Tracing Sources of Nitrate in Rivers of Southern Alberta Using Nitrogen, Oxygen, and Boron Isotopes As Co-Tracers

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2017 Tracing Sources of Nitrate in Rivers of Southern Alberta Using Nitrogen, Oxygen, and Boron Isotopes as Co-Tracers

Kruk, Mary Krystin

Kruk, M. K. (2017). Tracing Sources of Nitrate in Rivers of Southern Alberta Using Nitrogen, Oxygen, and Boron Isotopes as Co-Tracers (Unpublished master's thesis). University of Calgary, Calgary, AB. doi:10.11575/PRISM/28495 http://hdl.handle.net/11023/3699 master thesis

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Tracing Sources of Nitrate in Rivers of Southern Alberta Using Nitrogen, Oxygen, and Boron

Isotopes as Co-Tracers

Mary Krystin Kruk

A THESIS

SUBMITTED TO THE FACULTY OF GRADUATE STUDIES

IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE

DEGREE OF MASTER OF SCIENCE

GRADUATE PROGRAM IN GEOLOGY AND GEOPHYSICS

CALGARY, ALBERTA

APRIL, 2017

Abstract

The natural and anthropogenic sources of nitrate (NO3) in the Bow River, Oldman River, and select tributaries of the Oldman River (OMR) in southern Alberta were determined by analyzing

15 18 11 a combination of δ NNO3, δ ONO3 and δ B values in surface water samples and in

15 18 11 anthropogenic nutrient end-members. The δ NNO3, δ ONO3 and δ B values of surface water revealed Bonnybrook WWTP effluent has the greatest influx of nutrient load to the Bow River and downstream of Calgary agricultural return flow water is an additional source of nutrients.

Decreasing NO3 loads and NO3/B ratios downstream of Calgary indicated that there are also N-

15 11 removal processes occurring in this reach. The δ NNO3 and δ B values of the surface water in the OMR indicated a significant nutrient input by the Lethbridge WWTP effluent and tributaries displayed δ11B values heavily indicative of cow manure. Hence, boron isotopes are a useful co- tracer of nutrients in surface water.

Acknowledgements

I would like to thank my supervisor, Dr. Bernhard Mayer, for all the teaching, knowledge and guidance he provided during the course of this research project. I am truly grateful for all that I have learned through him and the spirit of independent research that he instilled. I am also thankful for the financial support of my project through the NSERC Discovery Grant.

To Mike Nightingale, I cannot express how thankful I am for all the hours of scientific discussion, the assistance in the geochemistry lab, and the light conversation provided when things were going awry in my lab work, it was a pleasure. Thank you to Maurice Shevalier with all of the logistical and technical support provided during my project, things would never have run as smoothly without this.

Thank you to the staff in the Isotope Science Laboratory: Steve Taylor, Andrew

Kingston, Jesusa Pontoy, Niloufar Nadari and to Kerri Miller for all the assistance with my isotopic analyses and to Farzin Malekani for assistance with chemical analyses. Thanks to Dr.

Michael Wieser for use of his laboratory and insight to my boron isotopic analyses.

To all of the students and research associates in the Applied Geochemistry Group, and to my friends from other research groups, I cannot express how thankful I am to have worked with you and to and built friendships with you. This truly shaped my graduate school experience and left a lasting impression on me. I am grateful for everything from getting assistance outside my area of knowledge, to coffee breaks, to lunch club, to rock climbing/skiing/hiking/swimming breaks, they were all much needed and I enjoyed every minute.

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To Veronique, Nadine, Jess, Christine, Friderike, Carmen and Paul who all assisted me with my field work, thank you so much for the help and the company. Veronique, thank you for being my fellow Bond Girl in trips to the Wastewater Treatment Plants.

Thank you to Susan Dooley and to all of those who work in the Department of

Geoscience main office for providing such efficient administrative support during my project.

Thank you to Dr. Benjamin Ellert from Agriculture and Agri-Food Canada for providing me with agricultural samples and the supplementary information, and to the research associates from

INRA and BRGM in France who provided insight to my boron ion exchange experiment.

Finally, I need to thank my family and friends who have given me life-long love and support, and to Paul who has only provided patience and encouragement during the ups and downs of my project.

Table of Contents

Abstract ...... ii Acknowledgements ...... iii Table of Contents ...... v List of Tables ...... viii List of Figures ...... xi

Introduction ...... 1

1.1 PROJECT RATIONALE ...... 1

1.2 INTRODUCTION TO ISOTOPE TRACING OF NO3 ...... 5 1.2.1 δ15N and δ18O of Nitrate ...... 5 1.2.2 δ11B as a Tracer of Nutrient Sources ...... 11

1.3 RESEARCH OBJECTIVES ...... 18

Study Area ...... 19

2.1 INTRODUCTION ...... 19 2.1.1 Climate and Hydrology of Southern Alberta ...... 23 2.1.2 Geology ...... 24 2.1.3 Soil Types ...... 25 2.1.4 Water and Land Use ...... 26

2.2 WASTEWATER TREATMENT PLANT PROCESSES ...... 29 2.2.1 Pre-Treatment ...... 30 2.2.2 Primary Treatment & Clarifiers ...... 31 2.2.3 Secondary Treatment & Clarifiers ...... 31 2.2.4 Tertiary/Advanced Treatment ...... 32

2.3 MANURE AND FERTILIZER ...... 32

Sampling and Analytical Methods ...... 35

3.1 FIELDWORK ...... 35 3.1.1 Overview ...... 35 3.1.2 Surface Water Sampling Sites ...... 37 3.1.3 Wastewater Treatment Plant Effluent Samples ...... 41 3.1.4 Surface Water Sample Collection & Preparation ...... 41

3.2 LABORATORY ANALYSES ...... 42

3.2.1 Physical and Chemical Data ...... 42 3.2.2 Solid Sample Preparation ...... 43 3.2.3 Boron Ion Exchange ...... 46 3.2.4 Isotope Ratios ...... 51

Discharge and Water Sources of Southern Alberta Rivers ...... 57

4.1 DISCHARGE DATA...... 57 4.1.1 Introduction ...... 57 4.1.2 River Discharge Data ...... 57

4.2 d2H AND d18O OF WATER ...... 62 4.2.1 Introduction ...... 62 4.2.2 Results ...... 64

Ion Chemistry of Southern Alberta Rivers ...... 74

5.1 INTRODUCTION ...... 74

5.2 MAJOR ION CHEMISTRY ...... 75 5.2.1 The Bow River ...... 75 5.2.2 The Oldman River & Tributaries ...... 81

5.3 NO3 CONCENTRATIONS ...... 87 5.3.1 The Bow River ...... 87 5.3.2 The Oldman River ...... 90

5.4 BORON CONCENTRATIONS ...... 93 5.4.1 The Bow River ...... 93 5.4.2 The Oldman River ...... 95

NO3 and Boron Budgets for Southern Alberta Rivers ...... 98

6.1 INTRODUCTION ...... 98

6.2 NUTRIENT FLUXES IN WWTP EFFLUENTS ...... 101

6.3 RIVERINE NO3 BUDGETS ...... 105 6.3.1 The Bow River ...... 105 6.3.2 The Oldman River ...... 107 6.3.3 WWTP influence on Riverine NO3 ...... 109

6.4 RIVERINE B BUDGET ...... 113

6.4.1 The Bow River ...... 113 6.4.2 The Oldman River ...... 115 6.4.3 WWTP influence on B riverine flux ...... 117

6.5 RIVERINE NO3/B RATIOS ...... 121

Nutrient Sources and Cycling Based on δ15N, δ18O, and δ11B Values ...... 123

7.1 INTRODUCTION ...... 123

7.2 RESULTS ...... 125 7.2.1 Isotopic Composition of Anthropogenic Nutrient Sources ...... 125 7.2.2 The Bow River ...... 131 7.2.3 The Oldman River ...... 135

7.3 DISCUSSION ...... 140 15 18 7.3.1 δ NNO3 and δ ONO3 Values Revealing Sources and Processes Affecting Riverine NO3 ...... 140 7.3.2 δ11B Values Revealing Sources and Processes Affecting Riverine Boron ...... 147 7.3.3 Combining δ15N and δ11B Values for Assessing Nutrient Sources ...... 153 7.3.4 Summary ...... 161

Chapter Eight: Conclusions and Future Work ...... 162

8.1 CONCLUSIONS ...... 162

8.2 IMPLICATIONS AND FUTURE WORK ...... 167

References...... 170

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List of Tables

Table 3.1: Surface water sampling sites GPS coordinates………………………………………38

Table 3.2: Sample list for the dried cow manure and mineral fertilizers analyzed in this study…………...... 44

Table 3.3: Boron extraction procedure using Amberlite IRA 743 resin in a gravity-flow column. ……………………………………………………………………………………………………50

Table 4.1: Summary of δ2HH2O and δ18OH2O values in per mil (‰) measured in river water and WWTP final effluent between June 2014 and October 2015 along the Bow and Oldman Rivers. …………………………………………………………………………………………...... 67

Table 5.1: Major cation, anion, and TDS data for the surface water samples from the Bow River 2014 sampling period. Bracketed numbers denote the river “reach”. “NA” indicates the site was not sampled during that specific campaign. ……………………………………………………..78

Table 5.2: Major cation, anion and TDS data for the surface water samples from the Bow River for 2015 sampling period. Bracketed numbers denote the river “reach”. ……………………….79

Table 5.3: 2014 major cation, anion, and TDS data for the surface waters from the 2014 Oldman River and tributary sampling sites. Tributary sites are highlighted in grey. Bracketed numbers denote the river “reach”. ………………………………………………………………………...84

Table 5.4: 2015 major cation, anion, and TDS data for the surface waters from the 2015 Oldman River sampling sites. Tributary sites are highlighted in grey. Bracketed numbers denote the river “reach”. ………………………………………………………………………...... 85

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Table 5.5: NO3 concentrations of the Bow River surface water samples during the 2014-2015 sampling campaign. Sites with concentrations “NA” were not sampled during the specific sampling campaign. …………………………………………………………...... 89

Table 5.6: NO3 concentrations of surface water samples from the Oldman River sampling sites, including tributaries (grey), during the 2014-2015 sampling campaign. “NA” indicates concentrations were not measured during that sampling campaign and “n.d.” indicates below detection limit. …………………………………………………………...... 92

Table 5.7: Boron concentrations in surface water obtained from the Bow River sampling sites during the 2014-2015 sampling campaign. “NA” indicates B concentrations were not measured during that sampling campaign. …………………………………………...... 94

Table 5.8: Boron concentrations of surface waters from the Oldman River sampling sites, including tributaries (grey), during the 2014-2015 sampling campaign. “NA” indicates B concentrations were not measured during that sampling campaign. ………………...... 97

Table 6.1: Nutrient concentrations of the effluents from the WWTPs. “NA” indicates the site was not sampled during that campaign and “n.d.” indicates below detection limit. …………………………………………………………………………………………………..102

Table 6.2: The average baseflow daily flow rate (October-March) for the sampled WWTPs…………………………………………………………………………………………103

