligands, including soluble organics and fluoride, decreasing its toxicity (Figure 3-42) (Baker and Schofield 1980; Driscoll et al. 1980). • Laboratory studies have shown significant reductions in fish survival at inorganic aluminum concentrations of 100 and 200 pg/L and greater for white suckers (Catostomus commersoni) and brook trout (Salvelinus fontinalis), respectively (Baker and Schofield 1982; Schofield and Trojnar 1980). Inorganic aluminum levels as high as 600 yg/L have « been measured in acidic Adirondack waters (Driscoll 1980b). Baker • and Schofield (1980) note that fry exposed soon after initiation of feeding and yolk sac absorption were more sensitive to elevated aluminum concentrations than were eggs and sac fry prior to yolk • absorption. They also found that the presence of aluminum actually • mitigated the toxic effects of low pH to fish eggs. The survival of brook trout and white sucker embryos through the eyed stage at pH • levels below 5 was significantly better in treatments with aluminum • than without. After hatching, brook trout fry were more susceptible to aluminum at the extremes of the pH range tested (4.2 to 5.5) than at intermediate pH levels (Figure 3-43). The greater susceptibility • of fry at these extreme pH values may reflect a dual mechanism of ^ aluminum toxicity. At low pH, aluminum (probably Al-'"*") appears to cause osmoregulatory stress and loss of salts from blood plasma H (Baker 1981; Leivestad and Muniz 1981). At higher pH values (5.5), | precipitation of Al(OH)3(s) damages the gills and leads to clogging by mucous (Baker 1981; Schofield and Trojnar 1980). Baker and mt Schofield (1980) also found that, at all stages, white suckers were • substantially more sensitive to low pH levels and elevated aluminum concentrations than brook trout. Schofield and Trojnar (1980) suggested that levels of aluminum, • rather than pH alone, may be the primary factor limiting survival of brook trout stocked in Adirondack lakes. Muniz and Leivestad (1980) • and Schofield and Trojnar (1980) suggested that mass mortalities of | fish, observed during episodes of acidification in the spring, were I I ------- 3-119 100 80 60 CO 3 CO 40 20 10 14 15 Time (days) Figure 3-42. Percent survival of brook trout fry plotted as a function of time in treatment waters at pH level 5.2 with no aluminum (control) or with 0.5 mg Al added per liter with no additional complexing agents (Al) or with 0.5 mg fluoride/litre (Al + F) or with 30 mg (Baker and Schofield 1980). ------- 3-120 + 100 C 1 C o 4-f o C 3 u_ CO 0} (0 ~G -100 3 CO C o 'w 0) 0) t_ O) o QC •5 -200 0> a o CO \ \ X 4.0 Figure 3-43. 4.5 5.0 5.5 PH Slope of the regression line of brook trout survival (arcsin transformation) as a function of total aluminum concentration at each pH level, plotted as a function of pH level. A positive slope indicates presence of aluminum improved survival: a negative slope indicates detrimental effects of aluminum (Baker and Schofield 1980). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-121 most likely a result of elevated concentrations of inorganic aluminum, mobilized from the soils by strong acids present in snowmelt water. The former study demonstrated that pH declines alone (to levels of pH 4.7-5.0) did not induce physiological stress in fish, as determined from changes in plasma chloride levels. However, associated increases in aluminum levels to 0.2 mg/L or more were found to be sufficient to induce severe stress and eventual mortality (Muniz and Leivestad 1980). Aluminum levels in streams in the Adirondack Region of New York State (Driscoll 1980b), in the Great Smoky Mountains National Park, U.S.A. (Herrmann and Baron 1980), and in the Muskoka-Haliburton area of Ontario (total aluminum levels from 1976 to mid 1978 ranged between 5 and 1000 yg/L in 60 streams) (Ontario Ministry of Environ- ment data from ongoing studies), fall within levels demonstrated to be lethal to fish in laboratory conditions. However, as the laboratory studies have demonstrated, the evaluation of aluminum as a toxic element in acidified waters is not a simple function of total concentration. In evaluating the survival of indigenous fish populations one must consider the form of aluminum, the level of hydrogen ion, the fish species present and their life history stage. Other metals besides aluminum also occur at elevated levels in acidic waters (Section 3.2.4). Harvey et al. (1982) reported increased lakewater concentrations of manganese were associated with decreasing pH for 50 lakes in the Wawa area of Ontario. They found Mn was elevated when pH values were less than 5.0 and reached very high concentrations in strongly acidified lakes. In the La Cloche Mountain lakes, Mn was correlated inversely with pH and Mn declined in acidified lakes in the Sudbury area following neutralization (Harvey et al. 1982). Manganese has been considered a relatively non-toxic element, and thus toxicological data are very limited. Lewis (1976) determined that manganese concentrations up to 770 yg/L had no effect on survival of rainbow trout in soft waters with pH levels 6.9 to 7.6. Concentrations of manganese in acidic waters have been measured up to 130 to 350 ug/L (Dickson 1975; Schofield 1976c). Available data suggest that manganese levels, by themselves, have no apparent adverse effects on fish, although Harvey et al. (1982) found elevated Mn concentrations in the vertebrae of white sucker (Catostomus commersoni) from acid lakes. Although laboratory bioassays examining effects of zinc on fish are numerous, none of these studies considered soft waters with pH levels below 6. Chemical models predict that as the pH level declines, an increasing proportion of the total zinc concentration should exist as the free aquo ion (Stutnm and Morgan 1970). For many metals, the free aquo ion (i.e., Me^"1") is considered the most toxic form (Spry et al. 1981). This has not been confirmed to be true for zinc but care should be taken in extrapolating bioassay data and maximum acceptable toxicant concentrations (MATC) determined for pH levels ------- 3-122 I I I above 6 to conditions in acidic waters. For the most part, however, lethal concentrations of zinc in bioassays are 10 times the zinc concentrations found in acidic lakes (Spry et al. 1981). Sinley et al. (1974) estimated that the MATC for rainbow trout (Salmo gairdneri) exposed to zinc in soft, circumneutral water was between 140 and 260 yg/L. Benoit and Holcombe (1978), in life cycle • experiments with fathead minnows (Pimephales promelas) in soft water, I determined that the threshold level for significant adverse effects on the most sensitive life history state was between 78 and 145 yg/L. Zinc concentrations in acidic waters range up to 23 to 56 yg/L • (Henriksen and Wright 1978; Norton et al. 1981a; Schofield 1978; Spry • et al. 1981). I In some regions, concentrations of cadmium, copper, lead and nickel (see Section 3.2.4) are also elevated in acidic lakes. Relation- ships between pH levels and cadmium, copper, lead, and nickel • concentrations, however, vary markedly between regions. High I concentrations of these metals probably result primarily from increased atmospheric loading and deposition, and occur principally in surface waters in close proximity to pollutant sources (e.g., I Sudbury, Ontario, Nriagu et al. 1982). Concentrations of some of I these metals in lakes in the vicinity of Sudbury have been demon- strated to have definite adverse impacts on fish and other aquatic • biota (Conroy et al. 1976; Yan and Strus 1980). Excluding lakes | within 50 km of Sudbury, acidic Ontario surface waters have concen- trations of metals ranging up to about 0.6 yg/L Cd, 9 yg/L Cu, _ 6 yg/L Pb and 48 yg/L Ni (Spry et al. 1981). Spry et al. (1981) • reviewed bioassay data available and noted no significant adverse effects on fish survival and reproduction at concentrations up to 0.7-11.0, 9.5-77, 13-253, and 380 yg/L for cadmium, copper, lead, and • nickel, respectively. In general, concentrations of metals in acidic I waters are below these "safe" concentrations (unless there is a local source of metal emissions). However: (1) most of these bioassays • were conducted in waters with pH levels above 6, and (2) the I possibility for synergistic effects has not been evaluated. In some regions, bioaccumulation of mercury in fish has been I correlated with low pH levels in lakes. These elevated levels of ™ mercury in fish may have adverse effects on consumers (e.g., man or fish-eating birds and mammals; Sections 3.7.12 and 5.2). However, no I data have been reported to indicate that this bioaccumulation has any I adverse effects on the fish themselves (Haines 1981c). Survival of fish populations in acidic waters is determined primarily I by levels of pH and inorganic aluminum (Baker 1982; Schofield and Trojnar 1980). Although concentrations of a number of metals are _ increased in acidic lakes and streams, definite effects on fish have • been demonstrated only for aluminum (except for lakes immediately ™ around Sudbury, Ontario). Other metals may play a lesser, but as yet undefined, role. I I I ------- I I I I I I I I I I I I 1 I I I I I I 3-123 3.7.9 Accumulation of Metals in Fish 3.7.9.1 Mercury There is substantial evidence of the effect of pH on mercury content in fish (Brouzes et al. 1977; Hakanson 1980; Landner and Larsson 1972). Bisogni and Lawrence (1973) and Jernelov et al. (1976) have argued that one reason fish in waters of low pH contain more methylmercury than fish in waters of comparable mercury contamin- ation, but higher pH, seems to be that more acidic waters retain the monomethyl-form of mercury in solution. It is, however, important to recognize that pH is not the only variable which determines the mercury burden in fish. Other factors include mercury availability, level of bioproduction (i.e., lake trophic state), lake flushing rates and lake/watershed drainage area ratio (Hakanson 1980). Few data exist to link mercury concentrations in fish to lake acidification. However, an increase in concentrations of mercury in fish from 1970 to 1978 is evident in some lakes in the Adirondack Mountains (Schofield pers. comm.). In Ontario, Suns et al. (1980) sampled young-of-the-year and yearling fish for contaminant studies. Their data (Figure 3-44) demonstrate increased mercury concentrations with decreasing pH in lakes in the Muskoka-Haliburton area. At any given pH level, however, the variation of mercury concentrations in fish is substantial. For lakes with similar pH, the mercury concentrations were higher in fish from lakes with a higher ratio of drainage area/lake volume. This result implies that a quantity of mercury from either direct atmospheric deposition or from watershed leaching is influencing the concentrations in fish. Data for 1981 are shown in Table 3-21 (Suns 1982). In 1980, the survey was extended to include adult smallmouth bass. Fish from six of the nine lakes studied had average mercury concentrations above the Canadian guideline (500 ng/g) for unlimited human consumption. In one lake mercury concentrations in fish exceeded the U.S. guidelines of 1000 ng/g (Suns 1982). Because of increased mobility and leaching under acidic conditions and/or deposition, it is possible that metals other than mercury may be accumulating in fishes. At present, however, the data base is extremely limited (Haines 1981c). In a survey of Ontario lakes by Suns (1982), yearling yellow perch were analyzed for body burdens of lead, cadmium, aluminum, and manganese. The data are shown in Table 3-21 and are summarized below. 3.7.9.2 Lead A significant (p less than 0.01; r = -0.74) correlation was found to exist between lead residues in perch and lake pH. Mean lead residues as high as 428 ng/g were found from Moot Lake (pH 5.5) and 403 ng/g from Fawn Lake (pH 5.4). ------- 3-124 200 180 160 140 o> c 120 CO ^ +* 0) o i 100 o I 80 60 40 20 13 4.5 LAKE #, NAME 1. Duck Lake 2. Little Clear Lake 3. Harp Lake 4. Bigwind Lake 5. Nelson Lake 6. Chub Lake 7. Crosson Lake 5.0 5.5 6.0 PH 6.5 A = 0.63 p < 0.05 11 TWP. LAKE #, NAME Minden 8. Dickie Lake Sinclair 9. Leonard Lake Sinclair 10. Heney Lake Oakley 11. Cranberry Lake Bowell 12. Healey Lake Ridout 13. Clear Lake Oakland 14. Fawn Lake 7.0 TWP. McLean Mo nek McLean Guilford McCauley Stanhope McCauley Figure 3-44. Mercury concentrations in yearling yellow perch vs. epilimnetic pH for selected lakes in Ontario (Suns et al. 1980). 7.5 I I I I I I I I I I I I 1 I 1 I I I I ------- 3-125 < Q 3 Z 0 p._.l i m M hj If 3 0 w en § U B t^ £-4 M ^ W 1 s~\ • cfl ^H ^J oo cd ON O 4-> 1 C 01 33 e o 5 &> o H M PM -H > s 5 0 W hJ J in Ed O P 0 M Z 4J M CO i-J i-l Pti C W S p-i Z-S M M cd C/3 <^J w c 1-0 0 o ^ crt O ss HM M-4 N H i, f ^ M VM? M *O 0) 01 t^ C ^3 i-l T3 -H ji a *j a c & o cd AJ (-1 cd oo 3 P 1-1 eg M -H Q U hJ 33 P3 M • ••••• •-i cxi co -d- m vo oo • o + | m r-. rx CXI + 1 CM vO m + 1 § i — i CO vO + 1 O CO 1 — 1 oo + 1 CO in i— H i-H vO >, tfl w O s • r-. O ^H i-H i — 1 + 1 +1 cxi m 1 — vO i— i r^ ^H O + 1 +1 m vo co en oo m i-H + 1 +1 in I-H vo r^ cxi .* CXI ON i-H + 1 +| O ON OO -H CXI -H S>-. J -H i-t O J4 cd oi u a ,s a • 0 00 ON -H r»- 0 + 1 H cd 01 iH CJ • cxi i — i •3- •-H + | vO vO in ex! + 1 vO , 01 0 0) di • CO r— ( CM i-H + | o -* «* CO i-H + 1 ON CO i-H cxi CO + 1 «* tf st CXI + 1 ON m fvl tN •<)• m g « fe • P^ i— 1 ------- 3-126 3.7.9.3 Cadmium 3.7.10 Effects on Fisheries in Canada and the United States 3.7.10.1 Adirondack Region of New York I I Although there is evidence that lead concentrations in water as low as 8 ng/L can cause neurological disorders in fish (Davies et al. 1976; Hodson et al. 1978; Holcombe et al. 1976), no data are • available to relate body-burden accumulations to any significant • biological response. I A statistically significant (p <0.05; r = 0.60) correlation exists I between cadmium residue levels and lake pH (Table 3-21). Little | reference material is available at this time to evaluate the environmental significance of these cadmium accumulations. However, *j a laboratory study using relatively hard water (pH 7.5; alkalinity • 980 yeq/L) showed that 80 yg/L killed 50% of the test population of young-of-the-year largemouth bass in 82 days. The same study _ discovered that 8 iig/L induced "abnormal behaviour" in the young fish I in 12-week exposure (Clearley and Coleman 1974). The young bass • average body-burden accumulations of cadmium were 38 ng/g after a four month exposure to a concentration of 80 yg/L. Although it is fl difficult to apply these laboratory data to field conditions, it is || apparent that cadmium residue accumulations in fish tissue from Ontario lakes, particularly in the more acidic lakes, were consider- • ably higher than accumulations observed under laboratory conditions • to cause biological effects. I 3.7.9.4 Aluminum and Manganese No correlations between lake acidity and mean residue accumulations • were apparent in the 1981 collections. It is likely that differences | in lake complexing capacities influence aluminum availability for uptake. Therefore factors other than pH and alkalinity will have to «| be considered to evaluate fully residue accumulations. • Moreau et al. (1982) compared the chemical content of opercula and scales of brook trout from lakes in Laurentian Park classified by • Richard (1982) as more acidic (Group 1, described in Section 3.7.10) • with the same calcified tissue from brook trout from three nonacidic lakes (Group 3, also described in Section 3.7.10). They reported ij that the content of manganese, zinc and strontium was significantly || higher in the calcified tissue of brook trout from the acidic lakes. 1 I The Adirondack region is one of the largest sensitive lake districts in the eastern United States, and it is also receives the highest • annual loading of wet sulphate. A recent inventory of Adirondack I waters classified lakes by type of fishery supported (Pfeiffer and I I ------- I I I I I I I I i i i I i i i i i i i 3-127 Festa 1980). These authors suggest that acidic deposition has exerted the greatest negative impact on the brook trout fishery. Brook trout are frequently the only game fish species present in the many small headwater ponds located at high elevations in the Adirondacks and particularly susceptible to acidic deposition. It is difficult to evaluate exactly how many fish populations have been lost from Adirondack waters as a result of acidification. The Adirondack region encompasses approximately 2877 individual lakes and ponds. Pfeiffer and Festa (1980) note that 180 Adirondack ponds that formerly sustained brook trout populations (either naturally or by stocking) no longer support such populations. It has not however been formally demonstrated that all (or most of) these populations extinctions occurred as a result of acidic deposition. For at least a few lakes (reviewed in Pfeiffer and Festa [1980]) historic records of fish population status, fish management procedures, and water chemistry do suggest that population declines were associated with a decrease in pH level and that alternative explanations for the loss of fish other than surface water acidification seem unlikely. Schofield (1976a) surveyed high elevation Adirondack lakes (total 214 lakes). For 40 of these lakes, historical data on fish and pH were available (Figure 3-32). In the 1930s, only 8% of these lakes had pH <5.0, 10% had no fish whereas in the 1970s, 48% had pH <5.0 and 52% had no fish. In some cases, entire fish communities consisting of brook trout, lake trout, white sucker, brown trout, and several syprinid species apparently have been eliminated over the 40-year period (Schofield 1976a, 1981, 1982). The present-day distribution of fish in Adirondack lakes and streams in relation to pH provides additional circumstantial evidence of the impact of acidification on fish. For high elevation lakes, Schofield (1976b, 1981, 1982) noted that the occurrence of fish was reduced at pH levels below 5.0 (Table 3-22 and Figure 3-32). Brook trout occur less frequently in lakes with pH <5.0, white suckers at pH <5.1, creek chub at pH <5.0, lake chub at pH <4.5 to 5.0, and brown bullhead at pH <4.7 to 5.0 (Schofield 1976b). About 50% of high elevation lakes had pH levels below 5.0 in 1975 and 82% of these acidic lakes were devoid of fish (Schofield 1976b). High elevation lakes, however, constitute a particularly sensitive subset of Adirondack lakes. It cannot be inferred that 50% of all Adirondack lakes have pH<5.0, nor that all lakes currently without fish once had fish and have lost their fish populations as a result of acidification. Indices of fish population status in Adirondack streams (sample of 42 streams) were also found to be positively correlated (p < 0.05) with pH measurements (Colquhoun et al. 1980). In addition to these observations of fish population status in Adirondack waters as related to acidity, Schofield and Trojnar (1980) examined the effect of water quality on fish stocking success. Poor survival of brook trout fall fingerlings stocked into Adirondack ------- TABLE 3-22. DISTRIBUTION AND FREQUENCY DURING SURVEYS OF BRACKETS pH <4.5 Total lakes 16 % of total 7.1 No fish 16 % 17.2 Fish 0 % Brook trout 0 f .80 Lake trout 0 % f Bullhead 0 % f White sucker 0 % f .15 Creek chub 0 % f Golden shiner 0 % 15.0 f .15 Common shiner 0 f Lake chub 0 % f Redbreast sunfish 0 % f Common sunfish 0 % f REFER TO 4.5-4.99 95 44.2 74 79.6 20 20.0 16(26) 19.5 .72 0(5) 8(8) 16.0 .40 3(1) 8.3 .28 0(7) 3(4) 15.0 .12 9(2) 9.1 .05 KD 14.3 .05 0 0(1) OF OCCURRENCE OF FISH SPECIES 3-128 COLLECTED 1 1 ADIRONDACKS LAKES >610 METRES ELEVATION. NUMBERS IN EXTINCT 5.0-5.49 36 16.7 2 2.1 25 25.0 18(1) 21.9 1.00 1(2) 7.7 0.4 11(1) 22.0 .44 7(1) 19.4 .73 5 18.5 .20 3 5.0 .09 0(1) 0 0 0 0 POPULATIONS 5.5-5.99 15 7.0 1 1.1 11 11.0 11 13.4 .77 4 30.8 .36 5 10.0 .45 8 22.2 .32 7 25.9 .64 1 40.0 .36 3(1) 27.3 .27 2 28.6 .18 0 1 16.7 .09 (Schofield 6.0-6.49 6 28 13.0 0 22 22.0 17 20.7 .89 2 15.4 .09 14 28.0 .64 7 19.4 .42 5(1) 18.5 .23 8 15.0 .16 1 9.1 .05 0 0 1 16.7 .05 1976b) .5-6.99 22 10.2 0 19 19.0 17 20.7 1.00 4 30.8 .21 9 18.0 .47 8 22.2 1.00 8 29.6 .42 3 10.0 .67 3 27.3 .16 1 14.3 .05 3 100.0 .16 2 66.7 .11 >7.0 3 1.4 0 3 3.0 3 3.7 2 15.4 .67 3 6.0 1.00 3 8.3 2 7.4 .67 2 3 27.3 1.00 3 42.9 1.00 0 2 66.7 .67 TOTAL 215 93 100 82 13 50 36 27 20 11 7 3 6 i 1 • 1 1 V i i •• i i i i i i 1 ------- I I I I t I I I 1 I I I I I I I I 1 I 3-129 lakes was significant, (p <0.05) associated with low pH levels and elevated aluminum concentrations. Schofield (1982) summarized available data relating water acidity and fish population status for the eastern United States. With the exception of studies in the Adirondack region, very few of these studies included comprehensive inventories of fish populations and no adverse effects of acidic deposition on fish have been definitely demonstrated. Discussions generally refer only to "potential impact". 3.7.10.2 Ontario More data on inland fisheries resource effects resulting from lake acidification are available from Ontario than from any other province in Canada. The case study of lakes in the La Cloche Mountain range by Beamish and Harvey (1972) is best known. These lakes have a naturally low buffering capacity and are only 65 km southwest of the Sudbury smelters. Some of the lakes had no fish populations at the time of the first survey, 1965-66; others had populations that were endangered, and still others were apparently in a healthy condition (Beamish 1976). The fish community of Lumsden Lake (one of 68 examined) has been studied for 14 years. The following chronology of fisheries losses has been assembled by Harvey (1980) from his studies with Beamish (Beamish and Harvey 1972), from provincial government fish capture records dating to the early 1960s, and from observations by local anglers and residents for some species prior to 1960: 1950s - 8 species present 1960 - last reported capture, yellow perch, Perca flavescens and burbot, Lota lota 1960-65 - sport fishery fails (pH 6.8, Sept. 1961) 1967 - last capture of lake trout, Salvelinus namaycush and slimy sculpin, Cottus cognatus 1968 - tagged population of white sucker, Catostomus commersoni disappears 1969 - last capture of trout perch, Percopis omiscomaycus and lake herring, Coregonus artedii 1971 - last capture of lake chub, Couesius plumbeus (pH 4.4, Aug. 1971) In their study, Beamish and Harvey (1972) also reported the loss of fish from nearby Lumsden III, Lumsden II and O.S.A. lakes. They ------- 3-130 I I 1 interpreted these observations as evidence that the factor(s) affecting the fishes of Lumsden Lake were probably widespread. They also noted that both sport and nonsport fishes had disappeared from the lakes, suggesting that overfishing was not responsible. The loss of populations of lake trout, lake herring, white suckers and other species was attributed to decreasing pH. Historical data available • for Lumsden Lake indicated that in one decade (1961-1971) the lake pH I had decreased from approximately 6.8 to 4.4. Measurements of pH from 1961 or earlier were available for eleven other La Cloche Mountain Lakes, and corresponding 1971 measurements for these lakes indicated 4 that pH had decreased one to two units in the decade. I Beamish (1974a) also examined fish populations in O.S.A. and Muriel Lakes. He found that few fish remained in O.S.A. Lake. While several species were present in Muriel Lake, only the yellow perch population appeared unstressed. A case history of another La Cloche « Mountain lake, George Lake, was compiled by Beamish et al. (1975) for • the years 1966 through 1973. They estimated that the pH of George * Lake decreased at an annual rate of 0.13 pH units. Coincident with the reduction in lake pH, populations of lake trout, walleye, burbot IB and smallmouth bass were lost in this period. In 1973, most brown m bullheads, rock bass, pumpkinseeds and northern pike did not spawn. Mountains and the concomitant loss of fish populations. He also examined other possible explanations for the response of fishes in — these lakes. He concluded that decreased pH appeared to be the • principal agent stressing the fish populations, as well as controll- ™ ing the concentrations of metals. Examination of the age distribution of white suckers in George Lake I in 1972 indicated no missing year classes and it was concluded that no major reproductive failures had occurred prior to 1972 (Beamish et • al. 1975). The pH of George Lake was measured colorimetrically In • 1960 as 6.5, ranged between 4.8 - 5.3 in 1972-73 and was 5.4 in 1979 (Harvey et al. 1981). In 1967, the white sucker population contained _ fish up to 14 years of age. By 1972, almost no fish were older than • 6 years. Sampling in 1979 revealed that 90% of the population was ™ composed of two- and three-year old fish (Figure 3-41). • Harvey (1980) also showed that the white sucker population of Crosson 0 Lake (pH 5.1; Muskoka-Haliburton) had a truncated age distribution with few fish older than five years (Figure 3-45) compared with the m age composition of white suckers in less acidic Red Chalk (pH 6.3) I and Harp (pH 6.3) lakes. Such a comparison must be viewed with caution due to the natural variability of age structure between lakes. However a change to a similar age structure patten was • observed, coincident with declining pH, in George Lake (Harvey et al. • 1981). Kelso et al. (1982) have recently reported on a survey of 75 | headwater lakes varying in size from 1.6 to 120 ha in the Algoma area I I ------- I I I I t I I I i i i I i i i i i i i 3-131 W il «*- o i_ d> E 3 Z 100 RED CHALK LAKE 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 Age in Years Figure 3-45. Age composition of the white sucker population of three lakes in the Muskoka-Haliburton Region of Ontario (Harvey 1980). ------- 3-132 of central Ontario. Most were found to be poorly buffered with 65% of the lakes having alkalinities less than 200 yeq/L, 26% less than 40 yeq/L and 8% less than or equal to 0 yeq/L. In 55 of the lakes sulphate concentrations were found to exceed bicarbonate. None of the eight lakes with alkalinity values less than zero were found to contain any sport fish, including brook trout, the primary sport fish in this area of the Province. Minns (1981) analyzed the Aquatic Habitat Inventory data base of the Ontario Ministry of Natural Resources (OMNR). This data base contains conductivity, pH, lake morphometry and fish species presence information for 6,393 Ontario lakes (as of September 1980, the time of analysis). The lakes contained in the data base were assumed to be representative of lakes in the area surveyed. Analysis of the data base for the presence of dystropic lakes indicated that very few were included and therefore their affect on the analysis would be minimal. Using relationships beween alkalinity, conductivity and pH, lakes were classified into categories in terms of their acidification status and the results were extrapolated to areas represented by the sample. Minns estimated that 1,200 lakes in the province are too acidic to sustain fish communities (lake pH less than 4.7) and approximately 3,500 other lakes are approaching that condition (lake pH 4.7-5.3). Most of these lakes are situated in watersheds in the region of Sudbury and are small (i.e., less than 10 hectares). Minns suggested that esocid and most percid communities are not currently at risk whereas the brook trout, lake trout and bass communities represent the most vulnerable resources. 3.7.10.3 Quebec Fisheries investigations in the province of Quebec have concentrated in the Laurentian Park. To determine the relationship between the level of acidity and fish productivity in these lakes, the Quebec Ministry of the Environment sampled 158 lakes in the area. Water samples were collected through the ice, three weeks after the beginning of snowmelt in March 1981. Most of the lakes sampled were headwater lakes ranging in size from 10 to 25 hectares, with brook trout populations. Richard (1982) classified the lakes into three groups using a multivariate analysis. The variables accounting for the greatest between group variance are described following: Group 1 2 3 Number of Lakes 23 65 57 pH 5.2 5.9 6.4 Alkalinity (WS/L) 8.5 45.6 130.6 HC03-/SO 0.1 0.6 1.8 Total ,2- Aluminum (yg/L) 230.0 143.8 71.2 I I 1 I 1 1 1 1 i i i I i i i i i i i ------- I I I I I I I I I I I I t I I I I I I 3-133 In each group of lakes the average annual yield, the angling effort and the mean weight of the fish caught (from detailed daily records prepared by all fishermen) were compared (Richard 1982). Only those lakes with nine years of continuous exploitation were included in the analysis (12 lakes in Group 1, 30 lakes in Group 2, 36 lakes in Group 3). During the last four years of study (1978-81) the mean yield from Group 1 lakes (the most acidic) was not statistically different from that of Groups 2 and 3. This conclusion was corrobor- ated by examination of data from 34 additional lakes that had been fished continuously for from four to six years (Richard 1982). Fisheries management practices within the Laurentian Park provide for closure to fishing when angling success was reduced as defined by a lower mean weight or lower number of fish caught, or when spawning habitat was disrupted. Forty-four lakes were not included in the analysis as they had been closed to fishing for one or more years preceding 1981. The 44 lakes which were closed to fishing included 43.5% of the most acidic lakes (Group 1) as compared with 36.9% of Group 2 lakes and 17.5% of the Group 3 lakes. This comparison suggests lower productivity in lakes in Groups 1 and 2, the more acidic and acid-stressed lakes, than in Group 3 lakes. Although the frequency of fisheries management problems was higher in the more acidic and acid-stressed lakes, one cannot assume a direct cause-and-effeet relationship with low pH, but only a general association between fish productivity, pH and the oligotrophic conditions of these waters. 3.7.10.4 Nova Scotia There are 37 rivers flowing through Nova Scotia for which there are records to verify that they are (or once were) Atlantic salmon rivers (Farmer et al. 1980). For 27 of these rivers, almost complete angling catch records are available (annual reports from federal fishery officers) from 1936. Of these 27 rivers, 5 have undergone major salmon stock alterations since 1936 by dam construction/ removal, and/or extensive hatchery stocking. Watt et al. (1983) examined the effect of low pH on angling by dividing the remaining 22 rivers into two groups, based on 1980 pH levels. For the 12 rivers presently at pH values greater than 5.0, only one shows a statistic- ally significant decline in angling success since 1936, another shows a significant increase, and 10 show no significant trend. Of the 10 rivers with pH values less than 5.0, 9 show significant declines, and one shows no significant trend. To combine the data so as to form averages for the two groups, the records were first normalized by expressing each river's angling catch as a percentage of the average catch in that river during the first five years of record (1936-40). These percentages were then summed and averaged for each of the two pH groups. The results (Figure 3-46) reveal virtually identical angling catches in the two ------- 200 O) £100 CO o> 80 0 O> co CD c 0} o CD D. CO CO CO o •*— CO o O> ) c 60 •S 40 20 3-134 I I I •£ 10 8 Mean for 12 rivers with pH>5.0(1980) Mean for 10 rivers with pH£ 5.0 (1980) j_ _L _L JL 1935 1940 1945 1950 1955 1960 1965 1970 Year Figure 3-46. Atlantic salmon angling data normalized to facilitate the comparison between high and low pH rivers. Each river's catch was expressed as a percentage of the mean catch in 1936-40 so as to give all rivers equal weighting, and the two groups were then averaged by year (Watt et al. 1982). 1975 19f I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-135 groups until the early 1950s; after which the angling catches in rivers of pH less than 5.0 declined, while the catch in rivers of pH more than 5 continued to show no significant trend with time. Factors other than pH (e.g., stream flows and sea survivals) also affect the angling success. Variation from these other factors should, however, affect both groups similarly. The apparent reason for the difference in angling success between the two groups of rivers is a difference in pH since the 1950s. Historical water chemistry data are available for some of these affected rivers from surveys performed in 1954 and 1955 (Thomas 1960). In the past 25 years, the pH of the Tusket River has decreased from an annual range of 4.9-6.1 to 4.6-4.9; the Roseway from a range of 4.4-6.4 to 4.3-4.5; the Jordan River from about 5.1 to a range of 4.4-4.6; the Medway River from a range of 5.5-6.5 to 5.1-5.8; and the Clyde River has decreased from 5.0 to 4.6. Alkalinity values were below zero in the Tusket, Clyde, Roseway and Jordan rivers in 1979-80 (Watt et al. 1983), but was greater than zero during Thomas' study 25 years earlier. Although Thomas (1960) sampled some of these rivers only once, his data on river pH suggest that salmon reproduction in a few rivers may have been adversely affected due to acidity by the early 1950s, consistent with the catch data presented in Figure 3-47. Within Nova Scotia, the pH of surface waters xs well correlated with geology (Watt 1981). Seasonal variation in the pH of those rivers is about 0.5 units, with the annual minimum occurring in mid-winter, and a maximum in late summer. At present there are seven rivers with pH less than 4.7 that previously had salmon but now have no salmon or trout reproduction; 11 rivers are in the pH range 4.7-5.0, where some salmon mortality may be occurring; and seven rivers are in the pH range 5.1-5.4, which is considered borderline for Atlantic salmon (Figure 3-48). Those rivers represent 2% of the total Canadian habitat potential for Atlantic salmon, and 30% in Nova Scotia. The numbers of salmon angled, recorded by Canadian federal fisheries officers since 1936 in six Nova Scotia rivers are illustrated in Figure 3-47. The Clyde River with a mean annual pH of 4.6 in 1980-81 has produced no angled salmon since 1969. Electroseining in the last several years also produced no salmon. The Ingram River with a mean annual pH of 5.0 (range 4.8-5.8) apparently still has a small reproducing population; it was at one time a good producer of Atlantic salmon. Federal fisheries officials consider this river to be in imminent danger of losing its remaining stock. This river has been identified by Canada Department of Fisheries and Oceans personnel as a candidate for liming in order to create a refuge for maintaining the gene pool of this stock. One of the Nova Scotia rivers "threatened" by pH declines, the Mersey, contains an Atlantic salmon hatchery. The Mersey watershed has poorly developed soils, and its underlying geology is Devonian granite. The mean total alkalinity of samples collected from the ------- 3-136 MIDDLE RIVER Q UJ _l O z i r ~~ii i 1935 40 45 50 55 60 65 70 75 80 Year z O < CO 300- 250- 200- 150- 100- 50- TANGIER RIVER DC til m 1935 40 45 50 55 60 65 70 75 80 Year 200-i 150- 10O- 50- SALMON RIVER 1935 40 45 50 55 60 65 70 75 80 Year INGRAM RIVER "IIIII 193540 45 50 55 60 65 70 75 80 Year EAST RIVER 40 45 50 55 60 65 70 75 80 1935 Year CLYDE RIVER 193540 45 50 55 60 65 70 Year 75 80 Figure 3-47. Angling records for six Nova Scotia Atlantic coast rivers with mean annual pHs (1980) <5.0 (Watt et al. 1983). ------- I I 3-137 I I I I I I I I I I I I I pH <4.7 (no natural salmon reproduction) pH range 4.7 - 5.0 (some mortalities likely) pH range 5.1 - 5.4 (fisheries threatened) pH > 5.4 (no immediate acidification threat) Figure 3-48. The Altantic salmon rivers of the Maritimes have been divided into 4 pH (estimated mean annual) categories based on significance to salmon reproduction. Present evidence indicates that salmon cannot reproduce at pHs below 4.7. Juvenile mortalities of 30% or more are expected in the pH range 4.0-4.7. Rivers in pH range 5.1-5.4 are considered threatened. Above pH 5.4 there is no immediate acidification concern with regard to Atlantic Salmon (Watt 1981). ------- 3-138 I river in 1978-79 was less than 10 yeq/L, while mean pH was 5.2 (range • of 4.9-5.4) (Farmer et al. 1980). In 1954-56 the river had a mean pH of 5.8, with a range of 5.4-6.6, and a mean total alkalinity of — 48 yeq/L CaC03, with a range from 20 to 88 (Thomas 1960). Mean • sulphate values have been estimated to have increased from 76 yeq/L in 1954-55 to 158 yeq/L in 1978-79. During the period 1975-78, mortality of Atlantic salmon parr reared at the Mersey hatchery I typically occurred during the third and fourth weeks after first • feeding. This higher-than-expected mortality was attributed to increased acidity in spring river water supplying the hatchery. In • 1979, by treating the water with CaC03, the salmon fry mortality was • reduced from 30% to 3% (Farmer et al. 1980). In 1980, the water was again treated and produced the same increase in survival of parr. ^ Farmer et al. (1980) noted that, even though all rivers classified as ™ presently unsuitable for salmon historically sustained Atlantic salmon populations, these rivers are all also naturally somewhat • acidic and historically had relatively low fish production. Of the | 20 readings of apparent water colour (rel. units) (an indicator of the presence of organic acids) presented for the 7 rivers classified m "unsuitable" by Farmer et al. (1980), 16 were 100. For "threatened" • rivers, only one of 21 readings was 100; the remaining readings averaged 69. For rivers classified neither "unsuitable" nor "threatened," and with pH readings above 5.5, the mean measure of • apparent colour was 44. High degrees of colour are largely attribut- • able to humates from peat deposits and bogs common in this area. Inputs from these materials probably contribute to the low pH levels • in "unsuitable" and "threatened" rivers. Historical records of pH in | these rivers do, however, indicate that acidity has increased in recent years. Watt et al. (1983) concluded "the Atlantic coast am rivers of Nova Scotia have become more acidic over the past 27 years • in response to increased acid loading in the precipitation." This increase in acidity has been clearly correlated with declines in populations of Atlantic salmon in the same rivers. • 3.7.10.5 Scandinavia 1 Hendrey and Wright (1976) reported that "acid precipitation has devastated the salmonid fish in southern Norway." Massive fish kills tm of adult salmon and trout have been reported in their river systems, • usually occurring during the spring snowmelt or after heavy autumn rains. An intensive survey of 50 lakes in southern Sweden showed that inland freshwater species are also threatened. The decreases in • pH have resulted in the elimination of Atlantic salmon from many • Norwegian rivers in the past 20 years. Scandinavian scientists have concluded that, directly or indirectly, the principal cause of the • fish losses is acidification of the waters, due to acidic deposition. || Portions of Canada's Atlantic salmon fishery appear to have declined as a result of acidification as has been experienced in Norway and . Sweden. • I I ------- I I I I I I I I I I I I I I I I I I I 3-139 3.7.11 Response to Artificial Acidification While we know that the end-product of acidification includes the disappearance of important fisheries, many of the early changes which occur in acidified ecosystems are relatively unstudied. Furthermore, it is not known whether declines in fish stocks are due singly or in combination to the toxic effects of hydrogen ion, to hydrogen ion and aluminum or other metals synergisms, to food-chain effects resulting from elimination of critical species of animals and plants or disruption of nutrient cycles. A whole-lake acidification experiment was done in Lake 223 in the Experimental Lakes Area, Ontario, in order to examine some of these possibilities. The pH of the lake was progressively lowered from a natural value of 6.5 to 6.9 (x = 6.7) to an average value of 5.1 by additions of sulphuric acid between 1976 and 1981. Detailed monitoring of chemical, physical and biological changes, as well as physiological and ecotoxicological studies, were done throughout this period. Earlier biological results were summarized by Schindler et al. (1980), Schindler (1980), Malley and Chang (1981), and Schindler and Turner (1982). Biological changes in the lake as it was artificially acidified and the pH thresholds at which these changes occurred are summarized in Table 3-23. The first changes which could have adversely affected lake trout and white sucker populations occurred in 1978-79, when populations of two species which are the usual prey of trout, fathead minnow (Pimephales promelas) and oppossum shrimp (Mysis relicta), collapsed. Despite these changes, no effects were detected in trout populations. A succession of strong white sucker year-classes in 1978-80 and greatly increased abundance of pearl dace were adequate food alternatives for trout. Apparently, the pearl dace partially occupied the vacated fathead minnow niche, while the primary food source of white suckers, benthic dipterans, increased in abundance (Davies pers. comm.). In addition, the appearance of excessive growths of Mougeotia in the littoral beaches probably provided excellent nursery areas for sucker fry, but increased water transpar- ency (Schindler 1980) perhaps made prey capture easier for trout. Even though many changes have occurred in lower trophic levels, juvenile and adult white sucker and lake trout populations have shown little indication of stress, except for recruitment failures in the very recent years of acidification, at pH values of 5.35 and below. Up to 1981, populations of both species increased, and their growth rates have remained high. Relative condition (a quantitative measure of fish fatness) has decreased progressively for trout from 1977 to 1980, and for white suckers from 1978 to 1980, but this would be expected due to the increased abundance of both species over the same time period. The relatively swift collapse of the fathead minnow population is due to two factors. Firstly, a recruitment (year-class) failure occurred ------- 3-140 CO § t Q Q ^ Q O 0 or T 3 >- 00 o 1- CJ — Lu Q O 12 y 3) Q CM CM LU -1 Z a LU co § CO u- 0 LU LU LU LU ~ _l E < £ 0 0 — 0 8 C — 0) CD CL O CO >- T3 or c § JC s: o Z3 CO CO v^ CM KN. LU CD - -D T) > >. 