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'               IMPACT ASSESSMENT
|                 WORK GROUP I
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•                 FINAL REPORT
                  JANUARY 1983
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                                        WORK GROUP  I
                                   Work Group  Co-Chairmen

                                    G.E.  Bangay,  Canada
                                 C.  Riordan, United  States
                                        February 1983
                                       IMPACT ASSESSMENT
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 'ง                                       FINAL REPORT
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                               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.

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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

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                                  TABLE OF CONTENTS
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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





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                    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

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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

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3-30
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                    TABLE OF CONTENTS (continued)                               •
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                                                       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
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             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              •


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                    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
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 3-107


 3-108


 3-109


 3-112


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 3-126


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              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 •ง- V\ a >• A •*• .-•ปซ-ปฃ•>  s\f IT o cซ <- *^ i—1-ป  f* ** Y^ *\ A f\       *\,~ 1 ft /i            ^"
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                   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

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_                           3.9.3.3   Summary                                 3-212
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                                   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

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                    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

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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



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                    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
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6-11
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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





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                    TABLE OF CONTENTS (continued)                              •


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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
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             7.4.1.3  Irreversibilities and the All  or                         •
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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

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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

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               2-4        Precipitation amount-weighted mean annual         2-14
                          pH in North America.
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                                         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.
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               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



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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


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                                                                   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
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               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.

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                                                                 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


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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


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                                                                xxv ii
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             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

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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

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                                                                  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.

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xxxii
                                                                               I
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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
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                                                                  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.

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                                                                     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
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                                                                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)

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                                                                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)
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                                                               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

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                                                              xxxviii
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(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




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SECTION I



SUMMARY

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                                                                  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.

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                                                                  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.
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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

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                                                                  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.

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                                                            1-4
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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.
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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).

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     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                 ซ
     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
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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

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                                                                  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
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     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            ซ
     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.
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     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.

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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               _
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                                                                  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.

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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           ซ
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         ซ
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
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                                                                  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.

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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            ซ
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              ซ
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.
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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.

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                                                                  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
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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.

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1.3.3   Effects on Soil                                                        •
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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

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                                                                  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.                                                        •
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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.              <•



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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,


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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

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                                                                  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.
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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.          ซ
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
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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.

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                                                                  1-20
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     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.
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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              ซ
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.
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     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

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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
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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.                            •
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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.                    •
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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 100ฐ 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.                           •
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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.
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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.

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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.
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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          ซ
     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.                                                          •
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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          ซ
the area  coverage required  would tend  to  be so  large as to be                 •
economically prohibitive.
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                                        SECTION 2



                                      INTRODUCTION
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                                                                  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

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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              ซ
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                 ฃ
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

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                                               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).

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                                                                  2-4
2.1.2   Hydrologic Cycle
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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              ซ|
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

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                                                                  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.

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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.







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                                                                                 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



(*>
(*>

(?)

<*)
<*)
<*>
ซ)
(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.

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                                                                  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

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                                                                 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).
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                                                                   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).

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                                                                  2-12
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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            ฃ,
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            ซ
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

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                                                                  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

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                      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 ปNADP
 QAPN   DMAP3S

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2-15
                                                                      Legend
                                                                      Canada   United States
                                                                     •CANSAP  ซNADP
                                                                     • APN     "MAP3S
                                                                     ปOME
                                                                          1980 CHH
          Figure 2-5a.   Precipitation  amount - weighted mean hydrogen ion
                        concentration  in  1980 ( pmoles per litre).

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                      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 ซNADP
   OAPN  QMAP3S
   *OME
                                               I

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                                                                  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.

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                                                                 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.
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2-19
                                                                       Legend

                                                                       Canada    United States

                                                                      •CANSAP  ซNADP
                                                                      • APN     "MAP3S
                                                                      • OME
                                                                           1980
           Figure 2-6a.  Precipitation amount - weighted mean sulphate ion
                         concentration for 1980 ( ymoles per litre).