Table 6.3: The calculated average baseflow flux values of NO3 and B from each WWTP. …...... 103

Table 6.4: NO3 Flux during peakflow (June) and baseflow (Sept) periods along the BRB. ….107

Table 6.5: NO3 flux during peakflow (June) and baseflow (Oct) periods along the OMR. Data for the measured tributaries are highlighted in grey. …………………………………………..109

Table 6.6: B flux during peakflow (June) and baseflow (Sept) periods along the BRB. ……...115

Table 6.7: B flux during peakflow (June) and baseflow (Oct) periods along the OMR. The measured tributary is highlighted in grey. ………………………………………………..……117

Table 7.1: δ15N, δ18O, and δ11B values of the major nutrient end-members within southern Alberta. …………………………………………………………………………………………129

15 18 11 Table 7.2: δ NNO3, δ ONO3 and δ B values of surface water samples collected along the Bow River. …………………………………………………………………………………………...133

15 + Table 7.3: Concentrations and δ N values of NH4 in surface water samples collected within the FC WWTP plume of the Bow River. …………………………………………………………..133

15 18 11 Table 7.4: δ NNO3, δ ONO3 and δ B values of surface water samples obtained along the Oldman River. ………………………………………………………………………………….136

15 18 11 Table 7.5: δ NNO3, δ ONO3 and δ B values of surface water samples collected from tributaries of the Oldman River. …………………………………………………………………………...138

List of Figures

Figure 1.1: The South Saskatchewan River Basin within Alberta, divided into its sub-basins: The Red Deer, the Bow, the Oldman, and the South Saskatchewan Rivers. ……………………..4

Figure 1.2: The major components of the nitrogen cycle within a terrestrial system (Robertson & Groffman, 2007). ………………………………………………………………………………....6

15 18 Figure 1.3: δ N versus δ O values of typical NO3 sources in aquatic receptor systems, modified from Kendall et al. (2007). ………………………………………………………………………11

Figure 1.4: Typical δ11B values of different freshwater contamination sources and background values, adapted from Komor (1997), Tirez et al. (2010) and Widory et al. (2004). …………….16

Figure 1.5: The B isotope variation within natural sources and geologic materials (Hoefs, 2009). ……………………………………………………………………………………………………16

11 15 Figure 1.6: δ B versus δ N values for NO3 – containing contaminant sources based on previously reported literature (Widory et al., 2004, 2005, 2013; Accoe et al., 2008; Seiler, 2005; Eppich et al., 2012). ……………………………………………………………………………..17

- Figure 1.7: The speciation of B(OH)3 and B(OH)4 at 25°C and 1 atm with increasing pH - defined by a) the coordination of B and b) the isotopic composition of B(OH)3 and B(OH)4 (Hoefs, 2009). …………………………………………………………………………………...18

Figure 2.1: The Bow River Basin with its major tributaries, divided by its natural sub-regions (BRBC, 2005). …………………………………………………………………………………..20

Figure 2.2: The Oldman River basin, divided into its natural sub-regions (OWC, 2010).……...22

Figure 2.3: The major types of water use in water licensing for the Bow and Oldman Rivers (BRBC, 2005; OWC, 2010)…………………………………………………………………...... 28

Figure 2.4: Fertilizer shipments to Canadian agriculture markets of the top four product types in the prairies, cumulative data (StatsCan, 2016b). ………………………………………...... 34

Figure 3.1: Overview of the Bow River sampling sites broken down into three reaches. Reach 1 represents the Bow River upstream of Calgary with five river sampling sites, Reach 2 represents the Bow River within Calgary with three river sampling sites, and Reach 3 represents Bow River downstream of Calgary with four river sampling sites. …………………………………………39

Figure 3.2: Overview of the Oldman River sampling sites broken down into two reaches. Reach 1 represents the Oldman River upstream and including Lethbridge with three river sampling sites and Reach 2 represents the Oldman River downstream of Lethbridge with two river sampling sites and three tributary sites……………………………………………………………………..40

Figure 3.3: Dimensions of the 2 mL Pierce centrifuge columns, with a total column capacity of 5.5 mL. The 50 µL of Amberlite resin forms an even bed at the base of the column. ………….48

Figure 4.1: The monthly average a) 2013 and b) 2014 discharge data for the Bow River at Environment Canada flow monitoring stations (Environment Canada, 2016b). The historical discharge data for Cochrane is only available until 2013. ………………………………………59

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Figure 4.2: Historical discharge data for the years 2013-2015 for A) the OMR at Lethbridge, B) the mouth of the OMR, and C) mouth of the Little Bow River, collected at Environment Canada flow monitoring stations (Environment Canada, 2016b). ……………………………………….61

Figure 4.3: δ2H (red) and δ18O (blue) values of river and WWTP sites with downstream distance of the Bow River. The river sites are plotted for both June (peak-flow) and September (beginning of base-flow) 2015. The points indicate sampling sites in Lake Louise (LL), Banff (BF), Canmore (CM), Cochrane (CR), Calgary (Calg.), Carseland (CS), Bow City (BC), Scandia (SD), and Ronalane (RL). ……………………………………………………………………………...69

Figure 4.4: δ2H and δ18O values of the mainstem river sites and the Lethbridge WWTP along the OMR. The river sites are plotted for June (peak-flow) and October (base-flow) of 2015. Points refer to sampling sites near Monarch (MN), Below Belly River (BL), Lethbridge (LB), Downstream Tributaries (DS), and Taber (TB). ………………………………………………...71

Figure 4.5: δ2H versus δ18O values comparing A) the BRB sites upstream and downstream of Calgary to the WWTP effluents, B) OMR sites upstream and downstream of the tributaries, the tributary values and the WWTP effluents.………………………………………………...... 72

2 18 Figure 4.6: δ HH2O versus δ OH2O values of surface water sampling sites for a) The Bow River sampling sites and from each WWTP effluent along the BRB and b) The Oldman River, the OMR tributaries and the WWTP effluent from Lethbridge. Also shown are the Local Meteoric Water Line for Calgary (LMWL Calgary) from Peng et al. (2004), the Local Evaporation Line (LEL) (Katvala, 2008), and the Shallow Groundwater Line (SGWL) (Cheung, 2009). ……….73

Figure 5.1: Piper plot conveying the major water types of Reach 1, 2, and 3 of the Bow River Basin based on % meq/L. ………………………………………………...... 80

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Figure 5.2: TDS of surface water samples versus distance along the Bow River for each sampling event. ………………………………………………...... 81

Figure 5.3: Piper plot conveying the major water types of Reach 1, Reach 2, and the sampled tributaries of the Oldman River Basin based on % meq/L. ……………………………………..86

Figure 5.4: TDS of surface water samples versus distance along the Oldman River (left axis) and tributaries (right axis) for each sampling event. ………………………………………………...87

Figure 5.5: NO3 concentrations (mg/L) in surface water samples along the Bow River with distance from the sampling point above Lake Louise. …………………………………………89

Figure 5.6: NO3/Cl ratios (using meq/L) along the Bow River with downstream distance during baseflow (September) sampling periods of 2014 and 2015. Plot begins below Bonnybrook

WWTP (205 km) where NO3 concentration begins to increase above background concentrations. ………………………………………………...... 90

Figure 5.7: NO3 concentrations (mg/L) along the OMR with distance from the Monarch sampling site. Blue markers indicate OMR mainstem sites and use the primary y-axis and red indicates the sampled tributaries that use the secondary y-axis. ………………………………..92

Figure 5.8: B concentrations (µg/L) for surface waters obtained along the Bow River with distance from the sampling point above Lake Louise. ………………………………………….94

Figure 5.9: B/Cl ratios (using meq/L) with downstream distance during baseflow (September) sampling periods of 2014 and 2015. Plot begins below Bonnybrook WWTP (205 km) where B concentration begins to increase above background concentrations. …………………………...95

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Figure 5.10: B concentrations (µg/L) of surface waters obtained along the OMR with distance from the Monarch sampling site. Blue markers indicate OMR mainstem sites and use the primary y-axis and red indicates the sampled tributaries that use the secondary y-axis. ……….97

Figure 6.1: NO3 flux (kg/d) in the Bow River. The flux values are based on baseflow values from the surface water sampling sites (circles) and the WWTP effluent (squares). The rectangle represents the agricultural return-flow non-point source of NO3 downstream of Calgary. Each riverine NO3 flux value has a 28% uncertainty and WWTP effluents a 11% uncertainty. ……………...... 111

Figure 6.2: NO3 flux (kg/d) in the OMR estimated using baseflow values of the surface water sampling sites (circles) and effluent from the Lethbridge WWTP (square). Both upstream and downstream of Lethbridge the non-point sources of nutrients, agricultural return-flow, are represented by rectangles. The three sampled tributaries downstream of Lethbridge are represented within agricultural return-flow. Each riverine NO3 flux value has a 28% uncertainty and WWTP effluents a 11% uncertainty...... 112

Figure 6.3: Flow chart of B flux (kg/d) along the Bow River during the baseflow period from the surface water sampling sites (circles) and WWTP effluent (squares). The agricultural return-flow (rectangle) is a non-point source of B downstream of Calgary. Each riverine B flux value has a 22% relative uncertainty and WWTP effluent a 14% uncertainty...... 119

Figure 6.4: B flux (kg/d) values of the OMR estimated using baseflow values of the surface water sampling sites (circles) and effluents from the Lethbridge WWTP (square). Both up- and downstream of Lethbridge the non-point sources of nutrients, agricultural return-flow, are represented by rectangles. The three sampled tributaries downstream of Lethbridge are represented within the agricultural return-flow. Each riverine B flux value has a 22% relative uncertainty and WWTP effluent a 14% uncertainty ...... 120

Figure 6.5: The NO3/B flux ratio during baseflow 2014 in the Bow River with distance from above Lake Louise. A ± 28% error is applied to this ratio...... 122

15 18 Figure 7.1: δ N (blue) and δ O (red) values of NO3 versus NO3 concentration for effluents from the WWTPs...... 130

15 18 11 Figure 7.2: δ NNO3, δ ONO3 and δ B values of surface water versus downstream distance from a point above Lake Louise along the Bow River. From 0km to 300km the δ11B value is within +8 ± 4‰...... 134

15 18 11 Figure 7.3: δ NNO3, δ ONO3 and δ B values of surface water versus downstream distance from the Monarch sampling site along the OMR...... 139

Figure 7.4: δ18O versus δ15N values of nitrate measured in the Bow River during baseflow, the WWTP effluents that discharge into the Bow River, local cow manure and fertilizer samples.