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CD O 0. 4- c *- •»- ID <0 4- -0 T3 U (D CD 3 in in •O ID 10 O CO CD L. l_ L. a. o o CD CD C or o — 10 10 4- 4- -g -S •D "P 0 CD JZ JZ in in ^ ^ JO JO 3 3 CL CL c c 3 3 in in ~ « ^ Sf ^ N^ L. 4- CD 8J^ o 1_ 3 4- in (D CO Ji 4- 10 — — C. **- : * 3 £ — 3 10 — •*- 10 CO > 4- t 1 3 4- "§ 3 L, L. or or *~ in CO ON I I I I i I I I I 1 I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-141 in 1978 (pH -5.8). This agrees well with the results of Mount (1973), who found that impaired reproduction of the same species occurred at this pH in laboratory studies done at a variety of pH values. Secondly, even under preacidification conditions, this species had a very short life span of three years in Lake 223. Even under natural conditions, during the second and third years of life an extremely high natural mortality rate occurred, over 50% per year (Mills pers. comm.), presumably caused in large part by trout predation. Very few individuals remained after the second year of life. Therefore, the failure of one year class in 1978 would leave few spawning adults (age 2 and 3) the following year. Population recovery was, therefore, almost impossible. The combination of successive year class failures in 1978 and 1979 assured the rapid disappearance of this species from Lake 223. The thresholds observed for disappearance of key species and appearance of others in Lake 223 agree well with observations made in other acidified lakes. For example, Mysis in Lake 223 disappeared in the same pH range as benthic crustaceans with similar food habits disappeared in Scandinavian lakes (0kland and 0kland 1980). Mougeotia epidemics in Lake 223 began at almost the same pH values as in Swedish waters (Hultberg pers. comm.). Recruitment failures in lake trout and white sucker began in the same pH range that year classes began to be absent in lakes near Sudbury and in Scandinavia (Harvey 1980; Muniz and Leivestad 1980; Raines 1981b,c). The Lake 223 results also demonstrate the danger of assessing biological damage from acidification solely on the basis of game fish populations. Major alterations to fish habitats and prey species occurred several tenths of a pH unit above where initial damage to lake trout was detectable, even with an extremely intensive study of the trout population. The predation habits of lake trout appeared to allow them to easily switch to pearl dace after the disappearance of the fathead minnows which had been their normal prey. In summary, the Lake 223 experiment clearly shows that alterations to aquatic food chains begin at pH values slightly below 6.0. The remarkable agreement between these whole lake experiments and observational studies in Scandinavia and eastern North America provides strong evidence that the observed declines in fisheries are caused by acidification and not by other ecological stresses. 3.7.12 Effects of Acidic Deposition on Birds and Mammals While birds and mammals are not affected directly by acidic depo- sition they are vulnerable to changes in their habitat caused by acidification, particularly to changes affecting the availability and quality of their food. Although adults may continue to find sufficient food in areas adjacent to their traditional nesting or ------- 3-142 I I breeding sites, they may be unable to obtain sufficient food to raise young. In Scandinavia there have already been reports of such effects on aquatic bird populations. Aimer et al. (1978) reported 4 that, "fish-eating birds, such as mergansers and loons, have been • forced to migrate from several acidic lakes, with decreasing fish stocks, to new lakes with ample food supply. In this way, many _ territories will become vacant and this will lead to decreasing • stocks." While the extent of the problem has not yet been documented ™ in Sweden, Nilsson and Nilsson (1978) found a positive correlation between pH and "water" bird species richness. "Water" birds were defined as those species dependent upon open water, and included a loon, and several species of waterfowl and gulls. From the results of this study it was suggested that a reduction in young fish, a very «| important food source for aquatic birds, may lead to low reproductive • success and local extinction in some bird species (Nilsson and Nilsson 1978). Eriksson et al. (1980) also proposed that reduced reproduction of fish in acidified lakes may decrease the availability • of fish of the size classes appropriate to young diving water birds. ™ I I Losses of other aquatic organisms such as clams, snails, and amphibians have been documented in acidified lakes and ponds (Section 3.7.6; Hagstrom 1977; Hall and Likens 1980; J. 0kland 1980; K.A. 0kland 1980). While wildlife are largely opportunistic feeders, « reductions of these organisms could affect the food availability for • many wildlife groups such as waterfowl and semi- aquatic mammals. The effects of changes in food and habitat will be difficult to witness in the short term but, in time, breeding densities may • decline and eventually productivity could fall in response to reduced 9 food availability. The diet of the common loon (Gavia immer) is approximately 80 percent ( fish, the remainder being made up of crustaceans, molluscs, aquatic insects, and leeches (Barr 1973). Because the food requirements of « loons while rearing young are high and many of their food organisms • are quite sensitive to acidification, the nesting densities of this * species may be reduced. In eastern Canada, the common loon nests on lakes throughout the susceptible terrain of the Precambrian Shield • (Godfrey 1966). In central Ontario and Quebec as well as in the 0 Adirondack Mountains of the northeastern U.S., a number of lakes have already been reported as devoid of fish as a result of acid loading • (Beamish 1976; Schofield 1976a). Studies in New York indicate that | loon productivity has remained high but nesting densities have declined in the Adirondack region (Trivelpiece et al. 1979). To • date, however, changes in loon populations in the Adirondacks have • been interpreted only with respect to human disturbance; the probable * role of food depletion has not been investigated. In Quebec, fish-eating birds were found more often on the nonacid lakes • (DesGranges and Houde 1981). The common merganser (Mergus merganser) | and the kingfisher (Megaceryle alcyon) were observed only on those lakes where the summer pH is higher than 5.6. In the vicinity of m Schefferville, Quebec, important differences in numbers and composi- • tion of lake-dwelling bird communities were found: a third as many I I ------- I I I I I I I I I I I I I I I I I I I 3-143 species and a quarter of the total number of aquatic birds were observed on lakes with pH less than 4.5 compared to lakes with pH greater than 6.0 (DesGranges and Houde 1981). The situation is less clear, however, for lakes of pH 4.5-5.5. It has been suggested that the biomass of some forms of benthic invertebrates increases with low to moderate inputs of acid because there are fewer fish predators (Henrikson and Oscarson 1978; Eriksson 1979; Henrikson et al. 1980). This may explain the larger number of invertebrate-feeding ducks which are found on moderately acid lakes in southern Quebec (DesGranges and Houde 1981) and in central Ontario (McNicol and Ross 1982). Insectivorous birds such as swallows, flycatchers, and kingbirds may be affected by lake acidity since this group of birds feed on emerging insects and it is during the emergence that many insects are most sensitive to high acid levels (Bell 1971). Because a number of species of aquatic insects emerge in early spring during the peak of acid input to lakes and ponds they are particularly vulnerable to the effects of acid loading. It is also in early spring that the birds have higher food requirements in nesting and raising young. In southern Quebec, the tree swallow (Iridoprocne bicolor) was more common during the breeding season in the vicinity of lakes of pH >6.0 while in northern Quebec this species was not observed in the area of lakes of pH <4.5 (DesGranges and Houde 1981). This was also the finding from the studies of insectivorous birds in the Killarney area of Ontario (Blancher 1982). The presence or absence of these birds will largely be determined by the biota of the nearby lakes. Effects of acidification on lower life forms such as microorganisms, essential to decomposition and nutrient cycling have been found (Hendrey et al. 1976; Leivestad et al. 1976). A loss in productivity at the base of the food chain due to decreased nutrient availability could result in progressively larger reductions at each succeeding trophic level. The implications for wildlife at the top of the chain are a critical loss in biological production and severely reduced carrying capacity of their habitat (Clark and Fischer 1981). Increased solubility and mobility of metals from sediments have been reported as a result of acidification (Schindler et al. 1980). The higher concentrations of metals produced in lake waters have important implications for biological organisms as described in previous sections. Studies by Nyholm and Myhrberg (1977) and Nyholm (1981) have implicated aluminum in the impaired breeding of four species of passerines. Reductions in the reproductive success of these birds was highly correlated with the distance of their nests from acid-stressed lakes in Swedish Lapland. Breeding impairment was manifested as abnormal egg formation producing thin and porous shells. In addition, clutch size and hatching success of the "affected" birds were reduced and egg weights were lower in the birds closer to the acid-stressed lakes. The link between the acidified lakes and the breeding impairment has been related to the high ------- 3-144 I I aluminum content of the limnic insects upon which the birds feed (Nyholm 1981). Birds feeding closest to the stressed lakes have the highest proportion of contaminated insects in their diets (Nyholm fl pers. comm.; Eriksson et al. 1980). Similar findings of decreased | egg size and weight were found for the eastern kingbird (Tyrannus tyrannus) in the Killarney area of central Ontario (Blancher 1982). n Although severe abnormalities in shell formation were not evident in • the eggs examined in this preliminary study, egg porosity as measured by the rate of water loss over the incubation period was negatively correlated with pH. • Elevated mercury levels have been found in fish in lakes with low pH in central Ontario (Suns et al. 1980). In the Bohuslan area of • Sweden, elevated levels of mercury were found in eggs of goldeneye | (Bucephala clangula) (Eriksson et al. 1980b). Raccoons (Procyon lotor) from the Muskoka area of Ontario support liver mercury levels M of 4.5 ppm, a concentration five times greater than specimens from an • area with nonacidified waters (Wren et al. 1980). Because neither of these areas receives point source inputs of mercury, the sources are believed to be leached from the watershed by acids or mobilized from • sediments. Methylation of mercury has been related to the process of • acidification and the formation of methyl mercury, a stable and soluble form which readily bioaccumulates, is believed to be favoured at low pH (Fagerstrom and Jernelov 1972). I Results of a preliminary study of metal accumulation in the tissues ^ of moose (Alces alces) have established an age dependent increase • in cadmium for tissues collected from 38 moose and 56 roe deer in ™ Sweden (Frank et al. 1981; Mattson et al. 1981). Aimer et al. (1978) reported a 10-fold increase in levels of cadmium in acidified lakes I on the Swedish west coast compared with those in nonacidified lakes • in the same region. Cadmium may be accumulated in large concentra- tions by some terrestrial and aquatic plants (Anderson and Nilsson M 1974; Hutchinson and Czyrska 1975), and therefore, metal contamina- • tion of wildlife feeding on these plants may be an indirect effect of acidic deposition. ^ A summary of potential effects on selected species of birds and ™ mammals dependent upon the aquatic ecosystem for their food and habitat is presented in Table 3-24. This summary is based solely • on feeding habits as research on the impacts of acidification on • vegetation structure and productivity relating to wildlife habitat is at a preliminary stage. • 3.8 CONCERNS FOR IRREVERSIBLE EFFECTS _ 3.8.1 Loss of Genetically Unique Fish Stocks " Loss of fish populations with specific gene characteristics from • lakes and rivers may be an irreversible process. Over several | thousand generations, most species appear to have evolved discrete I I ------- I I I I I I I I I I I I I I I I I I I 3-145 o fag 5 Q CO CO W ta o 0 Pi C= O O CO CO Cd JV1 (-VJ s o P O o o O fa Cu M Z EH M H CO PQ J 0 a H 3 rH 53 O P Cd Z P CJ H rH > CJ ^ ^ • CM 1 CO Cd jj S H T3 O O fa CO cu 0) O 4-1 i-l cfl 3 (3 O )-l CO 0) 01 4-1 Pi rH CO cu g o CO X-s c o X-N CJ CO rH 3 cd •H "O 0) O rH X-N x-v 0) rl M CO 43 01 CU 3 CJ C 4-1 cd cd 3 cu 01 bO 6 cd 13 cu •H -H i-l 53 rH < ^ CO cd ^ •H 43 H > (30) Cd (3 O 43 U O I-i CO N-/ -H 0) -H 13 CC M-i (3 C bO O cfl CU d O PH 3 -rl nJ ^ rH ^ PQ C >s 13 O 01 4-1 0) S )H Cd 4-1 B P. 0) rH O CO i-l 01 CJ O O PQ co oi rl d 01 O > N S rH cd M CO O CU 4-1 *s. 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The basic unit of a stock is the gene pool, which — is composed of a naturally sustained, genetically variable group of • individuals, adapted through evolution to specific lake conditions. * Surface water acidification is a stress that may reduce genetic variability in populations of native fishes in sensitive areas. As • an example, Beamish and Harvey (1972) documented the loss of gene | pools of fish in acidified lakes in Ontario. The Ontario Ministry of Natural Resources has attributed the extinction of lake trout • (Salvelinus namaycush) in 27 lakes in the Sudbury-Temagami area to • acidification (Olver pers. comm.). A naturally evolved complex of stocks appears essential to utilize • fully the productive capacity of waters. Therefore, it is important • to recognize and preserve stocks (Haines 1981c; Loftus 1976; Ryman and Stahl 1981). I Loss of discrete stocks may inhibit effective re-establishment of naturally reproducing populations in waters undergoing rehabilitation m and affect future opportunities for fisheries management. • I Evidence seems to be conflicting as to whether the geochemical alteration of watersheds due to acidic input should be viewed as • irreversible, and, if so, on what scale. Irreversibility can be | viewed most strictly as a failure to recover over geologic time; but, for natural resource systems, an incomplete recovery to a prestressed _ or undamaged state over a few decades, for all practical purposes, • may be regarded as irreversible. Although irreversible reduction in acid neutralizing capacity of K lakes and watersheds is one of the potential effects of acidic • deposition, our present information base is insufficient to determine its probability in impacted areas. • 3.8.3 Soil Cation and Nutrient Depletion _ The loss of soil cations, particularly Ca^+ and Mg2+, which can lead to decreases in soil fertility (Overrein et al. 1980), is another potentially irreversible consequence of watershed titration. • However, the extent to which these cation losses represent a m significant depletion of total available material is unknown. I The previous sections have discussed chemical and biological changes • observed in some surface water systems, including pH depression and I I ------- I I I I I I I I I I I I I I I I I I I 3-147 associated effects over long-term, annual, seasonal and event-related time series. Most of the results are consistent with the explanation that they result from acidity associated with the 804 2~ and NC>3~ ions originating from atmospheric deposition. This section will consider the significance of these levels of chemical alterations, with a comparision of the annual deposition that could be associated with acidification of the most sensitive streams and lakes. This analysis requires consideration, not only of trends in surface water and precipitation pH and sulphate concentration, but also of the frequency and severity of brief periods during which much of the response to the total acidic loading rate from runoff events is expressed. Emphasis has been placed on deriving as much information as possible from comparisons of observed water quality and biological effects in areas of varying deposition. These empirical observations integrate many "unknowns" regarding soil water interactions which are impli- citly taken into account by empirical comparisons. Loading rates estimated from conceptual models of aquatic systems are compared to the empirical observations. Such empirical approaches to support environmental management are common. For example, flood structure designs can be based on empirical relationships between discharge, precipitation and physical characteristics of the watershed (Chow 1964). Vollenweider and Dillon (1974) used an empirical modeling approach to set phosphorus loading criteria for eutrophication control in lakes and reservoirs, and these have proven effective. The following are the principal findings presented in previous sections important in evaluating aquatic effects related to measured acidic deposition: 1. Precipitation over most of eastern North America has hydrogen ion concentrations up to 100 times those expected for distilled water in equilibrium with atmospheric carbon dioxide. 2. Large quantities of sulphate and nitrate ions are deposited with ff1" ions in precipitation in eastern North America. 3. Lakes in eastern North America with low alkalinities are receiving elevated acid loadings. Such lakes, and their associated streams, may suffer low pH and elevated metal concentrations for short periods of time, particularly during snowmelt and other periods of heavy runoff. 4. Stressed fish populations have been observed in lakes that experience short-term low pH and elevated metal concentrations. Mortalities of adult fish have been observed in one study lake experiencing these conditions. 5. There are numerous examples of streams and lakes in Canada and the United States that have experienced and are probably now experiencing depletion of alkalinity. Fish populations that ------- 3-148 I I I survive short-term low pH conditions, will eventually be lost if alkalinity is depleted and pH values fall below critical levels causing successive reproductive failure. Long-term acidifica- m tion has caused losses of fish populations in some lakes and • streams. I 3.9.1 The Relative Significance of Sulphur and Nitrogen Deposition • to Acidification of Surface Waters Results presented in the previous sections have shown that four major | ions of concern in acidic precipitation, (H+, NH^"1", N(>3~ and SO^- have some potential for altering lake and stream water g| acidity. Soil and plant interactions with nitrate ions allow nitric • acid to be largely assimilated by the terrestrial portion of the watershed, except during periods of heavy runoff (Section 3.2.2) (McLean 1981). In contrast, in many regions with poorly developed • soils, that are limited in ability to neutralize acid, biological 9 uptake of sulphate is small in comparison to the mass balance of sulphur (Harvey et al. 1981). Christophersen and Wright (1980) reported that the sulphur export from a watershed in Norway was essentially the same as the total input over the period November 1971 to October 1978. In a number of areas studied, where there exist no _ significant terrestrial sources or sinks of sulphur, SO^" is a • conservative ion whose export to surface waters is directly related to deposition in precipitation. There are additional aspects to the issue of the dominant anion • associated with the acidification of surface waters. These include: 1) the relative magnitude of S0^~ and NC>3~ in the rain and f snow inputs, their variation during the year, and long-term trends; M 2) the relative magnitude of the biological interactions of * both anions in watersheds, as they are affected by biological activity at different seasons and by changes in B biomass over long periods; • 3) the production of alkalinity in terrestrial and aquatic • systems when NC>3 is assimilated by plants; and • 4) the contact time of precipitation inputs with the water- ^ shed. • Data presented in map form in Section 2 and other data presented by Galloway et al. (1980g), McLean (1981) and by Harvey et al. (1981) • indicate that acidic sulphur inputs exceed acidic nitrogen inputs V over eastern North America on an annual basis. The net yield of these anions to streams and lakes is predominantly SO^" on an annual • basis (Harvey et al. 1981). Because nitrate reaches surface waters • in small amounts relative to its loadings on an annual basis and does I I ------- I I I I I I I I I I I I I I I I I I I 3-149 not accumulate in surface waters, its influence on long-term surface water acidification is less than that of sulphate. Further evidence that nitrate deposition is not principally respons- ible for long term surface water acidification is given in Table 3-25. Data for 21 headwater streams in the Muskoka-Haliburton area of Ontario with a range of mean annual pH values from 4.08 to 6.18 show that as acidity increases, the relative importance of NOg declines. The acid (H+) concentration exceeds the N(>3~ concentration on a chemical equivalents basis for annual pH values of 5.5 or less, so that lower pH values cannot be explained by the presence of nitric acid. The E+/SO^~ ratios are also given for the same streams (Table 3-25). At lower pH values, H+/S042" ratios increase. The ratio is always less than one which indicates that the acid concentration can be explained by the presence of sulphuric acid. The SO^~/fiO^~ ratios range from 14 to 337 with a median value of 170, demonstrating the dominance of 8642- over NOg" in surface waters in the Muskoka-Haliburton region (Jeffries et al. 1979; Scheider et al. 1979c; and ongoing studies by Ontario Ministry of the Environment). Nitrate may be important on an episodic basis by adding to the pH depression caused by sulphate. At Sagamore Lake, New York, nitrate concentrations in the lake outflow increased during spring pH depression, while sulphate concentrations did not increase (Galloway et al. 1980g). Sulphate concentrations still exceeded nitrate concentrations on an equivalent basis, even during spring runoff. Uptake of nitrate ions by algae and aquatic plants results in the production of alkalinity in surface waters (Goldman and Brewer 1980). This has been shown to occur in one of the study lakes at Muskoka- Haliburton. Reported increases in lake pH from 5.1 to 6.6 over the summer were associated with decreases in nitrate concentrations by photosynthetic processes, and this was given as the explanation for the pH increases (Harvey et al. 1981). The evidence available, and the published interpretations of that evidence (Harvey et al. 1981; Overrein et al. 1980), lead to the conclusion that, for surface water systems, increases in acidity are the result of dilute solutions of strong acids reaching these waters. Further, Harvey et al. (1981) following extensive analysis of Canadian data and Overrein et al. (1980) following extensive research in Scandinavia conclude that most of the acidity is due to the changes observed in 804^" concentration attributable to sulphate and sulphuric acid deposition (Harvey et al. 1981; Overrein et al. 1980). Both sulphuric and nitric acid contribute acidity to surface waters during periods associated with pH depressions and fish stress. However, there is no strong evidence at present for anticipating any appreciable reduction in long-term lake or stream acidification from a reduction in nitrate inputs. In contrast, it is important to note there is a strong correlation between between sulphate deposition and surface water concentrations to suggest that a reduction in sulphate ------- TABLE 3-25. 3-150 MEAN AND RANGE OF pH VALUES, MEAN H+/N03 , H+/S042 AND S042~/N03~ RATIOS (calculated as ueq/L) FOR 21 HEADWATER STREAMS IN MUSKOKA-HALIBURTON, ONTARIO 1976- 1980 [Data is from an ongoing study, methods and study area as described in Jeffries et al. (1979) and Scheider et al. (1979b)] Stream Dickie 11 Red Chalk 2 Dickie 5 Dickie 6 Dickie 10 Chub 2 Dickie 8 Harp 6A Harp 5 Chub 1 Harp 3 Harp 6 Red Chalk 1 Red Chalk 3 Harp 3A Red Chalk 4 Jerry 3 Jerry 4 Harp 4 Blue Chalk 1 Jerry 1 Mean PH 4.08 4.30 4.34 4.35 4.59 4.82 5.03 5.19 5.34 5.41 5.64 5.77 5.81 5.95 5.95 5.96 5.98 6.07 6.08 6.16 6.18 Range pH 3.53-5.61 3.68-4.81 3.71-4.76 3.74-5.05 3.92-5.10 4.12-6.08 4.04-5.87 4.34-6.39 4.66-6.60 4.48-6.61 4.89-6.39 5.20-6.90 5.19-6.69 5.17-6.65 5.30-7.30 5.28-6.71 5.27-6.67 5.49-6.55 5.29-6.90 5.71-6.62 5.58-6.74 H+/N03 (yeq/L) 93.60 60.00 58.30 60.20 25.90 23.90 12.60 9.57 2.49 5.49 1.05 0.83 1.74 0.21 0.29 0.24 0.51 0.46 0.15 0.67 0.04 H+/SO|~ (yeq/L) 0.457 0.188 0.318 0.297 0.119 0.071 0.049 0.028 0.017 0.019 0.009 0.007 0.009 0.006 0.004 0.006 0.004 0.003 0.003 0.003 0.003 SO|~/N03 (ueq/L) 245 265 233 247 170 236 284 337 145 232 156 130 174 34 100 37 134 118 57 198 14 I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-151 loading to watersheds would reduce the sulphate concentrations and associated acidification of surface waters. 3.9.2 Data and Methods for Associating Deposition Rates with Aquatic Effects* The evidence available on the effects of acidic deposition on aquatic resources indicates that present loadings rates are in excess of the ability of watersheds to reduce the acidity for some lakes in some areas. This section will explore the association between loading levels of acids or sulphates and negative effects on the aquatic environment. In the following analysis, it is implied that sulphate deposition can be used as a surrogate for the acidifying potential of precipitation. The use of sulphate in precipitation as a surrogate for the acidi- fying potential of deposition should not be interpreted to mean that wet sulphate is the only substance potentially damaging to aquatic systems. It is recognized that dry deposition of sulphate and SC>2, and wet and dry nitrates contribute to the concentrations of acids. Sulphate in precipitation is reliably measured and therefore, is used here as a surrogate for the total sulphur deposition because dry deposition cannot be measured accurately. Similarly, this surrogate does not reflect the contribution of nitrate to acidity of precipi- tation. Surface water quality alterations fall into two categories: 1) short-term pH depressions during snowmelt or heavy rains, and 2) long-term reductions in alkalinity, with corresponding low pH values in surface waters throughout the year. The length of time it takes for a lake to become acidic (alkalinity reduced to zero or less) and the rate of change of water quality are among the least well-defined aspects of the acidification phenomenon. To date, the evidence available, based on sediment cores taken from several areas (Section 6.3.4), suggests that acidification has occurred and is occurring on the scale of decades. * It is the view of the U.S. members of the Work Group that the reliability of wet sulphate deposition values is uncertain and therefore, any attempt to use them for analysis must be done with great care. Examination of the data shows that: (1) limited deposition data are available, and (2) annual variability in wet sulphate deposition values can be large. ------- 3-152 I I Before the alkalinity of a lake or stream is totally depleted, it is very likely that the system experiences short-term pH depressions during periods of high runoff. Large temporal fluctuations in pH • levels may represent a transition phase in the process of • acidification. The phenomenon of short-term pH declines is probably more common than | long-term reductions in alkalinity (in terms of numbers of lakes and rivers affected in North America). The chemistry of these events is ^ fairly well defined. The biological consequences of these events are • known to be severe in some cases, but the relationship between short-term pH depressions and effects on aquatic biota are not fully understood. • "In the second stage, the bicarbonate buffer is lost during longer periods and severe pH fluctuations occur resulting in stress, • reproductive inhibition and episodic mortalities in fish populations | (transition lakes)" (Henriksen 1980). Damage to fish and other biota as a result of short-term exposures to low pH and associated high _ metal concentrations has been demonstrated to occur in both • laboratory and field studies (Section 3.7). Thus, summertime or ™ annual pH has questionable value for determining effects on organisms of H+ or metals over a few days. The timing magnitude and duration I of short-term increases in H+, associated with spring melt and V storm events must, therefore, be included in an evaluation of critical loading rate and episodic response relationships for streams • and lakes. • In summary, the short-term acute exposure or "shock" effects — (including responses to aluminum) can take place in two to four days • of exposure, with pH decreases in the order of 0.5-1.5 units; and • these shock exposures can be expected to occur in waters with a broad range of pH above the level at which chronic effects occur. B The second category, long-term acidification, has altered a large number of lakes in North America, but the percentage of lakes and a| rivers with mean annual alkalinity of zero or less remains small. • The biological responses to long-term acidification are, however, more clearly defined and generally more severe than for short-term pH declines. • The acidity and chemical composition of aquatic environments are affected by: (1) the acid neutralizing capacity of the basin; • (2) the geologic and morphologic characteristics of the basin; and •, (3) the acidity of the precipitation. Biological processes (e.g., production and decomposition) also have an effect on acidity. Models • used to simulate the geochemical processes and aquatic ecosystem • effects are not fully developed or validated at this time. Development and application of detailed models will require detailed information on basin geology, hydrology, and biotic interactions. • These are unlikely to be available soon for widespread application. • Therefore, at present, the relationships between acidic deposition I I ------- I I I I I I I I I I I I I I I I I I I 3-153 and aquatic effects can be determined only in a general way. Some data and phenomenological models exist that relate the behaviour of lakes and streams to acid loading. These empirical observations and models are discussed below. 3.9.2.1 Empirical Observations Observed sulphate loadings and corresponding chemical and biological observations for a series of study areas in North America and Scandinavia are available. The information in this section is drawn from a number of study areas within eastern North America which are located on the Precambrian Shield or on weathering resistant bedrocks. The surface water studies have been initiated for several reasons, have started at different times and are operated by different agencies. However, each project contributes information relevant to the acidification problem by comparison of results among and within the studies themselves. In general, each project involves some highly detailed work on a small number of watersheds and surface waters and less detailed work on a larger study set. Within a given study area, the surface waters and watersheds are usually chosen to cover as wide a range of water quality and geology as is available. The study area descriptions will give some appreciation for the extent of the data base used in the empirical derivation of loading versus chemical and biological effects. SASKATCHEWAN SHIELD LAKES More than 300 lakes in Northern Saskatchewan's Shield and Fringe Shield regions have been sampled to assess the sensitivity of lakes. Deposition Annual Precipitation Annual Runoff (kg S042-/ha.yr) (m) (m) 5a 1980 0.357 .100 - .200b a CANSAP Measurement. b Fisheries and Environment Canada 1978. ------- 3-154 Observed Characteristics EXPERIMENTAL LAKES AREA, ONTARIO I I I 1. Alkalinity = -18.20 + 0.92 (Ca + Mg) (n=281,r=0.97) Liaw (1982) indicating that the bicarbonate and Ca + Mg are related by a 1:1 relationship and sulphate contributes very little to the total ion balance. • 2. pH values range from 5.56 to 8.2, 39% <7.0 (Liaw 1982). I I The Experimental Lakes Area is situated in northwestern Ontario on Precambrian shield granite. Approximately one-half the area of Canada is Precambrian Shield. Within the study area there are about 1,000 lakes, of which 46 lakes in 17 drainage basins • have been set aside solely for experimental research. The • inflows and outflow of Rawson Lake (a control lake) are calibrated as well as 14 other watersheds. The project was initiated in 1969 and is continuing a wide range of whole-lake • chemical manipulations including the acidification of lakes with • monitoring of chemical and biological parameters including fish population studies. The results from this multi-faceted project • are published in many scientific journals including two special • issues of the Canadian Journal of Fisheries and Aquatic Sciences devoted entirely to the Experimental Lakes Area (1971, ^ Volume 28, Number 2 and 1980, Volume 37, Number 3). I I Deposition Annual Annual (kg S042~/ha.yr) Fraction Time Precipitation Runoff Period (m) (m) • 9.07a bulk 1972 0.69a 0.297a 10.8a bulk 1973 0.73a 0.354a g 5.9b wet 1980 0.51b 0.223a | 0.234a 0.15d • Sum of Cations for 31 lakes 217 yeq/L _+ 25 (Standard Deviation) I a Schindler et al. 1976; see also Figure 3-16. • b Barrie and Sirois 1982. | c CANSAP measurement. " Fisheries and Environment Canada 1978. • I I ------- I I I I I I I I I I I I I I I I I I I 3-155 Observed Characteristics 1. No long term acidification or biological effects observed in ten years of study (Schindler pers. comm; Can. J. Fish. Aquat. Sci. 37(3); Can. J. Fish. Res. Board 28(2)). 2. Sulphate export from the watersheds is about equal to the measured wet deposition (Schindler et al. 1976). 3. Lake alkalinity distributions for lakes in the Rainy River district have fewer low alkalinity values than four other Precambrian Shield areas in Ontario (Dillon 1982). 4. Lake pH values for a 109 lake survey ranged from 4.8 to 7.4 and averaged 6.5 (Beamish et al. 1976). 5. Lake sulphate concentrations ranged from about one-half to about equal to the bicarbonate concentrations (Beamish et al. 1976, Dillon 1982). 6. Filamentous algae are common in July and August but do not dominate the algal population (Stockner and Armstrong 1971) ALGOMA, ONTARIO The Algoma region of Ontario is an area of 862,000 ha in northcentral Ontario. From a chemical survey of about 85 lakes, Kelso et al. (1982) report results from 75 headwater, nondystrophic lakes with watersheds undisturbed by recent logging, fire or human settlement. Sampling was done in 1979-80 and included physical parameters, lake chemistry and phytoplankton analyses on the entire lake set with benthic invertebrate, sediment and fish tissue analyses done on subsets of the 75 lakes. The Turkey Lakes Project, situated within the Algoma region, is an ongoing calibrated watershed study of 5 lakes and 20 water- sheds, plus the outlet of the entire Turkey Lakes watershed basin. Initiated in 1980, intensive chemical, hydrological and biological studies are in progress including monitoring of precipitation, air quality, forest effects, ground and soil water, stream and lake chemistry. ------- 3-156 calculated from ion concen- tration data from Kelso et al. (1982) and precipitation data from Barrie and Sirois (1982) Barrie and Sirois 1982. Fisheries and Environment Canada 1978. Observed Characteristics I I Deposition Fraction Annual Annual • (kg S042~/ha.yr) Time Precipitation Runoff Period (m) (m) • 25 APN Turkey wet 1981 0.8a 0.50b Lakes Station, _ (Barrie, pers. • comm.) 28 1976 22 1977 32 1978 23 1979 _ I I Sum of Cations for 75 lakes 285 yeq/L _+ 125 (Standard Deviation) =™====™= I I 1. pH depression in streams during spring runoff up to 2.1 pH • units with minimum values as low as 5.0 in streams with summer alkalinities less than 400 peq/L (Keller and Gale 1982). • 2. Excess sulphate runoff is elevated about five times over the remote areas of northwestern Ontario and Labrador • (Thompson and Button 1982). Sulphate export from watersheds | exceeds wet deposition indicating possible dry deposition of sulphate. • 3. Of 75 headwater lakes surveyed, six had pH values of 5.3 or less and the lowest value was 4.8 (Kelso et al. 1982). 4. Sulphate ions are the dominant anions (i.e., exceed • bicarbonate) in lakes below pHs of about 6.5 (Kelso et al. 1982). • 5. In a survey of 31 headwater lakes (1.6-110 ha), the number of lakes devoid of the 8 fish species reported in the area _ was observed to increase with decreasing alkalinity. The • relationship between the presence of fish and pH in these ™ same lakes was weaker although a greater proportion of I I ------- I I I I I I I I I I I I I I I I I I I 3-157 lakes of pH <5.5 were fishless than lakes of pH > 5.5 (Kelso et al. 1982). These observations are consistent with the hypothesis that the biota in the surveyed lakes have been adversely affected by changes in lake chemistry but do not necessarily indicate causality (Kelso et al. 1982). 6. Aluminum and lead levels in 75 headwater lakes in Algoma were elevated in lakes of lower alkalinity; mean total aluminum levels of 53 yg/L was slightly greater than aluminum levels in Muskoka-Haliburton waters (Scheider et al. 1979a) and intermediate between concentrations found in severely affected and slightly affected systems in Canada and Norway (Kelso et al. 1982). MUSKOKA-HALIBURTON ONTARIO The study area in Muskoka and Haliburton counties of southcentral Ontario encompasses an area of about 490,000 hectares within which are its 8 intensive study lakes and 32 calibrated watersheds, some of which have been calibrated since 1976. The watersheds vary in water quality and from low to high pH. Twenty other lakes have been monitored on a seasonal basis for a varying number of years. Many concurrent chemical and biological studies are ongoing on the calibrated lakes as summarized in Harvey et al. (1981). The results of these studies have been reported in approximately thirty publications in the primary scientific literature. Studies of precipitation, deposition, air quality, soils, groundwater, forests and precipitation throughfall are all being carried out. A stream acidification experiment was started in 1982. Studies of pH effects on fish and fish populations have been intensified since 1979 by Harold Harvey of the University of Toronto. ------- Deposition Fraction (kg S042-/ha.yr) Sum a b c d 31 bulk 32 bulk 23 wet 29 wet 37 bulk 31 bulk 35 bulk 42 bulk 38 bulk Annual Precipitation On) 0.8a 0.8a 1.2a of Cations in Surface Barrie and Sirois 1982. Dillon et al. 1980. Time Period Aug76-Jul77 Aug77-Jul78 Aug76-Jul77 Aug77-Jul78 Jun76-May80 (Mean) Jun76-May77 Jun77-May78 Jun78-May79 Jun79-May80 Annual Runoff (»> 0.45C Waters 150-300 Fisheries and Environment Canada 1978. Ontario Ministry of Environment, ongoing 3-158 Reference Scheider et al. 1979a Scheider et al. 1979a Scheider et al. 1979a; Harvey et al. 1981 Scheider et al. 1979a; Harvey et al. 1981 Scheider & Dillon 1982 Unpublished4 Unpublished4 Unpublished4 Unpublished4 yeq/Lb studies . 1 1 1 1 I • 1 1 1 I • 1 1 1 1 1 1 1 ------- I I I I I I I I I I I I I I I I I I I 3-159 Observed Characteristics 1. Severe pH depressions in streams and lakes with values as low as 4.1 recorded (Jeffries et al. 1979). 