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                       2-20
       • 3.4
   • 7.8
              • 51
         • 10

          • 16
       ป3.8
                     .46
 |ป49

  • 48
              * 5ฐ

              • 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 ซNADP
   QAPN  DMAP3S
   *OME
                                               I

                                               I

                                               I

                                               I

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                                                                2-21
                                                            Legend

                                                            Canada   United States
                                                           •CANSAP ปNADP
                                                           • APN    BMAP3S
                                                           ปOME
                                                               1980 CNH4+
Figure 2-7a.  Precipitation amount  - weighted mean ammonium ion
              concentration for  1980 ( ymoles per litre).

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                     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 ปNADP
nAPN  QMAP3S
*OME

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                                                                 2-23
                                                             Legend

                                                             Canada   United States
                                                            •CANSAP ปNADP
                                                            • APN    "MAP3S
                                                            ปOME
                                                                1980
Figure 2-8a.  Precipitation amount - weighted mean nitrate ion
              concentration for 1980 ( ymoles per litre).

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                      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 ปNADP
QAPN   nMAP3S
4OME
              ป




              I

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                                                                     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

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                                                              2-26
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                                                             A 1980 % normal
                                                               precip
Figure 2-9.   Percent of normal precipitation  in North America
             in  1980.
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                                                                  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)

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                                                                 2-28
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                                                                               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.
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                                                                 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.

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                                                                 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               ซ
     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                   ซ
     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

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   SECTION 3




AQUATIC EFFECTS

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                                                                  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

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                                                                  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

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                                                                  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:

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                                                                  3-4
CaC03 + H2C03 —> Ca   +  2HC03                                  (1)
                                 or
(K-feldspar)
       30g + 2H2C03 +  12H20  —ป  2K  +  2HC03  + 6H4Si04          (2)
                                   m3Si3ฐ10(ฐH)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  ฃB~ 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

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                                                                   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).

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                                                                  3-6
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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 (ฐC)  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.
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                                                                  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.

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                                                                 3-8
 PLANT
       NHซ
         r
    R-C —I
NH/    ซ-
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



iซ~
        R   organic nitrogen
Figure 3-2.  Simplified nitrogen cycle showing chemical changes
             caused by plant and soil processes (from Reuss 1976).
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                                                                        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).

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                                                                 3-10

— c
1
SOIL S
2+
P^ *>
^^\ A
^2H *
/
2OH~^
H 2H +
aerobic ^^^
N ^- 	 k o
^ r VJ
2H +
0 ^ ฑ 	
2ซ 4
anaerobic
2-
f%
P4
2-
/^
ฐ4

Figure 3-3.  Simplified sulphur cycle  showing  chemical  changes caused
             by plant and soil processes  (modified  from Reuss 1976).
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                                                                  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

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                                                                  3-12
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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        ซ
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
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                                                                  3-13
    4  -
    5  -
    6  -
    7  -
    8  -
    9
                                PH
Figure 3-4.  Aqueous aluminum in equilibrium with gibbsite
             (after May et al. 1979).

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                                                                  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).
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                                                                  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

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                                                                  3-16
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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
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                                                                  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>ฐ|| •
"^x^r^j
dry deposition 't^
drainage water ฃM
*i^r










% 1
|^
>-ซ. _
^*!
^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

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                                                                  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
ซ   60
c
Si
"5
.2   50
3
a
UJ

ซ   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,
               ฃM+ is sum of all cations (Likens et al. 1980).
 I
 I
 I
 I
 1
 1
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1
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I

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                                                                  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

2.6ฐ 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

12.8ฐ 6.7 17.4 19.9 22.3 15.3
6.7ฐ 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
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18 0
14 0
04 0

6 33
1 46
5 -13

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.8

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.84

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.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
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5
1
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1 •
1

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0
1
9
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10 •
86
1
0
8 m
81

0 1
i •
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1

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were made *
specific year.
times precipitation
Johnson 1967).








for 1963-1964.