These isotope values are overlain by previously reported values for NO3 sources (Kendall et al., 2007)...... 145

Figure 7.5: δ18O versus δ15N values of nitrate measured in the OMR and select tributaries during baseflow, the WWTP effluent from Lethbridge that discharge into the OMR, local cow manure and fertilizer samples. These isotope values are overlain by previously reported values for NO3 sources (Kendall et al., 2007)...... 146

Figure 7.6: δ11B values of nutrient end-members and within the Bow River, OMR, and select tributaries measured in this study compared against the previously reported values for nutrient endmembers (blue box) (Barth, 1993; Komor, 1997; Tirez et al., 2010; Widory et al., 2004). The red point within the Bow River values indicates what is considered to be natural background δ11B values above Lake Louise. ………………………………………………………………..153

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11 15 Figure 7.7: δ B versus δ NNO3 values measured during baseflow in the Bow River as well as the local WWTP effluents, cow manure, and fertilizers. The overlaid boxes, indicating isotopic ranges of nutrient end-members, are a combination of the values measured in this study and previously reported values from the literature. ………………………………………………...159

11 15 Figure 7.8: δ B versus δ NNO3 values measured baseflow in the OMR and the selected tributaries as well as the local WWTP effluent, cow manure, and fertilizers. The overlaid boxes, indicating isotopic ranges of nutrient end-members, are a combination of the values measured in this study and previously reported values from the literature. …………………………………160

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Introduction

1.1 Project Rationale

Nitrate (NO3) constitutes a widespread contaminant in surface water and groundwater.

Contamination often stems from the overuse of nitrogen (N)-based organic and mineral fertilizers, N-containing animal manure in agriculture, and urban runoff and wastewater entering surface waters or aquifers (Benkovitz et al., 1996; Heaton, 1986; Howarth et al., 1996; Smil,

1999). Soil runoff and seepage supplies a baseline concentration of dissolved organic and inorganic nitrogen to surface and groundwater that undergoes nitrification to naturally produce

NO3 (Kendall, 1998). Streams and rivers collect and transport natural and anthropogenic-derived

NO3 coastward and excess loading causes such effects as eutrophication, hypoxia, and loss of biodiversity within lakes and rivers along its flowpath (Howarth et al., 1996; Peierls et al., 1991;

Rabalais et al., 2001). In the case of central and southern Alberta, Canada, all major river basins drain eastward, flowing to Lake Winnipeg and finally into the Hudson Bay in Manitoba (AMEC,

2009). Many instances of eutrophication within this river network have occurred in the past below wastewater outflows and through areas of high agricultural activity and the rapid eutrophication of Lake Winnipeg is an ongoing issue (Schindler et al., 2012; Sosiak, 2002).

Mitigation measures to reduce nutrients in surface water in Alberta include tertiary treatment within wastewater treatment plants (WWTPs) that includes both chemical or biological phosphorus and biological nitrogen removal and limits on fertilizer application (AAFRD, 2004;

Sosiak, 2002).

Furthermore, for many municipalities in Alberta, Canada, rivers are the main source of drinking water. Excess NO3 in drinking water may be linked to health impacts in humans such as methaemoglobinaemia in infants and cancers of the digestive tract (Powlson et al., 2008); thus

NO3 concentrations in drinking water sources should be closely monitored. The drinking water standard for NO3 states a limit of 45 mg/L (or 10 mg/L nitrate-N) set by the World Health

Organization and this maximum allowable concentration (MAC) is also followed by the

Guidelines for Canadian Drinking Water Quality (HealthCan, 2003; WHO, 1996).

The Bow, Oldman, and Red Deer River basins are the watersheds that comprise the South

Saskatchewan River Basin (SSRB) (see Figure 1.1). Urbanization and agriculture affect water quality in these three basins as wastewater effluent is discharged directly into the rivers in municipalities and each have reaches draining through cultivated lands, cattle pastures and areas with feedlots. These factors in combination with population growth and climate change have resulted in increased contamination and decreased water quality. In efforts to mitigate these effects the government of Alberta has released a series of plans for maintaining drinking water quality, a healthy ecosystem, and water supplies for a sustainable economy in the Water for Life

Strategy: Alberta’s Strategy for Sustainability, and the South Saskatchewan River Basin Water

Management Plan (Phase I and II) (Alberta Environment, 2003, 2006). Part of developing better management practices and remediation plans is identifying the sources of nutrients such as nitrogen and phosphorus and understanding the cycling that affects the concentrations of compounds formed, such as NO3, in these rivers.

Using flow and geochemical data to determine solute loads is an effective way to understand temporal and spatial trends in surface water. Alberta Environment conducts a long- term water quality monitoring program of the South Saskatchewan River Basin that dates back to

1970, as well as a newly implemented Tributary Monitoring Network established in April 2016, measuring the concentrations of a number of solutes to monitor water quality at monthly intervals (Alberta Environment, 2016, personal communications). This includes solute concentration measurements of nutrients such as total phosphorus and NO3. However, when characterizing sources of solutes being transported by a river network, concentration alone cannot define the sources of nutrients. To do this, a supplementary fingerprinting tool must be used, such as the use of stable isotope signatures. Stable isotope ratios measured in certain solutes can act as a signature to provide insight about the sources and processes that occur in freshwater aqueous environments. For this study, the Bow River Basin (BRB) and Oldman River

Basin (OMR) were used as case studies to investigate the potential sources and sinks of NO3 in southern Alberta rivers using the stable isotope compositions of nitrate and boron as co-tracers.

Figure 1.1: The South Saskatchewan River Basin within Alberta, divided into its sub- basins: The Red Deer, the Bow, the Oldman, and the South Saskatchewan River.

1.2 Introduction to Isotope Tracing of NO3

1.2.1 δ15N and δ18O of Nitrate

Provided there are traceable amounts of NO3 present in a water sample, stable isotope ratios of both N and O can be measured. Stable isotope ratios are usually expressed using the delta (δ) notation in units of per mil (parts per thousand, ‰) relative to an international standard:

δsample (‰) = [(Rsample- Rstandard)/ Rstandard] x 1000

Where R is the stable isotope ratio of the heavy isotope over the light, such as 15N/14N or

18O/16O of the sample and standard. When the δ-value of the sample is positive it indicates enrichment in the heavy isotope and when negative the heavy isotope is depleted in comparison

15 to the isotope ratio of the standard. Multiple studies have used stable isotope ratios of NO3 (δ N

18 and δ O) to estimate the origin and fate of NO3 in aquatic systems, as discussed below. Kinetic isotope fractionation during unidirectional processes occurs between N-compounds during biologically mediated reactions such as nitrogen fixation, assimilation, mineralization, nitrification, denitrification or physical reactions such as ammonium volatilization where most commonly the lighter isotope (14N) will preferentially partition into the product, leaving the substrate enriched in 15N (Kendall et al., 2007). Figure 1.2 depicts the nitrogen cycle in a terrestrial setting with the key transformation reactions. While most of these nitrogen transformations are associated with an N-isotope fractionation of -10 to -30‰ or more, nitrogen fixation is accompanied by a small N-isotope fractionation of between -3 to +1‰ and benthic

denitrification has small fractionation of -1.5 to -3.6‰ (Kendall et al., 2007; Sebilo et al., 2003).

Evidence of N-transformations in an aquatic ecosystem can be supported by these observed

15 18 16 isotopic shifts in δ N values in nitrogen species. Stable oxygen isotopes ( O/ O) of NO3 are also fractionated during denitrification, typically at a ratio of 1:1 to 2:1 (N:O ratio) depending on the environment and additional processes that may be occurring in parallel, causing the remaining oxygen to be enriched in 18O (Kendall et al., 2007). The isotopic fractionation during

N-transformations combined with where the nitrogen was originally sourced often leads to atmospheric deposition, natural soil nitrification, synthetic fertilizers, manure and sewage-

15 18 derived NO3 with distinct δ N and δ O values (Figure 1.3).

Figure 1.2: The major components of the nitrogen cycle within a terrestrial system

(Robertson & Groffman, 2007).

15 18 Figure 1.3 depicts the major NO3 sources and their ranges of δ N and δ O values based on a review by Kendall et al. (2007). Atmospheric-N (nitrate and ammonium) deposition has

15 δ N values of 0 ± 15‰. These N-isotope values overlap with NO3 derived from manure and sewage that is often characterized by δ15N values typically between +10 to +20‰, but can be

+ lower depending on the source. N-based synthetic fertilizers, both ammonium (NH4 ) and nitrate

15 (NO3) based, typically have low δ N values, 0 ± 4‰, but can have a total range of -8 to +7‰.

+ 15 NH4 in soil that is nitrified to produce soil nitrate has δ N values that can range from -10 to

+15‰, with most soils in the range of +2 to +5‰ (Kendall, 1998). Since the δ15N values are

18 overlapping, the δ O values of these NO3 sources are what further enable fingerprinting. Nitrate in atmospheric deposition has δ18O values of +60 to more than +90‰, but there have been studies that observe seasonal variation where δ18O values range both higher and lower (Elliot &

18 Brush, 2006; Kendall, 1998). The δ O values of NO3 from sewage and manure are typically less than +15‰ (Aravena & Robertson, 1998; Wassenaar, 1995). NO3-based synthetic fertilizers typically range in δ18O values around +22 ± 3‰ (Wassenaar, 1995). The δ18O values of nitrate from soil nitrification values typically range between -10 and +15‰, but can vary depending on

18 the δ O values of the water and atmospheric O2 at the site of nitrification (Aravena et al., 1993;

Kendall, 1998; Mayer et al. , 2001).

Isotopic studies of tracing nitrogen cycling and transformations in aquatic and terrestrial ecosystems have been conducted quite frequently with much success beginning in the 1960’s

(Hubner, 1981; Kendall, 1998; Letolle, 1980; Nadelhoffer & Fry, 1994). More specifically, the

δ15N value of nitrogen has been utilized as a fingerprint for the point and non-point sources of nitrogen in river systems (Fogg et al., 1998; Harrington et al., 1998). Starting in the early 90’s,

18 δ O began to be included in isotopic fingerprinting of NO3, specifically looking at the role of atmospheric deposition in watersheds (Kendall, 1998). Use of stable isotopes in N- apportionment in larger river basins has been successful in tracing sources and sinks of nitrogen in the Mississippi River (Battaglin et al., 2001; Chang et al. 2002; Kendall et al., 2003; Panno et al., 2006), large rivers in northeastern USA (Mayer et al., 2002), the San Joaquin River in

California (Kratzer et al., 2004), the Oldman River in Alberta (Rock & Mayer, 2004), and the

Seine River in France (Sebilo et al., 2006). There have been several research projects conducted at the University of Calgary Applied Geochemistry Group focused on characterizing sources of

NO3 contamination in freshwater systems in Western Canada using stable isotope tracing techniques. Chao (2011) combined ion chemistry with the isotope mass balances to determine sources of NO3 within the Bow River, Alberta and concluded that in the upstream reaches of the

BRB the dominant source of NO3 was due to nitrification in forest soils entering the watershed and downstream of Calgary between 84-92% of the NO3 load was input from wastewater

+ effluent. Nitrification of wastewater NH4 and denitrification trends were also determined using this isotope tracing technique. However, questions remained open from this study because of the

δ15N and δ18O values did not distinguish the proportion, if any, of agricultural or prairie sources from the urban inputs of NO3 downstream of Calgary. The wastewater-derived NO3 dominated the δ15N and δ18O composition downstream of Calgary and it could not be determined if any mixing of agricultural and urban nitrate sources were occurring using the N and O isotope ratios

15 18 alone. Rock (2005) also focused on characterizing NO3 sources using δ N and δ O in the

Oldman River Basin. Since the eastern portion of the watershed is mainly dominated by agriculture, the δ15N and δ18O values indicated a shift from soil nitrification to a manure- dominated NO3 source with downstream distance from west to east. In this study, it was found that the isotopic composition of riverine NO3 primarily reflected the original source signal rather than the N-transformation processes. Once again, the use of N and O isotope ratios of nitrate did not allow for deciphering any minor N-input into the OMR from human-derived waste, either municipal sewage effluent or septic waste.