2, Sulphate concentrations in lakes average about equal to the bicarbonate concentrations (Dillon et al. 1980). 3. Manganese concentrations are elevated to about 50 yg/L compared to about 3 yg/L at the ELA station (Dillon et al. 1980). 4. Aluminum concentrations (50 yg/L) are elevated over values at ELA (Dillon et al. 1980). 5. Clear Lake, for which there are historical records, has declined in alkalinity from 33 yeq/L in 1967 (Schindler and Nighswander 1970) to between 2 and 15 yeq/L in 1977 (Dillon et al. 1978), a reduction in alkalinity of greater than 50%. 6. Mercury concentrations are higher in fish from lakes with low pH than from higher pH lakes (Suns 1982). 7. Unusually dense and extensive masses of filamentous algae proliferate in the littoral zones of many lakes with pH values of 4.5-5.5 (Stokes 1981). 8. Chrysochromulina breviturrita, an odour causing alga has reached densities that have reduced the recreational use of lakes for periods of time during the summer (Nicholls et al. 1981). The species dominance appears to be a recent phenomenon (within the past decade). This alga has been shown to increase with decreasing pH in lake acidification experiments (Schindler and Turner 1982). 9. Elemental composition of fish bones reported by Fraser and Harvey (1982) showed the centrum calcium was reduced in white suckers from lakes of pH 5.08 (King) and 5.36 (Crosson) compared to lakes of higher pH in the same area. 10. The white sucker population in Crosson Lake (pH 5.1) showed a truncated age composition compared with the age composition of the less acidic Red Chalk (pH 6.3) and Harp (pH 6.3) lakes (Harvey 1980). 11. Adult pumpkinseeds (Lepomis gibbosus) and frogs have been killed around the edges of Plastic Lake during spring melt and acidification is the suspected cause. Inlet streams had pH values as low as 3.85 (Harvey and Lee 1981). ------- 3-160 LAURENTIDE PARK, QUEBEC Humid Alpine Lower Boreal Regions I I I Elevated dome dominating the surrounding plateau. Elevation varies from 500 to 1200 m asl with summit elevations of 1100 M to 1200 m. It is comparable to the entire Laurentian plateau, • although here there are very few lakes and the drainage pattern is characterized by deep dissecting river valleys such as the Jacques Cartier. • The frost-free season is generally 80 days or less with a growing season of about 140 days. Average annual rainfall, one • of the most abundant in Quebec, ranges between 1200 and p 1600 mm. On the upper slopes and summits, 85% of the surface is covered • with glacial till of which two-thirds is less than 1 m deep, while the other 15% consists of exposed bedrock (gneiss). Low-lying areas are, for the most part, blanketed by sandy • fluvio-glacial outwash deposits. A few organic deposits exist • and are generally shallow, digotrophic and treed. Ferro-humic podzols characterize the well-drained soils with little or no ortstein to be found on excessively to well-drained sand soils. I The region, as defined by Thibault (1980), confirms early work _ completed by Jurdant and others (1968, 1972). The limits I include all areas above 518 m. Jurdant (1968) and Lafond and * Ladouceur (1968) characterized a distinct peripheral-band in the central upland plateaus covered by balsam fir and black spruce • moss forests and occasionally white birch stands. Forest • regeneration after cutting or fire, is dominated by white birch rather than trembling aspen. The central plateau supports a • black spruce moss forest cover, but after cutting, regenerates • and develops into a balsam fir Hylocomium, Oxalis forest (Lafond 1968). _ The more exposed summits in the region such as Mount Blie in the * Malbaie watershed, support a scattered alpine cover dominated by a heath, moss, and sedge complex and occasionally lichens. B Humid Lower Boreal Region This region, the Laurential foothills, is found between 47°30' • and 50°00' N latitude and 67° and 75° W longitude. Mountainous topography characterizes the region. _ Average growing season is about 150 days with a total annual ™ rainfall between 900 to 1000 mm. Due to altitudinal variations, local climate conditions vary within the region. Lower • altitudes, especially in the southern sectors are not as cold or | as wet as conditions on the higher plateau, a difference of 200-300 degree days and an average rainfall 200-300 mm. • I ------- I I I I I I I I I I I I I I I I I I I 3-161 Near the foothills, crystalline Precambrian bedrock underlies the region. Hillsides are generally covered by a thin (less than 1 m) layer of till, with deeper deposits near the base and scattered deposits on the upper slopes and summits. Fluvio-glacial deposits characterize the valley floors of the region. Ferro-humic and humo-ferric podzols are the dominant soil formations. Rowe (1972), Jurdant et al. (1972) and work completed using provincial cover maps (MER-Ministe're d'Energie et des Resources) were used to define the region. The limits as defined by Thibault (1980) and Jurdant et al. (1972) regroup regions considered by Jurdant as part of a large balsam fir-white birch forest domaine. This domaine is characterized by a semi-dense forest cover (60% crown closure nature, tree height greater than 21 m) of balsam fir and black spruce associated with white birch and an absence of jack pine. Rowe's forest region and the MER information confirmed the region's limits. Mesic hillside conditions support balsam fir-black spruce mass as well as black spruce-balsam fir mass forest covers with white birch and white spruce associations. Pure black spruce stands preferred either dry sites or poorly drained hollows. White birch and to a lesser extent trembling aspen associated with black spruce, balsam fir and white spruce characterize the regeneration. Except for a few isolated areas, the meridional sugar maple, yellow birch, red maple, red pine, black ash and American elm are not to be found in the region. ------- 3-162 Deposition Fraction Time Period (kg S042-/ha.yr) Reference 40 30/6 mo 10/6 mo 35 wet Apr79-Mar80 wet Apr79-0ct79 wet Nov79-Mar80 wet 1980 22.2 wet 28Sep81-27Sep82 Interpolated* from Glass and Brydges 1982 Interpolated* from Glass and Brydges 1982 Interpolated* from Glass and Brydges 1982 Thompson and Hutton 1982; interpolated from Barrie and Sirois 1982 Grimard 1982 Annual Precipitation (m) 1.14a Annual Runoff (m) 0.95a a Ferland and Gagnon 1974. * Interpolations from existing deposition isopleth maps as a basis for estimating deposition values can be in error. I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-163 Major Cations in Peq/L Ca Mg Na K Cond Average 114.8 54.1 37.3 8.3 22.5 Standard Deviation 57.6 23.9 12.5 3.7 8.7 Observed Characteristics 1. The surface water pH is higher than the precipitation pH. The pH of 152 lakes sampled in the last week of March 1981 and the first of April varied between 4.7 and 6.6 with an average of 5.9 (Richard 1982). 2. The average content of sulphate in the lakes is of the order of 80 yeq/L (Bobe"e et al. 1982; Richard 1982) and it is higher or equal to bicarbonate. 3. The highest sulphate concentrations in lakes in Quebec and the greatest alkalinity differences were observed in the southwest. The lake water concentrations of sulphate and the alkalinity deficits decrease to the north and east (Bob€e et al. 1982). 4. There is a significant correlation (r = 0.76, p ^ 0.001) between pH and total aluminum of the 152 lakes of Richard (1982). 5. The Laurentide Park area is found in hydrographic regions 05 and 06 (Figure 3-13). Sulphate vs. £ [Ca] + [Mg] - [alk] for these two hydrographic regions is found in Figure 3-14. 6. Compared to the pH of 1938-41, there is a greater proportion of the lakes sampled 1979-80 in the classes of pH 4.40-5.09, 5.10-5.79 and 6.50-7.19 amongst 5 pH classes (Jones et al. 1980). Lakes in the two lowest pH classes showed reductions in pH; the higher pH class increased because of road salt and nutrient additions. The decline in surface water pH tended to occur in the southern part of the park. 7. In lakes continuously open to fishing for nine years prior to 1982, average annual angling yield, angling effort, and mean weight of fish caught in years 1978-81 were not significantly related to lake pH. Management policies within the Park provide for closure of a lake to fishing ------- 3-164 and 1.2 times higher in the population of the three more acidic group of lakes comparatively to the three non-acidic group of lakes (Moreau et al. 1982). I I when angling success is reduced below projected levels. The 44 lakes which were closed to fishing over the nine year period included 43.5% of the most acidic lakes • (Group 1, mean pH 5.2); as compared with 36.9% of Group 2 • lakes (mean pH 5.9) and 17.5% of the Group 3 lakes (mean pH 6.4). Although a direct cause-and-effect relationship between fish productivity and pH has not been established, the greater number of closures in the more acidic lakes suggests a lower productivity in these waters (Richard • 1982). • 8. The concentrations of manganese, zinc and strontium in the opercula of Salvelinus fontinalis are respectively 1.6, 1.3 • I I I NOVA SCOTIA The Nova Scotian River Study by Watt et al. (1983) encompassed the approximately 500 km long Atlantic coast of Nova Scotia which is underlain by granite on about one-half of the mainland. B This study of 23 rivers which historically supported salmon • fisheries reports results of monthly monitorings from June 1980 to May 1981, with certain rivers studied as long as 10 years. • An historical comparison of five of these rivers with data • collected in 1954-55 (Thomas 1960) pH, alkalinity, and major ion concentration data was made. Fisheries data for the past 45 _ years was available for 22 of the rivers and Watt et al. (1983) • related angling success to current water chemistry and ™ geological factors. Within Kejimkujik National Park, central Nova Scotia, an ongoing study involves three calibrated lakes. I Kerekes (1980) reported results for these lakes for the • June 1978 - May 1979 period. From this study a chemical budget is available for the Mersey River (the outflow of Kejimkujik • Lake), which is included in the fisheries data set of Watt • et al. (1983). I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-165 Deposition Fraction Time Period (kg S042-/ha.yr) Location 44 22 19 22-29 22 17 27 32 31 18.12 13.18 29.01 21.27 22.50 total wet wet excess Jun78-May79 wet and 1977-79 dry excess 1981 1980 1980 1979 1978 Feb78-Dec80 Nov77-Dec80 Oct77-Nov79 May78-Dec80 Oct77-Mar80 Kejimkujik, Kerekes (1980) Interpolated* from Figure 3, Underwood (1981) Kejimkujik3 Kejimkujikb Truroc Truroc Truroc East River St. Marysd Cobequidd Bridgetown^ New Rossd Kemptvilled Annual Annual Precipitation Runoff (m) (m) 1.2e 1978 1 mf 1.6e 1979 1.2e 1980 1.40 June 1978 - May 1979 1.46^ long-term average Sum of Cations for 41 lakes and rivers 59 _+ 17 ueq/L (Standard Deviation) a b c d e f Barrie pers. comm. Barrie et al. 1982. Truro CANSAP received a fair rating in the siting assessment (Vet and Reid 1982) and the station is being moved (Barrie pers. comm.). Underwood 1981 and Underwood pers. comm. Barrie and Sirois 1982. Fisheries and Environment Canada 1978. Interpolations from existing deposition isopleth maps as a basis for estimating deposition values can be in error. ------- 3-166 Observed Characteristics I I 1. Precipitation pH is generally lower than the pH of the • runoff water. High runoff is associated with the lowest pH • values in river waters. The lowest mean monthly values in rivers generally occur in winter (Watt et al. 1983). • 2. Sulphate is the dominant anion in three study lakes of pH 5.4, 4.8 and 4.5 (Kerekes 1980) and was highest in the _ two coloured lakes with lowest pH. • 3. Excess sulphate export from the watersheds are elevated above those of remote areas by a factor of about 4 • (Thompson and Button 1982) and sulphate export exceeds the • measured wet deposition indicating possible dry deposition. • 4. pH data are available for four rivers (corrected for flow) and 1980-81 values are less than 1954-55 by 0.24 to 0.79 _ units. The current bicarbonate concentrations are lower • and sulphate and aluminum concentrations are higher than ™ historical values (Watt et al. 1983). 5. Two rivers (St. Mary's and Medway) had the lowest pH values • and highest excess sulphate loads in 1973. Similar changes in pH and excess sulphate were noted for two Newfoundland • rivers (see Figure 3-30). | 6. Long-term (five years or greater) records for pH, calcium — and sulphate from eleven rivers in Atlantic Canada were • fitted by time series models. Five of eight sensitive * rivers decreased in pH and the other three did not change, while none of four insensitive rivers decreased. B Relationships between trends in pH and Calcium and sulphate 8 indicate that, conceptual models applied satisfactorily for pH and in only a limited number of cases for calcium and sulphate (Clair and Whitfield 1983). 7. Salmon catch data for 22 rivers which have not been _ affected by watershed changes or salmon stocking, have been • recorded from 1937 through 1980. As a group (n = 10), • rivers in the pH range 4.6 - 5.0 have reduced salmon stocks as reflected by a significant decline in angling catches over this time. Collectively, rivers with current pH values >5.0 do not show any significant trend in salmon catch over the past 45 years (Watt et al. 1983). The • absence or reduced abundance of Atlantic salmon in 17 • rivers was corroborated by electrofishing surveys in 1980-82 (Watt et al. 1983). I I I I ------- I I I I I I I I I I I I I I I I I I I 3-167 8. Diatom assemblages in four Halifax study lakes shifted toward more acid tolerant species between 1971 and 1980 (Vaughan et al. 1982). BOUNDARY WATERS CANOE AREA AND VOYAGEURS NATIONAL PARK, MINNESOTA The Boundary Waters Canoe Area Wilderness (BWCA), a wilderness unit within the Superior National Forest (Minnesota) and located along 176 km of the Minnesota-Ontario border. The area varies from 16 to 48 km in width. Over 1,900 km of streams, portages, and foot trails connect the hundreds of pristine, island-studded lakes that make up approximately one-third of the total area. Most of the BWCA is included within the Rainy Lake basin, except for the eastern section, which is part of the Lake Superior watershed. Of a park total of 88,800 ha, several thousand of the 34,700 ha of recreational water in the VNP were created by dams, leaving 54,080 ha of land. The park has 31 named lakes and 422 unnamed swampy ponds larger than 2 ha. The BWCA has a surface area of 439,093 ha patterned by 1,493 lakes greater than 2 ha, and over 480 km of major fishing and boating rivers in addition to numerous streams and creeks (Glass and Loucks 1980). Filson Creek watershed is approximately 13 km southeast of Ely, Minnesota. Filson Creek drains 25.2 km^ and flows north and west to the Kawishiwi River. Included in the watershed are Omaday and Bogberry Lakes and one tributary, designated South Filson Creek for this study. South Filson has a 6.3 km^ drainage area and no significant lakes. About 60% of Filson Creek watershed is covered by mixed upland forest, 30% by wetlands and lakes, and the remainder by planted or natural stands of pine. Wetlands surround the lakes. The precambrian bedrock is mostly troctolite (a pyroxene-poor, calcic gabbro) and other igneous rocks of the Duluth Complex. The northern 10% of the watershed is underlain by the Giants Range granite. A mineralized zone along the contact between the granite and the Duluth Complex contains copper and nickel sulfide minerals. The watershed has no carbonate rocks. Bedrock is at the land surface in about 10% of the watershed. Most of the watershed is covered by drift generally less than 1 m thick. Its mineral composition reflects the underlying bedrock types. The total thickness of drift and peat under the wetlands can exceed 15 m. The peat in most of the wetlands is fibric, herbaceous, and partly decomposed (sapric) below about 0.75 m (Seigel 1981). ------- 3-168 Deposition Fraction (kg S042-/ha.yr) 10-15 wet 13 wet 1 . 6 snow 17 bulk 17.2 wet Time Period 1976-78 1981 1978 (snow season) Nov76-0ct77 1980 Reference Glass and Loucks 1980 NADP 1981-83 (Marcel site) Glass 1980 Siegel 1981 NADP 1981-83 16.6 14.8 wet Apr78-May79 Wet Apr78-May79 (Marcell site) Total NE Minn., Eisenreich et al, 1978 Heiskary et al. 1982 (Hovland site) Observed Characteristics 1. No known chemical or biological effects in lakes (Glass 1980; Glass and Loucks 1980). 2. Most of BWCA lakes surveyed have pH values <6.0 and 36.5% had CSI >3 (Glass 1980; Glass and Loucks 1980). 3. Of the 290 sites sampled 50.5% had alkalinity values between 40-199 ueq/L no lakes had alkalinity values less than 40 yeq/L (Glass 1982; Glass and Loucks 1980). 4. Filson Creek watershed retained 10.6 kg S042~/ha.yr of 17 kg S042~/ha.yr bulk (Siegel 1981) 5. S042~ increased from 2 to 14 mg/L and [H+] from pH values of 6.6 to 5.5 during snowmelt (Siegel 1981). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-169 NORTHERN WISCONSIN Northern Wisconsin is a region in which a collapsing glacial mass left deep outwash sands and coarse tills interspersed with ice-blocks. The study area encompasses portions of seven counties in the Upper Wisconsin River Basin. Water covers 17% of the area. The area has had a 30% increase in population over the last decade, much of which has occurred along lakeshores. Although only 3% of the total land area is developed, approximately 40% of the lake shoreline is in residential land use. About 90% of the land surface in the region is now forested. A century ago the upland vegetation was dominated by white pine, hardwoods and hemlock, but most of it was removed during logging and subsequent burning in the late 1800s and early 1900s. Regrowth of aspen, birch, mixed hardwoods and a few conifers has taken place now, much of it since 1920. Black spruce is common on the wet, peat areas. The sands and sandy loams in the surface layers have produced mostly acid soils (commonly pH 4-5), with low cation exchange capacities (10 meq/100 g) and low base saturation (10-30%). The upland soils are primarily sands and sand loams with peatland soils in the depressions. Total concentrations of calcium and magnesium in these soils are typically 1-2 meq/100 g. The igneous and metamorphic bedrock underlying these northern Wisconsin counties is part of a southern extension of the Precambrian Canadian Shield. The principal bedrock type is granite, with lesser amount of diorite, schist, gneiss, quartzite, slate and greenstone. The bedrock is overlain by the glacial drift, the most recent of which was deposited during the Wisconsin glaciation. Drift thickness ranges between 10 and 70 m with an average slightly greater than 30 m. The drift is low in calcareous material, calcareous pebbles are found only in the deeper, older drift. Essentially all groundwater contributions to lakes and streams follows a path through the glacial drift. Because most of the lakes occur in pitted glacial outwash or end moraines, they are generally shallow, averaging about 10 m in maximum depth and rarely exceeding 30 m. Consequently, virtually all of the lakes in this study area are situated well above bedrock, encased in thick glacial deposits. The recent pH of the rainfall has averaged 4.6 annually compared with an estimated 5.6 in the middle 1950s. The climate is cool and wet, with mean July temperatures of 19°C and January temperatures of -11°C. The lakes commonly are ice-covered from late November to late April (Schnoor et al. 1982). ------- 3-170 (Trout Lake) 17 wet (71 cm) 1980 NADP 1981-83 (Trout Lake) 16 wet (84 cm) 1981 NADP 1981-83 22 bulk 1981 Becker et al. 1982 Precipitation Runoff (m) (m) .80 .30 I I Deposition Fraction Time Period Reference • (kg S042-/ha.yr) | 17 wet (68 cm) 1981 NADP 1981-83 I I (Spooner) • I Annual Annual • I I Observed Characteristics 1. Median alkalinity for 117 seepage lakes sampled was 39 yeq/L. Conductivity and colour for the same lakes was 21 I yS/cm and 8 Pt units (Eilers et al. 1982). For 409 total • sites, 25.4% had alkalinities <40 yeq/L and 22.7% had alkalinities between 40 and 199 yeq/L (Glass 1982). • 2. Two separate comparisons of present chemistry with the 500 Wisconsin lake survey of Birge and Juday (1925-41) have _ found that most lakes have significantly higher pH, I alkalinity and conductivity (Bowser et al. 1982; Schnoor et " al. 1982). Approximately 20% of lakes sampled had pH declines but the differences were not statistically • significant. • 3. Hydrologic type appears to control alkalinity. Median • values of pH (6.4) alkalinity (39 yeq/L) and conductivity • (21 ymohs) were found in seepage lakes (no defined inlet or outlet) (Eilers et al. 1982; Schnoor et al. 1982). _ I I I ------- I I I I I I I I I I I I I I I I I I I 3-171 ADIRONDACK MOUNTAINS OF NEW YORK "As a result of extensive glacial activity, the Adirondack region of northeastern New York State contains a vast and varied ponded water resource. The most recent count adapted from a 1979 inventory of the Adirondack ecological zone (Pfeiffer 1979) reveals that there are approximately 2,877 individual lakes and ponds, encompassing some 282,154 surface acres. The New York State portion of Lake Champlain, 97,000 acres, is purposely excluded from this summary since its low elevation waters are not considered to be representative of the Adirondack uplands. Average size of ponded waters included in this inventory approaches 98 acres and ranges from those of less than one acre to 28,000 acre Lake George." (Pfeiffer and Festa 1980) The Integrated Lake-Watershed Acidification Study (ILWAS) selected three forested watershed areas (Panther, Woods and Sagamore) in the Adirondack Park region of New York for field investigation. The watershed areas contain terrestrial and aquatic ecosystems having physical, chemical and biological characteristics which distinguish one area from another. Lake outlets are the hydrologic terminal points of all three watersheds. The study watersheds are within 30 km of each other. Runoff in Panther and Woods watersheds drains directly to the lakes without extensive steam development. Sagamore Lake receives the majority of its inflow through a drainage system of bogs and streams. All watersheds contain mixtures of coniferous and deciduous vegetation. Panther Lake sits on thick till rather than bedrock. The stratigraphy of the till is typically, from top to bottom, sand, sandy till, silty till, and clay till overlying bedrock. The till in Woods Lake basin is primarily sandy till with an average depth of three metres. Panther Lake basin has two till units, a sandy unit and a clay-rich unit; the two units together may be 60 m deep in places. Sagamore Lake basin has four units - a loose sandy unit, a more compact sandy unit, a silt-rich unit, and a clay-rich till. A thick sand deposit greater than 30 m deep, at the site of a glacial meltwater channel, is present near the inlet to Sagamore Lake. High runoff periods typically occur during snowmelt. A winter thaw has been observed in January and February. A larger spring melt occurs in March and April. During the summer and fall, occasional storms may also generate high runoff. ------- Deposition (kg S042-/ha.yr) 26.4 29 34-37 39-43 40.03 5.38 39.40 6.19 32.92 5.71 Fraction wet wet bulk bulk wet dry wet dry wet dry Time Annual Period Precipitation (m) 1981 1.02 1980 1965-78 1965-78 Jun78-May79 1.25 Jun78-May79 Jun78-May79 1.21 Jun78-May79 Jun78-May79 .98 Jun78-May79 3-172 Reference NADP 1981-83 (Huntington site) NADP 1981-83 (Huntington site) Peters et al. 1981 (Canton site) (Hinckley site) Johannes et al . 1981 (Wood ' s Lake - ILWAS) Johannes et al . 1981 (Panther Lake - ILWAS) Johannes et al . 1981 (Sagamore Lake - ILWAS) 1 1 I 1 1 _ • 1 1 1 1 • 1 1 1 1 1 1 ------- I I I I I I I 1 I I I I I I I I I I I 3-173 Summary of 13 Years (1965-1978) Precipitation Data (Mean + S.D.) (Peters et al. 1981) Precipitation Site (cm/yr) SO^2" (ueq/L) N(>3~ (yeq/L) Canton 94 _+ 8 0.104 Hh 0.057 0.033 + 0.034 Hinckley 129 + 52 0.084 + 0.039 0.027 + 0.025 8042~ concentration increased by 1-4%/yr, while H+ has remained unchanged. has increased by 4-13%/yr. and S042~ loads have increased [% slopes: 12-15% (N03~) and 0.5-0.7% (S042~) for the Canton and Hinckley sites, respectively] due partially to an increase in the amount of precipitation. Observed Characteristics 1. In the East Branch of the Sacandaga River, S042~ concentrations exceed HC03~ concentrations. USGS monitoring of the river from 1965 to 1978 indicate an increase in N03~ (4 peq/L.yr), a decrease in S042~ (4 peq/L.yr), and a decrease in alkalinity (83 peq/L.yr) (Peters et al. 1981). 2. In a 1975 survey of 214 Adirondack lakes at high elevations, pH ranged from 4.3 to 7.4. Fifty-two percent of the lakes had pH <5.0; 7% pH 5.5-6.0 (Schofield 1976c). 3. For a subset of 40 of these 214 lakes, historic data on pH and fish populations are available from the 1930s. Over this period, the number of lakes with pH <5.0 increased from 3 (out of 40) to 19. Likewise the number of lakes without fish increased from 4 to 22. In both surveys, none of the lakes with pH <5.0 had fish. 4. For 138 Adirondack lakes, a comparison of color-metric pH measurements for the 1970s vs. 1930s indicated a general decrease in pH (Pfeiffer and Festa 1980). 5. pH depressions in streams during spring snowmelt and periods of heavy rainfall have been observed (Driscoll et al. 1980; Galloway et al. 1980b). ------- 3-174 THE HUBBARD BROOK ECOSYSTEM, NEW HAMPSHIRE I I 6. Based on a comparison between lakes in the Adirondack region and within a given lake or stream monitored over time for a one- or two-year period, elevated aluminum • concentrations have been demonstrated to be associated with • low pH (Driscoll et al. 1980; Schofield 1976). 7. Current status of fish populations (presence/absence) in | Adirondack lakes and streams is clearly correlated with pH level. The occurrence of fish is reduced particularly at • pH levels below 5.0 (Colquhoun et al. 1980; Pfeiffer and I Festa 1980; Schofield 1976). In the 1975 survey of 214 • high elevation lakes, in 82% of the lakes with pH < 5.0 no fish were collected. For lakes with pH >5.0, about 11% I had no fish collected (Schofield 1976b). • 8. The New York Department of Environmental Conservation • reported (based on available data) that 180 lakes have lost | their fish populations (Pfeiffer and Festa 1980). Although no alternative explanations for this loss of fish are _ readily apparent, historic records are not adequate to I definitely establish acidic deposition as the cause. * 9. Survival of brook trout stocked into Adirondack waters was inversely correlated (p < 0.01) with elevated aluminum concentrations and low pH (Schofield and Trojnar 1980). I I The Hubbard Brook Experimental Forest (HBEF) was established in • 1955 by the United States Forest Service as the principal • research area for the management of watersheds in New England. The name of the area is derived from the major drainage stream B in the valley, Hubbard Brook. Hubbard Brook flows generally B from west to east for about 13 km until it joins with the Pemigewasset River, which ultimately forms the Merrimack River • and discharges into the Atlantic Ocean. Water from more than 20 I tributaries enters Hubbard Brook along its course. Mirror Lake, a small oligotrophic lake, discharges into Hubbard Brook at the lower end of the valley. The HBEF is located within the White B Mountain National Forest of north central New Hampshire. B Although the climate varies with altitude, it is classified as humid continental with short, cool summers and long, cold • winters. The climate may be characterized by: (1) change- | ability of the weather; (2) a large range in both daily and annual temperatures; and (3) equable distribution of • precipitation. HBEF lies in the heart of the middle latitudes I and the majority of the air masses therefore flow from west to east. During the winter months these are northwesterlies and during the summer the air generally flows from the southwest. B Therefore, the air affecting HBEF is predominantly continental. B However, during the autumn and winter, as the colder polar air I I ------- I I I I I I I I I I I I I I I I I I I 3-175 moves south, cyclonic disturbances periodically move up the east coast of the United States providing an occasional source of maritime air. The mean air temperature in July is 19°C and in January is -9°C. A continuous snowpack develops each winter to a depth of about 1.5m. Occasionally, mild temperatures in midwinter partly or wholly melt the snowpack. A significant microclimatologic feature of this area is that even the uppermost layer of the forest soils usually remains unfrozen during the coldest months because of the thick humus layer and a deep snow cover. The HBEF covers an area of 3,076 ha and ranges in altitude from 229 to 1,015 m. The experimental watershed ecosystems range in size from 12 to 43 ha and in altitude from 500 to 800 m. These headwater watersheds are all steep (average slope of 20-30%) and face south. The experimental watersheds have relatively distinct topographic divides. The height of the land surrounding each watershed ecosystem and the area have been determined from ground surveys and aerial photography. The geologic substrate, outcrops of bedrock and stoney till, in the Hubbard Brook Valley was exposed some 12,000-13,000 years ago when the glacial ice sheet retreated northward. Bedrock is derived from highly metamorphosed sedimentary rocks of the Littleton formation and the granitic rocks of the Kinsman formation. The bedrock of watersheds 1-6 is the Litleton formation, which in this area is made of highly metamorphosed and deformed mudstones and sandstones. It is medium to coarse grained and consists of quartz, plagioclase, and biotite with lesser amounts of sillimanite. Much of the area of the experimental watersheds is covered with glacial till derived locally from the Littleton formation. The geologic substrate is considered watertight and losses of water by deep seepage are minimal. Soils are mostly well-drained spodosols (haplorthods) of sandy loam texture, with a thick (3-15 cm) organic layer at the surface. Most precipitation infiltrates into the soil at all times and there is very little overland flow (Pierce 1967). This is because the soil is very porous, the surface topography is very rough (pit and mound, mostly from wind-thrown trees), and normally there is little soil frost. Soil depths are highly variable but average about 0.5 m from surface to bedrock or till. Soil on the ridges may consist of a thin accumulation of organic matter resting directly on the bedrock. In some places, impermeable pan layers at depths of about 0.6 m restrict vertical water movement and root development. The soils are acid (pH £ 4.5) and generally infertile. ------- 3-176 The vegetation of the HBEF is part of the northern hardwood ecosystem, an extensive forest type that extends with variations for Nova Scotia to the western Lake Superior region and southward along the Blue Ridge Mountains. Classification of mature forest stands as northern hardwood ecosystems rests on a loosely defined combination of deciduous and coniferous species that may occur as deciduous or mixed deciduous-evergreen stands. Deposition (kg S042~/ha.yr) Fraction 36.4 22 38.4 + 2.5 33.7 30.0 41.6 42.0 46.7 31.2 29.3 34.6 33.0 43.4 52.8 wet wet bulk bulk bulk bulk bulk bulk bulk bulk bulk bulk bulk bulk Time Annual Period Precipitation Reference (m) 1981 1980 1964-74 Jun63-May64 Jun64-May65 Jun64-May65 Jun64-May65 Jun64-May65 Jun64-May65 Jun64-May65 Jun64-May65 Jun64-May65 1973-74 1973-74 1.50 NADP 1981-83 (Hubbard Brook) .87 NADP 1981-83 (Hubbard Brook) 1.30 Likens et al. 1977a Likens et al. 1977a Observed Characteristics 1 . The external and internal generation of H"1" exerts nearly equal roles in driving the weathering reactions. Input of H+ is mainly in the form of H2S04 and HN03 (Likens et al. 1977a). 2. Average streamwater pH ~ 5. During snowmelt events pH depressions of 1.0 to 2.0 units have been reported (Likens et al. 1977a). 3. The Hubbard Brook ecosystem accumulated hydrogen, nitrate and ammonium ions over the period 1963-74. Over the same period there was a net loss of SO^- (Likens et al. 1977a). 4. Ca2+ and SO^- dominated the streamwater chemistry at the HBEF. SO^" was more than 4 times as abundant as the next most abundant anion which was N03~ (Likens et al. 1977a). I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-177 5. Elevated levels of Al are found in the headwater portions of streams in the HBEF. These levels are 2-29 times above levels in downstream waters. This effect was attributed to leaching of Al hydroride compounds from soils by acidic deposition (N.M. Johnson 1979). MAINE AND NEW ENGLAND The 97 lakes sampled by Norton et al. (1981a) ranged in pH from 4.25 to 6.99 (median = 6.40), in elevation from 12 to 1307 m (median = 154 m), surface area from <0.1 to 1098 ha (median = 56 ha), Secchi disc transparency from 2.5 to >5.0 m (median = 6.3 m), and water colour from 0 to 110 Pt units (median = 8 Pt units). The bedrock of the study area was noncalcareous and mostly granitic. As a result, the lake waters were of low alkalinity (0-360 yeq/L, CaC03; median = 64) and specific conductance (0-68 ymhos/cm at 25°C; median =29). The watersheds were almost completely forested; very little cutting had occurred in the few decades prior to sampling. Many of the lowland lakes (fJjOO m) had cottages along their shores and access roads in their watersheds; the high elevation lakes were pristine and accessible only on foot. In summary, the lakes were small to medium size, ologotrophic to mesotrophic with moderately to very transparent water, low to moderate concentrations of humic solutes, and low alkalinity and conductance, and with moderately disturbed to pristine watersheds. Haines and Akielaszek (1982) sampled a similar set of 226 headwater lakes and streams in the other New England states, including Maine. Deposition (kg S ha. 28. 25. 24. 17. 18 36. 22 38.4+2 35 04-2-/ yr) 0 31 80 22 4 .5 Fraction wet wet wet wet wet wet wet bulk (129.5 cm) wet Time Period 1981 1981 1981 1981 1980 1981 1980 1963-74 Annual Precipi- tation (m) 1.10 .87 1.15 1.10 1.50 .87 1.30 Reference NADP NADP NADP NADP NADP NADP NADP 1981-83 1981-83 1981-83 1981-83 1981-83 1981-83 1981-83 Likens et al (Acadia site) (Bridgton site) (Caribou site) (Greenville site) (Greenville site) (Hubbard Brook) (Hubbard Brook) . 1976, 1980 (Hubbard Brook) 1981 .74 NADP 1981-83 (Bennington VI site) ------- 3-178 Observed Characteristics SUMMARY OF EMPIRICAL OBSERVATIONS (kg S042~/ha.yr) NORTHERN 5 wet (1980) No chemical effects SASKATCHEWAN 17 bulk (1977) I I 1. Lakewater pH declines based on comparisons with historical • information (Davis et al. 1978) where 85% of 94 lakes studied | (Norton et al. 1981 and 64% of 95 lakes studied (Haines and Akielaszek 1982) were found to have lower pHs. • 2. Loss of alkalinity from lakewater in the New England states averaging about 98 )jeq/L for 56 lakes for which there was historical information (Haines and Akielaszek 1982). I 3. Paleolimnological confirmation of pH declines in lakes (Davis et al. 1982). Cores from New England acidic clear water • lakes (pH less than 5.5) with undisturbed drainage basins (5 | of the 30 lake samples taken over at least the last 50 years) show declines in sediment concentrations of Zn, Ca, Mg and Mn « starting as early as about 1880 suggesting increased leaching • of sediment delutus prior to entry into the lakes (Davis et al. 1982; Kahl et al. 1982) or reduced sedimentation rate. 4. Accelerated cation leaching from watersheds (Kahl and Norton • 1982). 5. Lakes of pH <5 are distributed throughout a range in | elevation from 10 to 1000 m. High elevation lakes (>600 m) tend to have low pH and alkalinity. All but two lakes having _ pH <5.5 were also less than 20 ha in surface area. I Alkalinity and pH also increased with stream order (Haines ^ and Akielaszek 1982). Of 226 lakes and streams sampled 25% had alkalinity 120 Veq/L, 41% were 1100 Meq/L and 50% were • 1200 Ueq/L (Haines and Akielaszek 1982). I SUMMARY Location Deposition Summary Effects • I ELA, ONTARIO 5.9 wet (1980) No effect I 9 and 11 bulk (1972-73) • MINNESOTA 10-15 wet (NovSO) No effect • I I ------- I I I I I I I I I I I I I I I I I I I 3-179 NORTHERN WISCONSIN ALGOMA, ONTARIO NOVA SCOTIA MAINE HUBBARD BROOK, NEW HAMPSHIRE MUSKOKA-HALIBURTON 16-17 wet (1981) 24.7 wet (1981) 22 wet (1981) APN (Kejimkujik) 17 wet (1980) APN (Kejimkujik) 22.5 wet (1977-80) (CANSAP-Kemptville) 13.2-32 (various years) (CANSAP - various N.S. sites) 17-28 wet (1981) 36 wet (1981) 22 wet (1980) 33-53 bulk (1963-74) 23-29 wet (1976-78) 31-42 bulk No effect pH depression 2.1 pH units Elevated excess sulphate relative to region not receiving acidic deposition More lakes of low pH than expected Relationship between fish and alkalinity Loss of Atlantic salmon populations (historic record). Historic record of decreased pH in river Evidence of slight pH decrease in lakes (historic records) No effects on Atlantic salmon No evidence of effects on fish in inland lakes Spring pH depressions No long term change in stream or lake pH 1963-present pH depressions Fish kill associated with pH depression in one lake Algal composition related to pH ------- 3-180 LAURENTIDE PARK, QUEBEC 22.2-40 wet (1977-80) ADIRONDACKS 32-40 wet (1978-79) 29 wet (1980) 34-37 bulk (1965-78) Indications of decrease in pH in some lakes, especially in southern region of park (increases in some lakes, especially along roads); indication of decline in angling success in lower pH lakes; pH depression in lakes in spring (Moreau et al. 1982) and lower pH in lakes in spring than summer (Bob€e et al. 1982) Evidence of pH declines and loss of fish populations over time Detailed studies of watersheds have been carried out in sensitive regions of North America and Scandinavia under a range of sulphate deposition rates. The results of watershed studies conducted in North America are described below. For those regions currently experiencing sulphate in precipitation loading rates of ^17 kg/ha.yr there have been no observed detrimental chemical or biological effects. For regions currently experiencing between 20 and 30 kg/ha.yr sulphate in precipitation there is evidence of chemical alteration and acidification. In Nova Scotia rivers, historical records of salmon population reductions as documented by 40 years of catch records have occurred as well as reductions in stream pH. In Maine there is evidence of pH declines over time and loss of alkalinity from surface waters. In Muskoka- Haliburton there is historical evidence of loss of alkalinity for one lake. There is documentation of pH depressions in a number of lakes and streams. Fish kills were observed during spring melt in one lake. In the Algoma region there are elevated sulphate and aluminum levels in some headwater lakes. For regions currently experiencing loading ^30 kg/ha.yr there are documented long-term chemical and/or biological effects and short-term chemical effects in sensitive surface waters. I I I I I I 1 I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 3-181 In Quebec, sulphate concentrations in surface waters decrease towards the east and north in parallel with the deposition pattern. Sulphate concentrations are equal to or greater than the bicarbonate concentration in lakes in the south west part of the Province. In the Adirondack Mountains of New York, comparison of data from the 1930s with recent surveys has shown that more lakes have been acidified. Fish populations have been lost from 180 lakes. Elevated aluminum concentrations in surface waters have been associated with low pH and survival of stocked trout is reduced by the almuninum. In the Hubbard Brook study area in New Hampshire there is pH depression in streams during snowmelt of 1 to 2 units. Elevated levels of aluminum were observed in headwater streams. 