1

1

-------
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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.                                                                       ซ
                                                                                I

                                                                                I

                                                                                I

-------
                                                                               3-25















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    "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

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                                                                  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

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                                                                  3-32
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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
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                                                                  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)

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                                                                  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).
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                                                                  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 50ฐN  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 52ฐN.  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 52ฐN 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.

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                                                                                               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


_!
•









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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




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                                                                                                 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.
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                                                                  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.

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                                                                  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
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                                                                  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

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                                                                  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               ซ
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

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                                                                  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

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                                                                  3-44
I
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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

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                                                                  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.
3.5ซ3   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

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                                                                  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.           ซ
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          ซ
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        ฃ
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
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                                                                  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	

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                                                                      3-48
  400
  300
  200
ง  100
                                   D •   ซ
                                             •  MANITOBA, SASKATCHEWAN

                                             A  LA CLOCHE MT LAKES

                                             •  SUDBURY AREA

                                             x  NWT, LABRADOR, NFLD

                                             O  MARITIMES

                                             O  MUSKOKA-HALIBURTON, ONT

                                             ฎ  ELA

                                             ฎ  MALI BURTON

                                             ฉ  OTTAWA R. dromoge

                                             ฉ  UPPER GREAT LKS. dromog.
                                                                        I
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  -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

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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 !• ป'H-V '• • A LA CLOCHE MT LAKES
* '•/•:' . •
A • .. j& ' :.vr ' ' SUOBURY AREA
.^5ฎ "•". •' * N.Wt, LABRADOR, NFLO.
ah ''Xv." "'" • D MAR1TIMES
**"y&' >^*'* '• * " ฐ MUSKOKA'HAI-I8URTON. ONT
~ ''/^ ' ' ' ' • ฎ ELA
„ / * * ฎ HALIBURTON
s O
~~* . / 9 OO @ OTTAWA R drainagt
/ <6 QcD D
B / ฎ< • ฎ UPPE" GREAT LKS drainage
"• /
/ ป 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).



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                                                                  3-50
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     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
Bobงe 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             ซ
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

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                                                                     3-51
         HUDSON

         BAY
                                                        GULF   OF

                                                       ST  LAWRENCE
                          03


                       UNITED  STATES
ONTARIO
Figure  3-13.   Hydrographic Regions of Quebec  (Bobge et al. 1982).

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3-52 |
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3-53
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HALIFAX
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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
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3-54
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                                                                  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.

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                                                                  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
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                                                                  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).

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                                                                3-60
      1500-
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                        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).
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                                                                 I
Figure 3-19.
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                  Minimum pH Observed in the Streams
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               Muskoka-Haliburton, 1976-80 (Scheider 1983).


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                                                                  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 50ฐ  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.

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                                                                  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

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                                                                   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

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3-70
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                                                            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).
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                                                          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).

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                                                                  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

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                  3-75
<20/jeq/L


20-200 ^eq/L


>200 jueq/|_
  states

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                                                                  3-76
differences, but these corrections  add uncertainty  to  some  of  the
comparisons.
3.6.3.1   Time Trends  in Nova  Scotia  and  Newfoundland
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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        ป
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

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                                                                  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.

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                                                                      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
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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^    Soฃ~*  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

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* Caution  is  required  in  interpreting SO^   trends  due to the
  interference  of  organic anions  in  measurement  procedures.                        •
                                                                                   I

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                                                                  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

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                                                                            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  <
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                                                       71
                                                       ?  a
                                                       o  ฐ-
                                                       S.  5
                                                                        -100
   6.6-
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                                                    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
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                                                                 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.
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                                                                  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.

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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



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1
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*


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1

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                                                                  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

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                                                                   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).
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                                                                 3-85
                         20
                          10
                     W
                     o>
                    x.
                     
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                                                                  3-86
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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
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                                                                  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).