15 18 Using the δ N and δ O values of NO3 in combination leads to clear distinction between

+ + atmospheric NO3 and NO3- fertilizer from nitrification of NH4 -fertilizer, soil NH4 , and manure/septic waste (Figure 1.3). There is also only a small overlap in δ15N values between

+ nitrification of NH4 fertilizer and NO3 derived from manure/human waste. Partially overlapping

+ + isotopic values between the nitrification of NH4 -fertilizer, soil NH4 , and NO3 from animal/human waste make distinguishing between these sources challenging in certain environments. Furthermore, there is often no isotopic distinction between NO3 derived from animal manure and sewage. Typically, there tends to be less than a 10‰ variation in δ15N values between different sources and N-isotope fractionation during biological cycling in aquatic

15 18 systems can further alter the original δ N and δ O values of the NO3 sources (Macko &

Ostrom, 1994; Rock & Mayer, 2004; Seiler, 2005). The challenges faced by this dual isotope fingerprinting method may be addressed by using an additional, more conservative geochemical or isotope tracer present in these NO3 sources. Ideally, a multi-isotope tracing method should be

15 18 used to complement δ N and δ O values and fully characterize NO3 sources and sinks in watersheds (Kendall et al., 2007). A good additional isotope tracer is one that is present in measurable amounts, acts conservatively in aquatic systems, and can differentiate between different sources based on their isotopic ratio. Studies elsewhere have found the stable isotope ratio of boron (δ11B) to meet these criteria (Accoe et al., 2008; Bronders et al., 2012; Eppich et al., 2011; Komor, 1997; Saccon et al., 2013; Seiler, 2005; Tirez et al., 2010; Vengosh et al.,

1994; Widory et al., 2004).

15 18 Figure 1.3: δ N versus δ O values of typical NO3 sources in aquatic receptor systems, modified from Kendall et al. (2007).

1.2.2 δ11B as a Tracer of Nutrient Sources

An ideal isotopic tracer of solutes is one that is present in detectable amounts and behaves conservatively in a system. Boron (B), in the form of either boric acid (B(OH)3) or the

- borate anion (B(OH)4 ), although not completely conservative, may act conservatively in many natural settings (Petelet-Giraud et al., 2009; Vengosh et al. 1994; Widory et al., 2004). The element boron has two stable isotopes, 11B and 10B, that have natural abundances of 19.9% and

80.1% respectively (Hoefs, 2009). Due to the relatively large mass difference between the 10B isotopes, boron has a large range of stable isotope ratios (11B/10B) observed in nature, with ratios typically expressed as δ11B. The two naturally occurring stable isotopes preferentially partition into separate species with 11B more abundant in boric acid and 10B more abundant in the borate ion (Barth, 1993). Ubiquitous in nature due to weathering of bedrock, B is commonly added to manufactured products such as detergents, cosmetics, synthetic fertilizers, and constitutes a portion of animal manure (Komor, 1997; Leenhouts et al., 1998; Petelet-Giraud et al., 2009). It is present in detectable concentrations in many aquatic systems, and the large variation in 11B/10B ratios may allow for distinguishing between sources of nutrients in aquatic systems.

The previously reported δ11B values for natural and anthropogenic B sources are summarized in Figure 1.4. Sodium perborate (“borax”), largely mined in the USA and western

Turkey, is a non-marine evaporite that is used commercially (Tirez et al., 2010). The δ11B values for sodium perborate range from approximately -5 to +10‰ and this is applied as the isotope signature for municipal sewage or industrial effluent (Barth, 1993; Tirez et al., 2010; Vengosh et al., 1994, 1999). Komor (1997) reports δ11B values of -2.0 to +0.7‰ for ammonium nitrate and urea fertilizer and 14.8‰ for phosphate fertilizer. However, in many fertilizer types the B concentration in N-fertilizers is below the detection limit (<0.1 µg/g) and a δ11B value could not be determined (Komor, 1997; Tirez et al., 2010; Widory et al., 2004). Widory et al. (2004), found δ11B values from approximately +20 to +40‰ for hog manure, +25 to +29‰ for cow manure, and approximately +11 to 14‰ for poultry manure based on samples taken from livestock in France. Komor (1997) found δ11B values in the range of +7 to +11‰ for hog manure

and +23‰ for cow manure from Minnesota, USA. The difference in δ11B values measured in cow and hog manure in France and USA can be explained by the fact that the B isotope ratio reflects the diet and physiology of the animal (Komor, 1997). Comparing these anthropogenic

δ11B signatures there is very little overlap in the B isotopic ratios of sewage effluent and animal manure making distinguishing between these two anthropogenic sources possible (Figure 1.4).

As for δ11B variation from natural sources, the main source in freshwater is due to weathering and leaching from geologic sources (Barth 1993). This allows for a large range of B isotope ratios from naturally-derived B, depending on the geologic environment; the δ11B values found in geologic sources range from less than -30‰ to greater than +40‰ and are broken down in

Figure 1.5 (Hoefs 2009). Rainwater or atmospheric deposition of B can range from -10 to +50‰ depending on the location (Widory et al., 2005). Thus, boron isotope ratios are only useful for differentiating anthropogenic from natural sources when well-defined within the study area.

11 15 18 When the δ B, δ N and δ O values of NO3 are used in combination, this allows for a more definitive isotopic distinction between sewage effluent, some mineral fertilizers, and livestock manure types as depicted in Figure 1.6.

Boron does not participate in biologic transformations that affect nitrogen compounds in water nor is it affected by redox reactions and hence B has the potential to act conservatively in natural surface waters (Leenhouts et al., 1998; Seiler, 2005; Widory et al., 2013). However, the borate anion has the tendency to adsorb onto clay minerals, iron, and aluminum oxide (Al2O3) surface sites, displacing hydroxyl groups, but this is strongly pH dependent due to boron speciation with a maximum abundance of the borate ion between pH 8.5 and 10 (Figure 1.7)

(Leenhouts et al., 1998; Palmer et al., 1987). During the adsorption process onto clay, a B isotope fractionation of -13 to -22‰ occurs, leaving the remaining borate enriched in 11B and the lighter 10B borate adsorbed onto the clay surface (Barth, 2000; Gonfiantini & Pennisi, 2005;

Palmer et al., 1987). Therefore, when water samples with a pH lower than 8 show an increase in

δ11B values moving downstream along a hydraulic flow path this likely implies mixing of B- sources rather than B-isotope fractionation due to adsorption (Barth, 2000). Isotopic fractionation of boron can also occur during salt precipitation in highly saline environments or boric acid volatilization in laboratory settings, neither of which apply to freshwater environments (Bassett et al., 1995; Seiler, 2005; Xiao et al., 1997). The B isotopic fractionation that occurs during volatilization was investigated by Xiao et al. (1997) and it was found that in waters with near neutral pH there is a -0.4 to -1.9‰ isotopic shift in the remaining boric acid residue during the volatilization. These reactions causing isotope fractionation are likely not of major significance to the watersheds of southern Alberta.

11 Using δ B as a co-tracer for NO3 sources has not been applied to any western Canadian watersheds. However, there are multiple studies using the multi-isotope tracing technique

11 15 18 employing a combination of δ B, δ N and δ O values of NO3 to trace NO3 in groundwater and surface water in France (Bronders et al., 2012; Tirez et al., 2010; Widory et al., 2004, 2005,

2013), Belgium (Saccon, Leis, Marca, Kaiser, Campisi, Bottcher, et al., 2013), Italy (Saccon,

Leis, Marca, Kaiser, Campisi, B??ttcher, et al., 2013), and southern United States (Eppich et al.,

2012; Komor, 1997; Leenhouts et al., 1998; Seiler, 2005; Vengosh et al., 1994). In these studies, the use of δ11B values typically allowed for a distinction between most livestock animal manure

types (cow, poultry, hog, horse), some fertilizers, and sewage effluent or industrial waste signals in ground- and surface water, an advantage over using N and O isotope ratios (see Figure 1.4 for

δ11B values of anthropogenic sources). The background δ11B values of surface and groundwater for these studies were specific to the geology of the study area, depending on the rock type and/or surficial material that interacts with the aquatic system. Therefore, it is important to establish baseline values in a study within a given area. When δ11B is used in combination with

δ15N and δ18O, often a mixing model was used to distinguish % source contribution derived from manure, sewage effluent, and synthetic fertilizer contribution to a watershed or aquifer. Using a

11 15 combination of δ B and δ NNO3 values made it possible in certain study areas to distinguish between these three anthropogenic sources (Figure 1.6) (Widory et al. 2004, 2005, 2013; Accoe et al. 2008; Seiler 2005; Eppich et al. 2012). Therefore, B isotopes should be considered as a valuable addition to a multi-isotope approach to tracing NO3 in aquatic systems.

Sewage Eﬄuent

Urea/Ammonium Nitrate Fer?lizer

Phosphate Fer?lizer

USA Europe Hog Manure

USA Europe Cow Manure

Poultry Manure

-10 5 20 35 50 δ11B (‰)

Figure 1.4: Typical δ11B values of different freshwater contamination sources and background values, adapted from Komor (1997), Tirez et al. (2010) and Widory et al.

(2004).

Figure 1.5: The B isotope variation within natural sources and geologic materials

(Aggarwal & You, 2016; Xiao et al., 2014).

11 15 Figure 1.6: δ B versus δ N values for NO3 – containing contaminant sources based on previously reported literature (Widory et al., 2004, 2005, 2013; Accoe et al., 2008; Seiler,

2005; Eppich et al., 2012).

- Figure 1.7: The speciation of B(OH)3 and B(OH)4 at 25°C and 1 atm with increasing pH

- defined by a) the coordination of B and b) the isotopic composition of B(OH)3 and B(OH)4

(Hoefs, 2009).

1.3 Research Objectives

The objective of this study was to determine the effectiveness of combining the boron

11 15 18 isotope ratios (δ B) with the isotopic composition of NO3 (δ N and δ O) as a tracer of nutrient sources in the surface waters of the Bow and Oldman Rivers. This was achieved by measuring

11 15 18 δ B alone and in combination with δ N and δ O of NO3 in potential urban and agricultural nutrient sources: sewage effluent, mineral fertilizer and cow manure. These isotope ratios were also measured in the receiving aquatic surface water systems and integrated with geochemical, hydrologic, and other isotopic data in the aquatic systems to trace the sources and fate of the nitrate in southern Alberta watersheds.

Study Area

2.1 Introduction

The Bow River (BRB) and Oldman River (OMR) were sampled and analyzed for this study to capture the major contributors and the fate of nutrients within a western Canadian watershed.