3.9.2.2 Short-Term or Episodic Effects While current and historical survey data may imply long-term trends, the samples usually represent only one or a few measurements at any one location and are usually collected only during the summer. This limited sampling period provides no record of pH and other chemical changes which take place in relation to seasonal cycles or major weather events. If short-term changes in water chemistry coincide with sensitive periods in the life cycle of fish, significant mortality and reduced reproduction can occur. Severe pH depressions in streams and small lakes due to snowmelt have been documented in a number of locales (e.g., Kahl and Norton 1982; Schofield 1973). The depression may be as much as 1-2 pH units. Much of the metal content of the snowpack is also released in early melting stages. Thus critical hydrogen ion and trace metal levels may be reached temporarily, even in waters with relatively high summer pH values. Leaching of metals from soils and sediments may be especially severe during this period, resulting in pulses of high concentrations of potentially toxic metals (e.g., Al > several hundred ppb [Kahl and Norton 1982; Schofield and Trojnar 1980]). The question has often been raised, "How long does it take before the lakes become acidic?" The previous sections have indicated relation- ships for lakes and streams which have already been acidified. However, the rate of change is one of the least well-defined aspects of the acidification phenomenon. The rate-of-change questions become less relevant in light of evidence that current acid loadings are causing damage to fisheries and other biota due to short-term exposures to low pH and associated high metal concentrations, as reviewed earlier in this report. The pH of lakes or streams tends to fluctuate considerably during the year, and average annual pH is a composite of these patterns. Thus, ------- 3-182 I organisms which may respond to extreme concentrations of H+ or metals over a few days. This, plus the known significance of brief • acute exposure (Spry et al. 1981), suggests that the magnitude and I duration of short-term increases in H+, associated with a defined "flushing event", could be used for further evaluating critical dose/ response relationships in stream ecosystems, and lakes. • Research on brook trout and white sucker by Baker (1981), Baker and Schofield (1980), on Atlantic salmon by Daye (1980) and Daye and • Garside (1975, 1977, 1980), and related research by Beamish and J Harvey (1972), Beamish (1974a, 1974b, 1976), and Harvey (1975, 1979, 1980) has provided a broad understanding of the response of several _ pH-sensitive fish species to both long-term and short-term elevated • H and aluminum exposures. Mortalities have been documented for chronic pH depression, and effects on egg viability, hatching success and adult survival for short-interval acute H+ and aluminum • exposures are reasonably well known (Baker and Schofield 1980). I Among the experimentally-based relationships developed by Daye, • Garside, Baker and Schofield is a recurring pattern (Loucks et al. • 1981): 1) the short-term acute exposure, or "shock", effects, • including responses to aluminum, can take place in two to ™ four days of exposure, with as little change as 0.5 to 1.5 units of the pH scale; and H 2) these shock exposures can be expected to occur in waters with a broad range of pH above the level at which chronic • effects occur. • Stream water chemistry studies from a number of locations (Table 3-26) show short-term pH depressions during snowmelt and storm • events (e.g., 1.0 unit on the Shavers Fork River in West Virginia B [Dunshie 1979]) and from 1.0 to 2.0 units in two watersheds being studied in the Adirondacks (Galloway et al. 1980b). A third lake • studied by Galloway et al. (1980b) at the Adirondack site had a mean | annual pH of about 4.8 and shows no pH depression during flushing event. Likens et al. (1977a) reported pH depressions of 1.0 to 2.0 ^ units for Hubbard Brook, New Hampshire. Outside the regions with • snow accumulation, the maximum pH decline during a flushing event appears to occur during major rainfall events following a rain free period (Dunshie 1979). fl Sulphate loadings associated with observed short-term pH declines and resulting biological effects are summarized in Table 3-26. In the • ELA, Ontario, annual loadings of sulphate in precipitation of about | 10 kg S042~/ha.yr have generally resulted in pH declines of only 0.2-0.3 units and no apparent biological effects. Depressions _ in pH of 0.3-1.0 units have been observed in northern Minnesota • streams receiving approximately 14 kg S042-/ha.yr. However, I I ------- I I I I I I I I 1 I I I I I I I I I I 3-183 TABLE 3-26. PERIODIC pH DEPRESSIONS OBSERVED IN STREAMS AND LAKES WITH DIFFERENT SULPHATE LOADINGS AND CORRESPONDING BIOLOGICAL EFFECTS. SURFACE WATER ALKALINITIES IN THESE AREAS ARE GENERALLY LESS THAN 200 yEQ/L. Location (kg Annual Sulphate Loading S042~/ha.yr Lowest PH Observed Largest Between pH Difference Spring pH Observed Biological Effect and summer or by wet deposition) w i nter val ues Tovdal R. Norway L. Timmevatten Sweden (1970) Sweden (1972) Hubbard Brook Experimental Forest Panther L. ILWAS Project New York 40 40 40 pH shock suspected but no field measurements taken during the fish kill 4.2 4.3 0.8 1.1 Harvey et a 1. 1982 Mills pers. comm. Keller and Gale 1982 Siege) 1981 Church and Galloway 1983 Fish kilI (sea trout)3 Wild population of minnows have disappeared*3 Caged sea trout and minnows experienced 68$and 59% mortality6 No biological studies No biological data available; fish population 1st from one lakeJ Muskoka-Hal i burton 30 4.1 1.1 Ontario (4 streams) ( lake outflows) 30 4.8 1.3 Plastic Lake 30 4.0 1 .7 Ontario Inlet Outlet 30 5.0 0.7 Shavers Fork W. Virginia 30 5.6 0.9 (stream) Algoma 5.0 2.1 Fi Ison Creek, 14 5.5 0.3-1 Northern Minnesota Experimental Lakes 10 4.5 has been 0.2-0.3 Area Ontario observed generally above 5 Evidence of fish population damage in areas lakes0 and actual algae species^ 100? mortality of caged rainbow trout' 13? mortality of caged rainbow trout^ Conditions caused by heavy rain; no biological studies6 No biological studies11 No biological studies' No apparent biological effects a Braekke 1976 b Hultberg 1977 c Harvey 1980 d Nichol Is et al . 1981 e riiinchia 1Q7Q ------- 3-184 I I the lowest pH reading recorded is 5.8 and no biological studies have been conducted. Galloway and Dillon (1982) have examined the relative importance of • sulphuric and nitric acids in causing alkalinity (and pH) reduction during snowmelt and conclude that a major portion of the reduction in alkalinity during snowmelt was attributable to nitric acid. Although • • itself showed little variation during snowmelt, its continued large presence in the stream was responsible for the alkalinity reduction in an indirect manner, namely by causing • long-term alkalinity reductions (as opposed to episodic). Thus, the | episodic reduction of alkalinity due to NOg" is added to the long-term reduction in alkalinity due to SO^". Jeffries et jm al. (1981) demonstrated that in Muskoka-Haliburton the increase in • hydrogen ion concentration in several streams during snowmelt was due to increases in both N(>3~ and SO^". 3.9.2.3 Sensitivity Mapping and Extrapolation to Other Areas of Eastern Canada J| TERRESTRIAL In order to identify the magnitude of the surface water acidification • problem our ability to extrapolate the results of the detailed watershed study areas to the remainder of eastern North America must be determined. Within eastern North America are hundreds of • thousands of lakes and streams and it is clearly impractical to • establish detailed or regional hydrochemical monitoring for them all. However, there is an urgent need to determine if the watershed study fij areas currently being monitored are anomalous in terms of their | geochemical characteristics or if, in fact, they are representative of conditions occurring over large areas of eastern North America. An early approach to this problem in Canada was to consider all of ™ the Precambrian Shield as "sensitive" and then assume any study area located anywhere on the Shield would be representative of over 75% of • eastern Canada (Altshuller and McBean 1979). This approach implied | that the Canadian Shield was a single granitic plate and not, as is the case, a number of complex geological provinces composed of a «m variety of rock types (including marble) and covered, in places, by • unconsolidated material of varying texture and carbonate content (Section 3.5). Areas outside the Shield, where hydrochemical changes have occurred (e.g., the Maritime Provinces), also exhibit a range of 9 soil and bedrock conditions. ™ A major drawback to more detailed analyses and extrapolation has been • the lack of the analyses of information on surficial and bedrock j| geological conditions for all of eastern North America, in a regional but detailed form. This has recently been alleviated for Canada with M bedrock sensitivity mapping of Shilts et al. (1981) which has been • incorporated into the bedrock-soil mapping composite of Lucas and I 1 ------- t I I I I I I I I I I I I I I I I I I 3-185 Cowell (1982). This mapping is discussed in more detail in Section 3.5. In order to determine the representativeness of three of the detailed watershed study areas, (Algoma, Muskoka-Haliburton and southwest Nova Scotia), the 65 classes of soil and bedrock characteristics mapped by Lucas and Cowell (1982) will be utilized. The basis for extrapolation is the 1:1,000,000 scale map shown in Figure 3-9 (in map folio). This mapping represents the most detailed compendium of soil and bedrock characteristics yet assembled for all of eastern Canada. Extrapolation has been carried out by reviewing the kinds of soil and bedrock terrains which form the geochemical templates of three of the watershed study areas and then determining how representative these areas are in eastern Canada. The 65 classes of terrain characteristics are listed in Tables 3-27 and 3-28. Each class is identified according to a two or three character alpha-numeric code which is defined in Table 3-8. Table 3-27 lists the area and percent cover of each class north and south of 52°N latitude for each province. [Figure 3-9 shows only the areas south of 52°N.] Table 3-28 summarizes the area and percent cover of each class for all of eastern Canada. This table indicates that 54% of eastern Canada is composed of bedrock types in combination with soil types which have a low potential to reduce acidity. These are predominately noncalcareous sands and sandy tills overlaying granitic-type bedrock. Within the area south of 52°N, 51% or 911,089 km^ is considered as having a low potential to reduce acidity of atmospheric deposition prior to entering surface waters. Terrain Characteristics of Three Specific Study Areas Terrain classes are based on bedrock geology, percent bedrock exposed, soil depth and soil texture or depth to carbonate (Table 3-8; Section 3.5). Table 3-29 shows the terrain classes for watersheds within which the study lakes and rivers occur. These results have been obtained by directly overlaying the watershed areas for Algoma, Muskoka-Haliburton and Nova Scotia onto Figure 3-9. By far the greatest proportion of each area is composed of terrain classes interpreted as having a low potential to reduce the acidity of atmospheric deposition (69 to 98%). The most complex area and the one with the greatest range of terrain conditions is Algoma which has up to 69% has a low potential to reduce acidity, 25% interpreted as having a moderate potential, almost 5% with a high potential to reduce acidity and less than 1% organic terrain (which has not been interpreted). Two low potential terrain classes dominate in each area. In Algoma and Muskoka-Haliburton these are the L3 (41.79% and 59.42%, respectively) and the L4c (21.05% and 32.25%, respectively) classes; in Nova Scotia these are the L4b (53.15%) and L4c (27.11%) classes. ------- 3-186 1 £ 1— I P M U U H P O P 53 W M Pd > O o &> H PM rc H P o ss W -, H H cS S W l-l P S s P£ W II a Fn 0 « Ed O O l-l W II o tn U ^ CM 1 W f£ H ^ n ua IH £ ~ in ic -o L. I/I U O. — *> Hi u go*- *L- C ID 3 r~ m "io <» - i- 1 IP L. f~ .Q * •o 1 fe £ z o -' ^ •^ i 0 £1 0 1 s CO* O 0_ « " O J) L «3 - z * "§ "o v Is "o *• i" z in « •o •55 P — •£ "~ ^ g i ?! 0 £ I5 Z a. •5* si 18 ; K •5 5 (t z R e 1" z *i IS ,s * "5 1 £ z in 0 °1 £" z R . !| i" |f m • t| °§ 1 «! 1 -I * H- "o — w 0 t * »I 1 O c ^ M 8 0 — » S i % "* ? S ** M % ° 1 W rsl (SI " *. 1 ~> £- CMe ^ 0 " M M 1 0 * w M % ^ S g : *. ^. S S S in s s g o -a- — r- -a- .0 (N 0* 0* 0* to* ua" ^ "*. ". ~. °. ^ *J °. 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CMC- g „ - SoS55S2^S25KSS 5 5 552 S =So.SXSS.SRSS S S rS 5 S S S S E2 P g inoOOOOOOOOOO-t*!^ r-- O OOO 0 0-0-*-000 ^ r- oOOOOO-CMO^CMO s|£38 = §|5|sss3 s a s?g 2 sssssg^gss g SsSsSsssSSs- KS »-- «SS S « R°"° -« = S 2 ' =K-SS _ 1 KJ^UOI^-CTJI ,gnm.Qij - fflo-o^ ^^c^o^CTi-^^^^^a^^^^^^.Q^.O^ £ ^cj^o^^^^o^^^^j ^ 1 E 1 ? 1 ?! 1 ^ 1 „' O tn 0 o - o o a K s s o K S CM K. (N 0 ON CM » __- ^ OD — oo in o S s i s R o i K o CD r- AJ S (SI* r- in g s; 3 s i - o-o-»o oooo iSSI" S5SR s £ "^ ° J?£^^(iS(ii^^^^ 0000000000 c o Q L. O K V - Q I 1 ------- I I I I I I I I I I I I I I I I I 1 I 3-187 TABLE 3-28. SUMMARY OF TERRAIN TYPES AND POTENTIAL TO REDUCE ACIDITY FOR ALL OF EASTERN CANADA Terrain Types (Potential to Reduce Acidity) Hla Hlb Hie Hie Hlf Hlg H1h Hli Hlj H2a H2b H3a H3b H3c Total High Potential Mia Mic Mid Mlf Ml i Mlj Mln Mlo Mlp Mlq Mir Mis Mlt South of (area = km2 43,632 65,690 8,105 1,004 109 7,989 5,305 8,959 1,405 1,283 3,237 15,933 61,555 83,914 308,390 86 20 174 48 3,956 698 52'N Latitude 1,779,436 km2) % of Zone 2.45 3.71 0.46 0.06 0.00 0.45 0.30 0.50 0.08 0.07 0.18 0.90 3.46 4.72 17.33 <0.01 <0.01 0.01 <0.01 0.22 0.04 North of (area = 1 km2 35,258 3,034 2,230 4,634 3,915 17,506 66,577 7,018 83 5,523 9,615 2,325 489 1,038 3,114 566 10,389 374 1,520 52 °N Latitude ,357,595 km2) % of Zone 2.60 0.22 0.16 0.34 0.29 1.29 4.90 0.52 0.01 0.41 0.71 0.17 0.04 0.08 0.23 0.04 0.77 0.03 0.11 Total for (area = km2 78,890 65,960 8,105 1,004 109 7,989 5,305 8,959 1,405 4,317 5,467 20,567 65,470 101,420 374,967 7,104 83 5,543 9,615 2,325 489 1,038 3,114 740 48 14,345 374 2,218 Eastern Canada 3,137,031 km2) % of Eastern Canada 2.51 2.10 0.26 0.03 < 0.01 0.25 0.17 0.29 0.08 0.14 0.17 0.66 2.09 3.23 11.95 0.23 < 0.01 0.18 0.31 0.07 0.02 0.03 0.10 0.02 <0.01 0.46 0.01 0.07 ------- 3-188 TABLE 3-28. CONTINUED Terrain Types (Potential to Reduce Acidity) M1u Mlv M2a M2b M3 M4a M4b M4c M5 M6a M6b M7a M7b M7c Total Moderate Potential Lib Lie Lid Lie L2a L2b L3 L4a L4b L4c L4d Total Low Potential South of 52'N Latitude North of 52'N Latitude Total for Eastern Canada (area = 1,779,436 km2) (area = 1,357,595 km2) (area = 3,137,031 km2) km2 82 982 13,662 1,564 117,987 9,382 7,749 10,104 32,237 18,023 83,473 46,831 345,058 143 5,064 914 705 2,110 369,467 11,226 109,262 386,090 21 911,089 % of Zone km2 <0.01 0.06 0.77 0.09 6.63 0.53 0.44 0.57 1.81 1.01 4.58 2.63 19.39 0.01 0.28 0.05 0.04 0.12 20.76 0.63 6.14 21.70 < 0.01 51,20 415 14 46,726 13,670 84,180 6,027 4,051 19,060 17,523 22,933 17,973 274,626 3,322 6,395 75,736 9,491 46,755 40 157,723 4,369 52,247 290,162 788,920 % of % of Zone km Eastern Canada 0.03 0.01 3.44 1.01 6.20 0.44 0.30 1.40 1.29 1.69 1.32 20.23 0.24 0.47 5.58 0.70 3.44 <0.01 11.62 0.32 3.85 21.37 58.11 415 14 82 982 60,388 15,234 202, 167 9,382 13,776 14,155 51,297 35,546 104,406 64,804 619,684 3,322 6,538 80,800 10,405 47,460 2,150 527,190 15,595 161,509 676,252 21 1,700,009 0.01 < 0.01 < 0.01 0.03 1.93 0.49 6.44 0.30 0.44 0.45 1.64 1.13 3.33 2.07 19.76 0.11 0.21 2.58 0.33 1.51 0.07 16.80 0.50 5.15 21.55 <0.01 54.19 I I I I I I I 1 I I I I I I 1 I I i i ------- I I I I I I I t I I I I I I t I I I I 3-189 TABLE 3-28. CONTINUED Terrain Types (Potential to Reduce Acidity) Ola Olb Olc Old 02a 02c 02d 03a 03c 03d Total Organic Terrain South of (area = km2 51,349 15,799 40,598 106,519 34 170 48 55 327 214,899 52'N Latitude 1,779,436 km2) % of Zone 2.89 0.89 2.28 5.99 <0.01 0.01 <0.01 <0.01 0.02 12.08 North of (area = km2 154,399 24,627 12,351 35,681 207 207 227,472 52*N Latitude 1,357,595 km2) % of Zone 11.37 1.81 0.91 2.63 0.02 0.02 16.76 Total for (area = km2 205,748 40,426 52,949 142,200 34 207 377 48 55 327 442,371 Eastern Canada 3,137,031 km2) % of Eastern Canada 6.56 1.29 1.70 4.53 < 0.01 < 0.01 0.01 0.01 < 0.01 0.01 14.10 ------- TABLE 3-29. Terrain Class L2b L3 L4a L4b L4c Total L M4b M7a M7b M7c Total M Hlb Hlc Hli Total H Ola Olc Old Total 0 Study Area TERRAIN CHARACTERISTICS OF WATERSHEDS STUDY AREAS OF EASTERN CANADA Algoma Muskoka-Haliburton km2 % km2 % 116.1 6.52 3,380.7 41.79 1,058.1 59.42 283.9 3.51 225.8 2.79 1,703.2 21.05 574.2 32.25 5,593.6 69.14 1,748.4 98.19 1,838.7 22.73 109.7 1.36 25.8 0.32 116.1 1.44 19.4 1.09 2,090.3 25.85 19.4 1.09 45.2 0.56 264.5 3.27 77.4 0.96 387.1 4.79 12.9 0.16 6.5 0.08 12.9 0.72 19.4 0.24 12.9 0.72 8,090.4 1,780.7 3-190 CONTAINING THE DETAILED Southwest Nova Scotia km2 % 154.8 1.36 6,045.2 53.15 3,083.9 27.11 9,283.9 81.62 1,419.4 12.48 1,419.4 12.48 509.7 4.48 161.3 1.42 671.0 5.90 11,374.3 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ------- I I I I I I I I I f I I I I t I I I I 3-191 It is assumed that these classes represent the terrestrial geochemical template for the three study areas. The other "low" potential classes are very minor in these watersheds and one would expect little or no effect of acidic deposition in basins dominated by "moderate" and "high" potential templates. In the Muskoka-Haliburton watersheds, nine of the lakes and associated tributary streams which have been monitored closely occur entirely within the L3 class. Detailed lake basin mapping by Jeffries and Snyder (1983) for 6 of the lakes indicate that this L3 class is predominately composed of their "Minor Till Plain" and "Thin Till" classes overlaying gneiss bedrock. These two surficial types represent between 84.3 and 94.0% of the basins of Red Chalk, Blue Chalk, Chub, Dickie, Harp and Jerry lakes. The three dominant terrain classes in these study areas (L3, L4b and L4c) are composed of the following: (1) L3 - shallow sands and acidic type rocks (granite, gneiss, quartzite or other alkalic rocks) which outcrop in 0-49% of the map area; (2) L4b - deep sands overlaying ultramafic, serpentine and noncalcareous sedimentary rocks outcropping in 0-49% of the unit; and (3) L4c - deep sands overlaying bedrock similar to L3. These classes represent dominant conditions in a map area. At this scale of mapping (1:1,000,000) other subdominant conditions probabl} occur. However, the evidence from more detailed mapping at Muskoka-Haliburton, as described above, indicates that the descriptions are representative. It should be noted further that the term "sands" refers to the matrix texture; the deposit it represents is most commonly a till or glacial-fluvial outwash which include larger sized fragments. Results of Terrain Extrapolation Table 3-27 provides the basis of extrapolation by province and Table 3-28 for all of eastern Canada. Terrain classes L3, L4b and L4c, which represent the major geochemical templates for the watershed study areas, are three of the four most common terrestrial types. In eastern Canada, they cover 17% (527,190 km2), 5% (161,509 km2) and 22% (676,252 km2) respectively (Table 3-28). They represent over 80% of the sensitive terrain types in Eastern Canada. Other classes which cover significant areas but are not represented in the study areas are H3c (deep clay overlying granitic rocks), Ola (organic deposits overlying carbonate rocks), Old (organic deposits overlying granitic rocks), and L2d (shallow sand overlying granitic rocks with 50-74% outcropping). Approximately one-fifth of eastern Canada (690,117 km2) currently receives loadings of about 20 kg/ha.yr or more of SO^2" in precipitation in 1980. Within this loading zone terrain classes L3, L4b and L4c cover 18% (127,237 km2), 6% (40,222 km2) and 22% (153,545 km2) respectively. In total, the three terrain types cover 46.52% of eastern Canada within the 20 kg/ha.yr, or higher, ------- 3-192 AQUATIC I I loading of SO in precipitation. This is an area of 321,004 km2 (125,192 mi2) which represents 99% of all those areas with the lowest potential to reduce acidity within this loading zone. fl These areas occur primarily on the Grenville Province of the I Precambrian Shield in southern Quebec and Ontario as well as in the Appalachian Region of New Brunswick and Nova Scotia (Figure 3-9). • These results indicate that over one-half of eastern Canada, is representative of terrain characteristics (Table 3-8) under which aquatic acidification effects have been observed. • From these results, it is concluded that terrain characteristics in the three watershed study areas are correlated with measured ft acidification effects, especially as expressed by alkalinity 0 measurements. These three study areas are not anomalous but are representative of larger portions of Eastern Canada as defined by mm these terrain characteristics. • 1 I As shown in the previous section the bedrock and surficial geology of the study areas are typical of large areas of eastern Canada. However specific watersheds with varying glacial deposits (kame, spillway, till, etc.) rock component hardness (i.e., resistance to ^ weathering) and varying hydrological characteristics result in W surface waters of varying alkalinity and total cation concentrations within each study area. Hydrochemical data from the Muskoka-Haliburton area of Ontario also ml compares closely with mapped terrain conditions. Average annual and spring T.I.P. alkalinity values for 9 lakes within the Muskoka- • Haliburton study area are all lower than 71yeq/L (Table 3-30). Five jj of these lakes are considered very sensitive on the basis of their alkalinity regime (<40 peq/L). The basins of all 9 lakes are _ composed primarily of shallow to deep (<2 m) sandy tills overlaying • gneiss (class L3 and L4c). In addition there is a close correlation between terrain class and alkalinity regime for a population of 141 lakes sampled throughout Haliburton County and Muskoka District. M Table 3-31 shows the occurrences of lake alkalinities grouped by v sensitivity classes, in each of the mapped terrain types. There is clearly a strong relationship with 77.5% of the lowest alkalinity J| lakes (0-39.9 and 40-199.9 yeq/L) occurring in terrain classes L3 and J L4c. It is not possible, at present, to extrapolate the results of Table 3-31 to all the areas of eastern Canada mapped in these two — terrain classes. I Further support for the representativeness of the study areas is drawn from the water quality data. Figures 3-49 and 3-50 show the 4| distribution of lake alkalinities for a series of geographical areas • on sensitive and moderately sensitive terrain. The data are taken I I ------- I I I I I I I I I I I I I I I 1 I I I 3-193 TABLE 3-30. AVERAGE ANNUAL OR SPRING TOTAL INFLECTION POINT ALKALINITIES FOR NINE LAKES IN THE MUSKOKA-HALIBURTON WATERSHED STUDY AREA (data from Ontario Ministry of Environment) Lake Time of Record Alkalinity mg/L yeq/L Harp Dickie Chub Red Chalk Blue Chalk Jerry Plastic Heney Crosson 1979-80 1979-80 1979-80 1979-80 1979-80 1979-80 Spring/79 Spring/79 Spring/80 3.32 0.762 0.798 3.15 3.53 3.31 0.62 +; 0.5 0.34 +_ 0.5 0.49 + 0.5 66.4 15.24 35.96 63.0 70.6 66.2 12.4 H- 10.0 6.4 + 10.0 9.8 + 10.0 ------- TABLE 3-31. 3-194 DISTRIBUTION OF 141 LAKE ALKALINITIES, GROUP BY SENSITIVITY CLASSES, IN VARIOUS TERRAIN TYPES OCCURRING IN HALIBURTON COUNTY AND MUSKOKA DISTRICT, ONTARIO Terrain Class L3 L4C L2d L2b Hlc Hli M4b M7c Old Map Area (km2) 4283.9 1645.2 206.5 141.9 109.7 51.6 25.8 45.2 25.8 Alkalinity 0-39.9 40-199.9 34 (24.1) 61 (43.4) 5 (3.5) 9 (6.4) 3 (2.1) 3 (2.1) 4 (2.8) 3 (2.1) 2 (1.4) Classes (yeq/L) 200-499.9 500 3 (2.1) 6 (4.4) 7 (4.9) 1 (0.7) Total 6535.6 46 (32.5) 78 (55.4) 11 (7.7) 6 (4.4) 1 I I I 1 I I I I 1 I I I I I 1 I I I ------- I I I I I I I I I t I I I I I I I I I M M _ra O >> +•• 'c O 03 0 0) JC ca 100 80 60 40 20 0- 100 80 60 40- 20 0 100 80 60 40 20 0 100 80 60 40 20 0 100 80 60 40 20 0 100 80 60 40 20 0 <0 ^0-39^^40-199 200-499 >500 3-195 BRUCE AND GREY COUNTIES n=10 <0 ' 0-39.9 40-199 200-4991 >500 <0 ' 0-39.9 40-199 200-499 >500 <0 0-39.9 40-199 200-499 >500 <0 '0-39.9 40-199 200-4991 >500 HALIBURTON COUNTY n=197 MUSKOKA DISTRICT n=159 KENORA DISTRICT (S. of 51° Lat.) RAINY RIVER DISTRICT n=99 Figure 3-49. ALGOMA DISTRICT n=449 <0 0-39.9 40-199 200-499 >500 Alkalinity (peq/L) Distribution of alkalinity values for lakes in six regions on Ontario. ------- 3-196 o •H tO 4J <§ 4-1 O C O CO CU 0> co ^ CU O 3 C •H to M-l o o CO • •H ^-N •a eg tX5 > -H •H 4-" C to o iH iH 3 i-l O in i CO I I I I 1 I I I I 1 I I I 1 1 I 1 I I ------- I I I I I I I I I I I I I I I I I I I 3-197 from Table 3-12. The percentage distribution of lake alkalinities are similar in all areas and contrast strongly with the alkalinities of 10 lakes in Bruce and Grey Counties which are located on calcareous till in southern Ontario (nonsensitive terrain). While the alkalinity distributions are similar, there are some important differences. The distributions for Haliburton, Muskoka, and Algoma have already been altered in that there is a greater number of lakes with low alkalinity than in the Kenora or Rainy River Districts or in background areas such as Northern Saskatchewan. Dillon (1982) further demonstrated the differences in alkalinity values for lakes in the areas of higher sulphate deposition (Muskoka- Haliburton and Parry Sound) by plotting the cumulative distributions (Figure 3-50). It is accepted that alkalinity distributions are already influenced by acid loadings in some areas and to reflect natural conditions the distributions should be shifted to the right as plotted in Figures 3-49. Within each study area, the number of lakes for which detailed data are available is small relative to the total number of lakes. Therefore, it is important to show that the intensive study lakes and rivers themselves are representative of the surface waters of the sensitive areas. There are a total of 18 calibrated study lakes at ELA (1), Algoma (5), Muskoka-Haliburton (8), Quebec (1) and Nova Scotia (3). The current alkalinities show 2 less than 0 peq/L, 7 in the 0-40 yeq/L range and 9 in the 40-200 yeq/L range. Lakes above 200 are not subjected to intensive studies since acidification effects are minimal. In addition, Ontario has extensive information on five calibrated lake studies near the point sources in Sudbury which is used to contrast effects of local sources and long range transport. Of the 22 rivers in Nova Scotia used in analysis of salmon catch data, current alkalinities range from less than zero (acidic) to 173 yeq/L (Figure 3-47). The study lakes and streams are located in areas with terrain characteristics and have alkalinity values similar to other sensitive areas in Canada. Therefore, the effects observed in the study lakes and rivers in response to specific loading rates should be similar in other water bodies in these sensitive areas. Similarly, loading rates protective of these study lakes should be protective of other sensitive waterbodies. Possible Magnitude of Effects The Canadian members of the Work Group have concluded that an indication of the extent of the current water quality effects may be derived for all of Ontario using the information presented in Section 3.6.1. The Precambrian area east of Algoma contains some 50,000 lakes (Cox 1978). The distribution of alkalinity values for lakes in districts within the 20 kg/ha.yr wet SO^~ deposition isopleth (from Table 3-12) indicates that about 20% or about 10,000 lakes have alkalinity values and acid loadings that are combining to ------- 3-198 3.9.3 Use of Acidification Models I I currently cause pH depression to values (less than 5.5) likely to be causing biological damage. Cox (1978) has indicated that the lake counts underestimate the • number of lakes with surface areas less than one hectare by as much as a factor of three so the 50,000 and 10,000 numbers are both underestimates. I 1 The data for the 57 headwater streams in Muskoka-Haliburton show that 65% experience minimum pH values less than 5.5 and 26% have minimum pH values less than 4.5 (Figure 3-21). Although the total number of miles of streams within the 20 kg/ha.yr wet SO^- deposition isopleth is not known and quantitative extrapolations are not « possible, it is clear that many miles of stream water must also • currently experience pH depressions to levels that can potentially cause biological damage. There is a larger area of lakes underlain by Precambrian rock in V Quebec and the Maritime provinces where the acid loadings are at least as much as those at Algoma. While specific lake count data are • not available, it is likely that many thousands of lakes are || currently receiving acidic deposition. In both Ontario and Quebec many more thousands of lakes are slightly • less sensitive to acidic deposition and may experience biological * damage in the future if the acid deposition continues. _ Precambrian areas of eastern North America is measured in the tens of thousands with even more sensitive to effects in the future. • The U.S. members of the Work Group believe the statements in this section cannot be supported by the facts. The combined analysis of _ lake survey data, terrestrial mapping data and deposition data is an • interesting methodology. Pending validation, the U.S. members have ™ too many concerns about the influence of uncontrolled variables to consider its use more than speculative. One important variable is I the level of dry deposition from local sources which can affect the • representativeness of the survey lakes. Other factors which may determine the overall neutralizing capacity of a watershed system in • addition to terrain class include elevation, hydrologic routing time, \l lake morphometry and vegetative cover. We therefore cannot support the conclusions in this section in the absence of further _ methodological validation. • I A number of process-oriented (mechanistic) models have been developed (or are under active development) that simulate in detail the flow • of acidic precipitation through terrestrial systems and the resulting • chemical response of surface waters. These models have the potential I I ------- I I I I I I I I I I I I I I I I I I I 3-199 to predict stream and lake responses (e.g., pH depressions) to episodic events, but most of them are not suitable for predictions of long-term ecosystem responses. Examples of these process-oriented models include the ILWAS model (Chen et al. 1982), the Birkenes model (Christophersen et al. 1982), and the trickel-down model (Schnoor et al. 1982). Each of these models has achieved some success in relating short-term variations in water chemistry of small drainage basins to hydrology and chemistry of precipitation. These models, while calibrated for specific watersheds have not been validated on a temporal or spatial scale that permits their general application with significant confidence. More global modeling efforts, such as those of Hough et al. (1982), and USFWS (1982) have formulated detailed mechanistic submodels but have not developed them to the level of working codes. Thus, prediction of the dynamic response at the aquatic regime to the atmospheric loading remains to be achieved at this time. However, several efforts towards development of empirical or semi-empirical steady state models relating aquatic chemistry to the atmospheric loading stress have advanced to the point that response estimates are possible within the limits of assumptions of the models. Three important general points must be made about these models. First, validation (especially for surface waters in North America) remains to be achieved. Second, each of these models is based upon individual, specific sets of asumptions regarding their application. Application of these models is therefore limited by the degree to which these assumptions are met. Third, these models are not dynamic and therefore, determination of the rates of reaction between sulphate deposition and lake water pH based on the models is not possible. The models rely on steady state conditions. With these important points in mind, potential use of these models for quantitative estimates of the relationship of SO^- deposition to lake pH is discussed below. The earliest empirical acidification model was developed by Aimer et al. (1978) and modified by Dickson (1980, 1982), who related lake pH and excess SO^" load (concentration of excess SO^" multiplied by annual runoff) for arbitrary classifications or groupings of Swedish lakes. Since this relationship is, in effect incorporated by Henriksen (1979, 1980, 1982) in his model, it will not be discussed in detail here. 3.9.3.1 The "Predictor Nomograph" of Henriksen Henriksen (1979, 1980) has studied atmospheric and edaphic influences on the chemistry of oligotrophic lakes in Scandinavia and has developed empirical formulations relating these influences to acidification. He has derived an acidification "indicator," a quantitative acidification "estimator," and an acidification ------- 3-200 I I "predictor nomograph" (Henriksen 1979, 1980). Of these formulations, only the "predictor nomograph" is intended for use as a predictive tool. • Henriksen (1980) developed his "predictor nomograph" of freshwater acidification based on the hypothesis that "acidified waters are the M result of a large scale acid base titration." He compared the • concentration of "Ca* + Mg*" with lakewater 804* concentrations (* indicates "excess concentration" — that above contributions from seasalts) in the pH range 5.2-5.4 and 4.6-4.8 using data from 719 • lakes in southern Norway (Wright and Snekvik 1978) and obtained ™ "highly significant" linear correlations. The line for the pH range 5.2-5.4 agreed very well with a theoretical titration nomograph of l| bicarbonate concentration vs. (H+ added), assuming that bicarbonate j| concentration is directly proportional to (Ca* + Mg*) concentration and that (H+ added) is proportional to 804* concentration. The _ line for the pH range 4.6-4.8 did not agree with such a theoretical • bicarbonate titration nomograph, but Henriksen (1980) argued that his deviation was readily explained by the effects of dissolved aluminum leached from soils. To complete his predictor nomograph, Henriksen I (1980) added a Ca* concentration axis parallel to the (Ca* + Mg*) I axis and a precipitation pH axis parallel to the 804* axis (Figure 3-51). The former was derived from correlations of Ca* • concentrations and (Ca* + Mg*) in lake waters; the latter was derived • by combining: (1) a correlation of 864* concentration in lake water to 804* concentration in precipitation, and (2) a correlation of ^ 864* concentration in precipitation to H+ concentration in V precipitation. Henriksen (1982) added an axis of 864* in • precipitation parallel to the axis of 804* in lakewater based on his 1980 regression. • Henriksen (1979, 1980) derived his predictor nomograph for pristine, oligotrophic lakes in areas with granitic or siliceous bedrock types • and thin podsolic soils. In these lakes that have been receiving • acidic deposition, 864^" is the major anion. Prior to the advent of acidic deposition, Ca2+ and HC03~ were the dominant _ ions in these lakes. Lakes used to develop the relationships had low fl concentration of organic acids. The lakes ranged in area from ' 0.1 to 30 km2 and in 90% of the lakes the Ca+2 concentration was less than 80 yeq/L. None of the lakes was on a major river H (i.e., had very large watersheds) (Wright and Snekvik 1978). I Henriksen (1980) verified the predictor nomograph with an independent • data set from an October 1974 survey of 155 Norwegian lakes (Wright • and Henriksen 1978). These lakes ranged in area from 0.25 to 1.0 km2, occurred at the head of undisturbed watershed drainage basins, and constituted 5% or more of their watersheds (Wright and • Henriksen 1978). Henriksen (1980) found that for over 85% of the ™ lakes, the nomograph correctly predicted a pH "grouping" (pH<4.7 — "acid lakes", 4.7 pH <5.3 — "transition lakes", pH>5.3 — • "bicarbonate lakes"). He also found that the nomograph was valid for | 18 "large lakes" in southern Norway. I I ------- | 3-201 1 1 • 300- ™ § 200- 3- ^H ** I *<» + §*« 0 _ 100- 1 vv 1 0J 250- 200- 3 150- "17 CD 3- *«" 0 100- 50- 0- &'' X X X X / 1 X ft,.1 x HCOo- Lakes xx ^xx O ' Vx" X ^^ X^ x- ^/^ ^^ .X ^x^ / ^^ Acid Lakes >x ^^ //^ ////^ 1 1 1 1 1 1 1 1 1 1 1 0 100 200 • SO4* in Lakewater, fyieq/L) i i i i i i i i i i i i i i 10 50 100 SO4* in Precipitation, fyieq/L) 1 i i i i i i i i 7.0 5.0 4.7 4.5 4.4 4.3 4.2 4.1 4.( _ pH of Precipitation 1 • Figure 3-51. Nomograph to predict the pH of lakes given the sum of ^ nonmarine calcium and magnesium concentrations (or nonmarine calcium concentration only) and the nonmarine • sulphate concentrations in lake water (or the • weighted-average hydrogen ion concentration in precipitation) (Henriksen 1982). 1 1 ------- 3-202 I I Henriksen (1980) concluded that the nomograph could successfully predict lake pH changes in response to changes in the pH of the precipitation of the particular composition for that area and, if the m titration process of lake acidification is reversible, the nomograph • could be used to indicate the amount of decrease in precipitation acidity necessary to restore acid lakes to bicarbonate lakes. _ A number of assumptions and cautions pertain to the use of the ™ predictor nomograph. One assumption initially inherent in the predictor nomograph was that increases or decreases in the acidity of 4 precipitation do not affect the rate of leaching of Ca^+ or | Mg2+ from soils. As Henriksen (1980) noted, this is a matter of debate (e.g., see Aimer et al. 1978; Dillon et al. 1979) and a m question that "certainly deserves further attention." If, for • example, increased acidity of precipitation does cause increased cation leaching from soils (instead of decreased lake pH), then the titration hypothesis on which the nomograph is based is violated and • extrapolations from the precipitation pH axis will be incorrect. ™ Henriksen (1982) has performed further research on this particular 4 problem. Using data from lakes in Norway, Sweden, Canada, and the 41 U.S., he: (1) compared historic and recent concentrations of (Ca* +Mg*), and (2) evaluated ranges of (Ca* + Mg*) concentrations • in lakes in similar geologic settings over a gradient of acidic • deposition. In some cases he found that (Ca* + Mg*) concentrations increased in conjunction with higher levels of acidic deposition. In other cases he found no such concurrent increases. For the data he V examined the maximum value of a "base cation increase factor" for the • lake waters would be about 0.4 yeq (Ca* + Mg*)/yeq 804* (Henriksen 1982). Thus, estimates of the effect of changes in acidic deposition • on the chemistry of lake waters still require knowledge of the degree | of increase of base cation concentrations, ranging from 0 ueq (Ca* + Mg*)/yeq 804* to roughly 0.4 yeq (Ca* + Mg*)/yeq 804*. • This applied to certain lakes in Sweden, Norway, and North America • where there was enough historical information to make an estimate. However, he does state (p.38, Henriksen 1982) for Lake Rishagerodvatten, Sweden, the factor was 0.63, and the Birkenes model • (Christophersen et al. 1982) predicts an increase factor of about • 0.55. Dickson (1980) showed increases greater than 0.4 for some Swedish west coast lakes. • The increase factor represents possible responses of the watershed system to acidic deposition. It reflects the geologic and hydrologic ^ sensitivity of the system. The lowest limit of the increase factor • is zero, which refers to a system with little base exchange capacity * in the organic soil, quartz (Si02) sands for the mineral soil, and/or a lake in which precipitation does not flow through soils. • Perfect seepage lakes without any drainage area other than lake area V would qualify as systems with near-zero increase factors based on the lack of flow through neutralizing soils. The maximum upper limit • would be a watershed with calcareous soils or bedrock which would | serve as a perfect buffer and yield an increase factor of 1.0 yeq (Ca + Mg)/peq 8042~. _ I ------- I I I I I I I I I I I I I I I I I I 1 3-203 The leaching of aluminosilicate minerals in response to hydrogen ion attack has been studied in the laboratory. Wollast (1967) found a dissolution increase factor of 0.33 initially with respect to hydrogen ion attack in 5% K-feldspar solutions. Furrer and Stumm (1982) found a 0.4 increase factor in the dissolution of A^C^. The factors that control a watershed's neutralizing capacity, and hence the cation increase factor, are not well known and are critical. A second caution noted by Henriksen (1980) is that the predictor nomograph should not be applied to waters containing high concentrations of organic acids. Not only may the organic acids affect lake pH in a manner independent of precipitation acidity, but also ionic Ca^+ and Mg^+ may be overestimated inasmuch as analyses for these ions include Ca^+ and Mg^+ bound to organics (Henriksen 1980). A final point to note is that the derivation and verification of this model is based upon the premise that the observed data represent steady state conditions, both for concentrations and pH in deposition, and concentrations and pH in lake water. A key question is whether the "predictor nomograph" is applicable to sensitive lakes in northeastern North America. Relationships between Ca* and (Ca* + Mg*) and between concentrations of these cations and 80^* may be different in regions of varying geochemistry in North America. Furthermore, the empirical relationship between SO^* in lake waters and 804* in precipitation (as well as the relationship between SO^* in precipitation and pH of precipitation) may vary in different geographical regions. Therefore, for more accurate predictions it would be appropriate to develop region-specific regression relationships and predictor nomographs like Henriksen's from data bases for the regions of interest. Such studies would be a useful extension of Henriksen's model and should be pursued. Church and Galloway (1983) examined data from two small oligotrophic headwater lakes in the Adirondacks and found, using only the (Ca* + Mg*) and lake water (SO^*) axes, that the nomograph correctly predicted the pH for all 66 measurements in a "bicarbonate lake" and 71% of 78 measurements for an "acid-transition lake". However, they also found that the relationship between the precipitation pH axis and lake water (864*) axis for the Adirondacks differs significantly from the relationship for southern Norway. This is possibly due to the different contributions of nitric and sulphuric acids to precipitation acidity or to the presence of other cations in precipitation in the two geographic regions. The variation of nitric and sulphuric acid contribution to acidity of precipitation has been further shown by Barrie (1982). Because as shown in Section 3.9.1, nitrate has only minor influences on long-term acidity of the aquatic regime in comparison with sulphate, only the relationships to sulphate loading are considered in this section. For water pH values less than 4.7, the presence of aluminum or of other buffering apparently becomes important as shown by Henriksen (1980, 1982) and ------- 3-204 I I may affect regression lines. However, we are more concerned with the "transition" sector of the Henriksen nomograph. Raines and Akielaszek (1982) examined data from 122 New England lakes • in relation to the predictor nomograph. The nomograph correctly predicted 6 of 7 lakes in the pH range <4.7 but incorrectly predicted flj that 19 other lakes with higher pH values fell in this grouping. The £ nomograph correctly predicted 5 of 14 lakes that fell in the pH range 4.7 - 5.3 but incorrectly predicted that 32 lakes not in this range M had such pH values. Of the 101 lakes in the pH range >5.3, the • nomograph correctly predicted 60%. For those New England lakes the nomograph predicted the pH of acidic • lakes correctly but frequently predicted lower pH values than were • actually observed in higher pH lakes. These differences may occur because the relationships of calcium, magnesium, and sulphate are M different in New England than they are in Norway, where the model was • developed. Application of the predictor nomograph in New England should be based on empirical relationships that exist in this region. — Presently the relationship between lake sulphate concentration and m atmospheric sulphate deposition has not been established for this ™ region. Keeping in mind the important limitations and assumptions inherent in || its use, we have attempted an application of this approach to estimating the effects of SO^" deposition on the chemistry of M lakes in northeastern North America. Numerous lakes in Norway have • calcium concentrations less than 50 yeq/L, and Bobe"e et al. (1982) found that 7.5% (15) of 199 lakes sampled on the Precambrian Shield of the Province of Quebec had calcium concentrations less than • 50 yeq/L. Raines and Akielaszek found that 11% (25) of 226 lakes and ™ streams in New England had calcium concentrations less than 50 yeq/L. This indicates that such a limit would include all except the more M sensitive waters. From the regressions given by Henriksen on: | (1) the relationship of both (Ca* + Mg*) vs. alkalinity and (Ca*) vs. alkalinity I and thus (Ca*) vs. (Ca* + Mg*)I (Henriksen 1980), and -mt (2) the relationship of strong acidity to 804* and (Ca* + Mg*)I • both with and without increased leaching of base cations (Henriksen 1982)1, we can roughly estimate a 804* concentration that yields a pH of 5.3 in surface waters having initial Ca* concentrations of • 50 ueq/L. Using the regression given by Henriksen (1980) on the w relationship of lake 804* concentration to 804* concentration in precipitation and assuming an annual rainfall of 100 cm, we can then fl estimate loading rates consistent with maintenance of a pH of 5.3 or || greater (pH 5.3 is the upper limit of Henriksen's transition zone). The results of such calculations and the regression equations used • are given in Table 3-32. Estimated loading values of wet sulphate • deposition that will maintain lakewater pH at values 2:5.3 range from approximately 26 kg/ha.yr (assuming no increased leaching of base cations) to approximately 43 kg/ha.yr (assuming leaching of base • cations of 0.4 times the change in excess sulphate concentration (see • Henriksen 1980) and an initial lake 804* concentration of 0 yeq/L). • I ------- I I I I I I I I I I I I I I I I I I I TABLE 3-32. 