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                                                                  3-88
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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

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                                                                 3-89
       6
ฐ- 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).

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                                                                  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

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                                                                  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).

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                                                                  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

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    300
    250
    200
T3

D)
*ซ,
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

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                                                                  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 Bobฃe 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

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                                                                  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

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                                                                                                                            3-96
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                                                                  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. ฑ....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).
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3-99


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                                                                  3-100
3.7   ALTERATION  OF  BIOTIC  COMPONENTS IN AQUATIC SYSTEMS RECEIVING
      ACIDIC DEPOSITION
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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           ซ•
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                 jง
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

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                                                                    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).

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                                                                 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).
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                                                                     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).

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                                                                  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          ซ
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              ซ
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

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                                                                  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

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                                                                  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
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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,           ซ
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

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                                                                  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

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                                                                  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
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                                                                  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.

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                                                                      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
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                                                                  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).

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                                                                  3-112
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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

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                                                                  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.

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                                                                  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
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                                                                  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).

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3-116


35
30
25
x:
0)
b 20
"o
0>
1 15
z
10
5
0

PATTEN LAKE

-
-

c$$w 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

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                                                                   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).

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3-118
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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            ซ
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

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                                                                 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
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   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).
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                                                                  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

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                                                                  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
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3-125
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-------
                                                                  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          ซ|
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

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                                                                  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




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1
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                                                                  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

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                                                                  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          ซ
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

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                                                                3-131
           W
           il
           ซ*-
           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).

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                                                                  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
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                                                                  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

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  200
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                                                                         3-134
                                                                                   I


                                                                                   I


                                                                                   I
•ฃ  10
    8
            Mean for 12 rivers with pH>5.0(1980)

            Mean for 10 rivers with pHฃ 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

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                                                                  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

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                                                                      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).


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                                                                     3-137
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                                                  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).

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                                                                  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

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                                                                   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

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3-140





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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

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                                                                   3-142
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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         ซ|
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,         ซ
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           ซ
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

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                                                                   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

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                                                                   3-144
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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

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                                                                   3-146
3.8.2   Depletion of Acid Neutralizing  Capacity
3.9   ATMOSPHERIC  SULPHATE  LOADINGS  AND THEIR RELATIONSHIP TO
      AQUATIC ECOSYSTEMS
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stocks adapted to similar, yet discrete  and  specific,  habitats
(Loftus 1976).  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

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                                                                   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

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3-148
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     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

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                                                                   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

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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
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                                                                   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.

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                                                                  3-152
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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

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                                                                  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.

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                                                              3-154
 Observed Characteristics
 EXPERIMENTAL LAKES AREA, ONTARIO
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 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

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 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.                                •
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                                                             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.

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                                                                  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
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  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                                            _
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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

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                                                             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.

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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
(ป>

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





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1

1



1
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•
1

1


1
1

1
1
1

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                                                             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).

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                                                             3-160
LAURENTIDE PARK, QUEBEC

Humid Alpine Lower Boreal Regions
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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 47ฐ30'          •
and 50ฐ00' N latitude and 67ฐ and 75ฐ 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.                 •
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                                                             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.

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                                                            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.
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                                                             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. ฃ [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

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                                                             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
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     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         •
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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).
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                                                             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.

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                                                             3-166
Observed Characteristics
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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).
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                                                             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).

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                                                            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).
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                                                             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 19ฐC  and  January
temperatures of -11ฐC.  The lakes commonly are ice-covered from
late November to  late  April (Schnoor et al. 1982).

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                                                            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
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Deposition       Fraction   Time Period       Reference                 •
(kg S042-/ha.yr)                                                        |


   17            wet (68 cm)     1981         NADP 1981-83              I
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                                              (Spooner)                 •
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               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

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                                                             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.

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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
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1
1
_
•

1

1




1

1

•


1
1
1
1
1
1

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                                                             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).