The headwaters of the Bow River are found on the Albertan side of the Continental

Divide in the Rocky Mountains, mainly fed by glacial melt-water and snowpack originating at

Bow Lake during peak flow in the spring, and later in the year by an increased amount of baseflow from aquifers (BRBC, 2005). The river flows southeast through Banff National Park, meanders through the foothills and finally to the prairies of Alberta (AB) as seen in Figure 2.1.

With downstream distance the river widens and decreases in gradient through the prairies before it meets at the confluence with the Oldman River, where they form the South Saskatchewan

River (SSR) (Figure 1.1). The Bow River beginning at Bow Lake has an elevation of 1,900 m and flows down to an elevation of 740 m above sea level (asl) at its confluence with the Oldman

River, for a total elevation change of approximately 1,060 m and total length of 645 km (BRBC,

2005). Average annual recorded discharge (1964-2011) near the mouth of the Bow River is

2,776 million m3, making it the sub-basin with the largest contributor to flow in the SSRB

(BRBC, 2005; Environment Canada, 2016b). The Bow River is also the most regulated watershed in Alberta, with eleven hydroelectric facilities, several weirs, irrigation canals, and withdrawal sites for agriculture (AMEC, 2009). The city of Calgary, located on the Bow River,

approximately at km 200, discharges the most wastewater effluent and stormwater to the river.

There are also wastewater effluent contributions from the municipalities of Lake Louise, Banff, and Canmore (AMEC, 2009).

Bassano Dam

Figure 2.1: The Bow River Basin with its major tributaries, divided by its natural sub- regions (BRBC, 2005).

The Oldman River headwaters are made up of three rivers: The Oldman, Castle, and

Crowsnest in the region of Crowsnest Pass, AB, fed by snowmelt from the southern Rocky

Mountains during spring peakflow and increasing groundwater flux in the summer and fall

(AMEC, 2009; OWC, 2010). These rivers merge together at the Oldman Reservoir to become the single Oldman River mainstem. Past the city of Lethbridge, the Oldman River mainstem is joined by the Belly and St. Mary Rivers (AMEC, 2009) as seen in Figure 2.2. It follows a course south of the Bow River, flowing through foothills and eastward into the prairie grassland through

Lethbridge, before the Bow and Oldman Rivers combine northeast of Taber, AB (OWC, 2010).

The OMR has an elevation of 2,100 m asl at the source to 700 m near the mouth, for an elevation change of approximately 1,400 m and a total length of 363 km (OWC, 2010). The Oldman River mean annual discharge (1964-2010) at the mouth of the sub-basin ranges from 1,410 million m3 to 7,100 million m3 during flood years (OWC, 2010). In addition to the natural tributaries, there is an extensive network of storage reservoirs, canals, and pipelines that are part of the irrigation infrastructure (AMEC, 2009; Saffran, 2005). The grasslands portion of the Oldman Basin is considered the most intensive agricultural region in Canada, composed of irrigated cropland, livestock operations and feedlots (AMEC, 2009; Saffran, 2005). Treated wastewater effluent from the city of Lethbridge is discharged into the Oldman River.

Figure 2.2: The Oldman River basin, divided into its natural sub-regions (OWC, 2010).

2.1.1 Climate and Hydrology of Southern Alberta

Southern Alberta has a sub-humid to semi-arid climate, with marked climatic variation between mountain and prairie regions. The temperature in the southern Rocky Mountains are highest in July with a daily maximum temperature of +20 to +25 °C and coldest in January when the average overnight low is -15 °C within the valley bottoms (Environment Canada, 2016a).

Lower temperatures and greater amounts of precipitation, mostly in the form of snowfall in the winter months, are found with increasing elevation in the Rocky Mountains. The range of maximum temperatures throughout the year is about 37°C between summer highs and winter lows (Environment Canada, 2016a). Annual rainfall/snowfall ranges are about 470-600 mm/250-

330 mm around the bottom of Bow River valley and 460-550 mm/200-230 mm around the bottom of Oldman River valley in the Rocky Mountains (Environment Canada, 2016a; Holland

& Coen, 1983). The predominant wind direction is southwesterly, perpendicular to the alignment of the mountain ridges (Gadd, 1995).

In the Alberta prairies, average temperatures range from daily maximums of 23-26 °C in summer to daily minimums of -12 to -15 °C in winter months (Environment Canada, 2016a).

Annual rainfall/snowfall decreases substantially moving west to east across the province of

Alberta, with approximately 330 mm/130 mm around Calgary and 280 mm/110 mm around

Lethbridge (Environment Canada, 2016a). Winds are normally westerly in direction

(Environment Canada, 2016a).

A portion of the flow from each the Bow and the Oldman Rivers stems from mountain precipitation, accumulating as snowpack during the winter and providing water during peakflow

in the spring and summer (BRBC, 2005). Therefore, winters with little snowfall can cause low flow in these rivers, or winters with high snowfall can cause flooding. Groundwater is also an important water source for these rivers and during baseflow the Bow and Oldman rivers are mainly groundwater fed (AMEC, 2009; BRBC, 2005).

2.1.2 Geology

Mountain Geology

From Crowsnest Pass and northward, the Rocky Mountains (also known as the “Front

Ranges”) are composed of dominantly Middle Cambrian to Permian (525-245 Ma) limestone, dolomite and shale with some minor gypsum and anhydrite deposits (Hamilton et al., 1999).

Near Lake Louise some of the oldest bedrock outcrops occur composed of argillite, argillaceous sandstone, conglomerate, quartzite, and quartz sandstone with shale and limestone lenses, dating back to the lower Paleozoic to Neoproterozoic (650-540 Ma) (Hamilton et al., 1999). As the

Front Ranges meet the foothills, the geology is dominated by dolomite, limestone, dolomitic siltstone and silty dolomite (Hamilton et al., 1999). During the Pleistocene epoch (2.5 Ma to 12

Ka), colloquially referred to as the Ice Age, the Cordilleran Ice Sheet originated and descended from the Rocky Mountains into British Columbia and the western edge of Alberta, carving and eroding the mountains into what we see today (Stalker, 1977).

Prairie Geology

Upper Cretaceous (65-100 Ma) formations of calcareous to feldspathic sandstone, bentonitic mudstone, and lenses of siltstone, mudstone and coal beds dominate the underlying bedrock of the rolling plains of Alberta (Hamilton et al., 1999). During the Pleistocene Epoch, a series of glacial and interglacial periods carved and eroded the mountain and prairie landscape

(Stalker, 1977). The Laurentide Ice Sheet from the east covered most of Alberta during the last glaciation somewhere between >20,000 to 12,000 years before present, and left behind the glacial deposits we find in the prairies today (Levson & Rutter, 1996). The Cretaceous bedrock is now overlain by varying thicknesses of till, glaciolacustrine, glaciofluvial, and wind-blown deposits left behind from the rich glacial history of the last glaciation, known as the Late

Wisconsinan (Levson & Rutter, 1996).

2.1.3 Soil Types

Throughout the Rocky Mountains, soil horizons are typically found as thin veneers over glacial deposits in the valleys, with lots of exposed rock. The most extensive soil type is

Brunisolic, but Luvisolic, Chernozemic, Cryosolic, Regosolic, Gleysolic and Organic are scattered throughout (Agriculture Canada, 1991; Holland & Coen, 1983). Brunisolic soils are typically forest soils that are developed enough to exclude them from the Regosolic category (an undeveloped soil), but lack the degree of horizon development of other soil orders (Agriculture and Agri-Food Canada, 1998). Chernozemic soils are most common in the foothills and through the prairies around Calgary and Lethbridge, characterized by well to imperfectly drained soils

that are darkened by the presence of organics due to decomposition in the grassland (Agriculture and Agri-Food Canada, 1998). There is also a minor presence of Solonetzic soils, that are very hard when dry and swell up to a sticky, impermeable mass when wet. They occur on saline parent materials in association with Chernozemic soils (Agriculture and Agri-Food Canada,

1998).

2.1.4 Water and Land Use

2.1.4.1 Water Use

In addition to maintaining requirements for aquatic life, the Bow and Oldman River basins are used to sustain the needs of a growing population in Southern Alberta. Both have almost pristine headwaters in the west and with downstream distance to the east water use increases. These rivers are used for (BRBC, 2005; OWC, 2010):

• Hydroelectric generation

• Water licensing and allocations (e.g. for irrigated agriculture, among others)

• Effluent dilution

• Recreation

Water licensing and allocation is broken down in Figure 2.3, with agricultural irrigation accounting for more than 75% of the water use in the BRB and over 90% in the OMR (AMEC,

2009; BRBC, 2005). The agriculture water allocation defined separately from irrigation in Figure

2.3 refers to water allocation for livestock operations. In the BRB the irrigation districts are

divided into the Western (WID), Eastern (EID), and Bow River Irrigation District (BRID). On the OMR, there are 9 irrigation districts that cover 11% of the entire watershed that include the

Taber Irrigation District (ID), Lethbridge Northern ID, United ID, Mountain View ID, Raymond

ID, Magrath ID, Leavitt ID, St. Mary River ID, and Aetna ID (Rock, 2005). On the BRB there are eleven hydroelectric power generation stations operated by TransAlta Utilities, making it the most regulated river in Alberta (BRBC, 2005). These hydroelectric stations influence the timing and magnitude of stream flows in the BRB that in turn affects the river water quality and ecosystem characteristics. The city of Calgary is the largest municipal water user on the BRB, sourcing its water from the Bearspaw and Glenmore Reservoirs of the Elbow and Bow Rivers, respectively (BRBC, 2005). Lethbridge is the most substantial municipal water user of the OMR, sourcing its water directly from the river. Urban areas typically return almost all their water through wastewater effluent discharge back to the river. There are two types of wastewater conveyance systems: sewer lines that carry waste to WWTP for treatment before discharge back into the river, and storm sewer lines that carry runoff either directly into the river or into a treatment facility (BRBC, 2005). When stormwater is treated, it is piped through to the WWTP facility and otherwise it is discharged back into the river with or without the use of screens, wet ponds, or wetlands to improve water quality (BRBC, 2005). More detail on the WWTP treatment processes can be found in Section 2.2.

Figure 2.3: The major types of water use in water licensing for the Bow and Oldman Rivers

(BRBC, 2005; OWC, 2010).

* Other for Bow River refers to smaller licenses for golf courses, parks, water management, waterfowl projects, household, and storage purposes. For the Oldman River it refers to water management, lake stabilization, and wildlife enhancement

2.1.4.2 Land Use

Land use apportionment along the BRB is divided into the upper reaches in the Rocky

Mountains with exposed rock and forest cover that remains mainly pristine, the foothills that are mainly forest covered with some sparse amounts of cropland, and downstream of Calgary where agricultural use dominates with cropland and grassland (BRBC, 2005). The dominant agriculture is dryland, irrigated crops (cereal, forages, oil seed, speciality crops), livestock operations and feedlots. Livestock includes primarily cattle, but also chickens, pigs and some sheep, however number for livestock within the BRB are not available. Downstream of Calgary there is also some industrial development accounting for approximately 3% of water use, mainly oil and gas

exploration and development, with the area below Bassano Dam being the most developed.