3-205 CALCULATION OF WET SULPHATE LOADINGS CONSISTENT WITH pH 5.3 OR GREATER IN LAKES WITH INITIAL CALCIUM CONCENTRATION OF 50 yeq/L OR GREATER (Regressions are from Henriksen [1980, 1982]) All Units (yeq/L) (except where noted) Ca*i (Ca* + Mg*)i (Ca* + Mg*) (so4*)w p (S04*) (S04*)£ (kg/ha. yr) No Leaching of Base Cations 50 70 70 81 53 26 Condition Leaching (according .4 50 70 128 146 87 43 of Base Cations to Eqn (4) below) .2 .1 50 50 70 70 121 114 138 130 83 78 41 39 Ca*i (Ca* + Mg*)± (Ca* + Mg*)p (so4*)w (S04*)p (so*)L concentration of excess sulphate in lake water prior to "acidification" (i.e., initial S04* concentration) initial excess calcium concentration initial sum of excess calcium plus excess magnesium concentrations predicted sum of excess calcium plus excess magnesium concentration final concentration of excess sulphate in lake water concentration of excess sulphate in precipitation areal wet sulphate loading assuming annual rainfall of 100 cm Equations Used in Calculations = 1.32 (Ca*)± + 4.3 (adapted from Henriksen 1980) = [1.01 (Ca* + Mg*) + 1.81/0.9 (assuming no leaching of base cations; Henriksen 1982) = [1.01 (Ca* + Mg*)p + 1.8J/0.9 (assuming maximal leaching of base cations; Henriksen 1982) = (Ca* + Mg*)± + 0.4 ( S04*)w (Henriksen 1982) = (Ca* + Mg*)i + 0.4 (S04*w - (1) (Ca* + Mg*)± (2) (S04*)w (3) (S04*)w (4) (Ca* + Mg*) (5) (Ca* + Mg*)p Substituting Equation 5 into Equation 3 and solving for (S04*)w yields (6) (S04*)w (7) (S04*) (8) (S04*)£ = 2.04 (Ca* + Mg*)± + 3.64 - 0.82 (S04*)i = KS04*) + 191/1.9 (Henriksen 1980) = (S04*) /2 (assuming 100 cm annual rainfall) ------- 3-206 I I 3.9.3.2 Cation Denudation Rate Model (CDR) Thompson (1982) developed a model relating the pH of a river to the • atmospheric loading of excess sulphate and the rate of cations from a watershed via runoff (the Cation Denudation Rate or CDR). This model is designed to apply to areas with acid-resistant bedrock, fl till, and soils and relatively unbuffered surface waters. V In most fresh waters the sum of base cations (Ca+2, Mg+2, Na+, K+) • closely approximates the sum of anions HC03~ and SO^" after • correction for seasalt or road salt contributions. Thompson (1982) noted that if excess sulphate concentration is plotted against the ^ sum of the cation concentrations, a series of lines can be generated, • each line representing constant bicarbonate concentration. If the ™ partial pressure of CC>2 (Pco2) ^n tne surface water in question is constant, then each line also represents constant pH. • This model may be applied to either rivers or lakes (Thompson 1982; fl Thompson and Hutton 1982). If a runoff value of 1 m/yr is assumed and the concentrations of terms in the axes of Figure 3-52 are ^ multiplied by this value, the axes become loading rates, and the • figure becomes a plot of cation denudation rate (CDR, meq/m^.yr) versus the discharge rate of excess sulphate (Thompson 1982). If all the atmospheric sulphate deposited on the watershed is contained in • runoff and if we assume that all non-seasalt sulphate comes from W atmospheric loading, then the abscissa is equivalent to atmospheric loading of acid sulphate. Note that if wind-blown dust has neutralized some of the sulphuric acid in atmospheric deposition, the loaing terms in Figure 3-52 must be corrected for these neutral salts. Thus, according to the model, if CDR, runoff, excess sulphate _ load, and Pc02 are known, mean pH can be estimated. • An example of how model calculations are made is given below. If the rate of excess SO^" loading is less than the CDR by 20 meq/m^.yr • (i.e., the HCC>3~ residual equals 20 meq/m2.yr), the model estimates • that the resultant runoff water (assuming a yield of 1 m/yr) will have a mean pH of 5.6 (Figure 3-52). As the rate of excess • 504^" loading approaches the CDR, the runoff water will approach a f pH of 5.1 (at which HC03~ alkalinity is totally exhausted). Data for very soft water rivers in Nova Scotia and Newfoundland that have _ mean runoff rates near 1 m/yr are shown in Figure 3-52. These rivers • have a total CDR ranging from 55 to 200 meq/m^.yr. In 1973 at * least three of these rivers received SO^" loads exceeding their CDR and had median pH values less than 5.1. M At first glance the CDR model appears to be quite similar to the predictor nomograph of Henriksen. The CDR model is developed • strictly from consideration of charge balance, however, whereas the • predictor nomograph is strongly dependent on empirical observations. Thompson (1982) explicitly assumes that CDR is independent of acid — loading; that it varies only with discharge. The recent data review • by Henriksen (1982) shows that CDR cannot be considered to be ™ I I I ------- I I I I I I I I I I I I I I I I I I I 3-207 CO r2.5 PC02=10 RUNOFF = 1m/yr 100 200 ACID LOAD or EXCESS SO4 (meq/m2. yr) (|aeq/L) 2- Figure 3-52. The model plot - pH predicted for consideration of the sum of cations and sulphate (modified from Thompson 1982). ------- 3-208 I I constant in all cases. Thompson et al. (1980) compared data from between 1954-55 and 1973 for very soft water rivers in southern Nova Scotia. They found a lower pH and higher excess 804^" • concentrations in the most recent data but did not find significant | changes in major cation loads. A way in which the CDR model is similar to Henriksen's predictor • nomograph is that it does not apply in situations where organic acids strongly influence pH. The CDR differs, however, in that it does not consider the possible effects of buffers other than bicarbonate. • Also, PCOZ roust be known to estimate pH with the CDR model. As is commonly known, Pcc-2 varies significantly in surface waters. A Raines and Akielaszek (1982) applied the CDR model to data from 122 New England lakes. The CDR model gave better results than the • predictor nomograph (discussed above). Predicted pH agreed very well • with measured pH at values <_6.3. However, this model also predicted lower pH than was measured for many lakes with pH >6.3. As discussed above, estimates of the relationship between sulphate H deposition rates and surface water pH may be made. As an example, the Roseway River, Nova Scotia (Figure 3-53) has a CDR of • 56 meq/m^.yr. If all of the assumptions noted above hold and if • the acidification process is reversible, then a reduction of the sulphate loading rate to 40 meq/m^.yr (20 kg SC>42-/ha.yr) might be expected to return the river to an annual pH of roughly 5.3. • A significant problem exists with such a prediction. The Roseway River has strongly coloured waters, as do the Mersey and Medway Rivers (also shown in Figure 3-53). As Thompson (1982) notes, the pH II values of these rivers "have been thought to be dominated by • naturally- occurring organic acids." Thompson (1982) feels that "their low pHs can be explained quite well on the basis of simple • inorganic chemistry." No chemical data (e.g., Gran titrations for • weak and strong acids) were presented to confirm this. If the pH values of these rivers were controlled by naturally-occurring organic _ acids, reduction of excess sulphate deposition would not result in • the increases in stream water pH predicted. ™ Figure 3-54 and Table 3-33 were calculated based on the Thompson • (1982) model. If 80 yeq/L of cation concentration (roughly f equivalent to 50 ueq/L Ca2+ as used in Table 3-33) is used as a criteria for basin sensitivity to acidification, maintenance of the m basin water to a mean pH >5.3 should be possible with sulphate • loadings of 35 kg S042~/ha.yr given a runoff of 100 cm/yr. Other combinations of sulphate deposition and runoff are shown on Table 3-33. It should be noted that any retention of sulphate within • the watersheds or increased leaching of base cations would violate 9 assumptions in the model, causing the above loading estimates to be too low. • I I ------- I I I I I I I I I I I I I I I I I I I 3-209 200 CT 0) § rr Q <•> 150 100 RIVER CDR EXCESS SO42~ MEDIAN pH Wallace 203 Meteghan 129 Le Have 126 Pipers Mote 101 St. Mary's 100 Tusket 75 N£.Pond 71 Medway 71 Mersey 66 Roseway 56 O 4.3 Figure 3-53. 50 100 150 Excess SO4 meq/m2- yr Cation Denudation Rate Model applied to rivers of Nova Scotia and Newfoundland (Thompson 1982). ------- 3-210 I I 2 400 CO CO £ 360 3 0 C C o ** CO l_ •frrf c 0) o c o o c o 4~ CO o to maintain Aquatic Regime at: __ pH 5.3, i.e., HCOg = 10jjeq/l_ pH 5.8, i.e., HCO = 32jueq/|_ 320 Runoff 30 cm/yr TJ 0 o 280 CD i_ i_ O 3 240 £ 200 3 CO o 160 « •S 120 - 80 40 0 8 12 16 20 24 28 32 36 40 44 Excess Sulphate (kg SCL2 /ha - yr) Figure 3-54. Relation of excess sulphate and cation concentration for pH 5.3 and 5.8 for basin runoff of 30, 50 and 100 cm/yr. The model was developed for an area with 100 cm runoff. It has not been corroborated for areas with lower runoff (derived by the Work Group from Thompson 1982). I I I I I I I I ------- I I 3-211 TABLE 3-33. ACIDIFICATION SENSITIVITY OF SURFACE WATERS RELATED TO SULPHATE LOADING FOR TWO pH OBJECTIVES AND THREE RUNOFF Thompson (1982)] Cation Concentration ( jjeq/L) pH Objective 300 5.3 5.8 200 5.3 5.8 100 5.3 5.8 50 5.3 5.8 Runoff (cm/yr) 30 44 40 28 25 13 10 6 3 50 50 50 47 42 22 17 10 5 100 50 50 50 50 45 34 20 9 • VALUES [derived by the Working Group from CDR model, I I I I I I I I I I I I I I I I The model was developed for an area with 100 cm runoff. It has not been corroborated for areas with lower runoff. ------- 3-212 3.9.3.3 Summary I I The application of two simplified models to the problem of relating • wet deposition of sulphate to lake pH has been discussed in this • section. Before any environmental or water quality model can be used to make estimates with specified confidence of future conditions in a • particular geographic region, the applicability of that model for • that region and conditions must be verified. This process of verification is just beginning for Henriksen's predictor nomograph M and CDR model to northeastern North America. Until such verification • (and perhaps, model adaptation) is achieved, quantitative predictions based on these models must be viewed with caution. I 3.9.4 Summary of Empirical Observation and Modelling to normal or altered fluxes in the hydrologic regime; the regional responses in lake chemistry; the basin characteristic which influence ^ sensitivity to acidification; evidence of changes or trends in • surface water quality in sensitive regions; evidences of alteration ™ of biological components; and finally, the methodologies which are available to assist in estimation of target loadings by atmospheric B deposition which would be consistent with protection of the ecosystem • to a degree acceptable to society. Because environmental concerns are of rather recent recognition and those which have been recognized M are most often related to more intense urban contamination, long term • records of verified significance are available in only a few cases from which firm conclusions can be drawn relating to acidification of remote ecosystems. A deterministic knowledge of the inter- • relationships of the bio-hydrogeochemical system and of its responses ™ to altered precipitation chemistry is not yet available, therefore rendering precise predictive modelling of system responses, as yet, fl| unattainable. These limitations have been thoroughly reviewed in I recent summaries of the Associate Committee on Scientific Criteria for Environmental Quality, National Research Council of Canada m (Harvey et al. 1981) and by the Committee on the Atmosphere and the I Biosphere, Board on Agriculture and Renewable Resources Commission on Natural Resources (NAS 1981) and are further detailed in this report. However, these learned summaries of present knowledge have all • indicated strong evidence of significant ecosystem deterioration due 9 to past and present levels of acid precipitation loading and thus indicate the urgent need to use this present knowledge to arrive at A best estimates of levels of acid loadings which can be tolerated. | While this chapter has considered only the aquatic portions of the « ecosystem, it would appear that because of the interactions with • other components, protection of the aquatic regime would, to a large degree, result in protection of the total environment. This sub-section therefore, reviewed the information and methodologies • presented earlier with respect to their utility in producing • estimates of loading/response relationships. I I ------- I I I I I I I I I I I I I I I I I I I 3-213 As developed in previous sections, acidification of aquatic regimes can be related to proton (H+) loading, concentration of IT1", i.e., of precipitation, or to the constituents of the loading which determine the acidity (i.e., the major ionic species). The anthropogenic loadings add to and interact with the natural components to an extent that also influences the factors available for effective control. Evans et al. (1981), after reviewing the extensive evidence of dose response acidification relationships and considering the empirical model approach of Henriksen (1980) have proposed that "an annual volume weighted H+ concentration of 25 yeq/L (pH 4.6) in precipitation appears to be a critical threshold." These authors have reached their conclusions through the basic consideration of H+ exchange in the reaction processes and through general empirical observations of dose response in sensitive regimes. However, as reviewed in earlier sections, the biosystem response and ability to assimilate nitrate, ammonia or sulphate (the primary acidifying ions of precipitation) differ and therefore the acidifying potentials of these ions differ. In addition, Stensland (1979) and Barrie (1981) have shown that the ionic concentrations of precipitation over eastern North America varies as to relative contribution to its acidity in both space and time. Thus the H+ concentration cannot be considered to have a unique relation to the acidity controlling ions nor has it a unique relation in its dose response in the bio-hydrogeosystem. Thus, neither H+ concentration of precipitation nor H+ loading rates form acceptable criteria for target loadings in relation to protection of aquatic ecosystems from acidification. Henricksen (1980) has argued that surface water acidification can be accounted for as the titration of bicarbonate waters and replacement of bicarbonate by sulphate in the ionic charge balance. He found good empirical agreement between sulphate loadings and observed acidification in widely diverse areas without consideration of any nitrogen species. His relationship to precipitation pH, as cited by Evans et al. (1981), was empirical and based upon Norwegian precipitation and was not an integral part of the argument. It should be stressed here, that while Henricksen's model has a basis in chemical equilibrium, as shown by Thompson (1982), it is in fact a "phenomenological" model which derives from actual dose response observations. A range of sulphate loading vs bio-geo-system responses observed in eastern North America are summarized in Table 3-26 and Summary Table (p. 3-178). This includes several cases relating to episodic event pH changes. While the number of cases are small and statistical significance cannot be assigned, the identified cases of surface water acidification and observed biosystem effects all fall within regions of sulphate deposition of greater than 17 kg S042-/ha.yr. There appear to have been no reported cases of identified acidification which cannot be related to organic sources in areas of less than this level of sulphate deposition. ------- 3-214 I I The Canadian members of the Work Group consider that this evidence, often circumstantial but not inconsistent with theory, leads to the approach best able to provide estimates of target loadings of • sulphate in relation to surface water acidification. It is well gj recognized that this estimate is of limited accuracy in terms of predicted ecosystem response and must surely be subject to later ^ re-evaluation as more information is developed from scientific study. • The empirical observations presented in Table 3-26 and Summary * Table (p. 3-178) immediately suggest a target loading of sulphate which could be accepted but is only poorly defined in terms of • geosystem parameters. The Henricksen-Thompson model permits a I quantification of the target loadings in terms of the geochemical r)~i- Q I Jbasin sensitivity parameter CDR (Ca* + Mg^ ) as an approximation or • unaltered alkalinity as suggested in Section 3.9.3 and further • developed in this section. As pointed out by Henricksen (1980) this model will have, perhaps, significant errors below the titration end — point for alkalinity due to other buffers but should apply with • sufficient accuracy for estimates of the loadings of sulphate which ™ would control the aquatic regime acidity above this transition pH. The CDR serves as the basic geosystem sensitivity criteria in this • model and thereby links the basin hydrology and the sulphate loading fl to sensitivity to acidification. CDR and cation concentrations are related through the hydrological runoff. M Information derived from the Thompson (1982) model may, within the limitations cited, be used to estimate target loadings of ^ sulphate (Figure 3-54 and Table 3-33). Thus if 200 yeg/L of cation • concentration (also unaltered—at/cai-inity; is used as a criteria for ™ basin sensitivity to acidification, protection of the basin water to a mean pH of 5.3 would be indicated for sulphate loadings fl 47.5 kg S0^2~/ha.yr if runoff of 50 cm/yr occurred. For a m 30 cm/yr runoff the protection would only tolerate a loading of 28 kg SO42~/ha.yr. Thus the criteria of 200 yeg/L total M cations or unaltered alkalinity is a reasonable choice of threshold • of sensitivity to acidification over much of eastern North America where runoff may be near 50 cm and sulphate loadings exceed 40 kg S042~/ha.yr (see Figure 2-6b). • A target loading of 15-20 kg SO^2~/ha.yr would, by this model, serve to maintain surface water pH greater than 5.3 on an annual H basis for basins having cation concentrations of 200 yeg/L or greater ^ even in areas of low runoff. More sensitive basins in low runoff areas could not tolerate this level of loading and maintain a pH m greater than 5.3. • The estimates of dose-response relationships presented here do not account for the episodic events discussed earlier which may, in some • ecosystems, be cause for more concern than that based on the mean ™ acidity. The estimates do not consider any time response and must therefore be limited to steady state conditions. Rate of response of • I I ------- I I I I I I I I I I I I I I I I I I I 3-215 any basin to changes in precipitation loading, either quantity or quality, must relate in general to the water residence time. Other factors such as ionic migrations in the soils are not considered. Thus no response rates or equilibrium times are implied, in any sense, by these loading estimates. In the watershed studies summarized above, sulphate in precipitation was used as a surrogate for total acid loading. Sulphate in precipitation is reliably measured. It is recognized that dry deposition of sulphate and sulphur dioxide, and the wet and dry deposition of nitrogen oxides, nitric acid, particulate nitrate and ammonia, as well as other compounds also contribute to acidic deposition. Based on documented effects, wet and dry deposition of sulphur compounds dominate in long-term acidification. Based on the results of the empirical studies, interpretation of long-term water quality data, studies of sediment cores and models that have been reviewed, we conclude that acidic deposition has caused long-term and short-term acidification of sensitive surface waters in Canada and the U.S. The work group also believes on the basis of our understanding of the acidification process that reductions from present levels of total sulphur deposition in some areas would reduce further damage to sensitive surface waters and would lead to eventual recovery of those waters that have already been altered chemically or biologically (Loss of genetic stock would not be reversible.) The U.S. members conclude on the basis of modelling and empirical studies that reductions in pH, loss of alkalinity, and associated biological changes have occurred in areas receiving acidic deposition, but cause and effects relationships have often not been clearly established. The relative contributions of acidic inputs from the atmosphere, land use changes, and natural terrestrial processes are not known. The key terrestrial processes which provide acidity to the aquatic systems and/or ameliorate atmospheric acidic inputs are neither known nor quantified. The key chemical and biological processes which interact in aquatic ecosystems to determine the chemical environment are not known or quantified. Based on this status of the scientific knowledge, the U.S. Working Group concludes that it is not now possible to derive quantitative load ing/effects relationships. 3.10 CRITICAL RESEARCH TOPICS The following topic areas represent issues in which there are major information gaps, and which should be addressed by research programs, in both the U.S. and Canada, at the earliest possible date. ------- 3-216 3.10.1 Element Fluxes and Geochemical Alterations of Watersheds I I Three areas of research are needed here, all requiring relatively • intensive study of both terrestrial (geochemical) and aquatic B (hydrologic) components, mostly focused around calibrated watersheds of comparable research design and intensive data quality assurance. — • B 1. The four ions of primary concern regarding acidification are hydrogen, ammonium, sulphate, and nitrate. Each ion reacts differently with the soil matrix and vegetation. It is necessary, therefore, to define , in specific terms , the fate and effect on surface water acidification of hydrogen, ammonium, sulphate and nitrate ions originating as atmospheric input. m Comparison of results from calibrated watersheds with different soil and vegetation conditions is urgently needed. This report indicates that priority may have to be given to sulphur B emissions control, drawing heavily on evidence that nitrogen 9 deposition does not contribute significantly to long-term surface water acidification, even though it contributes to precipitation acidity and pH depression during snowmelt or runoff events. The long-term necessity for a sulphur control priority needs to be established beyond doubt, as soon as M possible, in order to minimize the risk of making costly errors B in a control program. I I 2. Acidic deposition results in mobilization of metals, such as B aluminum, iron, zinc and manganese, from the soil particles in B watersheds. Further work is needed to define the amounts and species of metals leached from watersheds and their biological consequences. 3. There is evidence that groundwater is being acidified, and that M metal concentrations are elevated, in areas where snowmelt gains B direct access to sandy subsoils with low acid neutralizing capacity. The effect may be seasonal, with pH values recovering during the summer, as neutralization slowly takes place. B Further surveys are needed to establish the extent and B characteristics of groundwater modification over time and across geographical gradients in acid loadings. • 3.10.2 Alterations of Surface Water Quality • Two major areas of information needs have been identified in the extent and periodicity of surface water quality effects: 1. The geographical extent of surface water acidification is not B yet fully documented in North America. Obvious data gaps exist in the central, southern and western U.S. and in parts of • Canada. In addition, reliable data on time-trends in water £ quality appear to be sparse throughout North America, although I I ------- I I I I I I I I I I I I I I I I I I I 3-217 some data have not yet been evaluated. Much of the new data needed can be obtained as part of the long-term monitoring program described below. The critical need is to begin long- term water quality measurements, in a carefully selected range of aquatic environments, as soon as possible. 2. One of the most common manifestations of acidic deposition observed in eastern North America is periodic pH depression in streams and lakes, due to snowmelt or heavy rain. Since periodic low pH is a current problem for biological resources, and likely to remain so until acid deposition is reduced, the quantitative relationship between acid deposition and short- period pH depression should be determined for a broad spectrum of aquatic environments. A dose-response relationship for episodic acute exposures to H+ and aluminum will be a major element in defining acceptable acid loadings. 3.10.3 Alteration of Biotic Components Effects on the biological components of aquatic ecosystems are known only partially. Five research topics are identified: 1. It is essential that the biological responses to various water chemistry changes induced by acidic deposition, be evaluated in considerable detail to define dose-response relationships further. Studies of dose-response relationships in aquatic ecosystems should include surveys of phytoplankton, macro- phytes, zooplankton, benthos and amphibians. Several species among these groups are quite sensitive to changes in pH. Of particular importance to the dose-response relationship is quantification of response data from indigenous species which may be vulnerable to low pH or elevated aluminum, and the pH at which effects are expressed. Special attention needs to be given to determining the pH at which species unique to certain areas are harmed and begin to show some failure in reproduction. In addition, community-level attributes of aquatic systems are likely to be sensitive to acid-induced stresses, but are difficult to determine; nevertheless, they should be understood fully. These include plankton species composition, predator- prey relationships, and trophic-state modification of lakes due to altered nutrient cycles. 2. Damage to fish populations is of particular concern because the loss of fish breaks a major link of the water/terrestrial food chain. Sport fishing is an important industry in most of the areas affected by acidic precipitation and reduction in fish supply could have serious economic consequences. Mechanisms by which low pH and high metal concentrations affect fish should be studied to improve general understanding of the toxicity phenomenon and to improve the ability to predict future effects ------- 3-218 and if so, whether there has been any reduction in spawning success for fish species in those tributaries. I I if acidic deposition continues. Fish sensitivity to H+ and metal ions should be determined, by direct bioassay, at different stages in the life cycle, concentrating on fl reproduction and recruitment. Behavioural or physiological • changes (e.g., blood ion levels) known to be affected by sublethal acid and metal concentrations should also be • evaluated. Long-term monitoring should include fish population • data, as well as other measures of biological productivity. 3. Further study is needed to define the biological effects and • tolerances for periodic pH depression in streams and lakes. • Current work should be extended , to include the Great Lakes tributaries draining Precambrian areas. All such potentially B sensitive areas in the U.S. and Canada should be surveyed, to | determine whether low pH and high metal concentrations occur, • • Mercury concentrations in fish and other wildlife may be increased by the acidification process and/or by increased atmospheric emissions. Increased effort should be placed on measuring existing mercury concentrations and time trends throughout the wildlife food chain, as a function of lake and stream pH values. Laboratory and field studies are needed to establish the biological significance of various mercury concentrations in indigenous species of fish, birds and • mammals. I I 5. When aquatic and/or terrestrial productivity is affected, the effect is often evidenced through the entire food chain. Thus, • there is reason to believe that acidification will have an • adverse effect upon food availability to the higher trophic levels of the food chain, including aquatic birdlife and • mammals. The long-term effects of habitat damage on the | populations of wildfowl and other wildlife should be better defined, and the losses of habitat should be quantified. • 3.10.4 Irreversible Impacts 1. Geochemical and hydrologic principles suggest that the processes W of sulphate accumulations, and associated acidification of soils and surface waters, represent a large-scale titration of • available acid neutralizing capacity. There is evidence that f| the capacity of watersheds to provide neutralization of acids may become depleted, over long periods. Therefore, further work _ is needed to define the rate of acidification of surface waters, • develop predictive models to quantify watershed capacity to neutralize acid over the long term, and to anticipate recovery following abatement. • I I ------- I I I I I I I I I I I I I I I I I I I 3-219 The studies should include measurements on the rates of acidification of lake and stream sediments. The results of such studies are needed to assist in setting acid loading tolerances which will be protective of the aquatic environment in the long term. 3.10.5 Target Loadings and Model Validation Much uncertainty remains as to the quantification of sulphate deposition level ("target loadings") consistent with no further significant degradation of natural resources. Two areas of research are needed: 1. Several relationships, based on field environmental data, have been used to develop descriptive and predictive models of the acidification process. Dickson's relationship, the Henriksen nomograph, and the episodic receptor/dose relation, appear to be potentially useful empirical models which warrant comparative analysis with similar background data bases. Efforts should be made to conduct additional validation of existing and emerging model descriptions of the process of acidification. 2. Relatively detailed simulation models of the acidification process, and its effects, are being developed by several research groups. These should be evaluated, using watershed data bases from a number of intensive study sites in sensitive areas, as identified in this report. If important data are presently missing at these sites, they should be added to the measurement program, or if certain summaries are not being made, these should be added. The need is to have the most complete, quantitative long-term dose-response models evaluated fully and compared with the more empirical field relationships now in use. In support of this validation process, every effort should be made to maximize the use of existing information from all sources. Reasonable validation of both types of models will require considerable new research. Study areas for evaluating atmospheric transport models (see Work Group II report) and loading predictors should coincide with detailed studies of sensitive receptor areas. Locations which already have some data, and which should be considered, include: Experimental Lakes Area - Ontario Boundary Waters Canoe Area Wilderness - Minnesota Algoma Area Watershed Study - Ontario Dorset-Haliburton Study Area - Ontario ILWAS Project - New York Laurentide Park (Lac Laflamme) - Quebec Kejimkujik Park - Nova Scotia Hubbard Brook - New Hampshire ------- 3-220 Northern Highlands Lakes - Wisconsin Coweeta - North Carolina Andrews - Washington North Cascades - Washington I I I 3.10.5.1 Long-Term Data Collection and Monitoring • The present limited ability of the scientific community to assess critically the extent of impacts from elevated acidity in H precipitation, and from other components of atmospheric deposition, | is a consequence of few reliable baseline observations on sensitive aquatic environments. This lack of systematic data arises, • primarily, because many studies and monitoring programs were planned • to define the influences of local anthropogenic development and are, therefore located near these influences. Because acidification is of greatest importance in remote areas unaffected by local discharges, • very few areas exist with any long-term baseline information. ™ Filling this information gap as quickly as possible should be a B priority in both the U.S. and Canada. This information is needed so | that positive, definitive analyses of ecosystem response to the changes in atmospheric deposition can be carried out, with extensive M verifications. Unless a monitoring program is in place and providing • a documented time-series of system properties, there will be no significant capacity to quantify the results of either emission reductions or increases. • While a variety of data needs have been implicit throughout the aquatic effects section, certain classes of long-term measurements are needed at selected sites. Included are the following four: 1. Since a major component of aquatic research is the calibrated « watershed, long-term studies of these systems should be • intensified with the general objective of improving the estimates of rates of changes in water quality and biological effects relative to acid loadings (i.e., dose-response • relationships), improving the understanding of the relative •• influence of sulphur and nitrogen loading; and establishing better measures of lake sensitivity, so that the present and potential extent of the problem can be more clearly defined. 2. Analyses should be undertaken of all available baseline studies, • including regional monitoring of surface water quality, • plankton, fauna, soil, and vegetation records. 3. Criteria for selection of streams and lakes for new monitoring • of water quality and biota should include factors related to • alkanity sources, lake morphometry, watershed morphometry, groundwater inputs, vegetation cover (i.e., age of forest and • community structure), surface water chemistry, groundwater £ chemistry, and type of biotic community (cold water, warm water I I I I ------- I I I I I I I I I I I I I I I I I I I 3-221 etc.). The regions and lakes chosen for analysis should range from very sensitive, through moderately sensitive, to "tolerant" (reference lakes), although a geographic grid of comparable sites should also be developed. Data collected should include chemical and biological parameters identified as susceptible to change. 4. Experimental manipulations should be carried out, using adjacent watersheds with small lakes. Watershed-level experiments should include "simulated acid precipitation" additions of ff1", SO^-, NH^, N03~, etc., so that long-term recovery, following termination of acid additions, can be investigated. ------- 3-222 1974. Effects of acidification on Swedish lakes. Ambio 3:30-36. 1982. Effects on fish of metals associated with I I 3.11 REFERENCES Abrahamsen, G. 1980. Effects of acid precipitation on soil and • forest. 4. Leaching of plant nutrients. In Proc. Int. Conf. • Ecological Impact of Acid Precipitation, eds. D. Drablos and A. Tollan, p. 196. SNSF - Project, Sandefjord, Norway, 1980. Abrahamsen, G.; Horntvedt, R.; and Tveite, B. 1977. Impacts of acid • precipitation on coniferous forest ecosystems. Water, Air, Soil Pollut. 8:57-73. • Ahern, A., and Leclerc, J. 1981. 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Department of the Environment, Halifax, N.S. . 1981. Acidic precipitation in Nova Scotia. N.S. Department of the Environment, Halifax, N.S. 17 pp. Urquhart, C., and Gore, A.J.P. 1973. Redox characteristics of four peat profiles. Soil Biol. Biochem. 5(5): 659-672. U.S. Department of Agriculture (USDA). 1978. 1977 National resources inventory. Soil Conservation Service, USDA, Washington, DC. U.S. Department of Commerce (USDC). 1979. Census of agriculture, 1978. Preliminary file, Technical documentation, Bureau of Census, USDC, Washington, DC. U.S. Fish and Wildlife Service (USFWS). 1982. Results of a modelling workshop concerning acid precipitation. Natural Power Plant Team and Western Energy and Land Use Team, Kearneysville, WV. U.S. Geological Survey (USGS). 1970. The national atlas of the United States. Washington, DC. Vaughan, H.H.; Underwood, J.K.; and Ogden, J.G. III. 1982. Acidification of Nova Scotia Lakes I: Response of diatom assemblages in the Halifax area. Water, Air, Soil Pollut. 