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                                                             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

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                                                             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 19ฐC and in
January is -9ฐC.  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  ฃ  4.5) and generally
infertile.

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                                                             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).
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                                                             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 25ฐC; 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)

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                                                                  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         ซ
        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

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                                                                  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

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                                                                  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.
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                                                                  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,

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                                                                  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

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                                                                                            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

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                                                                  3-184
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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              ซm
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

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                                                                  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 52ฐN latitude  for each province.  [Figure 3-9 shows only the
areas south of 52ฐN.]   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 52ฐN, 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.

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3-186
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                                                                                        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 ฐN 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

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                                                                                     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

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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


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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





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                                                                  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,

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                                                                  3-192
AQUATIC
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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

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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

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                                                                  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

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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)
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       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 51ฐ 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.

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3-196
         o
        •H

         tO
        4J

        <ง

        4-1
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         O
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        •a  eg
            tX5
         
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                                                                  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

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                                                                  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                ซ
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

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                                                                  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

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                                                                  3-200
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"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

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| 3-201
1
1
• 300-


™ ง 200-
3-
^H **
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X
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HCOo- Lakes xx ^xx
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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

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                                                                  3-202
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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

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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

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                                                                  3-204
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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        ฃ
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

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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*)ฃ (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*)ฑ

(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*)ฑ + 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*)ฑ + 0.4 (  S04*)w (Henriksen 1982)
                        = (Ca* + Mg*)i + 0.4 (S04*w -
(1)  (Ca* + Mg*)ฑ

(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*)ฃ
                        = 2.04 (Ca* + Mg*)ฑ + 3.64 - 0.82 (S04*)i
                        = KS04*)  + 191/1.9 (Henriksen  1980)
                        = (S04*) /2 (assuming 100 cm annual rainfall)

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                                                                  3-206
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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
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                                                  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).

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                                                                  3-208
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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

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                                                                     3-209
            200
         CT
         0)

        ง

        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
                  Nฃ.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).

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                                                                        3-210
                                                                                    I
                                                                                    I
2  400

CO
CO
ฃ  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
ฃ 200

3
CO
o 160

ซ
•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).
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              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,


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             The model was  developed  for  an area with 100 cm runoff.  It has not
             been  corroborated  for  areas  with lower runoff.

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                                                                  3-212
3.9.3.3   Summary
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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            ซ
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.
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                                                                  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.

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                                                                  3-214
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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          •
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                                                                  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.

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                                                                  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
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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             ฃ
     quality appear to be sparse throughout North America, although
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                                                                  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

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                                                                  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
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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.                                                    •
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                                                                  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

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                                                                  3-220
          Northern Highlands Lakes               -  Wisconsin
          Coweeta                                -  North Carolina
          Andrews                                -  Washington
          North Cascades                         -  Washington
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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           ซ
     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              ฃ
     chemistry, and type of biotic  community (cold water, warm water
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                                                                  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.

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                                                                 3-222
     1974.  Effects of acidification  on  Swedish  lakes.   Ambio
     3:30-36.
              1982.  Effects on  fish  of metals  associated  with
                                                                              I
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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.  Etude  comparative de trois  modeles
     de prevision de la sensibilite" des milieux lacustres:  Le               M
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     (abstract)
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     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.

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                                                                 3-258
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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
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Watt, W.D.; Scott, D.; and Ray, S.   1979.  Acidification and  other
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Watt, W.D.; Scott, D; and White, W.J.  1983.  Evidence of acidifi-
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Wetzel, R.G.  1975.  Limnology.  Philadelphia:  Saunders Co.                    ฃ
                                                                               I
Wiederholm, T., and Eriksson, L.  1977.  Benthos of  an acid lake.
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Yan,  N.D.,  and Miller, G.E.  1982.  Characterization  of lakes near
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      pp. 127-137.

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      to experimental alterations  of pH.  Environ. Conserv.  5:93-100.