Municipal development accounts for only 1% of the Bow River Basin.

In the OMR, the western part is dominantly forest covered. In the eastern part of the basin agricultural activities increase greatly along with some municipal (1%) and commercial (1%) use. Agriculture is the main land use in the OMR basin (the main crops being cereal and canola), and the rest divided between grassland, forested area, and native prairie land (OWC, 2010). As part of the agriculture within the OMR basin there are approximately 580 livestock feeding operations that includes cattle, poultry and hog. With a total of 1.24 * 106 cattle, the OMR basin provides about 75% of the province’s total beef cattle for slaughter (AAFRD, 2000; Rock &

Mayer, 2006).

In terms of nitrogen contribution to these watersheds, Chao (2011) found that urban wastewater effluent from Calgary is the major anthropogenic contributor to the BRB and Rock

(2005) reported that manure, followed by mineral fertilizer, are the leading anthropogenic contributors along the OMR.

2.2 Wastewater Treatment Plant Processes

Centralized WWTP facilities are the most common mode for treating sewage within municipalities. Southern Alberta has some of the highest quality wastewater treatment processes in the country that include primary, secondary, and tertiary/advanced treatment. The WWTPs that discharge into the BRB are (from up-to downstream) Lake Louise, Banff, Canmore, and in

Calgary, Bonnybrook, Fish Creek, and Pine Creek. The Lethbridge WWTP is the only major

sewage treatment plant that discharges into the OMR. On average the mean daily flow rates to the WWTP facilities within Calgary are over 300 ML/day for Bonnybrook, between 75-100

ML/day for Pine Creek, 30-50 ML/day from Fish Creek, and for the Lethbridge WWTP between

35-40 ML/day. The average flow rates of the WWTPs upstream of Calgary along the BRB are variable with peak flows during tourist seasons but the facilities at Lake Louise treat approximately 4 ML/day of sewage, Banff 7-10 ML/day, and Canmore 5-10 ML/day. Most of these plants follow a standard biological three-phase treatment procedure; this includes primary treatment by sedimentation, secondary treatment by a microbial mediated aeration tank, and tertiary treatment by Ultra-Violet radiation. The exception is the Fish Creek WWTP that does not use biologic N-removal and therefore much of the N is discharged to the BRB in the form of

+ - NH4 (ammonium) opposed to NO3 . The general sequence of treatment is described in the following sections (City of Calgary, n.d.; City of Lethbridge, n.d.).

2.2.1 Pre-Treatment

This is screening to remove large objects, debris, grit, and oily scum using 6mm screens.

This is performed to avoid shock loadings, to condition wastewater for subsequent treatment processes and to protect downstream mechanical equipment.

2.2.2 Primary Treatment & Clarifiers

This primary treatment physically separates solids from wastewater using settling

(sedimentation) tanks where solids settle to the bottom and oils, fats and grease are skimmed off at the top. The solids and sludge are pumped to digesters for anaerobic decomposition and the remaining primary effluent is sent to the bioreactors.

2.2.3 Secondary Treatment & Clarifiers

A bioreactor (aeration tank) uses an activated sludge process by injecting air into the tank containing a high concentration of microbes that accelerate biological degradation of organic matter. The water is separated into a series of aerobic, anoxic and anaerobic tanks where water flows through in series and allows the different microbes to react with organics, phosphate, and ammonia.

Phosphorus is removed by chemical precipitation with Alum (Al2(SO4)3) or using enhanced biological phosphate removal and nitrogen is removed by a nitrification/denitrification process:

+ - + Nitrification: NH4 + 3/2 O2 à NO2 + 2H + H2O (Nitrosomonas bacteria)

- - NO2 + 1/2O2 à NO3 (Nitrobacter bacteria)

+ - + Total: NH4 + 2O2 à NO3 + 2H + H2O

- - Denitrification: 4NO3 + 5CH2O à 2N2 + 5CO2 + 3H2O + 4OH

Pathogens are also reduced during this denitrification process. After the aeration tank, water is transferred to a secondary settling tank or clarifier where after settling any microorganisms remaining are recycled back to the bioreactor. Specifically, at the Pine Creek plant, water flows through cloth-media disk filters where solids, phosphorus and algae are reduced even further before moving to disinfection. The remaining effluent is then sent to the filtration and U.V. disinfection building for tertiary treatment.

2.2.4 Tertiary/Advanced Treatment

The water is exposed to Ultra-Violet (UV) radiation to kill remaining pathogens by causing them to be unable to reproduce. After tertiary treatment, the post-UV final effluent is discharged into the river. The sludge produced from the secondary treatment is shipped away for agricultural land application.

2.3 Manure and Fertilizer

Fertilizer samples used for this study were previously collected by the Applied

Geochemistry group (University of Calgary) from fertilizer sources throughout Alberta. Dried and coarsely ground cow manure samples were kindly provided by Dr. Benjamin Ellert of

Agriculture and Agri-Food Canada (AAFC) from their Lethbridge Research and Development

Centre.

Cow manure used as organic fertilizer or from feedlots is abundant in the reaches of the

Bow River downstream of Calgary and throughout the studied reaches of the Oldman River

(BRBC, 2005; OWC, 2010). Manure contains nitrogen mainly in the form of urea (CO(NH2)2),

+ that is converted to ammonium (NH4 ) once it is introduced into the subsurface and eventually undergoes nitrification (Robertson & Groffman, 2007). Most manure is mixed with straw or wood chip bedding, that also contains nitrogen.

The most common mineral fertilizer used in the Canadian prairies is Urea, followed by monoammonium phosphate, ammonium sulphate, and urea ammonium nitrate (StatsCan,

2016b). Figure 2.4 shows 2011-2015 Canadian agriculture shipment data for these fertilizer types. Once applied, these fertilizers will also undergo nitrification in situ through the microbial oxidation of ammonium (Robertson & Groffman, 2007). Ammonium nitrate fertilizer is not applied in the Canadian Prairies.

Figure 2.4: Fertilizer shipments to Canadian agriculture markets of the top four product types in the prairies, cumulative data (StatsCan, 2016b).

Sampling and Analytical Methods

3.1 Fieldwork

To complete this project, surface water samples of the Bow and Oldman Rivers were collected from 12 mainstem BRB sites, 5 mainstem and 3 tributary sites of the OMR. They were sampled during peak-flow (June) and baseflow (September/October) over a 16-month period.

3.1.1 Overview

Surface water sampling of the Bow River occurred between June 2014 and October 2015 and of the Oldman River between October 2014 and October 2015. During this time both rivers were sampled in the early summer (June) during peakflow and in the fall (September/October) during baseflow for a total of 4 sampling events along the BRB and 3 along the OMR. The baseflow dataset of the Oldman River was collected after October 15 each year to ensure irrigation canals had been drained for the season. The treated wastewater effluents that are discharged into the Bow River from WWTPs in Lake Louise, Banff, and Canmore as well as

Bonnybrook, Fish Creek and Pine Creek within Calgary were sampled in February and October

2014. The final effluent discharged into the Oldman River from the Lethbridge WWTP was sampled in October 2014. In February 2016, supplementary final effluent samples from the Lake

Louise, Banff, Canmore and Lethbridge WWTPs were collected in order to complete laboratory analyses.

The first sampling season (summer/fall 2014) was used as reconnaissance for assessing nitrogen compounds and boron concentrations in river water. The sampling sites within

Banff National Park (around Banff and Lake Louise) could not be accessed until September 2014 after obtaining a research permit to sample within the National Park. Additional sampling sites around Lake Louise, Banff, and Canmore were sampled in September 2014 to assess changes in nutrient inputs around these municipalities. During one sampling event in September 2014 a canoe was used to access points directly within the Bow River wastewater plumes of the

Bonnybrook, Fish Creek and Pine Creek WWTPs in Calgary to capture elevated ammonium

(NH4) and boron concentrations within the plumes. This method was logistically difficult for most sites along the two river basins and hence this was not further pursued after this sampling event. During the second sampling season (summer/fall 2015), sampling sites were refined so that only the sampling sites that captured marked fluctuations in nutrient concentrations were visited.

Sampling sites along the BRB and OMR were chosen based on sampling locations from the previous studies by Chao (2011) and Rock (2005). Single sampling points along the rivers and tributaries were assumed to be representative of the chemical and isotopic conditions of the river at that location. Sample sites were chosen based on accessibility by a combination of car and by foot. Water was sampled from the edge of the river using a sampling pole at a point where there was steady flow and assumed good mixing of solutes, a common practice for surface water studies (Chao, 2011; Stednick, 1991; USGS, 2006). Although cross-sectional sampling of

a river at different points and depths is the most representative method for capturing data, this method is time-consuming and impractical for the purposes of this study.

3.1.2 Surface Water Sampling Sites

The locations of the final sampling sites with their locations and number of times sampled can be found in Table 3.1. The 2015 surface water sites of the Bow River used to obtain river water samples for analysis are above Lake Louise (0 km), above Banff (56 km), below

Banff (64 km), above Canmore (79 km), Cochrane (179 km), below Bonnybrook (205 km), below Fish Creek (220 km), below Pine Creek (228 km), Carseland (302 km), Bow City (482 km), Scandia (513 km) and Ronalane (571 km) for a total of 12 surface water sites (Figure 3.1).

This constitutes the entirety of the Bow River before its confluence with the Oldman River. The

Oldman River surface water sampling sites were Monarch (0 km), below Belly River (15 km),

Pavan Park (62 km), Downstream Tributaries (113 km), and Taber (140 km) for a total of 5 main-stem surface water sites. The three tributaries sampled along the Oldman River were Haney

Drain (96 km), Battersea Drain (98 km), and Little Bow River (110 km). Select tributaries along the OMR were chosen for sampling based on results from Rock’s OMR study (2005) where tributaries in the eastern part of the OMR had significant NO3 concentrations (on average >1

15 18 mg/L) and δ N and δ O values that were found to primarily represent NO3 sources. The isotopic signal from these tributaries contributed to the signal of the mainstem OMR. On the contrary, Chao (2011) found the tributaries along the BRB had considerably low NO3 concentrations (typically <1 mg/L) and this did not influence the NO3 isotopic signal of the

mainstem since it was dominated by sewage effluent signal from Calgary. Therefore, sampling of the tributaries along the BRB was not repeated. These sites constitute a portion of the Oldman

River centering on the Lethbridge WWTP, up- and down-stream into agricultural areas (Figure

3.2).

Table 3.1: Surface water sampling sites GPS coordinates.