18:353-61. Vet, R.J., and Reid, N.W. 1982. A comprehensive evaluation and integration of selected wet deposition data from Canada and the U.S.A. prepared for the Canada - U.S.A. Memorandum of Intent Work Group II. Concorde Scientific, Downsview, Ont. Vollenweider, R.A. 1975. Input-output models with special reference to the phosphorus loading concept in limnology. Schweiz. Z. Hydrologie 37:53-84. ------- 3-258 I Vollenweider, R.A,, and Dillon, P.J. 1974. The application of the I phosphorus loading concept to eutrophication research. Publ. NRCC No. 13690, Environmental Secretariat, National Research Council Canada, Ottawa, Ont. • Watt, W.D. 1981. Present and potential effects of acid precipitation on the Atlantic salmon in eastern Canada. In ll Proc. Conf. Acid Rain and the Atlantic Salmon, ed. L. Sochasky. J| Special Publication Series of the International Atlantic Salmon Foundation, Number 10, Portland, ME., 1980. •• Watt, W.D.; Scott, D.; and Ray, S. 1979. Acidification and other chemical changes in Halifax County lakes after 21 years. Limnol. Oceanogr. 24:1154-1161. jL Watt, W.D.; Scott, D; and White, W.J. 1983. Evidence of acidifi- cation in some Nova Scotia rivers and of impact on Atlantic • salmon, Salmo salar. Can. J. Fish. Aquat. Sci. (in press) £ Wetzel, R.G. 1975. Limnology. Philadelphia: Saunders Co. £ I Wiederholm, T., and Eriksson, L. 1977. Benthos of an acid lake. Oikos 29:261-267. Wiklander, L. 1973/74. The acidification of soil by acid W precipitation. Grundforbattring 26:155-164. Wilson, C.V. 1971. Le climat du Qu€bec, atlas climatique. Service | de la M£t£orologie, Ministere de 1'Environnement, Quebec, P.Q. pp. 74. _ Wilson, D.E. 1979. The influence of humic compounds on titrimetric determinations of total inorganic carbon in freshwater. Arch. Hydrobiol. 87:379 • Wiltshire, J.F., and Machell, J.R. 1981. A study of acidification in sixteen lakes in mainland Nova Scotia and southern New • Brunswick. Preliminary Report, Environmental Protection • Service, Atlantic Region, Environment Canada, Halifax, N.S. Wollast, R. 1967. Kinetics of the alteration of K-feldspar in M buffered solutions at low temperatures. Geochim. et Cosmochim. ^ Acta. 31:635-648. Wren, C.; MacCrimmon, H.; Frank, R.; and Suda, P. 1980. Total and I methylmercury levels in wild mammals from the Precambrian Shield area of south-central Ontario, Canada. Bull. Environ. Contam. Toxicol. 25:100-105. 1 I I I ------- I 3-259 W Wright, R.F.; Conroy, N.; Dickson, W.T.; Harriman, R.; Henriksen, A.; and Schofield, G.H. 1980. Acidified lake districts of the I world: a comparison of water chemistry in southern Norway, southern Sweden, southwestern Scotland, Adirondack Mountains of New York, and southeastern Ontario. In Proc. Int. Conf. _ Ecological Impact of Acid Precipitation, eds. D. Drablos and A. • Tollan, pp. 377-379. SNSF - Project, Sandefjord, Norway, 1980. Wright, R.F.; Dale, T.; Gjessing, E.T.; Hendrey, G.E.; Henriksen, A.; f Johannessen, M.; and Muniz, I.P. 1976. Impact of acid precipitation on freshwater ecosystems in Norway. Water, Air, n_JT T» _ 1 1 _ i- ^ . /. O1 I. f\f\ I I I I I I I I I I 1 I I I Soil Pollut. 6:483-499. Wright, R.F.; Dale, T.; Lysholm, C.; Storen, E.; Henriksen, A.; Hendrey, G.E.; Gjessing, E.T.; Johannessen, M. 1977. Regional surveys of small Norwegian lakes, October 1974, March 1975, March 1976, and March 1977. IR 33/77, SNSF - Project, Oslo, Norway. 467 pp. Wright, R.F., and Gjessing, E.T. 1976. Acid Precipitation: changes in the chemical composition of lakes. Ambio 5:219-223. Wright, R.F., and Henriksen, A. 1978. Chemistry of small Norwegian lakes with special reference to acid precipitation. Limnol. Oceanogr. 23:487-498. Wright, R.F., and Snekvik, E. 1978. Acid precipitation: Chemistry and fish populations in 700 lakes in southernmost Norway. Verh. Int. Verein. Limnol. 20:765-775. Yan, N.D. 1979. Phytoplankton of an acidified, heavy metal- contaminated lake near Sudbury, Ontario; 1973-1977. Water, Air, Soil Pollut. 11:43-55. Yan, N.D., and Miller, G.E. 1982. Characterization of lakes near Sudbury, Ontario. In Studies of lakes and watersheds near Sudbury, Ontario. Tech. Report, Ontario Ministry of Environment, Toronto, Ont. Yan, N.D., and Stokes, P.M. 1976. The effects of pH on lake water chemistry and phytoplankton in a La Cloche Mountain lake. In Proc. llth Can. Symp. Water Pollution Research in Canada, pp. 127-137. . 1978. Phytoplankton of an acidic lake and its responses to experimental alterations of pH. Environ. Conserv. 5:93-100. Yan, N.D., and Strus, R. 1980. Crustacean zooplankton communities of acidic, metal-contaminated lakes near Sudbury, Ontario. Can. J. Fish. Aquat. Sci. 37:2282-2293. Zeman, L.J. 1975. Hydrochemical balance of a British Columbia mountainous watershed. Catena 2:81-94. ------- I I I I I I I I I I I I I I I I I I I SECTION 4 TERRESTRIAL IMPACTS ------- I I I I I I I I I I I I I I I I I I I 4-1 SECTION 4 TERRESTRIAL IMPACTS 4.1 INTRODUCTION A number of air pollutants generated by various sources cross international, state and provincial boundaries. The main pollutants which are potentially harmful to terrestrial ecosystems are oxides of sulphur (SOX), oxides of nitrogen (NOX), particulates, and secondary products, such as oxidants and acidic deposition. There are also smaller amounts of heavy metals, several of which have potentially toxic significance after accumulation. Sulphur dioxide (802) is emitted at phytotoxic concentrations by a large number of mainly anthropogenic sources, including power plants and smelters. Most of this S02 is deposited in dry forms near the sources, though some is transformed chemically in the atmosphere to other sulphur compounds. A moderate amount of S02 remains widely distributed in the atmosphere. In areas remote from sources, the concentration of S02 near the ground is close to background levels, and not likely to cause adverse direct effects. However, S02 is transformed in the atmosphere through a series of reactions into sulphuric acid (H2S04) thus contributing to the formation of the secondary pollutant, acidic deposition. Similarly, NOX gives rise to nitric acid (HN03) and are likewise precursors of acidic deposition. Ozone (03) is also an indirectly emitted secondary pollutant formed in the atmosphere in the presence of sunlight, after chemical transformations of nitrogen dioxide and reactive hydrocarbons. In summary, acidic deposition and ozone, although secondary in nature, are usually considered to be long-range transported pollut- ants as they frequently occur in relatively high concentrations at distances hundreds of kilometres from the source of their primary precursors. Because ozone is a strong oxidizer, oxidative decay usually is rapid in polluted atmospheres and therefore decreases in concentration during late afternoon and evening as sunlight intensity decreases. However, ozone can persist overnight in rural areas or at altitudes where there are low concentrations of reactive components (Jacobson in press). Improved understanding is needed of the ecological effects of the phytotoxic primary and secondary pollutants on terrestrial eco- systems. Field observations and laboratory studies have provided detailed descriptions of the visible injury symptom syndrome produced by ozone. Several review articles and chapters have provided excellent descriptions of these symptoms (Brandt and Heck 1968; Hill et al. 1970; USEPA 1978a). Field studies including the use of field chambers (Heagle et al. 1973; Thompson and Taylor 1966) and those with plots located in a natural ozone gradient have ------- 4-2 I I demonstrated that chronic ozone exposures suppress growth and reduce yield, often in the presence of little or no visible injury symptoms. A more detailed description of the response of plants to acute and • chronic exposures to ozone is presented elsewhere (NAS 1977). • It has been more difficult to determine the adverse or beneficial • effects of acidic deposition on plant communities. Although | simulated rainfall experiments have produced some direct effects on plants exposed to higher than normal hydrogen ion (H+) loadings, ^ direct effects have not been documented conclusively in the field for • vegetation exposed to ambient precipitation (Jacobson 1980). However, some studies have demonstrated the direct effects of acidic deposition on soils (Cronan et al. 1978; Dickson 1978). A Indirect effects of acidic deposition (i.e., acting through soil, other organisms) and its implications are even less well known. • Increases in acidic deposition could result in accelerated changes in • the natural evolution of soils, leading to alterations in soil fertility over the long term. These changes in soil chemistry could ^ have detrimental implications for long-term sustained forest I productivity, and also must be considered in association with aquatic ™ sensitivity. This section on terrestrial effects of transboundary air pollutants 0 is presented in four parts: (1) effects on vegetation; (2) effects on wildlife; (3) effects on soil; and (4) sensitivity assessment. Where M possible, the information on acidic deposition and combinations of • these pollutants has been partitioned and further subdivided into agricultural crop and forest effects. 4.2 EFFECTS ON VEGETATION I I 4.2.1 Sulphur Dioxide (S02) 4.2.1.1 Introduction Sulphur dioxide is an air pollutant of concern to vegetation having most often been recognized for inducing direct foliar effects to plants growing proximal to major point sources of emission. The • phytotoxicity of this gas has been studied extensively around V long-term sources such as Sudbury, Ontario (Dreisinger and McGovern 1970; Linzon 1971) and the districts of Fox Creek and West • Whitecourt, Alberta, (Legge et al. 1976). Controlled long-term • exposure studies have recently been completed as part of the Montana Grasslands Studies (Lee et al. 1978; Preston 1979). This pollutant _ has also been considered of great importance to the vegetation within • the heavily industrialized areas of Great Britain (Cowling and Koziol * 1978) and central Europe (Guderian 1977). Sulphur dioxide is not found on a regional basis at concentrations ^ sufficient to cause direct injury to most plant species. Long-term, I i ------- I I I I I I I I I I I I I I I I I I I 4-3 low-dose studies have demonstrated direct effects to lichen communities (Hawksworth 1971) and indirect effects to several plant species (Keller 1978, 1980; Laurence 1978). Likewise effects may result from lower doses of pollutants in combination with special reference to 03 and S0£ in mixtures (Heagle and Johnston 1979; Reinert and Nelson 1980). Several reviews of the effects of SC>2 on vegetation are available (Guderian 1977; Jacobson and Hill 1970; Linzon 1978; Rennie and Halstead 1977; Treshow 1970; USEPA 1973, 1978b). 4.2.1.2 Regional Doses of S02 As presented in Table 2-3 of Section 2 (Rasmussen et al. 1975), estimates of global background concentrations of SC>2 in gaseous form should be expected within a range of approximately 0.5-5.0 yg/m3 (0.0002-0.002 ppm S02 at STP) with expected residency times of these concentrations to last from one to five days. Regional S02 emissions are shown in Figure 4-1. Mueller et al. (1980) reported on atmospheric pollutant data collected during the period August 1977 - October 1978 for an area covering much of the eastern half of the United States (Figure 4-2). Monthly 1-hr averages varied from 5-40 yg/m3 (0.002-0.015 ppm 802). The highest annual average SC>2 concentrations occurred along the Ohio River Valley; averages ranged from 0.019-0.029 ppm S(>2. The maximal 1-hr concentrations were from 0.11-0.19 ppm S02 and occurred in the same area during October 1978. Hourly deposition values of 1.5-2.3 ppm S02 are common near large emission sources (USEPA 1978a). In the northeast alone, anthropogenic sources exceed all others by a factor of 12.5. Within this region, S02 levels annually average 16 yg/m3 (0.006 ppm S02) (Shinn and Lynn 1979) which is several times that recorded in pristine areas. Therefore, it is reasonable to assume that at the present time concentrations of S02 seldom reach direct foliar injury thresholds for vegetation growing in forested areas or in areas of significant agricultural production. Duchelle and Skelly (1981) reported S02 concentration ranges of 0.001-0.002 ppm/hr S02 during the summer seasons of 1979 and 1980 within the Shenandoah National Park in Virginia and did not consider this pollutant of importance to vegetation in the area. Distribution of even these low doses of SC>2 (and N02) over the major portion of eastern United States corresponds well with known ozone occurrences (USEPA 1978b). 4.2.1.3 S02 Effects to Agricultural Crops There are several possible responses to S02 and related sulphur compounds: (1) fertilizer effects appearing as increased growth and ------- 4-4 I >10,000 1000.1-10,000 x3 100.1-1000.0 &x :-:+\ 10-100.0 (ANNUAL EMISSIONS IN g/s) 200 0 200 400 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 Figure 4-1. Magnitude and distribution of sulphur dioxide (802) emissions in eastern North America. Data from SURE II data base and Environment Canada (Environment Canada 1981d). 1 I I I ------- I I I I I I I I I I I I I I I I I I I 4-5 SO2(ppm) Aug. 1977 Jan./Feb. 1978 SO2 (ppm) Jul. 1978 SO2 (ppm) Oct. 1977 1978 SO2 (ppm) Oct. 1978 Figure 4-2. Geographic distribution of monthly arithmetic means for S02 (Mueller et al. 1980). ------- 4-6 acids and proteins. The rate of entry is particularly important to determining toxicity. Plants have an inherent, and apparently I 1 yields; (2) no detectable responses; (3) injury manifested as growth and yield reductions without visible symptom expressions on the foliage or with very mild foliar symptoms that would be difficult to • perceive as air pollution incited without the presence of a control • set of plants grown in pollution-free conditions; (4) injury exhibited as chronic or acute symptoms on foliage with or without • associated reductions in growth and yield; and (5) death of plants £ and plant communities. Sulphur.dioxide passively enters plants via stomata as part of normal • gas exchange during photosynthesis. Many factors govern stomatal • opening and closing including light, relative humidity, CC>2 concentration and water stress. Sulphur dioxide uptake and ingress ft may also be limited according to plant genetics, previous exposure to V) SC>2 (Jensen and Kozlowski 1975) and subsequent biochemical and/or physiological alterations within exposed plants. Sulphur dioxide has • been shown to increase or decrease stomatal resistance and this may • directly affect potential for the photosynthetic performance (Hallgren 1978). Based on the available literature, it is difficult to assess the relationship of SC>2-induced biochemical and/or I physiological changes at the cellular level in relation to subsequent • effects on photosynthetic activity or resultant growth and yield. Sulphur dioxide, upon absorption is further oxidized to 863 and 50^2- ancj subsequently is incorporated into S-containing amino » *- A- A species dependent, capacity to absorb, detoxify, and metabolically ™ incorporate SC>2 and some plants may absorb low concentrations of S02 over long time periods without injury. • Atmospheric S02 can have beneficial effects to agronomic vegetation (Noggle and Jones 1979). Sulphur is one of the elements required for • plant growth and Coleman (1966) reported that crop deficiencies of S • have been occurring with increasing frequency throughout the world. Several studies using SC>2 as a nutrient supply for S requirements • of plants have been accomplished under varying degrees of soil- • sulphur availability (Cowling et al. 1973; Faller 1970; Noggle and * Jones 1979). The results of these and other studies leave little doubt that application of S as a nutrient via SC>2 fumigation of • plants grown on borderline or S-deficient soils will lead to V increased productivity. The interpretation of studies demonstrating such beneficial effects • must be evaluated in light of their single influence to one crop. Long-term natural ecosystem studies showing similar positive effects — for the entire ecosystem have not been accomplished. Since these • agronomic and natural ecosystems are often physically proximal to one ~ another, further research is needed on the potential influence of S compounds to each singly and collectively. • I I ------- I I I I I I I I I I I I I I I I I I I 4-7 Acute foliar in j ury occurs following high-dose exposures and the rapid absorption of a toxic dose of SC>2 results at first in marginal and interveinal areas having a dark-green, watersoaked appearance. After desiccation and bleaching of tissues, the affected areas become light ivory to white in most broadleaf plants. Some species show darker colours (brown or red), but there is characteris- tically an exact line of demarcation between symptomatic and asympto- matic portions of leaf tissues. Bifacial necrosis is common. In monocotyledons (e.g., corn, grasses) foliar injury occurs at the tips and in strips along the veins (Malhotra and Blauel 1980; USEPA 1976). Plant injury that is visible but does not involve collapse and necrosis of tissues is termed chronic injury. This type of visible injury is usually the result of variable fumigations consisting of both short-term, high-concentration or long-term, low-concentration exposures to In broadleaf plants, chronic injury is usually expressed in tissues found between the veins, with various forms of chlorosis predomi- nating. Chlorotic spots or chlorotic mottle may persist following exposure or may subside and disappear following pollutant removal or as a result of changing environmental conditions (Jacobson and Hill 1970). The presence of acute or chronic foliar injury is not necessarily associated with growth or yield effects. Furthermore, when present, the degree of foliar injury may not always be a reliable indicator of subsequent growth or yield effects. The uniformity of exposure to even the low doses of 862 experienced by crops growing under field conditions presents difficulty in measuring 'treatment1 effects due to the lack of a set of control (nonpollutant exposed) plants. Artificial systems must therefore be used under more controlled laboratory and field situations. The more ubiquitous exposure to known phytotoxic concentrations of 03 must also be recognized and singly evaluated. Yield effects in the absence of foliar symptoms have been reported for soybeans by Sprugel et al. (1980) and Reinert and Weber (1980) under field conditions using a zonal air pollution delivery system and using chamber exposures. Both reports, however, used doses more typical of point sources of emission and would therefore not be considered comparable to regional conditions of exposure. No studies consider all the potential variables that can effect plant response. This is not a possibility for a single study and is especially true for field studies (which are most relevant) where many environmental variables cannot be controlled. From the data available, we can conclude that growth and yield effects are not necessarily related to foliar injury. Depending upon the plant affected, the environmental conditions, and the pollutant exposure conditions, one may observe yield effects without injury, injury without yield effects or more direct correlations between injury and yield. ------- 4-8 I I The primary focus of dose-response studies should be to develop useful generalizations of the relationship between meaningful parameters of plant response and measurable indices of exposure dose. • The relationship between exposure dose and the amount of pollutant I entering the plant may be significantly influenced by environmental factors controlling the rate of pollutant flux into plant leaf _ tissues (see Figure 4-3). The dose of 862 must be considered in • relation to known concentrations under field conditions since both * the regionally expected dose and the phytotoxicity of S02 are comparatively low (e.g., ozone dose and phytotoxicity are relatively 4 high). • The role of short-term fluctuations in S02 may be of particular m importance in areas proximal to point sources of SC>2 (Mclaughlin V and Lee 1974). Here concentrations may fluctuate widely during exposure and damage to vegetation may be closely associated with short-term averages (1 hr) or even peak concentrations. McLaughlin • et al. (1979) studied the effects of varying the peak to mean S02 ™ concentration ratio on kidney beans in short-term (3 hr) exposures to SC>2. They found that increasing the peakrmean ratio from 1.0 B (steady state exposure at 0.5 ppm for 3 hr) to 2.0 (3 hr exposure | with peak = 1.0 ppm) did not alter post fumigation photosynthetic depression. However, further increasing the ratio to 6.0 (1 hr _ exposure with peak =2.0 ppm) tripled the post fumigation • photosynthetic depression. Total dose delivered in the three exposures was 1.5, 1.8, and 1.1 ppm respectively. Clearly the quantity of S02 to which the plants are exposed may have a very I different effective potential as the kinetics of the exposure are ^ changed. Data on S02 effects on plant growth and yield in most cases provide f the most relevant basis for studying dose-response relationships. As a whole-plant measurement, plant productivity is an integrative ^ parameter which considers the net effect of multiple factors over • time. Productivity data are presently available for a wide range of ™ species under a broad range of experimental conditions. Because results would not be expected to be closely comparable across these • sometimes divergent experimental techniques, data have been tabulated flj separately for only controlled field exposures (Tables 4-la and 4-lb). • Relatively few crops of economic importance have been studied under field conditions utilizing various field exposure systems. Of the ^ seven "studies" reviewed in Tables 4-la and 4-lb, dose exposure to • induce a yield effect was 0.09 ppm S02 for 4.2 hr average ^ fumigation period on 18 days scattered from July 19 through August 27 of the soybean growing season (Sprugel et al. 1980). Five studies • indicated no effect following various exposure regimes, and one study | (Neely and Wilhour pers. comm.) reported increased yields (27% and 8%) of winter wheat cv. Yamhill following exposure dose of 0.03 and ^ 0.06 ppm S02 for 24 hr/day for the entire growing season, • respectively. I I ------- I I I I I I I I I I I I I I I I I I I 4-9 POLLUTANT CONCENTRATION NUMBER OF EXPOSURES CLIMATIC FACTORS EDAPHIC FACTORS BIOTIC FACTORS PLANT RECEPTOR MECHANISM OF ACTION DURATION OF 'EACH EXPOSURE -GENETIC MAKEUP STAGE OF PLANT DEVELOPMENT EFFECTS ACUTE CHRONIC SUBTLE Figure 4-3. Conceptual model of the factors involved in air pollution effects (dose-response) on vegetation (Heck and Brandt 1977). ------- 4-10 to CD ^ 5 S CO o (J Q LU Lu Z ~ i CO 1 LU CN § e CO 1- o LU U- U_ LU Q. O 5 fe cc 13 CO • ID 1 LU _J m c 0 4? 0 o: 0 4- L. O Q. 0 ce in t 0 **- s- LU 4- C 8 • — £ c D) CO 0 3 8 Q. X LU §• L. 0 • ID ^_ O CD ID 1— 0 CJV X — T3 JZ — C S— ID 4- L_ c *"-» M- 3 — O « 4- L. 0 C 4- *^ 0) (0 C — ID — O O .- 0. O l_ — N» . 4- O in " c -4- TJ O 0 VI 0 C 0 CN V O m •— 4- >4- C Vi. 0 O O — 0 — x- 4- 4- 4- 4- * (D S 0 — l_ — T3 4- O O 0 C S- 0 ID « 0 in O 4- "O — •— c ^^ t- ID >- >. ID C O. (0 13 o) — -a •—in .c in TO T3 4- 0 c in 589!- >- ID vj .C VO "i a. 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O JZ -> e*, CM p~ to - O L. • — O O •*- 73 t O C E m c m ID o (DC • L. c^ E •— (/} o c in VO 4-3 CT> O U) 3 CMtO ID CO r~ O C CO E CO 3 O 3 O CO E CD (D 4- t-a.3r~ •— 0 ID CT> CM^M-CM IOECOCM in c — (D 0 Q) ^ ^ • o > co o X X o o in ^t 1 1 t t 0 0 CO CO c c ID ID L. L. . in c c 0 O O in — — O 4-4-73 CD ia 4- 4- L. C C 3 0 E 0 0 ID O CO CO CD 73 73 C L. 1_ CD II ID (D 4- 73 73 3 • C C — S (0 ID — • 4- 4- O CD CO CO Q_ ID -Q O 73 ------- 4-12 I I Tables 4-la and 4-lb also reviewed a large number of studies which were conducted using various greenhouse or exposure chamber techniques and exposure of agronomic or horticultural crop plants. f| Conclusions indicated difficulty in determining the significance of | results of such studies in relation to actual similar fumigations under field conditions. Doses used for exposure treatments were m usually considered to be in excess of expected doses for ambient I field exposures. Acute foliar effects have not been reported in long-term studies using less than 0.15 ppm S02 for 24 hr/day for 7 days. • In greenhouse experiments conducted in England using ryegrasses, yield losses were measured following long-term exposure to low levels fl of SC>2. In one study (Bell and Clough 1973), perennial ryegrass | experienced a 52% reduction in dry weight after exposure to a mean concentration of 0.067 ppm S02 over a 26~wk period. At the end of M the study the plants were smaller and chlorotic in comparison to the H control plants exposed to air that was purified by both activated charcoal and an absolute filter. In the other study (Crittenden and Read 1978), shoot dryweight of Italian ryegrass was reduced by 30 to • 40% after 8-10 wk of exposure to 0.02 to 0.03 ppm S02, and was • reduced about 10% after 5-wk exposure to air containing 0.004 to 0.02 ppm S02« The Italian ryegrass plants did not display visible • symptoms of air pollution injury in either the exposure chamber or | the control filtered air chamber. In spite of differences due to exposure regimes, techniques, and • species, certain generalizations can be made with respect to average and outer-limit responses of the plants under study. These have been made in the form of correlations of yield response with total • exposure dose in part-per-million hours (ppmh). The latter data were W calculated as the product of exposure time and SC>2 concentration and transformed to log values. For experiments employing controlled • exposures under field conditions (Tables 4-la and b), data are ^ graphed in Figure 4-4 (McLaughlin 1980). For the 36 data points shown, exposure dose ranged from 0.24 to 259 ppmh. No effects on — yield were detected in any of the six studies at doses _>_ 6 ppmh. • Yield losses occurred in 26 cases at levels ^_ 6 ppmh, while no effects and positive effects were noted in two cases each at levels _> 6 ppmh. A linear regression of yield on dose for all studies • reporting yield losses showed strong positive correlation (r = 0.75) • of yield with dose and took the form: Yield loss = -13.6 + 23.8 (log dose) | r2 = 0.53 (Significance = >_ 0.001) This correlation excludes four data points, two with no effects and • two with positive responses. All were studies with wheat reported by * Neely and Wilhour (pers. comm.). Data from studies reporting no effect or a positive effect are however all plotted in Figure 4-4. fl Calculation of the phytotoxic potential for regional scale S02 exposures involves many assumptions regarding toxic and nontoxic m I ------- I I I I I I I I I I I I I I I I I I I 80 60 40 LU CO Z o a. CO LU DC Q 20 _i LU 20 40 4-13 REGRESSION LINE: % YIELD LOSS —13.6 + 23.8 (LOG DOSE) r2*0.53 P>F< 0.001 22 DATA POINTS I I 0.1 1.0 10.0 EXPOSURE DOSE(ppmh) 100.0 Figure 4-4. Regression of yield response vs. transformed dose (ppnh) for controlled exposures using field chambers (zero and positive effects excluded from regression analysis) (after McLaughlin 1980). ------- 4-14 4.2.1.4 S02 Effects to Forest Vegetation I I components of the total dose to which vegetation is exposed. Obviously not all, but probably most exposures to S(>2 on a regional scale are below levels producing phytotoxic reactions. An important H aspect of evaluating the likelihood that plants will be negatively | influenced by S02 exposures is the determination of what components within a plant's total exposure history are phytotoxic. Mclaughlin M (1980) recently examined USEPA (1978b) data on regional S02 concen- • tration averages. Using the assumption that only the upper 10% of all S02 exposure days would have S02 concentrations high enough to cause stress to vegetation, and that only daylight exposure • (8 hr/day) during the active growing season (6 mo/yr) would be • effective, he calculated that the average potentially phytotoxic dose within designated air quality control regions would range from 0.9 • ppmh (Region IX) to 5.5. ppmh (Region VIII). Maximum doses (highest Jf reporting stations within regions) ranged from 2.6 ppmh to 27 ppmh, thus pointing once again to the potential injury to vegetation grown ^ within smaller areas of high S02 point source emissions. • I The effects of S02 on broadleaf tree species and similar types of native vegetation closely resemble those as described for agronomic • crops. • In conifers, acute injury on foliage usually appears as a bright — orange red tip necrosis on the current-year needles, often with a • sharp line of demarcation between the injured tips and the normally • green bases. Occasionally, the injury may occur as bands at the tip, middle, or base of the needles (Linzon 1972). A Recently incurred injury is light coloured but later bright orange or red colours are typical for the banded areas and tips. As needle M tips die, they become brittle and break or whole needles drop from I the tree. Pine needles are most sensitive to S02 during the period of rapid needle elongation but injury may also occur on mature needles (Davis 1972). • Chronic effects of S02 in conifers are generally first expressed on older needles (Linzon 1966). Chlorosis of tissues starting at the • tips progresses down the needle towards the base (i.e., symptoms || progress from the oldest to youngest tissues). Advanced symptoms may follow, involving reddening of affected tissues. Continued chronic M injury to perennial foliage of coniferous trees results in premature • needle abscission, reduced radial and volume growth, and early death of the trees (Linzon 1978). Forest trees vary considerably in their sensitivity to S02 doses W and Jones et al. (1973) evaluated the response of numerous species growing near point sources in southeastern U.S. (Table 4-2). Visible symptom expression only occurred on the most sensitive species at I I I ------- I I I I I I I I I I I I I I I I I I I 4-15 TABLE 4-2. SULPHUR DIOXIDE CONCENTRATION CAUSING VISIBLE INJURY TO VARIOUS SENSITIVITY GROUPING OF VEGETATION3 (Jones et al. 1973) Maximum Sensitivity grouping average concentration Sensitive Intermediate Resistant (ppm S02) (ppm S02) (ppm SC^) Peak 1.0-1.5 1 hr 0.5-1.0 3 hr 0.3-0.6 Ragweeds Legumes Blackberry Southern pines Red and black oaks White ash Sumacs 1.5-2.0 2.0 1.0-2.0 2.0 0.6-0.8 8.0 Maples White oaks Locust Potato Sweetgum Upland cotton Cherry Corn Elms Dogwood Tuliptree Peach Many crop and garden species a Based on observations over a 20-year period of visible injury occuring on over 120 species growing in the vicinities of coal- fired plants in the southeastern United States. ------- 4-16 I I doses of 0.30 ppm/3 hr thus once again pointing to the smaller area of source influence on direct foliar injury. Dreisinger and McGovern (1970) indicated a somewhat similar injury threshold (i.e., 0.26 ppm • S02/4 hr) for visible foliar injury to the most sensitive vegetation I to S02, but doses were still above ambient concentrations as expected on a regional basis. • A few major investigations of the effects of S02 on tree species growing under natural conditions have been reported (Dreisinger 1965; ^ Dreisinger and McGovern 1970; Linzon 1971, 1978). These reports • indicated that a pollution (802) gradient existed within the ™ designated study area near Sudbury, Ontario, and effects correlated well with this gradient. Chronic effects on forest growth were fl prominent where S02 air concentrations during the growing season | averaged 0.017 ppm S02, and were only slight in areas receiving 0.008 ppm S02 (Linzon 1978). In Czechoslovakia, Materna et al. m (1969) reported the occurrence of moderate chronic injury to foliage • of spruce trees at Celna, under the influence of an average annual concentration of S02 at 0.019 ppm. ^ Table 4-3 summarizes the results of tree studies that have utilized • artificial exposure chamber systems under laboratory conditions. Only two studies (exposures) used doses close to ambient concentra- • tions (Houston 1974); however, the use of selected clones of known ^j sensitivity to S02 hinders further field speculation from this study. The remainder of the studies presented in Table 4-3 have used M doses above expected occasional exposures under field conditions. • Concentrations of 0.25 ppm S02 for 2 hr were required to induce slight injury to several pine species (Berry 1971), but overall trends for increasing foliar injury do not follow increasing dose for • conifers per se. Smith and Davis (1978) exposed several conifers W (pine, spruce, fir and Douglas fir) to doses of 1.0 ppm S02 for 4 hr or 2.0 ppm S02 for 2 hr and only pines developed necrotic tips • at the 2.0 ppm dose. Likewise, Keller (1980) found only trends in f reduced photosynthesis in Norway spruce at S02 doses of 0.05 ppm S02 for 10 wk exposure with significant effects noted at 0.10 and M 0.20 ppm S02 over the same period. • 4.2.1.5 S02 Effects to Natural Ecosystems I Ecosystems are basically energy processing systems whose components have evolved together over a long period of time. They are composed • of living organisms together with their physical environmental • conditions. Ecosystems respond to environmental changes or perturba- tions only through the response of the organisms of which they are composed (Smith 1980). The living (biotic) and nonliving (abiotic) • units are linked together by functional interdependence. 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D) II ^ JZ • in 4- o 0 »*- M- 0 XI "0 '>- 1_ O \ X) c ID 1_ 10 "o 4- XI c 3 O 4- >* x> in in 0 4- 10 O XI c X o • -o 07 t o QL 0 U 4- o 0 4- 4- 0 4- C (D O M— C cn in L, O 4- C 0 C 'E o CL 4- in O s X) c o 4- ID 4- 0 l_ CL L. (D 4- c XI c (0 c O3 in 0 XJ >~ XI 3 4- in i_ 0 o. 0 L. CL -3 & ID c O 4- 10 L. 0 XJ in 8 L, O M- in 4- 10 0 > ID O 0 ------- 4-24 I I ecosystems are studied that are not observable when individuals, populations or communities are studied. Natural ecosystems are seldom, if ever, exposed to a single air «• pollutant. Therefore, the responses observed under ambient conditions cannot conclusively be attributed to a single substance • such as sulphur dioxide alone. Consideration of low SC>2 doses on a | regional basis presents even further difficulties in discerning effects induced by this pollutant. . Questions relating how sulphur deposition from anthropogenic emissions is incorporated and distributed by aquatic and terrestrial ecosystems is not fully resolved. The issue is critical since • ecosystems subject to excess nutrients or toxic materials do not • commonly distribute them uniformly throughout the system but rather preferentially sequester them in specific pools or compartments. In m addition, sulphur dioxide as a gas can cause injury to the vegetative • components of specific and local ecosystems so that energy flow and the cycling of other nutrients as well as sulphur may be disrupted if ^ the pollutant is at sufficient concentrations. • Specific studies of the more detailed effects of SC>2 on natural systems have been conducted proximal to point sources of high 862 B emissions and include studies in the vicinity of the Kaybob gas V plants (Fox Creek, Alberta) (Winner et al. 1978)5 West Whitecourt gas plant (Whitecourt, Alberta) (Legge et al. 1976) and the Sudbury, • Ontario smelter district (Dreisinger and McGovern 1970; Linzon 1971). • Additionally, a series of designed studies using ariticial sources of S02 have been conducted in the Montana grasslands (Preston 1979). The results of these studies, particularly the West Whitecourt and • Montana grasslands studies, document the usefulness of addressing ecosystem level responses to S02 from a multidisciplinary approach • incorporating investigations of physiology, autecology, synecology, | geochemistry, meteorology and modelling. The results confirm that producers are sensitive to direct S02 effects as evidenced by mm S02~associated changes in cell biochemistry, physiology, growth, I development, survival, fecundity, and community composition. Such responses are not unexpected. An equally important point of agreement among the different research efforts is the potential for • ecological modification resulting from either direct S02 effects on • nonproducer species or direct changes in habitat parameters, which in turn affect an organism's performance. Changes in biogeochemistry, • particularly in the soil compartment, are notably responsive to • low-dose S02 exposures. A major conclusion of the Montana grasslands studies indicated that at S02 levels above 0.02 ppm (52 _ yg/nH), induced changes occur in the performance of producers, • consumers, and decomposers. Many of the responses are individually * small, but collectively over time they gradually modified the structure and function of the grasslands. The significance of these B changes to the long-term persistence of the ecosystem remains | controversial (Preston 1979). I I ------- I I I I I I I I I I I I I I I I I I I 4-25 Direct effects of SC>2 on individuals within natural plant communities are most noted within the lichens. Sulphur pollution not only has caused the depletion of lichen vegetation in certain areas, but also has resulted in changes in the distribution of different species (Hawksworth et al. 1973). Epiphytic lichen communities have been mapped within several regions of North America. In a rural area of Ohio surrounding a coal-consuming power station (emitting 1025 tons S02/day), the distribution of two corticolose lichens, Parmelia caperata and I>. ruderta, was markedly affected by elevated S02 levels (Showman 1975). In regions experiencing an annual S02 average exceeding 0.020 ppm, both species were absent. The distribution of more resistant lichens was not noticeably affected until S02 levels exceeded 0.025 ppm (annual average). Somewhat lower levels were projected by LeBlanc and Rao (1973) to effect the ability of sensitive lichen species to survive and reproduce; acute and chronic symptoms of S02 toxicity in epiphytic lichens occurred when annual averages of SC>2 exceeded 0.03 and 0.006-0.03 ppm respectively. The network of biotic-abiotic interactions, which is characteristic of managed and natural ecosystems, leads to the hypothesis that S02 effects on producers must have repercussions to other trophic levels. Demonstration of such responses, however, is difficult experimen- tally, and an accurate assessment of the specific importance of S02 in eliciting these responses is complicated by the often complex relationships between producers, consumers, and decomposers. More subtle effects may occur in areas of low S02 (0.05 ppm annual average) deposition by shifts in soil microfloral populations thus further influencing plant rhizopheres leading to subsequent ecosystem alterations (Legge et al. 1976; Wainwright 1979). Induced changes in natural ecosystems should not be evaluated on a positive or negative basis. Change as induced by anthropogenic sources of 862 must be considered as an alteration of natural processes. For example, natural ecosystems evolved on sulphur- deficient soils have done so within the imposed constraints per se. Although atmospherically derived sulphur may not be sufficient to cause injury, the prolonged input of sulphur may relax the constraints of a limited sulphur supply thereby inducing shifts in species composition. 4.2.2 Ozone (03) Ozone air pollution injury was first reported by Richards et al. (1958) and during the subsequent years a diverse array of visible injury symptoms was described on a wide variety of crop, ornamental and native vegetation. Numerous review chapters and journal articles contain detailed descriptions of these symptoms (Brandt and Heck 1968; Hill et al. 1970; NAS 1977; USEPA 1978a). Characteristics of the injury symptoms and extent of injury are influenced by climatic ------- 4-26 4.2.2.1 03 Effects to Agricultural Crops I I and edaphic conditions, genetic variability, characteristics of the pollutant dose, and by interactions between the pollutant and other _ air pollutants or other environmental factors (NAS 1977). Injury • symptoms described by the various researchers have included: ™ bleaching, bifacial necrosis, general chlorosis, chlorotic mottling, chlorotic streaking, topical necrosis such as "fleck" and "stipple," I and pigmented leaf tissue (Hill et al. 1970; NAS 1977; USEPA 1978a). • In addition to the development of visible injury symptoms, exposure to atmospheric ozone can: (1) suppress photosynthesis; (2) stimulate • respiration; (3) inhibit carbohydrate transport; (4) change membrane • properties; (5) alter metabolite concentrations; (6) alter symbiotic associations; and (7) alter host-parasite interactions. — Prior to 1970 most 03 research dealt with observed foliar symptoms • resulting from acute (short-term), artificially controlled, dose-response studies. In the 1970s, the research approach shifted • toward chronic (long-term) studies providing a more realistic | estimate of natural plant response. The results of several such studies are summarized in Table 4-4. These studies formed the M foundation for quantification of dose-response relationships that • provided a more realistic basis for the assessment of losses under field conditions. A number of assessment techniques (e.g., open-top chambers, protective sprays) were utilized in several major studies • designed to pursue this objective. V The National Crop Loss Assessment Network (NCLAN) (Heck et al. 1982) • utilized open-top chambers and controlled 03 concentrations. Its f purpose was to provide standardized crop dose-response data which could be utilized in the development of reliable regional scale loss m assessment calculations. • I Foliar responses of crops to artificial 63 exposure have been well documented and used in the development of species and varietal • sensitivity listings and the preparation of predictive dose-response £ curves (Larsen and Heck 1976; Linzon et al. 1975). 