Yan,  N.D.,  and Strus,  R.   1980.   Crustacean  zooplankton communities
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     SECTION 4

TERRESTRIAL IMPACTS

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                                                                  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

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                                                                  4-2
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                                                                             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,
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                                                                  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 S0ฃ  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

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                                                                            4-4
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                                                                       >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

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                                                                           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).

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                                                                  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           ฃ
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
                                                               ป *-  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.                                    •
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                                                                  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.

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                                                                  4-8
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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.
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                                                              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).

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                                                                  4-12
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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 S02ซ  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

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                                                                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).

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                                                                  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

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                                                                  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.

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                                                                  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.  Processes           •
necessary for the existence of all life, the  flow of  energy and
cycling of nutrients are based on relationships that  exist among the          •
organisms within the system (Billings 1978; Odum 1971; Smith 1980).           |
Because of these relationships, unique attributes emerge when
                                                                               I

                                                                               I

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                                                                  4-24
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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               ซ•
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

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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         ฃ
curves (Larsen and Heck 1976;  Linzon  et al. 1975).   However,  these
data may not be reliable for estimating the total  effect  on crop              _
productivity (e.g., yield, quality).  Most  information now indicates         •
that the severity of foliar symptoms  is not a  reliable index  of crop         ™
growth or yield effects (Reinert  1980) as there  is uneven competition
among several sinks that receive  photosynthate.  Also,  compensatory           H
responses to ozone can produce  rapid  recovery  from injury (Jacobson           |
in press).  Studies with soybeans (Tingey et al.  1973), tomatoes
(Oshima et al. 1975) and alfalfa  (Tingey and Reinert 1975) all               M
support this concept.  The exceptions to this  general  finding are            •
cases where the harvested  product is  the foliage  and where foliar
injury development coincides with the rapid growth of  the harvested
product (Linzon et al. 1975).                                                 •
                                                                              I

                                                                              I

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                                                                  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

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                                                                  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

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                                                                  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.

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4-34



















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                                                                  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               ซ
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              ซ
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

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                                                                  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

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                                                                  4-40
4.2.3.1  Acidic Deposition Effects  to  Agricultural  Crops
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(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

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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.

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                                                                  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

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                                                                  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).

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                                                                  4-44
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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

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                                                                  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.

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                                                                  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

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4-47








































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                                                                  4-50
4.2.4.3  S02 - 03 - Acidic Deposition Effects
I
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expressed as dry mass/seed was 9% less than that  of  plants  exposed  to
ambient air; higher doses reduced yield by 10% and  15%.                       _


4.2.4.2  S02 - N02 Effects

The possibility of adverse effects occurring on plant  life  due  to  the         |
interaction between atmospheric  S02 and N02 needs consideration.
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

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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

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4-52
                                                                               I
                                                                               I
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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
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                                                                  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

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                                                                  4-54
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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  irฑ 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

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                                                                  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.

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                                                                  4-56
1
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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        
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                                                                  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

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                                                                  4-58
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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                ป
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           ™
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                                                                  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.

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                                                                  4-60
4.4.3   Influence of Soil Biota and Decomposition/Mineralization
        Activities
1
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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             ฃ
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                    ฃ
(1980) show ammonification and nitrification  may decrease  markedly  in
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                                                                  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,

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                                                                  4-62
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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             ซ
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.              ซ
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             ฃ
cumulative levels of heavy  metals  in  the  soils have  exercised
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                                                                  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

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                                                                  4-64
soil solutions than in acidic deposition  falling  in  Berkeley,
California.
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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.
                                                                               ฃ
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                  ฃ
organisms or ecosystems  (e.g., forests, fish, soil,  bedrock, aquatic
ecosystems).  Those directed toward aquatic effects  have emphasized           ซ
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

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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.

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4-68

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      4
               pH
   pH
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                    6

                    5

                    4

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    > 7

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SOIL 5

     4

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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—ป>
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                          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