Number of Times Site Name Watershed Km Longitude Latitude Accessed Lake Louise Above Bow River 0 51°26'32"N 116°12'48"W 3 Banff Above Bow River 56 51°10'13"N 115°35'57"W 3 Banff Below Bow River 64 51°10'33"N 115°30'30"W 3 Canmore Above Bow River 79 51°10'35"N 115°37'77"W 4 Cochrane Bow River 179 51°18'32"N 114°48'71"W 4 Bonnybrook Below Bow River 205 50°98'67"N 114°02'68"W 4 Fish Creek Below Bow River 220 50°90'68"N 114°01'15"W 4 Pine Creek Below Bow River 228 50°84'22"N 113°95'47"W 4 Carseland Bow River 302 50°83'13"N 113°41'74"W 4 Bow City Bow River 482 50°43'09"N 112°22'67"W 4 Scandia Bow River 513 50°24'64"N 112°07'68"W 4 Ronalane Bow River 571 50°04'81"N 111°59'03"W 4 Monarch OMR 0 49°45'0"N 112°59'56"W 3 Below Belly River OMR 16 49°47'24"N 113° 7'21"W 3 Pavan Park OMR 52 49°45'2"N 112°51'20"W 3 Haney Drain OMR - Trib 96 49°51'26"N 112°37'31"W 3 Battersea Drain OMR - Trib 98 49°51'49"N 112°34'16"W 3 Little Bow River OMR - Trib 110 49°53'16"N 112°28'32"W 3 Downstream Tributaries OMR 113 49°52'10"N 112°27'37"W 2 Taber OMR 140 49°48'54"N 112°10'17"W 3

38 sampling sites, andR sites, sampling fivewithofrepresentsriversamplingsites,Calgary Calgary 2 river thewithinReachupstream RiverBow three with OverviewofsamplingRiverBow3.1: down three into 1 the sites Figure reaches.Reachbroken RiverBow represents the eachBowfourwithriversamplingsites.representsofCalgary 3 the downstream River

riversamplingsitesandLethbridge two withtr of three RiverandincludingReach represents Lethbridge upstream River2 and the Oldman downstream Figure 3.2: Overview of the Oldman River sites broken down into two reaches.ReachrepresentsOverviewof Figuredowntwo Oldman3.2:1 into the sites Riverbroken the ibutary sites.ibutary

3.1.3 Wastewater Treatment Plant Effluent Samples

Final post-UV effluent from the WWTPs of the towns of Lake Louise (LL), Banff (BF),

Canmore (CM), and Bonnybrook (BB), Fish Creek (FC) and Pine Creek (PC) within Calgary were sampled directly before being discharged into the Bow River, with locations represented in

Figure 3.1. Lethbridge (LB) WWTP final effluent was sampled before being discharged into the

Oldman River, with location represented in Figure 3.2.

3.1.4 Surface Water Sample Collection & Preparation

Temperature, pH, and electric conductivity were measured in-stream at each sampling site using an Orion 4 Star portable pH meter (Thermo Scientific) and a Russell RL060C portable conductivity meter (Thermo Electron Corporation). The pH meter was calibrated before each field day using pH buffers of 4.0, 7.0, and 10.0 (VWR International). At each sampling site between 0.5 to over 2.0L of water was collected, filtered, and preserved as necessary into appropriate bottles. The samples were filtered using a Nalgene hand-pump filtration device with

0.45 µm nitrocellulose filters (Millipore Corporation). The filtration device was triple rinsed with sample water between each sample location. For collection and storage, samples for cation and anion concentrations and for isotopic analysis were collected in polyethylene Nalgene bottles

(VWR International), leaving headspace at the top for expansion if the samples were to be frozen. Bottles were triple rinsed with filtered sample water before filling. While in the field samples were kept at approximately 4°C in cooled containers until they were returned to the laboratory and either processed within 72 hours or refrigerated until further processing. For anion

and NO3 isotope analysis, one 60 mL bottle was frozen until ready to be processed. For cation analysis, samples were collected in a 60 mL bottle and acidified to a pH < 2 with Omnitrace concentrated nitric acid. For boron concentration and isotope analysis a 1 L bottle was used.

From the same bottle, stable isotope ratios of water and alkalinity were measured. When NH4

15 concentrations and δ N-NH4 were analyzed in 2014, an additional 1 L of water was filtered and acidified to pH < 2 using concentrated analytical grade sulphuric acid.

3.2 Laboratory Analyses

3.2.1 Physical and Chemical Data

A Varian 725-ES Inductively Coupled Plasma Optical Emission Spectrophotometer

(ICP-OES) within the Applied Geochemistry Group (AGg) at the University of Calgary was used to measure chemical concentrations of major cations (Ca2+, Mg2+, K+, Na+, Si4+, Sr3+) of water samples. The spectrophotometer was calibrated (r2 > 0.99) via serial dilution of a certified standard or set of standards for each major cation (BDH Ltd.). Analytical precision and accuracy

+ for analysis is typically ± 5 %. Boron concentrations, measured from the B ions of the B(OH)3

- or B(OH)4 compounds, in water samples were initially measured using the ICP-OES, but since many samples had concentrations near the ICP-OES detection limit for B (5 µg/L), a greater sensitivity was needed to accurately determine B concentrations. For final values, samples were sent to the SWAMP laboratory at the University of Alberta for analysis on the Thermo Scientific iCAP-Quadrapole Inductively Coupled Plasma Mass Spectrometer (Q-ICP-MS) that has a

detection limit of 0.005 µg/L for boron (Thermo Scientific) with an accuracy better than 92 ±

11%. A Dionex ICS 2000 ion chromatograph (Dionex Corporation) was used for anion analysis

- 2- - of major anions (Cl , SO4 , NO3 ) in water samples. A 25 µL aliquot of each sample was injected onto an Ion Pac AS18 anion column (Dionex Corporation) and then separated by isocratic elution using 35.0 mM potassium hydroxide with a flow rate of 1.1 mL/min and a column temperature of 30°C. Analytical precision and accuracy for analysis is better than 5%. An Orion

- 960 Titrator (Thermo Electron Corporation) measured the bicarbonate (HCO3 ) total alkalinity using titration on a 10 mL aliquot of sample with 0.01M sulphuric acid (Fisher Scientific).

+ During baseflow sampling in 2014 water samples were analyzed for ammonium (NH4 ) using a

Dionex DX-120 ion chromatography system (Dionex Corporation) with an analytical precision of ± 5%.

3.2.2 Solid Sample Preparation

The dried manure samples of stockpiled manure in straw bedding (SMS), stockpiled manure in woodchip bedding (SMW), composted manure in straw bedding (CMS), and composted manure in woodchip bedding (CMW) were analyzed for both B and NO3 concentrations and isotopic compositions in solution. Several types of mineral fertilizers used in agriculture such as ammonium sulphate ((NH4)2SO4), ammonium phosphate (NH4*H2PO4), urea

((NH2)2CO), and lawn fertilizers (NH4*H2PO4+other) were also analyzed for B concentration and isotopic composition in solution. The nitrogen content and isotope data from these fertilizers was taken from a previous study conducted by the Applied Geochemistry Group at the

University of Calgary (unpublished data). A summary of the manure and fertilizer samples is provided in Table 3.1. Due to the extremely low boron concentrations in urea and ammonium nitrate fertilizers (see Section 7.2) only one urea fertilizer sample was successfully analyzed for boron content. For each of the solid samples analyzed, simple procedures modified from Komor

(1997) were used to prepare them for chemical and isotope analysis. The samples were prepared in order to analyze water-soluble ion chemistry and isotope ratios that would be representative of sources entering the studied watersheds.

Table 3.2: Sample list for the dried cow manure and mineral fertilizers analyzed in this study.

Sample Sample Type ID Cow Manure Stockpiled Manure SM Composted Manure CM Mineral Fertilizer

(NH4)2SO4 BJMJ-8 (NH4)2SO4 BJMJ-9 NH4*H2PO4 BJMJ-12 NH4*H2PO4 BJMJ-16 NH4*H2PO4 BJMJ-21 (NH2)2CO BJMJ-37 NH4*H2PO4+other BJMJ-41 NH4*H2PO4+other BJMJ-47

3.2.2.1 Manure Samples

For chemical and isotopic analysis, 1.5 g of dried cow manure was weighed and mixed with 25 mL of deionized water in a 50 mL Teflon centrifuge tube. The centrifuge tubes were left on an automatic shaker for 4 hours then centrifuged for 30 mins at 2500 rpm. The samples were then filtered using 0.45 µm Acrodisc GHP membrane syringe tip filters (PALL Life Sciences).

The filtered solutions were analyzed for NO3 and B concentrations and isotopic composition within the week of making up the solution or else refrigerated at 4°C until analysis. For NO3 analysis the dried manure samples were first autoclaved to kill any denitrifying bacteria that may have affected nitrate concentrations when mixed into solution.

3.2.2.2 Fertilizer Samples

Based on the fertilizer with the highest solubility, 5 g of fertilizer was added to 25 mL deionized water in a 50 mL Teflon centrifuge tube. The centrifuge tubes were left on an automatic shaker for 4 hours before being filtered using 0.45µm Acrodisc GHP membrane syringe tip filters (PALL Life Sciences). The filtered solution was then analyzed for B concentration and isotopic composition. The N- data for the fertilizers was taken from a previous study (Applied Geochemistry Group, unpublished data).

3.2.3 Boron Ion Exchange

This procedure was modified from the method developed by Lemarchand et al. (2002) for the separation of boron from natural river water samples, with input from researchers at the

Nancy Lorraine French National Institute for Agricultural Research (INRA) and the French

Geological Survey, Bureau de Recherches Géologiques et Minières (BRGM). With the use of boron-specific resin, Amberlite IRA 743 (Amberlite resin), boron was separated and concentrated from Bow and Oldman river water samples, WWTP effluent, organic-rich cow manure extracts, and mineral fertilizer solutions to measure accurate δ11B values. The ion exchange procedure is based on altering pH values. Amberlite resin is a tertiary amine and a weak base with an acid dissociation constant (pKA) of ≈ 7. Consequently, at pH values neutral to alkaline (7 to 9) boron in the form of boric acid (B(OH)3) is strongly fixed onto the alcohol groups of the glucamine and at pH values <7 acts as an anion exchanger, exchanging the

B(OH)3- ion (Lemarchand et al., 2002). Due to the nature of trace element experiments, extra precautions had to be taken to ensure as little boron contamination as possible. Laboratory glassware was avoided since most are made of borosilicate compounds that contain leachable boron. Therefore, for both field and laboratory components there were exclusive use of plastic and Teflon ware. Stringent measures were taken to maintain the integrity of low boron samples

(US EPA, 1994).

The modified procedure for extracting and concentrating boron in water samples that was used in this study is as follows:

Preparation

Using a mortar and pestle Amberlite resin beads were crushed to 100-200 mesh size.

They were then wetted with deionized water to create a paste and 50 µL of resin was pipetted into a 2 mL polypropylene Pierce Centrifuge Column (Thermo Scientific). The gravity flow-type column was loaded with deionized water to ensure all the resin beads settled as an even bed on the bottom of the column. Figure 3.3 outlines the dimensions of the column. Using an Ismatec

MCP Standard peristaltic pump device (Cole Parmer Instrument Company) the beads were flushed with 0.5 M HCl for 2 hours or more. The Amberlite resin beads extract boron more efficiently as they become conditioned with multiple uses, so beads were preserved in 0.1 M HCl between experiments since the resin loses its exchange properties if dried out. Several columns were prepared for separate solution types including WWTP effluent, solutions of mineral fertilizer, cow manure solutions, and for river water samples. At this point the Amberlite resin is prepared for boron extraction from the samples.