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L. •— 0 >• — 0 _ (U S 2 E s 1 o i L. * : o\ « vO ••• CM 4- 2 4-2 L. c1 >- "° 01- 0 — T3 4- O)4- 2 (D 2 t) 0 >- — >- O 4- l_ O L. L. Ul T3 t- 13 CM*— * m* — CM — in in ID ID "O ^ VO ^D CM CN :*. >^ ID ID "O "O ^^ s^ « ^^ ^^ O O »— » — o o *~r *-^ vo vo o\ o\ *~ *~ 0 C 0 — c Q. ~ Q. ID 80 +- L. •"• 0 ^ r> 2 C 0 ^ CL S I I I I I I I I I I I I I I I I I I I ------- I I I I I I I I I I I I I I I I I I I 4-31 Although the adverse effects of 03 exposure on crop yield or productivity have not been as extensively documented as has been the case with foliar injury, there are nevertheless numerous reports on this topic. Any assessment of yield or quality parameters under field conditions is complicated by the ubiquity of ozone exposure, the effect of meteorological variables on ozone distribution within crop canopies, and the difficulty in establishing ozone-free control plots. Numerous biotic (pathogen, genetics) and abiotic factors (i.e., RH, light, and soil moisture) within the environment must also be taken into account. These difficulties have been partially overcome by recent progress which has been made in the development of field assessment techniques for plant growth and productivity (Reinert 1980). These include open-top field chambers, pollutant exclusion methods, open-air fumigations, ambient air pollutant gradients and chemical protectants. Experimental studies with field grown crops have demonstrated yield reductions in a large number of ozone-sensitive crops: beans (Heggestad et al. 1980), potatoes (Heggestad 1973), grapes (Thompson et al. 1969), corn (Heagle et al. 1972) and others (Heggestad 1980; Jacobson in press; Reinert 1975). In general the studies have shown that decreased yield of susceptible species occurs with average ozone concentrations of between 0.05 and 0.1 ppm for 6-8 hr/day during the growing season (Heck et al. 1977). In a 5-yr study in Maryland (1972-79), typical yield reductions were 4, 9, 10, 17 and 20% respectively for field grown (open-top chambers) snap beans, sweet corn, potatoes, tomatoes and soybeans (Heggestad 1980). The first report from the NCLAN project (Heck et al. 1982) appears to provide good agreement with earlier dose-yield response data (Heagle and Heck 1980) and with yield losses in the various crops as follows: soybean 10%, peanut 14-17%, a single turnip 7%, head lettuce 53-56% and red kidney bean 2%. The yield reductions were equated with seasonal 7 hr/day mean 03 concentrations of 0.06-0.07 ppm compared to the 0.025 control value. In the earlier study (Heagle and Heck 1980) employing open-top chambers with 03 dispensing capabilities, an annual U.S. crop loss estimate assuming a seasonal 7 hr/day mean 03 concentation of 0.06 ppm in all crop production areas was calculated at $3.02 billion (5.6% of the national production). In a subsequent manuscript Heck (1981) pointed out that it is a weak assumption that crops in all parts of the United States are in a sensitive state during much of the growing season and the values should be reduced by 50%. This would bring the estimate of 63 crop losses in the U.S. to between$1 billion and $2 billion or 2-4% of total production assuming all areas were at concentrations of 0.12 ppm for 1 hr. As most sections of the country are above the current standard, the national losses are probably higher than the above values (Heck 1981). There are limitations in assessing 63 impact on crop species, in that a majority of presently operating 03 monitors in both the U.S. and Canada are in urban locations. They therefore may not represent ------- 4-32 I I levels to which rural vegetation is exposed. However, some indica- tion of the occurrence of 03 in rural areas along the U.S./Canada border is given in Table 4-5. The Ontario rural data (Table 4-6) I have been summarized to provide some indication of the potential for • adverse crop effects (growing season daytime basis) and can be compared directly with the 03 data (Table 4-7) for urban locations • in the National Air Pollution Surveillance Network (NAPS) in Ontario, | Quebec and New Brunswick. It is apparent from these urban and rural data that the southern • portion of the Province of Ontario is most adversely affected by ozone in Eastern Canada. This finding is corroborated by numerous reports of ozone-related crop injuries in this area (Cole and Katz • 1966; Curtis et al. 1975; Hofstra et al. 1978; Ormrod et al. 1980) • and by the absence of any documented injurious effects to sensitive agronomic or forest species in Quebec or the Maritime provinces. I In Ontario the first indication of transboundary ozone movement across Lake Erie was documented (Mukammal 1960) following extensive • work on the relationship between the incidence of weather fleck on • tobacco and meteorological conditions associated with the buildup of ozone. Since then a number of large-scale meteorological investi- gations (Anlauf et al. 1975; Yap and Chung 1977) have documented • these early findings and have shown that high ozone levels generally • are associated with regional southerly air flows which have passed over numerous urban and industrialized areas of the U.S. and which, • as they move across the lower Great Lakes, undergo rapid dispersion • as they encounter unstable conditions near the northern shore of Lake Erie. Contributing to these influx patterns are the more localized _ downwind urban effects which can add to the already high background • levels. ™ In an effort to estimate the severity and extent of plant injury or B yield loss resulting from exposure to ambient ozone in southern | Ontario, a summary has been prepared for all major crop species on the basis of documented research reports of yield or productivity • losses in Ontario or the northeastern U.S. and on unpublished • documents by government agencies or university departments working under assessment mandates or research contracts. On the basis of these findings and 1980 economic values it is estimated that the • average annual loss for ozone-sensitive Ontario crops based on 1980 • economic values is in excess of$20 million (Pearson 1982). An example of the types of work which were considered in the assessment • of crop loss is shown for one of the most sensitive species, white | bean. In 1961, bronzing and rusting of white bean foliage was reported • (Clark and Wensley 1961) throughout southwestern Ontario and the resultant defoliation and pod abortion was estimated to have resulted in a loss of approximately 600 pounds of beans per acre (45% yield • loss) in severely affected fields. Following extensive field work in • 1965 and 1967 the disorder was found to be associated with the I I ------- I I I I I I I I I I I I I I I I I I I 4-33 TABLE 4-5. THE NUMBER OF TIMES IN 1980 and 1981 THAT OZONE CONCENTRATIONS EXCEEDED THE USEPA STANDARD OF 0.12 ppm ALONG THE U.S./CANADA BORDER3 Station Allen Park Detroit Detroit Essexville Livonia Macomb Co . Marquette Co. Port Huron Port Huron Southfield Warren Lake Co . St. Louis Co. Berlin Amherst Erie Co. Essex Co. Monroe Co. Niagara Co. Niagara Falls Rochester Wayne Co . Berea Cleveland Conneaut Elyria Elyria Painesville Toledo Toledo Westlake Burlington Burlington State MI MI MI MI MI MI MI MI MI MI MI MI MI NH NY NY NY NY NY NY NY NY OH OH OH OH OH OH OH OH OH VT VT 1980 1 2 6 1 1 6 0 5 - 0 0 0 - - 0 - 5 1 5 2 1 2 0 0 1 0 2 1 0 3 0 0 0 1981 1 0 4 - 1 6 0 7 7 0 0 0 0 0 0 1 7 0 1 0 1 1 1 0 2 - 1 - 5 2 0 0 0 Only data from the U.S. counties touching the international boundary were used. Data were compiled by Rambo and Patent (pers. comm.). SAROAD data base covers all of calendar year 1980, but only includes January to September of 1981. ------- 4-34 •_ CO VO i— ON •l o i 1- z o § UJ X 1- 13 z CO o _ h; 8 i < Ct 3 ce z CO Q Z UJ ft p^ UJ s ° 1- 31 o >; Q Z o CO UJ to 1 fe CO vo 1 UJ m 1- ^_^ ID •o ^ j: o •— s^ (D •— in .E L 0 ID C • L_ 3 S- ~3 10 - +- vD ID c Q 3 to 0 0) >- •— _» l_ I- 3 3 Q i? 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X +- -^ E 0 "1 13 0 ^- ECO. — 0 ^ X 0 IO O TD — ID ID -H > ID •ws. ° U ID 0 *~ -^ 0 •z. c c .2 .2 +- +- SID •(- O C/N _J ^ 2 •— O t CM O — o co in t o o o o o o o o o o CM o to r> o O O CM O O vO O ON ^" O o o r~ •* ^f — 10 — «— O *~ to to CM m i^ in ^- i^ ON o ON ON ON r^~ oo r** r^ r^ r^ oo ON ON ON ON ON _ _o ^ — I- CM ID ID — S +- <£> CCO O ° ""' 0 'J — •* CM — \r\ CNJ r^ ^~ ro •* — ^t CM — o o o o o O in •* co vo CM O CM O O VO T t CO IO in — co to to CM O CM CM — 1^ — 10 ON 1^- o r~ m ^c ^ ON CO 00 ON ON *£> f^- CO ON O r^ ^ (^ ^ oo ON ON ON ON ON in 5 o IO o — O — CM tO O to ID — O O O — — +- ID O O O O O c 4- *O vo vO vO vO o c o o o o o o 1- ON IO IO O to oo ^ ^ o m •~ *— r- CM o — CM o o o o o t^ >* •— — CO IO •— CM O O r^- CM f^ to vo CM ^J- <— — •* — "• r— IO CM CM CM IO r^ o vo CM co r- to co o ^ vO 1^ CO ON O ON ON ON ON ON in 0 c ~ ID O •- £ — O °v~ u to ID (D — O 4- VO . 8 ° m tj >^« ON ON ^ r^ r^ r^ oo ON Oi ON ON ON 1_ O — * — 0 — 0 — vo in r~ ^- CM in ^ ^ vo m ON ON ON ON ON VO r~ 00 ON O ON ON ON ON ON •25 I- ON C ID O O 4- vO •D c O o ^ —1 1 I I I I I I I I I I I I I I I I I I I ------- 4-37 t~* >- ID T3 \ 1_ _C O ^~ Su* CD > in 3 — l_ O ID C 0 — >- • JZ E 0 • ID Q. UJ ID 4- in | 3 a • 3 E < ID "0 C CJ\ ID g > 1- 3 4- ~~3 ID • •I- CD ID C D = -7 cTo — 1_ 1_ 3 3 Q O X C 0 X> CD in ID CD in CD 3 — ID > _ < m Ml 3 O 0 CN I in >- L CD CD ID O I- > Q -4- 6 •— E 4- t- O 3 O • 1- O 0 O CD • •— m o c z O 0 in Ml 3 O O X in «* >- 0) ID I- > a o — t- +- H— ^ o o o • CD • oo in o c 2 O o o C CN o — +- (D 1_ +~ c 0 o o o 2 o * £ oa. ro— o. O ^ CD E 1— •4- O O •*«. oo" C >- o 1- •(- 3 (D O L. X -1- —• C ^ E CD E ECO. .- 0 ^ X O ID ^ K1 0 T3 — ID ID +- > ID Q b% 1_ ID CD >- • 0 z C C O 0 +- +- (D ID 0 -1- O -i cn OL < 2 ~ ^ o r-. ro ^j- o o> o ^o r^ *~ "~ "~ o\ r~- csi o «t •- O O O 0 *o •— 10 fo in i~ f r>j •— — o oo — o o • • • • • oo in o o CN r~ ^f — CN MI "~ "~ "~ *— "~ in 10 ^- oo o • • • • • O CN in o\ M^ C^ C7V O* CO O vo r^ co o o i~- i^ i^ r^- oo CT\ a\ o o*\ o\ •^ r— ^— ^ ^ o ^• — o l_ 1- CN O ID O in •*- 10 xi c o C 0 ~ 3E • Ul C O 4- ID 4— in ID CD > ID ------- 4-38 4.2.2.2 03 Effects to Forest Vegetation I I occurrence of elevated levels of atmospheric ozone pollution (Weaver and Jackson 1968). The symptoms first appear sometime between « flowering and normal senescence, a critical period in the development • of yield potential. They appear as a bronze-coloured necrotic stipple which, as it becomes more severe, results in premature leaf drop and reduced seed set. • In an effort to assess and compare the annual severity of ozone injury on sensitive white beans, Ontario government personnel have • conducted visual assessment surveys throughout the major production • areas in southern and southwestern Ontario since 1971. These studies ruled out any varietal resistance and confirmed that the bronzing _ disorder was widespread throughout all the bean production areas • (Pearson 1980). • Studies utilizing chemical protectants against ozone injury have fl helped to provide information on the economic relevance of the || bronzing disorder in Ontario. In one case a 13% yield increase was associated with the reduction in bronzing severity (Curtis et al. M 1975), while in another study, yield increases of up to 36% (27% • yield reduction) were realized (Hofstra et al. 1978). In 1977 and 1978 yield increases with antioxidant protection were not • as high (Toivonen et al. 1980) due to climatic problems. The overall • response in these years was 16% and 4% increase in yield respectively due to antioxidant protection. On the basis of these values and considering the uniformity of cultivar sensitivity, the average annual loss for this crop was estimated at 12% (Pearson 1982). I I As in the case of agricultural crops, economic evaluation of the I effect of pollutants on forest productivity is ultimately contingent • upon the establishment of dose-response relationships. Consideration must be given to pollutant loadings and then quantitative measure of • growth-suppression or yield-depression. • There are different considerations in evaluating the effects of 03 _ and acidic deposition on forest trees than for agricultural crops. • Most forest tree species are long-lived, perennial plants that are ™ not subjected to fertilization, soil amendments, cultivation, extensive pest control or other cultural practices that agricultural • crops receive. Their size also precludes pollutant exclusion | (chambers) studies or protective sprays limiting the assessment of growth or productivity losses to visual observations of growth « characteristics. This must then be related to ozone dose information • (i.e., pollution gradients) where available. In general, many tree species indigenous to North America are • classified as susceptible to 03 damage (Davis and Wilhour 1976; ' Skelly 1980). Direct injury to tree foliage by 03 has been I I ------- I I I I I I I I I I I I I I I I I I I 4-39 demonstrated repeatedly in experiment situations (Table 4-4), and in nature as well. Concentrations of 03, at least in some forested areas, are sufficient to cause injury (Miller and McBride 1975; Skelly 1980). These effects of 03 can alter the productivity, successional patterns, and species composition of forests (Smith 1980) and enhance activity of insect pests and some diseases (Woodwell 1970). The current status concerning 03~induced effects on Temperate and Mediterranean forest tree species, communities and ecosystems has been summarized (Skelly 1980). It is possible that primary productivity, energy resource flow patterns, biogeochemical patterns and species successional patterns may all be challenged by oxidant air pollution. 4.2.3 Acidic Deposition Various types of injury listed below may result from direct exposure of plants to acidic deposition (Cowling 1979; Cowling and Dochinger 1980; Tamm and Cowling 1977): 1) Damage to protective surface structure such as cuticle; 2) Interference with normal functions of guard cells; 3) Poisoning of plant cells, after diffusion of acidic substances through stomata or cuticle; 4) Disturbance of normal metabolism or growth processes, without necrosis of plant cells; 5) Alteration of leaf- and root-exudation processes; 6) Interference with reproductive processes; 7) Synergistic interaction with other environmental stress factors; 8) Accelerated leaching of substances from foliar organs; 9) Increased susceptibility to drought and other environmental stress factors; 10) Alteration of symbiotic associations; and 11) Alteration of host-parasite interactions. In contrast to results with 63, experimental studies with simulated acidic deposition have produced both positive and negative results ------- 4-40 4.2.3.1 Acidic Deposition Effects to Agricultural Crops I I (Jacobson 1980; Shriner 1978). Increases and decreases in yield, as well as no significant effects, have been found. These results depend upon concentrations of acids, plant species and cultivars, I pattern and timing of rain applications, and soil, environmental, and I cultural conditions (Irving and Miller 1980; Lee et al. 1980). Each species may thus have unique patterns of physiological and genetic • responses to the potentially beneficial and detrimental components of I acidic deposition. I Experimental studies with plants grown under controlled (or I semicontrolled) conditions have demonstrated that visible foliar • symptoms can be produced on certain crops, when pH of applied simulated rain is 3.5 or less (Table 4-8). Field-grown plants may be • less susceptible to the development of foliar symptoms than plants I grown under controlled or semicontrolled conditions (Jacobson 1980; Shriner 1978). Further, as with 03 and SC>2, foliar symptoms may _ not correlate closely with yield reductions (Lee et al. 1980). • However, recent evidence suggests that generalizations concerning • effects on crops from experiments with 03 alone or with acidic deposition alone, may underestimate the interactive effects of • sequential exposures to these two pollutants (Jacobson et al. 1980). | Further research is needed to determine if acidic deposition enhances the likelihood of actual yield reductions in areas also experiencing H repeated exposures to elevated concentrations of 03. I In studies with soils and in studies on aquatic systems focus has often been on relationships with mean annual deposition rates. • Characteristics of individual rain events may have greater • significance in producing direct effects on agricultural crops than average annual rates. Although annual pH values of rain are as low • as 4.0 in eastern North America, concentrations of H+ ions (and | 30^2- an(j N03~ ions) may be ten times greater than average during individual events. The one (or several) most acidic event(s) • of a growing season may have greater significance for production of • direct effects on annual crops than average deposition rates. The potential for crop damage in the field from acidic deposition is I further amplified substantially by agricultural practices. Economic • constraints in any given area and year tend to result in the exposure of extensive areas of a given crop in a relatively uniform state of • plant development. The onset of the cycle of flowering physiology, • pollen dispersal and fertilization, and photosynthetic partitioning, could all be potentially susceptible to extensive damage over vast areas. • To evaluate the economic cost of acidic deposition on agricultural crops, answers to several questions are needed. Which crops are • actually benefited by components of acidic deposition? Which crops | are most susceptible to reductions in yield by exposure to acidic I I ------- I I I I I I I I I I I I I I I I I I I 4-41 TABLE 4-8. REPRESENTATIVE TOLERANCE LIMITS OF SELECTED PLANTS TO SIMULATED ACID PRECIPITATION3 Plant Species Birch Wi 1 low herb Scots pine Mosses Lichens Sunflower, bean Hardwoods Rad i sh Beet Carrot Mustard greens Sp i nach Swiss chard Tobacco Lettuce Cau 1 i f 1 ower Brocco 1 i Cabbage Brocco 1 i Potatoes Potatoes Alfalfa Kidney beans Oak Conifer seed 1 ings Mosses Chrysanthemums Juniper Yel low birch Pol lutant Concentration pH 2.0 - 2.5 pH 3.0 pH 4.0 pH 2.7 pH 2.5 pH 3.5 pH 4.0 pH 4.0 pH 3.5 pH 3.5 pH 3.0 pH 3.0 pH 3.0 pH 3.5 pH 3.5 pH 3.2 pH 2.0 pH 2.0 - 3.0 pH 3.0 pH 4.0+ pH 2.3 - 3.0 Species Effect Foliar lesions Reduced N fixation rate Fol iar damage Fol iar damage Fol iar damage Fo 1 iar damage Reduced yield Foliar damage and reduced marketabi 1 ity Fol iar damage Reduced yield Fol iar damage reduced yield Increased yield Fol iar damage Increased yield Inhibition of parasitic organisms Fol iar damage Desiccation, death Fo 1 i ar damage and increased phosphate uptake Growth decreased Fo 1 i ar damage Reference Abrahamsen et al . 1976 Denni son et al . 1976 Evans et al . 1977 Haines and Waide 1980 Lee et al . 1980 Shriner 1976 Strifler and Kuehn 1976 Teigen et al . 1976 Tukey 1980 Wood and Bormann 1976 The average precipitation pH in eastern North America is currently greater than or equal to pH 4.0. Individual storm events may have episodes where the pH drops into the range of pH 3.0 to 4.0. ------- 4-42 4.2.3.2 Acidic Deposition Effects to Forest Vegetation I I deposition? Unfortunately, only preliminary indications are avail- able in response to these questions (Lee et al. 1980). Accordingly, the dose-response function needs to be provided with many more I quantitative dose descriptions that relate to yield effects under B actual growing conditions. Information on the influence of other parameters on these dose-response functions also needs to be • provided. These factors include patterns of rainfall occurring as | they interact with stage of crop development, soil nutrient and water supplies, and deposition of particulate matter from the atmosphere. _ Further clarification also is needed of the possible modifying • influence of NC>3~ and S0^~ as nutrients in leaf tissue in response ™ to acidic rainfall events. Finally, the critical factors determining plant susceptibility, expressed as yield reductions, need further I definition to enhance extrapolation from a few of the most economi- | cally important crop species and cultivars to describe the response of the entire ecosystem. • When this information is provided, it may then be possible to make reasonable and reliable estimates of the economic impact of acidic deposition on agricultural productivity. I I Effects of acidic deposition on forest trees involves several considerations differing from those relating to agricultural crops. K Trees are perennial plants with long lifetimes. Thus, there is • greater concern with the cumulative impact or repeated exposures to acidic deposition. Furthermore, forests are usually in areas where soil nutrient supplies are limited, and are generally not supplied I with fertilizers or lime. Forests present large surface areas for I interception of gaseous and particulate pollutants from the atmos- phere, and these substances eventually move to the soil. Finally, • the composition of precipitation as it passes through the forest f system, the properties of soil, and characteristics of streams and lakes in watersheds are partially affected by the nature, age, and • condition of forests. Consequently, the effect of acidic deposition • on forests could also have important secondary impacts which are ™ initiated by direct effects on trees. The historic pattern of forest growth as revealed in the growth • rings may show "direct" evidence of the effects of acidic deposition. Based on substantial analysis of growth rings of Scots pine and • Norway spruce trees that grow in spatially-intermixed "more I susceptible" and "less susceptible" regions in south Sweden, Jonsson and Sundberg (1972) concluded that "acidification cannot be excluded as a possible cause of the poorer growth development, and may be • expected to have had an unfavourable effect on growth within the more • susceptible regions." This is a controversial study because other Scandinavian researchers have not been able to uncover similar • trends. For example, in a large study in Norway, Strand (1980) was | unable to "find definite evidence that acidic deposition has had an I I ------- I I I I I I I I I I I I I I I I I I I 4-43 effect on the growth of the trees". Studies of a similar type in North America have been limited in scope. Cogbill (1976), having examined historic patterns of growth rings in two forest stands (one a beech-birch-maple woods in New Hampshire and the other a spruce woods in Tennessee) observed that "no regional, synchronized decrease in radial increment was evident in the two mature stands studied." However, Johnson et al. (1981) noted both an abnormal decrease in growth of pitch pine on the New Jersey pine barrens, and a strong statistical relationship between stream pH (an index of precipitation pH) and growth. The relationship of these findings to other possible incitants (i.e., disease, insects, ozone) should be more fully explored. Experimental evidence from studies of the action of simulated acidic deposition on tree parts does indicate that under regimes of high acid dosing, direct damage (i.e., foliar lesions) can be produced (Table 4-8). A potential impact of acidic deposition may occur indirectly through the soil and may become involved in the complex natural circulation of elements upon which forest vegetation depends, (i.e., the nutrient or biogeochemical cycle). Rodin and Bazilevich (1967) describe this cycle of elements as "the uptake of elements from the soil and the atmosphere by living organisms, biosynthesis involving the formation of new complex compounds, and the return of elements to the soil and atmosphere with the annual return of part of the organic matter or with the death of the organisms." Interrelationships in the cycle are such that a change in one part of the system, if not counter- acted, could ultimately produce changes throughout. Generally, forests are relegated to soils which are of low fertility or, for some other reason, unsuited for agricultural use. In contrast to agricultural practice, amendments (i.e., fertilizers or lime) are rarely used in forestry practice. Deficiencies of nitrogen (N) are common in forests of the temperate and boreal regions. Appreciable responses to N-fertilizer have been reported frequently, particularly for conifers on upland sites in both the acidic deposition zone of eastern Canada (Foster and Morrison 1981), and in Scandinavia (Malm and Holler 1975; Moller 1972) . In a small number of fertilizer field trials carried out with conifers in Canadian forests, phosphorus (P), potassium (K), calcium (Ca) or magnesium (Mg) fertilizers did appear to elicit responses, though only when demand for N was first met (Foster and Morrison 1981; Morrison et al. 1977a,b). Growth of red pine and other conifers has been shown to be limited by K and Mg deficiency in restricted areas of New York State (Heiberg and White 1951; Leaf 1968, 1970; Stone 1953), and Quebec (Gagnon 1965; Lafond 1958; Swan 1962). ------- 4-44 I I The generally-held association of base-rich with more fertile soils and base-poor with less fertile soils (well-demonstrated in agricul- tural situations) has been investigated with forest species and soils • in only a limited number of instances. Pawluk and Arneman (1961) I associated better growth of jack pine on sites in Minnesota and Wisconsin with several soil factors which could be considered acidic deposition sensitive, including cation exchange capacity (CEC), I exchangeable K and percent base saturation. Also, in northern B Ontario Chrosciewicz (1963) associated better growth of jack pine with soils rich in basic minerals (and presumably richer in exchangeable bases). Hoyle and Mader (1964) noted a high degree of correlation between Ca content in foliage and height growth of red pine in western Massachusetts. Lowry (1972) across a wide range of mm sites in eastern Canada noted with black spruce relationships between H site index and foliage content (including N, P, Ca, and to a lesser extent Mg concentrations). Studies of forest soils (Lea et al. 1979) indicate that Ca and Mg • levels can be leached following applications of acidic deposition simulants. Leaching of these elements from forest soils, as a result of high S04 mobility (Mellitor and Raynal 1981), may lead to a chronic decrease in nutrient status of certain soils. Since nutrient availability is a significant growth-limiting factor • for many forest ecosystems, the concern is that acidic deposition * will interfere with uptake and cycling of various elements. First, acidic deposition may promote increased leaching of essential foliar I constituents (e.g., K, Ca and Mg) as a function of both acid- • related surface disintegration and mass exchange by HT1" ions. Both wet and dry deposition undergo chemical alteration directly on • the surface of the leaves and indirectly within the cellular tissue. The nature of the leachate or throughfall depends upon plant _ characteristics such as tree species, leaf morphology, stand I characteristics (e.g., age and stocking), and site conditions • (e.g., precipitation rate, distribution and chemical composition). Input/output analyses and element budgets with particular reference H to acidic deposition, have been described by various authors (Lakhani | and Miller 1980; Mayer and Ulrich 1980; Tukey 1980). Generalizations are difficult, because of the wide range of environmental (i.e., • soil, water, and climate) conditions. I Not all elements are leached equally and although all plant parts can be leached, young leaves are less suceptible to leaching than mature • foliage (Tukey 1980). Some elements (e.g., K) leach readily from • both living and dead parts, while others (e.g., Ca) leach more slowly. • Some researchers have found that throughfall from deciduous forests exhibit increased pH and higher Ca and Mg concentrations when • compared to the incident precipitation. In other instances the I opposite has been found. In studies of two hardwood species (i.e. , I I ------- I I I I I I I I I I I I I I I I I I I 4-45 sugar maple and red alder), little difference in throughfall chemistry was reported (Lee and Weber 1980). Stemflow from birch species shows increased acidity, relative to the incident precipita- tion (Abrahamsen et al. 1977). Beneath coniferous canopies, through- fall pH generally decreases relative to precipitation in open areas even though concentrations of Ca and Mg as well as many other dissolved ions may increase (Horntvedt and Joranger 1976). This ion enrichment is due to both washout of dry deposits and canopy leaching. It has been reported that 90% and 70% of the H+ in precipitation was retained in the forest canopy in New Hampshire northern hardwood (Hornbeck et al. 1977) and Washington Douglas-fir (Cole and Johnson 1977) forests, respectively. Leaching of low molecular weight organic acids from the canopy may decrease the pH of throughfall (Hoffman et al. 1980). Spruce canopies may filter dry pollutants from the atmosphere better than deciduous canopies. This cleansing action is partially attributed to the presence of spruce needles throughout the winter, during which S02 is dissolved in water films adhering to their surfaces. Subsequent removal of these deposits accounts for part of the difference in chemical composition of the throughfall. In summary, several processes may be affected when rainfall passes through a forest canopy. Substances residing on and in foliage are removed. These processes occur with both acidic and nonacidic deposition. Certain elements are leached more rapidly than others, especially when rainfall is acidic. There are also differences between species and stages of leaf development in rates of leaching. Leaching results in a marked change in the chemistry of precipitation before it reaches the soil. Dry deposits removed from leaf surfaces and substances lost from foliar tissues may neutralize or enhance acidity and the concentration of inorganic substances may increase considerably. More rapid transfer of elements to the soil provides opportunities for enhanced uptake and recycling by trees. Moreover, soil processes may also be affected by these deposits. Several pathways exist by which changes to precipitation occurring in the forest canopy can affect the chemistry of water transported through the terrestrial ecosystem and into streams and lakes. These are discussed further in other sections of this report. Acidic deposition may affect health and/or productivity of forest or other vegetation through indirect channels, or through effects on nutrition. Research efforts are just beginning to evaluate the possible role of acidic deposition in the predisposition of trees to disease infection and insect attacks. Further, the behaviour of plant litter and soil-occurring facultative saprobes, which may exhibit plant pathogenic tendencies under acidification, requires evaluation. ------- 4-46 4.2.4 Pollutant Combinations I I For much of the northeast and midwest sections of the United States • where acidic rainfall events and low dose SC>2 trends have been | recorded, ozone air pollution also occurs on a concomitant basis. Sulphur dioxide, NOX, and particulate emissions may be of "local" • importance to vegetation, but mesoscale background concentrations of these pollutants are well below known thresholds for inducement of direct vegetation effects. From these background concentrations, • long-term accumulation by plants of sulphates and nitrates and the • related potentially beneficial or detrimental effects are poorly defined. • Extrapolation from results of single pollutant effects on vegetation under ambient field conditions must be approached with caution. ^ Reactions in pollutant combinations may be additive (sum of effects), I less than additive (antagonistic), or more than additive ™ (synergistic). In addition to pollutant combinations under controlled conditions, the interaction of constantly changing • environmental factors and fluctuating pollutant doses must be further • evaluated before a conclusive statement of the importance of such interactions can be made. Reinert (1975) and Reinert et al. (1975) have prepared the most recent reviews of this area of investigation. I 4.2.4.1 S02 - 03 Effects I The most frequently occurring pollutant combination of significance to plant life must be considered as 03 and S02• However, few I studies have utilized doses which would be considered as even close | to ambient except as they pertain to areas affected by point sources of emission of S02- Studies using combinations of 03 and S(>2 • are presented in Table 4-9. As indicated, only the study of Houston • (1974) used doses of SC>2 approaching regional expected averages. He used mixtures of S02 and 03 in doses to simulate actual field conditions and reported that even the lowest concentrations of 03 fl (0.05 ppm) and S02 (0.05 ppm) for 6 hr in mixture caused more H serious damage than that resulting from either pollutant alone at similar concentrations. Studies by Tingey et al. (1971a,b, 1973), • Tingey and Reinert (1975), and Neely et al. (1977) used doses | reasonably expected in smaller areas such as the Ohio Valley (Mueller et al. 1980). Doses used in other studies used less realistic doses _ for either S02 or 03 and the results are of little value in • estimating field effects on a regional basis. — A recent study by Reich et al. (1982) utilized a linear gradient • field exposure system of S02 and 63 over soybeans exposed during • pod fill. Low dose exposure combinations averaged S02 at 0.040 ppm and 03 at 0.034 ppm for 5.5 hr per day for 12 days. Yield • I I ------- I I I I I I I I I I I I I I I I I I I 4-47 CO z ^ QL a o LU LU CO z O o Q •* CM 8 Ll_ o CO LU O X u. o to o LU Lu- ll. LU • ON •* LU _J m i— 0 o c 0 0 ce 0 4~ ID 0 > (D 0 73 t M- M- - C O in o 0 0 O> M- ID LU — O Ll_ -Q c 0 O L. — 3 4- O "O Q. c X O LU O ID 0 L. in £ O — o-l- X LU (D O C E cS a in 4_ Q) c — (0 O — 0 Q_ 0. 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Tingey et al. (1971a) demonstrated experimentally that a gaseous m mixture of 0.10 ppm S02 and 0.10 ppm N02 caused synergistic • effects with more than 5% leaf injury being induced on 5 of 6 plant species treated in 4-hr exposure periods. The symptoms of injury produced by a mixture of S02 and N02 can resemble those caused by • ozone which may make diagnoses in the field difficult. * I The currently available literature concerning the interactive effects _ of acidic deposition and gaseous air pollutants on terrestrial • vegetation is extremely limited, consisting of only three separate ™ studies. Shriner (1978) examined the interaction of acidic deposition and S02 or 03 on red kidney bean (Phaseolus vulgaris) M under greenhouse conditions. Treatments with simulated rain at pH • 4.0 and multiple 63 exposures resulted in a significant reduction in foliage dry weight. Simulated precipitation and sulphur dioxide m in combination did not affect either photosynthesis or biomass • production. Troiano et al. (1981) exposed two cultivars of soybean to ambient photochemical oxidant and simulated rain at pH 4.0, 3.4, _ and 2.8 in a field chamber system. The interactive effects of • oxidant and acidic deposition were inconclusive with seed germination ™ being greater in plants grown in the absence of oxidant at each acidity level. Irving and Miller (1980) also examined the response 4 of field-grown soybeans to simulated acidic deposition at pH 5.3 and P 3.1 in combination with sulphur dioxide and ambient ozone concentra- tions. No interactive effects of acid treatments with S02 on m soybean yield occurred. However, sulphur dioxide alone resulted in a • substantial yield reduction. Changes in such things as soil chemical properties nutrient • recycling resulting from acidic deposition do not occur rapidly. • After more than a decade of research in Scandinavia, the observed changes in chemical properties of forest soils that can be attributed M to acidic deposition still remain undetermined (Overrein et al. | 1980). It is therefore unlikely that interactive effects of acidic deposition and gaseous pollutants on plants involving changes in soil ^ properties will become evident within a single growing season. I A physical and chemical potential exists for interaction of various forms of wet fall and dry fall (including gases and trace metals) at, • on, or within leaf surfaces. However, very few studies have • addressed these interactions and the significance of the observed I I ------- I I I I I I I i I I I I I I I I e i i 4-51 phenomena remain inconclusive (Fuzzi 1978; Gravenhorst et al. 1978; Penkett et al. 1979). 4.3 EFFECTS OF ACIDIC DEPOSITION ON TERRESTRIAL WILDLIFE Although direct effects of acidic deposition are not likely, terrestrial wildlife may be indirectly affected in three ways: (1) contamination by heavy metals mobilized under acidic conditions; (2) loss of essential nutritional components from food; and (3) reduction in food resources. While the sensitivity of organisms such as plankton and fish to metals released in acid waters has been established (Baker and Schofield 1980; Marshall et al. 1981; Muniz and Leivestad 1980), the potential for accumulation and subsequent effects on terrestrial animals is less well understood. Metal contamination and reduced size of roe deer (Capreolus capreolus) antlers from an industrialized area of Poland has been reported recently (Jop 1979; Sawicka-Kapusta 1978). Acidification and sulphurization of roe deer browse (Sawicka-Kapusta 1978) was suggested as the cause of the high metal levels (Jop 1979). Such a means of contamination has been demonstrated in southeastern Denmark where cadmium and copper in epiphytic lichens and mosses were compared with those from northwestern Denmark (Gydesen et al. 1981). Epiphytes from the southeastern areas of Denmark which received elevated metal deposition in bulk precipitation showed metal levels 1.5 times higher on average. The same trend was found in the kidneys of cattle feeding in these areas. While direct deposition to plant surfaces may be partially responsible, plant uptake of some metals such as cadmium increases as soil pH decreases (Andersson and Nilsson 1974), and high plant metal content is another route of contamination. In Sweden, moose (Alces alces) closer to sources of anthropogenic sulphur supported higher tissue levels of cadmium and the body burden increased with age (Frank et al. 1981; Mattson et al. 1981). The mechanism of contamination was not explored and could be via terrestrial or aquatic vegetation. The availability of essential elements in wildlife nutrition may be affected by sulphur deposition and soil pH. Selenium, for example, is an essential element for vertebrates (Stadtman 1977). Selenium deficient conditions lead to degeneration of major body organs such as the liver, kidney and heart (Harr 1978; Schwarz and Foltz 1957). Most importantly from the viewpoint of ranchers, muscular dystrophy (known as white muscle disease) has been caused by selenium deficiency and reported in sheep, cattle, swine and horses (Harr 1978; Hidiroglou et al. 1965; Muth et al. 1958). The occurrence of white muscle disease in North American livestock is correlated to the concentration of selenium in forage (Allaway and Hodgson 1964). Lameness and poor growth and reproduction in domestic animals have resulted from selenium-deficient diets (Harr 1978). In poultry, edema (abnormal excess accumulation of fluid in connective tissue or ------- 4-52 I I I body cavities) has been related to selenium deficiency (Harr 1978; Patterson et al. 1957). Many of the soils in the temperate region of eastern North America • are low in selenium and hence produce forages which contain low selenium concentrations, frequently less than the 0.1 ppm minimum • level for animal health (Kubota et al. 1967; Levesque 1974; Winter • and Gupta 1979). Although selenium deficiencies in livestock have been associated with forages grown on soils naturally low in _ selenium, many incidences of deficiency have been attributed to the V high agricultural use of sulphate fertilizers (Allaway 1970; Davies 9 and Watkinson 1966). Calves reared on hay grown in Kapuskasing, Ontario, developed muscular dystrophy due to selenium deficient I conditions in the soil there (Lessard et al. 1968). flj Due to the correlation between soil selenium levels and concentra- M tions in plants grown on these soils (Muth and Allaway 1963), there w is evidence that wildlife forage plants are similarly selenium deficient. This was the finding in a study of moose browse plants in Alaska (Kubota et al. 1970). Moreover, selenium deficiency symptoms M have been reported for several wildlife species, (e.g., the prong- IP horn; Antelocapra americana) (Stoszek et al. 1978). Mountain goats (Oreamnos americanus) from an area where selenium levels in forage are low and where white muscle disease occurs in livestock, revealed symptoms of white muscle disease upon being stressed by handling (Herbert and Cowan 1971). It is suggested that the symptoms in wild — populations may well be masked by predation (Herbert and Cowan 1971). • The net effect of selenium deficiency diseases in wildlife would be ™ an increased susceptibility to predation as well as reduced productivity and survival of young. • Recent increases in anthropogenic sulphur emissions have caused concern regarding the influence on selenium availability in • vegetation. Selenium concentrations in plants in heavily industrial- • ized Denmark have decreased over the past decades (Gissel-Nielsen 1975). Experimental applications of S02 and SO/ to plants ,— and soils have demonstrated that selenium levels are depressed by • both the presence of sulphur and reduced soil pH (Shaw 1981a,b). ™ Because excessive sulphur and sulphate cause uptake of selenium to be reduced in plants (Davies and Watkinson 1966; Gissel-Nielsen 1973; fi Shaw 1981a), the impact in areas of low selenium soils could be | substantial. Furthermore, the solubility of selenium declines with pH, rendering selenium less available to be taken up by plants in M acid soils (Geering et al. 1968; Johnson 1975). | Sulphur and its compounds have a further depressing effect upon selenium in the animal itself. Excessive sulphur in the diet can • lead to increased elimination of selenium from the body (Harr 1978), • compounding deficiency conditions. 9 I I ------- I I I I I I I I I I I I I I I I I I I 4-53 Other essential elements in animal nutrition such as calcium, magnesium and sodium are similarly released from soils upon acidifi- cation (Abrahamsen 1980; Rorison 1980; Stuanes 1980). Accordingly, such elements will be less available for uptake by plants, resulting in lowered concentrations in plant tissues (Beeson 1941). Soil acidification similarly causes leaching of phosphorous which, if reflected in the vegetation (Rorison 1980), could have significant effects on wildlife nutrition. Lucas and Davis (1961) summarize the influence of pH on the availability of 12 plant nutrients. Aside from nutrient content, the availability of food resources may decline due to acidic deposition affecting the entire food web including wildlife. For example, caribou (Rangifer sp.) may be affected due to the sensitivity of lichens to sulphur and acid compounds (Lechowicz 1981; Puckett 1980; Sundstrom and Hallgren 1973). The importance of lichens in the winter diet of Canadian caribou herds is well documented (Kelsall 1960, 1968; Thompson and McCourt 1981). Thompson and McCourt (1981) reported that 67% of the diet of the Porcupine Caribou Herd of the Yukon consists of lichens. The George River caribou herd of Nouveau Quebec and Labrador is the largest in North America (Juniper 1979; Juniper and Mercer 1979; Mallory 1980) and may rely heavily on the carrying capacity of their winter range. Much of this area lies in the zone of acidic deposition (Figure 8-lb). Exposure of the primary caribou lichen (Cladina stellaris) to simulated acidic deposition with pH 4.0, reduced maximum photosynthesis by 27% and slowed recovery from dormancy after wetting by 14% (Lechowicz 1981). These results suggest that acidic deposition reduces the growth and productivity of this lichen (Lechowicz 1981). The significance of reduced lichen productivity to the population dynamics of these caribou herds is uncertain, because the degree to which they are food-limited is unknown. Another example of potential food loss involves herbaceous ground cover. Trees have tap roots in deep soil layers that are less susceptible to acidification, while plants draw their moisture and nutrients from the upper layers of soil making them more exposed to the effects of acidic deposition (Clark and Fischer 1981). Application of sulphuric acid in quantities corresponding to 100 kg/ha.yr killed much of the ground vegetation consisting mainly of mosses, lichens, and a species of dwarf shrub (Tamm et al. 1977). Therefore animals which feed on such vegetation may be affected by food loss. 4.4 EFFECTS ON SOIL Soils vary widely with respect to their properties (i.e., physical, biological, chemical and mineralogical), support different vegetation communities, are subjected to different cultural practices, are situated in different climatic zones, and are exposed ------- 4-54 I 1 to a broad spectrum of acid loadings making it difficult to generalize from findings indicated in this report. Further, there are various offsetting mechanisms, influencing effects of increased • precipitation acidity which vary with soil properties, vegetation • types, climatic regimes and cultural practices. Water moves through soils by uniform capillarity and gravitational processes. Also, A considerable moisture flow may be directed overland or may be <| channelized in the soil in root channels reducing the opportunity for equilibration. A high stone content can concentrate leaching effects ~ to a smaller soil volume than in nonstony soils. Thus, theoretical • calculations have to take into account particular ir± situ character- ™ istics. In the discussion which follows, the documented and hypothesized 9 effects of acidic deposition on soils are described under the following headings: • 1. Effects on Soil pH and Acidity. 2. Impact on Mobile Anion Availability, Base Leaching, and Cation V Availability. ™ 3. Influence on Soil Biota and Decomposition/Mineralization fl Activities. I 4. Influence on Phosphorus Availability. <^fl 5. Effects on Trace Element and Heavy Metal Mobilization and Toxicity. — I 4.4.1 Effects on Soil pH and Acidity acidifying sulphate fertilizers brings about appreciable soil acidification, along with other changes in soil chemical and • biological properties (Glass et al. 1980). The more striking of • these changes are reductions in exchangeable bases, increases in soluble aluminum and manganese levels, shifts in optimum conditions ^ for bacteria and mycorrhizal fungi, and reductions in soil micro- • faunal populations. Some of these undesirable changes have also been • shown to occur in the proximity of strong point emitters of sulphur dioxide (Freedman and Hutchinson 1980; Nyborg et al. 1976; Strojan • 1978), so concern is well-founded that the range of soil changes | outlined in Table 4-10 could occur to a greater or lesser extent over more widespread geographical areas. M I The process of soil acidification primarily involves the replacement of exchangeable basic cations (Ca, Mg, K, Na, NH^"1") by H+ and, at lower pH ranges, Al^"*" ions. The chemistry of soil acidification • is relatively well understood, at least in states other than strong • I I ------- I I I I I I I I I I I I I 1 I I I I I 4-55 TABLE 4-10. ACIDITY RELATED REACTIONS INFLUENCING AVAILABILITY OF SEVERAL ELEMENTS Element(s) Type of Reaction N Chiefly biological - biochemical; nitrifying bacteria decline with declining pH, thus ammoniacal-N predominates over nitrate-N; reduces mineralization. P Phosphate fixation reactions. K, Ca, Mg Chiefly mass displacement of absorbed bases by H and Al^+ ions. Fe, Mn Chiefly dissolution of hydroxides in acid solution; organic status, redox important particularly for Fe. Al pH regulated solubility of Al-oxy and hydroxy compounds. ------- 4-56 1 i acid as described by Jenny (1961), Wiklander (1973/74, 1975, 1980a), Bolt and Bruggenwert (1978), Bache (1980), and Nilsson (1980). In the most strongly acid soils, there is evidence that aluminum becomes • very mobile without there being any associated notable change in pH ™ (Cronan and Schofield 1979; Norton et al. 1981; Ouellet 1981; Ulrich et al. 1980). The common range of pH for soils in humid regions is fl about pH 5.0 to 7.0, with the preferred range for cultivated soils || being pH 6.0 to 7.0. Many forested soils, particularly under coniferous cover, fall within the range pH 4.5-5.5 in the mineral • horizons, with surface organic layers commonly exhibiting pHs in the • range 3.5 to 4.5. The numbers of field situations where investigators have been able to M compare present with former soil pH values are extremely limited. 9 However, Linzon and Temple (1980) report a lowering of soil pH in the brunizolics, but not podzols, of south-central Ontario after 18 years ft of pollutant deposition. Ulrich (1980b) and his colleagues (Ulrich \| et al. 1980) working in the more heavily polluted parts of central Germany report a long-term fluctuation of pH in the surface humus — under beech and spruce. The pH values do not show a steady decline, Tm but rather show cyclic variation between 4.2 and 3.8. This parallels ™ deacidification and acidification phases alternating between cooler, moister summers and warmer, drier ones. From 1969 to 1980 under M beech, and from 1973 to 1980 under spruce, there were substantial • increases in the amounts of soil aluminum mobilized. These increases were associated with the continued entrapment and deposition of acid • sulphate pollutant. • Various field and laboratory experiments of a simulation nature have _ also been set up to examine the effects of acidic deposition on soil fl acidity. Results indicate that artificial acidic deposition at pH<4 ™ can lead to measurable decreases in soil pH (Abrahamsen et al. 1976; Bjor and Teigen 1980; Stuanes 1980). For example, simulated acidic 4 deposition inputs of pH 4.0 and below to spruce podzol soils in ^ Norway caused soil acidification of the 0, A, and B horizons (Abrahamsen et al. 1976). In some cases, the soil pH depression over ------- I I I I I I I I I I 1 I I I 1 I I I 1 4-57 pH, 4.9), carbonic acid contributed approximately twice as much H"1" to the soil as did precipitation. However, a drop in pH to 4.0 (about 30 times more acid than normal) occurs in the most heavily impacted areas of eastern North America (Cogbill and Likens 1974) and results in H+ inputs far in excess of that produced by carbonic acid. In the more acid soils having a pH of less than 5.5 (e.g., podzols developed under coniferous forests), organic acids contribute significantly to natural soil acidification. It is not as yet known what role anthropogenic acidification will have in these ecosystems. Presumably, extremely acid soils will experience the least change in pH, but changes in the ionic make-up of the soil colloidal complex and ionic mobility may take place. Sollins et al. (1980) proposed a comprehensive scheme for calculating H+ ion budgets in forest ecosystems, based upon measured mass balances of cations and anions within the nutrient cycles. Andersson et al. (1980b) used this model to obtain H+ ion budgets for forest ecosystems in Sweden, West Germany, and Oregon. In the heavily impacted Soiling site in West Germany, their analysis shows that atmospheric H+ ion inputs are small (approximately 10%) compared to net internal flows. Ulrich (1980b), using essentially the same approach, stressed input-output balances to assess the long-term net acidification of soils caused by internal compensations of H+ production and consumption and uptake and mineralization processes. He also pointed out important spatial considerations within the soil profile. For example, ammonium mineralization that consumes hydrogen ions might occur in litter layers while ammonium that produces hydrogen ions may occur in mineral soil layers at the same time. Some indication of orders of magnitude of H"1" ion contribution by softwood versus hardwood forest and their relationship to anthropogenic loading, were provided (Ulrich 1980b). Total H+ ion input was determined as about ca 0.81 keq/ha, of which 0.79 keq/ha was considered man-made. A beech canopy generated an additional ca 0.58 keq/ha and a spruce canopy, an additional 2.28 keq/ha. This evidence suggests that as mean pH of rainfall declines below pH 4, its contribution to the H+ ion balance is not insignificant even in comparison to spruce forest H+ ion production. Thus, the process of podzolization is hastened. As noted earlier, the adverse effect of soil acidification results chiefly from the influence of changed pH on other processes (e.g., soil biochemical reactions and N availability, organic matter turnover, mobilization of trace elements, and transformation of clay minerals) . 4.4.2 Impact on Mobile Anion Availability and Base Leaching Acidification and soil impoverishment involves the displacement of basic cations (i.e., K, Ca, Mg, Na) from exchange surfaces, their replacement by H"1" and Al^+ ions, and the establishment of new exchange/solution equilibria (Wiklander 1973/74). Under natural ------- 4-58 1 I conditions, two sets of processes seem to be involved: (1) there are exchange processes whereby H"1" ions displace base cations from the _ exchange surface, and (2) there are the processes whereby the • exchanged ions are transported within the soil column under the ™ influence of mobile anions (Johnson and Cole 1976, 1977). _ capacity or CEC and the relative degree of saturation of the CEC with bases or base saturation). In humid regions, the total permanent and m pH dependant CEC of a productive soil under cultivation might range I from 15 to 30 meq/100 g. In the surface horizons this might be higher and in the subsoil it may be lower. To illustrate this, in coniferous podzols the CEC of the humus layer may be high while I beneath it values decrease abruptly with depth. It is presumed that Qi the loss, particularly of those base cations of nutritive value (chiefly K, Ca, and Mg), could be accelerated under acidic deposition, with attendant adverse impacts on forest growth. I Various "simulated acidic deposition" leaching experiments are » described in the literature (Abrahamsen et al. 1976; Abrahamsen and M Stuanes 1980; Lee and Weber 1980; Morrison 1981; Overrein 1972; * Roberts et al. 1980; Singh et al . 1980). In some controlled irrigation experiments, Ca and Mg appear to be the most affected and • K the least affected (Abrahamsen 1980; Hovland et al. 1980; Ogner • and Teigen 1981; Wood and Bormann 1976). To some extent, this may reflect the relative amounts of these cations on soil exchange sites, • but the rate of increase in K depletion seems to be consistently • below that for Ca or Mg under acid irrigation as well (Abrahamsen 1980; Ogner and Teigen 1981; Wood and Bormann 1976). The relative ^ lack of response in K may also be due to the greater plant fl requirements for K, as opposed to Ca or Mg, and possibly also to ~ fixation of K in 2:1 clays. 9 In some cases, the accelerated cation leaching has led to net depletion of available cations in the rooting zone. Significant reductions of base saturation percentage were noted in the 0 and A • horizons in Norwegian spruce podzol soils, following applications of • simulated acidic deposition with a pH of 3.0 or lower (Abrahamsen 1980). ^ Soil acidification and decreases in base saturation do not always ™ occur concurrently. Under natural soil acidification by humic acids, production of humus increases CEC, but does not increase the cation B content (Konova 1966). Soil pH and base saturation will thus | decrease without a corresponding reduction in exchangeable base content (Ulrich 1980a). Similarly, with anthropogenic acidification, • soil pH and base saturation may decrease, with no corresponding net I nutrient loss. This occurs if the soil is actively adsorbing both H+ and SO^-, which would increase CEC over time (Johnson and Cole 1977). In addition, decreases in base saturation and pH in • soils subjected to leaching losses of base cations can be offset to ™ I I ------- I 1 t I t 1 I f I I 1 I I 1 I I I I I 4-59 some extent by acid-induced increases in soil weathering (Johnson 1980). Much of the potential impact of atmospheric deposition stems from the input of the mobile S0^~ anion to soils. Whereas the mobility of bicarbonate or organic anions may be severely limited in many acid or clay-rich northern soils, SO^" anions may be very mobile in these same soils. It has been shown that atmospheric H2S04 inputs overwhelm natural leaching processes, in some New Hampshire Spodosols, causing perhaps a threefold increase in the natural rate of cation denudation, and marked increases in the leaching of soluble inorganic Al. In New Hampshire subalpine coniferous soils, anthro- pogenic 504^" anions supplied 76% of the electrical charge balance of the leaching solution, while A1.3+ and H+ were the dominant cations in solution (Cronan 1980; Cronan et al. 1978; Cronan and Schofield 1979). In contrast, some soils (chiefly those rich in Fe- and Al-sesquioxides) exhibit a substantial capacity to adsorb S0^~, and thus demonstrate a considerable initial resistance to base leaching by anthropogenic ^864 (Johnson and Cole 1977; Johnson and Henderson 1979; Morrison 1981; Roberts et al. 1980; Singh et al. 1980). This generally implies that the effect of acidic deposition on soil cation leaching is highly dependent upon the mobility of the anion associated with the acid, whether it be 864^", N03~, or an organic anion (Cronan 1980; Johnson and Cole 1980; Seip 1980). This is due to the requirement for charge balance in the soil solution, a necessary condition that precludes the leaching of cations without associated mobile anions. Soils low in free Fe and Al, or high in organic matter (the latter appears to block sulphate adsorption sites, [Johnson et al. 1979, 1980]) are therefore generally susceptible to leaching by H2S04 (e.g., Cronan et al. 1978). Where SO,2- adsorption does occur, (e.g., in the highly weathered soils of Tennessee), S accumulation could initially be beneficial in three ways: (1) prevent cation leaching by H2S04 by immobilization of the 804^" anion; (2) create new cation exchange sites; and (3) release OH~ from adsorption surfaces (Johnson et al. 1981). It follows, however, that once 804^" exchange sites become fully occupied, cation leaching could commence. On Walker Branch Watershed, 48% of total S to input accumulates in the soil, whereas only 13% accumulates in vegetation (Johnson and Henderson 1979; Shriner and Henderson 1978). Along the same lines, one might expect the N03 in acidic deposition to contribute to net cation leaching only in those systems where N03~ is mobile. Because of the N-limited status of many forests, most N03 tends to be assimilated by plants during the growing season, thereby not contributing to cation leaching. ------- 4-60 4.4.3 Influence of Soil Biota and Decomposition/Mineralization Activities 1 1 t It has been postulated that atmospheric deposition of strong acids may adversely affect soil biota and decomposition activities either directly through soil acidification or indirectly through trace tt metal mobilization and toxicity. Laboratory experiments and • observations on soils in close proximity to pollutant sources provide information on changes which occur in soil biota, as a result of — increased acidic deposition. Observations indicate decreases in j total numbers of soil bacteria and actinomycetes, and some relative ™ increase in presence of fungi; although, under conditions of very high loading, fungi have been reported less abundant. Generally, H total numbers of enchytraeids have not been affected (except under £ extreme conditions), though differential species responses have been reported (Abrahamsen et al. 1976, 1977; Alexander 1980; Baath et al. * 1980). '| The available evidence on the effect of acidity on organic matter breakdown and soil respiration is not conclusive (Rippon 1980; Tamm ^B> et al. 1977). However, decomposition experiments suggest that acidic ™ deposition may retard organic matter decomposition. Studies (Baath et al. 1979, 1980; Francis et al. 1980; Lohm 1980; Tamm et al. 1976) ft have noted decreased decomposition or carbon mineralization in soils || and litter exposed to artificial acidic deposition inputs at pHs below 3.5 to 3.0. Meanwhile, other studies have shown little or no ^ effect (Abrahamsen 1980; Hovland et al. 1980). Clearly, the results • are partly dependent on soil type and severity of the simulated ™ acidic deposition treatment. In some soils, there are indications that acidic deposition may alter • humic/fulvic acid dynamics. While moderate acidity may aggregate humic acid particles, it may lead to dissolution and mobilization of • fulvic acids. In soils like Podzols, which contain appreciable • quantities of fulvic acid, substantial losses could occur in moderate acidic leaching. Besides carbon cycling, there is concern that acidic deposition may * have adverse effects on N cycling patterns and processes. In this case, there are actually two potential sides to the issue: (1) the • possibility that acidic deposition may decrease N mineralization and | availability, and (2) the possibility that atmospheric inputs of anthropogenic N compounds may provide a fertilizer effect by increas- M ing the amount of available nitrogen. Tamm (1976) predicted short- • term increases in N availability and tree growth, due to net N losses from ecosystems. In Germany, Ulrich et al. (1980) resampled soils ^ over a 13-year period and showed significant accumulations of N-poor • organic matter in the forest floor of a 120-year old beech forest. w This was interpreted as a condition which could lead to internal H+ ion production, immobilization of N, and mobilization of soluble • Al3+. other studies, by Francis et al. (1980) and Alexander £ (1980) show ammonification and nitrification may decrease markedly in I I ------- I 1 I I I I I I I I i I i i i i t i i 4-61 soils exposed to artificial rain at pHs approaching 3.0. However, several studies have demonstrated increased N availability, at least during the initial stages of H2S04 input (Abrahamsen 1980; Ogner and Teigen 1981; Roberts et al. 1980), and this has produced minor growth increases in situations where N is limiting (Abrahamsen 1980; Tamm and Wiklander 1980; Tveite and Abrahamsen 1980). Whether this increase in N availability is due to change in microbial activity, or to the acid catalyzed hydrolysis of labile soil N, is unknown as yet. Norwegian studies show that both N availability and N03~ leaching were stimulated by H2S04 inputs-. This strongly suggests that contrary to earlier predictions (Tamm 1976), nitrification can be stimulated by acid inputs as well. This has definite negative long-term implications for forest N and cation status, if NC>3~ production exceeds plant uptake, resulting in net ecosystem N and cation loss. 4.4.4 Influences on Availability of Phosphorus Like N, phosphorus (P) is an essential element for plant life. In soil, P occurs in both inorganic and organic compounds. It is utilized from the soil solution by plants chiefly, though not entirely, as the (inorganic) orthophosphate anion. For perennial plants, including trees, P is assimilated through the intermediary mechanism of a mycorrhizal root association (Bowen 1973; Fogel 1980; Hayman 1980). The availability of P to plants is determined to a large extent by the ionic form in which it is present. In soil solutions of low pH, available P is present largely as H2P04"; as pH increases, HP042~ predominates. In strongly acid soils, H2P04~ ions may react with soluble Mn, Al and Fe compounds and be mostly precipitated as the insoluble and nonavailable metal hydroxyphosphate (Hsu and Jackson 1960). Also, under conditions of increasing acidity, H2P04~ tends to react with the insoluble oxides of Fe, Al and Mn, and in more weathered soils it may become fixed on silicate clays, through the process of anion exchange. 4.4.5 Effects on Trace Element and Heavy Metal Mobilization and Toxicity A further effect of acidic deposition or increased soil acidification is an increased solubilization of heavy metals in the soil system. This can arise from the increased solubilization of metals that are already present in mostly insoluble or nontoxic forms or it may arise from metals being deposited along with an acidifying pollutant. Thus, at low concentrations naturally present Mn and Fe serve as essential nutrient elements for the growth of higher plants and except in alkaline or calcareous soils are usually present in adequate available amounts. However, at high concentrations these metals and Al can cause nutritional imbalance and growth impairment. Different plant species vary in their susceptibility to heavy metals, ------- 4-62 I 1 an example being Al. Barley, sugar beet, corn and alfalfa are very sensitive, whereas ericaceous shrubs and conifers appear much more tolerant. Soil characteristics also affect tolerance/susceptibility fl including soil pH, Ntfy"1" compared with NC>3~ nutrition, Al-exclusion V processes, Ca and P status, and organic-Al complexation (Foy et al. 1978). Striking examples of the effect of soil pH on the solubility * of Mn and Al are given in Glass et al. (1980), concentrations rising • very rapidly in each suite of soils when pH values moved from 5 to 1 In acid forest soils that support highly productive forest in eastern • Canada it is not unusual to have a pH gradient down the profile of from 3.0 to 4.6. Associated with such values are exchangeable Ca A values falling from 2.0 to 0.15 meq/lOOg and exchangeable Al values "p falling from 7.0 to 0.40 meq/lOOg (Anonymous 1979). Of a very much wider Ca/Al ratio, however, is the Soiling soil profile in Germany « described by Ulrich et al. (1971), where exchangeable Ca and Al • concentrations are respectively 0.2 and 4.7 meq/lOOg, and where Gb'ttsche's beech studies in the acidification year of 1969 are plotted to reveal the remarkable correlation between the seasonal • increase in soluble soil Al concentrations and the dramatic increase w in fine-root mortality (Ulrich 1980b). Indeed, this correlation and other studies have encouraged Ulrich (1981) to advance his ecosystem • hypothesis explaining the widespread "die-back" of fir in Europe. « Fine roots are killed by high soil Al concentrations or high Al/Ca ratios with a subsequent invasion of the damaged tree tissues by rot ^ fungi. There is evidence to indicate that increased amounts of • aluminum can be mobilized in the soil and passed on to water bodies * (Abrahamsen et al. 1976; Cronan and Schofield 1979). It is not clear whether the allegedly toxic concentrations present in the • loess-derived forest soils of central Germany can also be expected to • arise in the glacial till-derived soils of Scandinavia or northeastern North America (Tyler 1981). 1 Soil acidification in environments where there is also appreciable deposition of heavy metals is the second area of concern. Heavy ^ metals arise from various industrial activities, including fuel • combustion (Hansen and Fisher 1980; Watanabe et al. 1980). The scale * of emissions and airborne transportation has caused increasing attention to be directed to the amounts of different elements being I deposited in remote rural areas. Thus, at the Soiling site in m central Germany, for an open-site wet deposition of 23.8 kg of sulphur per hectare per year, there is an accompanying 10 kg of • nitrogen, 10.4 kg of calcium, 1.9 kg of magnesium and 1.1 kg of • aluminum (Ulrich 1980b). In south-central Ontario recent comparable figures are 10 kg for S, 6 kg for N, 5 kg for Ca, and 0.7 kg for Mg - (Scheider et al. 1979). For the same locality figures for elements • more commonly understood as "heavy" are 0.46 kg for aluminum, 0.54 kg ™ for iron, 0.095 kg for zinc, 0.132 kg for lead, 0.033 kg for copper and 0.022 kg for nickel (Jeffries and Snyder 1981). These authors 4| also point to the much higher deposition rates near smelters where £ cumulative levels of heavy metals in the soils have exercised I 1 ------- I I I I t I I I I I I I I I 1 1 I I I 4-63 pronounced toxic effects upon the vegetation (Hutchinson and Whitby 1974, 1976). However, there is danger in extrapolating from such heavily polluted local situations onto the more diffuse regional scale without taking into account the different parameters. Nevertheless, if the Soiling site is taken as exemplifying the more diffuse rural situation, Heinrichs and Mayer (1977, 1980) found that the beech and spruce forests act as filters for atmospheric substances. Some elements (e.g., sulphur, lead, mercury, bismuth and thallium) are largely accumulated in the upper part of the soil profile but the complex biogeochemical picture that emerges suggests that far more needs to be known in other locations on the fate of deposited metals having potentially significant physiological and toxicological roles (Andersson et al. 1980a; Bradley et al. 1981; Smith and Siccama 1981). There is a rapidly expanding literature focusing on the soil behaviour of heavy metals derived from town wastes (Leeper 1978) and much of this related to pH-dependent considerations (Hatton and Pickering 1980; McBride and Blasiak 1979) and metal-organic compound complexes (Bloom et al. 1979; Marinsky et al. 1980; McBride 1980) should be applicable to the acidic deposition problem. The dissolution and mobilization of many other trace metals in soils is also affected by acidic deposition and decreasing soil pH. Recent studies in the Adirondack Mountains of New York have determined from acidic leaching experiments on native bedrock that this process is an important contributor of Cu, Pb, and Hg in addition to Al (Fuhs et al. 1981). The trace metals Cu, Pb, Hg, Cd, and Zn were leached rapidly upon exposure to acid while Al and other major metals were leached more gradually. Leaching of soils and bedrock by long-term acidic deposition has resulted in soil impoverishment for metals such as Mn and Zn in New Zealand (Norton et al. 1981). Other studies have demonstrated accumulations of trace metals in soils. Norton et al. (1980) found Pb and chemically similar metals accumulating in soils while Al and Mn were being leached. Leaching occurred in the upper soil horizons resulting in potential impoverishment for shallow rooted plants. Deeper rooted plants, on the other hand, are subjected to potentially toxic concentrations of dissolved metals. Tyler (1978) also showed that Pb is not readily leached from surface soils by acidic deposition inputs. Although the solubility of this element increases with decreasing pH, most soils contain sufficient organic matter to tie up the Pb as insoluble organic - Pb complexes in the soil matrix. Mobility and transport within the soil horizons and direct atmospheric deposition is responsible for the accumulation of metals in the soil. For example, concentrations of Cu and Ni increased in soils with proximity to the Sudbury, Ontario smelter (Heale 1980). Studies of metal deposition in the Walker Branch Watershed in Tennessee, found that soils efficiently retained Pb, Cd, and Cu, and less readily accumulated Cr, Mn, Zn, and Hg (Andren et al. 1975). McColl (1980), however, found that the concentrations of Mn, Fe, Cu, and Zn were all greater in ------- 4-64 soil solutions than in acidic deposition falling in Berkeley, California. I 1 I In restricted areas, vegetation may be stunted or absent due to toxicity of metals such as Ni (Foy et al. 1978). In a well-known study on serpentine-derived soils in Czechoslovakia, Nevmec (1954) M attributed the failure of pine plantations and various hardwood • species to excessive levels of Ni , Cr, and Co. Plantation failure was considerably reduced by fertilization with lime and diabase dust. Around Sudbury, Canada, Ni and Cu added from atmospheric deposition "W from smelters are maintained in acidified soils in concentrations • sufficiently high to be toxic to vegetation (Hutchinson and Whitby 1974, 1976). Thus, any possibility of mobilization of trace metals 1^ through decreasing soil pH by acidic deposition has implications for ^ forest productivity. Accumulations of trace metals from atmospheric deposition can contribute to this problem. £ 4.5 SENSITIVITY ASSESSMENT Several sets of sensitivity criteria have been proposed and used to M define geographical regions most susceptible to acidic deposition effects (Johnson and Olson in press). Each set of criteria is based ft upon a different philosophy and is aimed at different target £ organisms or ecosystems (e.g., forests, fish, soil, bedrock, aquatic ecosystems). Those directed toward aquatic effects have emphasized « bedrock geology (Hendrey et al. 1980; Norton 1980) or bedrock geology • and soils in combination (Cowell et al. 1981; Glass et al. 1982; see ™ Section 3.5). Those directed toward terrestrial effects have emphasized cation exchange capacity and base saturation (Klopatek • et al. 1980; McFee 1980a,b; Wang and Coote 1981). • Terrestrial sensitivity has been defined in terms of forest Ife productivity (Cowell et al. 1981; Table 4-11) and in terms of soil || acidification (Wiklander 1973/74, 1980b; Table 4-12). In both cases effects in the soil body were emphasized. Cowell et al. (1981) ^ regarded low pH soils as the most sensitive based on the assumption ^m that these already had the smallest reserve of nutrient cations. * Thus, any additional loss of forest nutrient cations, however small, would be significant to forest productivity in acid soil systems • (even though these soils were less sensitive to acidification). This *i sensitivity assessment concentrated on the upper 25 cm of the soil profile where, at least in boreal ecosystems, nutrient cycling is • most efficient. Acid soils are known to actively adsorb SO^", • hence reduce cation mobilization, and are considered less sensitive than nonsulphate-adsorbing soils (Johnson and Cole 1977; Singh et al. ^ 1980). This contrasts with the sensitivity concept suggested by • Wiklander (1973/74, 1980b) whereby noncalcareous, moderately acid * sandy soils (pH 5-6) with low cation exchange capacities are considered most sensitive. Wiklander (1973/74, 1980b) derived these I criteria from laboratory studies in which he found that the cation 9 displacing efficiency of H+ was greatly diminished as base I I ------- I 1 I i i i i i i i i i i i t i i i i 4-65 1 h- to -^ LU CO - 1— s 8 h- CO n i g 01 CO CO LU CO CO o l__ CO l_ y _ 1 — oc LU |— 5 a 8 CO CO <- Q z- CO •— ce co O ON 1— — o U- — ro _i < 4- — CD cc 1 CO — LU CD ce a ct: o LU CJ i— — . V *~ 1 LU CD £ >- — _ h- co LU CO i X LU ce 1 8 j CO g O U- ^ ce i- co LU Ct LU I— >- c o io" -i- 5 — — Q. L. CD O L. CD 4- . 04- c E *ft m 4- U O inom roo TJ ro oo • *. 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JS bO bO -r-l •H rH a to CU J2 bO •H cfl pr| j ^ 0) (-< bO O ^J o cd CU t-3 CU 43 bO bO M •H CO ffi (_J (~j bO i-H •iH Cfl <~1 Q iH C7 Cfl cu S ^ 4J CO d •r-l cfl bO C cfl O CU -H bO bO 4J a c c •rH Cfl CU -4 XI 4J CU O CU MH 1 r4 MH a 3 Q. + PQ a 1 4-1 bO •^ 4J rH XI co bo •H V* M CO CU > 1 4-1 A bO rH 1 co i-H v | cu >, fl M O 0) 2 > CU rH CO I-l CU ^Hi ""C V | -H CO e o o CU 4-> cfl -H CU V j T? O s cu rH C! V 0 CO 4-1 CJ CU 14^ MH CU rH 0) iH CO O a >-< co < CU \ > a S T3 0 I 1 I i i l i l I I i i i i t i i l I ------- I I I I I I I I 1 I 1 I I I I I I I I 4-67 saturation and pH decreased. Thus, for a given H+ input, very acid soils will yield fewer cations and are classed as less sensitive than moderately acid soils. Moderately acid soils with low cation exchange capacity (i.e., less buffering by exchange sites) will experience more rapid pH-change than very acid soils with the same exchange capacity. This concept of assessing soil acidification potential, in which the most sensitive soils are those experiencing the greatest change in their inherent properties, is specifically a soil sensitivity evaluation. No cause-effect relationships with vegetative or aquatic systems are specified. 4.5.1 Terrestrial Sensitivity Interpretations Acidic deposition may cause increases as well as decreases in forest productivity (Abrahamsen 1980; Cowling and Dochinger 1980). The net effect on forest growth depends upon a number of site specific factors such as nutrient status and amount and composition of atmospheric acid input. In cases where nutrient cations are abundant and S or N are deficient, moderate inputs of acid may actually increase forest growth. At the other extreme, acidic deposition in sufficient amounts may reduce productivity on sites with adequate N and S but deficient in cations. Other detrimental effects to forest productivity include changes in soil, microorganisms and Al toxicity. These effects are increased (Ulrich et al. 1980) with increased acidification. However, there is insufficient empirical evidence establishing cause-effect linkages between forest productivity and acidic deposition. It is not certain which ecosystem factors are most significant with respect to forest systems and thus it is not presently possible to map forest productivity sensitivity at any scale. Sensitivity assessment for this section, therefore, will concentrate on soil characteristics and how pH, CEC and sulphate adsorption properties hypothetically relate to different effects. Terrestrial ecosystem effects to be considered are: loss of base cations, soil acidification and Al solubilization. Table 4-13 and Figures 4-5 and 4-6 depict hypothetical sensitivities (Johnson and Olson in press). For nonsulphate-adsorbing soils (Figure 4-5), it is assumed that each equivalent of incoming H+ causes the leaching of an equivalent of some cation (including H+ or Al^+) through the forest soil. Case 1. For soils with pH >6, H+-base cation exchange is likely to be nearly 100% efficient (Wiklander 1973/74, 1980b), and thus soils are very "sensitive" to base cation loss (Figure 4-5a). If the soil with pH >6 has a high CEC (i.e., a large reserve of exchangeable cations and hence a large buffering capacity), it will take a very long time for a given acid input to acidify it. This is depicted by the width of the CEC box in Figure 4-5a. Thus, such a soil is thought to have low sensitivity to acidification and Al mobilization. ------- 4-68 C ,-J cd M 0 (3 co o CO Z C MX: > o M >-} o a co o EH I-l CJ <4H W [JLj *^J pL| CU W *iH 2; -H O T) M O H a M ^ CO On'cd W XI W "^^^ cr CJ O> M - — o O C"H W ^^' CO CO CO H 0) iJ S l-i *H M 4J EH cd M iH O CO -H -rl 6: 04-1 W CO -r-l CO TJ •rl rH 0 M OH EH CO r-r"! OH a! W EH CO CO o ^14-14-14-1 o) cd cd cd *c ^ Tj ^i M tn bO O O 0) 0) CU -H rJ 8 T) ^3 TJ PS 1000 3 g g g O rH 3 4-1 4-1 CO x! X! X! X! cd cd C bO bO bO bO M M O -H -iH i-l -i-l ------- I I I I I I I 1 I I I I I I 1 I >7 6 SOIL 5 4 pH pH >7 6 5 4 3-4 > 7 6 SOIL 5 4 3-4 pH Figure 4-5. NONSULPHATE-ADSORBING SOILS (a) CATION EXCHANGE CAPACITY HIGH 2H+ 100 %BASE SATURATION 804" CATION EXCHANGE BASE CATIONS (b) CATION EXCHANGE CAPACITY HIGH SO?" 100 % BASE SATURATION 0 LOW •4—»> H+AI3+ CATIONS; V///// CATION EXCHANGE H+,AI3+ BASE CATIONS (c) CATION EXCHANGE CAPACITY 2H+ 100 % BASE SATURATION H+.AI34" LOW //////A/// BASE CATIONS CATION EXCHANGE H+AI3 + BASE CATIONS V 4-69 INPUT SOIL INTERACTIONS OUTPUT 804" INPUT SOIL INTERACTIONS SO?" OUTPUT So|" INPUT SOIL INTERACTIONS OUTPUT Effects on base cation loss, soil acidification and Al->+ solubilization for nonsulphate-adsorbing soils having (a) moderate to high pH ( >6), (b) moderate pH (5-6) and (c) low pH ( <5) (Johnson and Olson in press). ------- 4-70 I 1 Case 2. If the soil with pH >6 has a low CEC (area depicted to the right of the dashed line in the CEC box in Figure 4-5a), it will take less time to deplete the exchangeable cation reserves and, therefore, • a low-moderate rating is arbitrarily assigned to acidification and Al f mobilization in soils to differentiate it from case 1. As in case 1, H+-cation exchange is nearly 100% complete, so that soils are ^ "sensitive" to base cation loss. • Case 3. If a soil has pH 5-6 (i.e., a moderate base saturation), H+-cation exchange will be nearly as complete as in cases 1 and 2 • while cation reserves (at a given CEC) will be lower (Figure 4-5b). H For the high CEC case, a moderate rating is assigned to acidification and Al mobilization in terrestrial ecosystems. As in cases 1 and 2, • soils are "sensitive" to base cation loss. V Case 4. In this case, the total reserves of base cations are low _ yet H+-cation exchange is nearly 100% efficient (Figure 4-5b) and I thus the soil is highly sensitive to base cation loss and ™ acidification. Once base cations are depleted and the soil is acidified, Al may become mobilized; thus, a moderate rating is I assigned to soil Al mobilization. • Case 5. In soils with pH <5 (i.e., low base saturation), H^-base • cation exchange is less efficient and therefore soils are only