Figure 3.3: Dimensions of the 2 mL Pierce centrifuge columns, with a total column capacity of 5.5 mL. The 50 µL of Amberlite resin forms an even bed at the base of the column.

Extraction

Several ion exchange columns were set up to run multiple samples at one time. Following the list of iterations in Table 3.2, the columns were properly conditioned before samples were passed through the gravity flow columns and a final volume of 500 µL HCl, containing the boron eluted from the Amberlite resin beads, was collected in a 2 mL microcentrifuge vial to be analyzed for δ11B values. The pH of the sample being loaded was maintained between 7 to 9, but can be as low as 5.5, to prevent potential precipitation of hydroxide or carbonate species

(Lemarchand et al., 2002). This ideal pH range also ensures the most efficient rentention of B onto the Amberlite resin. To test this SRM 951 standards diluted to 10 µg/L (an average boron concentration value for river water) as well as samples of WWTP effluent, mineral fertilizer solution, and manure solution, were run through the Amberlite resin ion exchange process. The final extracted 500 µL HCl was diluted to 7 mL with deionized water in order to run on the ICP-

OES. This determined a final boron concentration after the extraction to ensure a procedure efficiency of better than 90%. The ICP-OES was used for testing the extraction efficiency since the equipment conveniently resides in-house at the University of Calgary Applied Geochemistry laboratory and could be run at a high frequency. These values were not recorded as final B concentrations (see Chapter 5).

Table 3.3: Boron extraction procedure using Amberlite IRA 743 resin in a gravity-flow column.

Volume Solution Iterations (µL) Strength Solution Column Cleanup: 5 100 0.1 M HCl

5 100 H2O 5 100 0.5 M NaOH Column Conditioning:

5 100 H2O Load sample: pH adjusted between 7 to 9 with NaOH Washing Procedure:

5 50 H2O 5 50 0.6 M NaCl

5 50 H2O Elution: 1 100 0.5 M HCl 4 100 0.1 M HCl

Fertilizer samples

The high concentrations of salts in the mineral fertilizer solution samples resulted in high cationic charge of the sample that can interfere with boron isotopic analysis. Therefore, fertilizer solutions were passed through an AG50W-X8 (100-200 mesh) Polyprep Biorad cation exchange column with 2 mL resin before the Amberlite resin boron exchange. The cationic resin was conditioned with 3 resin volumes of 0.2 N HCl before loading the fertilizer sample solutions.

The fertilizer sample was acidified with HCl a to 0.2 N solution, loaded onto the cationic resin where the cations were exchanged with H+ and the eluted solution that contains boron is collected in a 15 mL centrifuge tube. The resin was then washed with 3 resin volumes of 0.2 N

HCl and the washed solution was collected in the same centrifuge tube as the eluted fertilizer sample. Before loading this sample onto the Amberlite resin, the collected solution was adjusted to a pH of 7 to 9 using NaOH.

This ion exchange process ensured boron concentrations were high enough to measure an accurate isotopic value and there was no interference by other ions during isotopic analysis.

3.2.4 Isotope Ratios

The O and H isotope ratios of water, the N and O isotope ratios of nitrate, the N isotope ratios of ammonium, and the isotope ratios of boron (B) were all determined using isotope ratio mass spectrometry. Isotope ratios are expressed using the delta (δ) notation (see Section 1.2.1) where R is 2H/1H, 18O/16O, 15N/14N, or 11B/10B. δ18O and δ2H values are reported with respect to

Vienna Standard Mean Ocean Water (V-SMOW), δ15N values are reported relative to AIR, and

δ11B values reported relative to NIST SRM-951 (boric acid).

3.2.4.1 δ2H and δ18O of Water

Isotope abundance ratios of H and O of water were measured using laser spectroscopy.

Approximately 750nL of water was injected into a Los Gatos Research ‘DLT-100’ instrument.

Water is vaporized and expanded into a laser cell and measured directly by ‘Off-Axis Integrated-

Cavity Output Spectrometry’ (Sturm & Knohl, 2009). The measurement accuracy for δ2H and

δ18O values is better than ±1.0‰ and ±0.1‰, respectively.

3.2.4.2 δ15N and δ18O of Nitrate

- The isotopic composition of NO3 in water samples is measured using the standard denitrifier method (Casciotti et al., 2002; Sigman et al., 2001). Tryptic soy broth amended with

10 mM potassium nitrate, 1 mM ammonium sulphate, and 1 mL/L antifoaming agent was inoculated with the P. aureofaciens bacteria and cultivated for 6-10 days to ensure complete consumption of O2 in the headspace and the amended nitrate. The bacterial culture was split into

2 mL aliquots in 20 mL headspace vials, crimp-sealed with Teflon-backed silicon septa and

- purged for 3 hours with N2 gas. Water samples with dissolved NO3 (30 nmol) were then injected into the sample vials and left overnight to allow for complete conversion of nitrate to N2O before the addition of 0.1 mL 10 N NaOH to stop bacterial activity and scavenge CO2. A constant sample size of 30 nmol was targeted for each batch of water samples to simplify the corrections

15 18 of the δ N and δ O values. For isotopic analysis, N2O was stripped from the sample using a helium carrier gas. The sample was carried through to a U-tube sitting in a liquid nitrogen bath, the stopcocks of the U-tube were closed and the U-tube was removed from the bath. The sample is then placed in-line with a Thermo Finnigan MAT Precon device (Thermo Scientific) where it is sent through a chemical trap for water and CO2, the N2O was cryogenically focused, and the

N2O peak was chromatographically separated from any remaining CO2. Subsequently, the sample was sent to the Thermo Finnigan Delta V (Pyro) isotope ratio mass spectrometer (IRMS)

(Thermo Scientific) for measurement of the 15N/14N and 18O/16O ratios. The accuracy is better than ±0.5‰ for δ15N and ±1‰ for δ18O.

3.2.4.3 δ15N of Ammonium

15 + The analysis of δ N-NH4 of water samples was based on the method developed by

Sebilo et al. (2004). It began by preparing filter packages for ammonium diffusion. 30 µL of 8N

H2SO4 was pipetted onto a glass-fibre filter (APFD, 25 mm, Millipore) cut to approximately 15 mm x 5 mm. This filter was then wrapped and enclosed within a hydrophobic filter (‘Mitex’,

PTFE, 46 mm diameter, Millipore) to form a filter package. The filter package is impermeable to water but will allow NH3 gas to diffuse through and be trapped on the glass-fibre filter paper.

The filter package was placed in a 250 mL incubation bottle containing a magnetic stirrer to

+ which a water sample with 150 µg of NH4 -N was added. If necessary deionized water was added to obtain a volume of between 100-150 mL. If the volume of sample needed for 150 µL of

+ NH4 -N exceeded 150 mL, the sample was split between multiple incubation bottles and the filters used were combined into a single tin cup at the end for isotopic analysis. After adding the filter packages, 2 mL of 5 N NaOH was introduced and the bottle was quickly and tightly closed.

+ The incubation bottles were slowly stirred for 1 week at room temperature to ensure all NH4 is volatilized to NH3 gas, occurring as a result of a change of pH to >12 when NaOH is added.

When this was complete, the filters were removed from the bottles and put into a vacuum-drier for a minimum of 24 hrs. When completely dry, the glass-fibre filters were removed from the hydrophobic filter package and packed into large tin cups to be thermally decomposed in an elemental analyzer (Elementar vario ISOTOPE cube) and coupled to a Thermo Finnigan Delta V isotope ratio mass spectrometer (EA-CF-IRMS) (Thermo Scientific) for analysis of δ15N of the

N2 gas produced. Analytical precision is better than ±0.5‰.

3.2.4.4 δ11B of Boron

Initially, water samples were prepared and analyzed for δ11B values on a Triton Thermal

Ionization Mass Spectrometer in negative ion mode (N-TIMS) (Thermo Scientific). This method measures the isotopic ratios of BO2- ion by heating an Re filament loaded with 2-3 nanograms

(ng) of B from the water sample in an ion source to approximately 970°C, measuring the signals and taking measurements of ratios between mass 42 (10B16O16O) and mass 43 (11B16O16O)

(Duchateau & De Bievre, 1983; Hemming & Hanson, 1994). The advantage to this method is high analytical sensitivity, allowing measurement of boron in the ng range, ideal for low concentration river samples. Disadvantages are there the length of time needed to analyze each sample, and the analytical precision of the method, which is stated at ± 0.7-2.0‰, but with the

Triton instrument at the University of Calgary Department of Physics and Astronomy precision was ± 2.0‰ or greater. Also, there is potential for isobaric interference on mass 42 by the CNO- ion, if there are large amounts of organics in the sample, however this was not observed during this study (Hemming and Hanson, 1994). The low precision achieved with the instrumental setup and other technical difficulties made analysis of δ11B in the water samples difficult and unreliable.

A Neptune Multi-Collector Inductively Coupled Plasma Mass Spectrometer (MC-ICP-

MS) (Thermo Scientific) at the University of Calgary, Department of Physics and Astronomy, was used to complete analysis of δ11B values with analytical techniques modified from Guerrot et al. (2011). Analysis begins when approximately 100 µL/min of water sample is taken up into a double-pass cyclonic quartz spray chamber with a low-flow glass microconcentric nebuliser. The

mass spectrometer then takes isotopic measurements of the 11B/10B ratios of the B+ ions. MC-

ICP-MS requires correction for instrumental mass bias, especially when measuring boron which has a large memory effect, using sample-standard bracketing (Aggarwal et al., 2003; Vanderpool et al., 1994). This means that the mean 11B/10B ratio of NIST SRM-951 preceding and following two consecutive samples or standards is taken and used to determine the δ11B value of the sample or standard, check the accuracy and reproducibility of the ratios as well as the quality of boron purification in the samples. It assumes linear mass bias between the NIST SRM 951 reference material and the sample. It is also subject to matrix-induced effects of the sample

(Barth, 1993; Holcomb et al., 2015). Therefore, it is important that the samples are purified using the Amberlite resin ion exchange procedure previous to running isotopic analysis. Also to mitigate these mass bias effects, all standards are made up in the same matrix as the sample (0.1

M HCl) as well as the spray chamber is coated with 0.1 M HCl before introducing each sample.

A limiting factor to the MC-ICP-MS method is the large memory or blank effects associated with δ11B analysis, requiring prolonged washout to avoid cross-contamination (Aggarwal et al.,

2003; Vanderpool et al., 1994). A Triton X-100 solution (Sigma-Aldrich) in 3% HNO3 was used to help washout boron between samples. This is an organic detergent that helps to solubilize boron and remove it from the spray chamber of the MC-ICP-MS. Another limiting factor is that, compared to N-TIMS analysis, this technique requires greater sample throughput of approximately 100µL/min of sample, with 500 ng/mL B, each run compared to just a few µL’s of sample used for each run with N-TIMS. An advantage of this method is the isotope measurements have a precision of 2‰ or better, calculated by replicate analysis of standards and

samples. Another advantage is analysis of a single sample takes about 5-7 minutes including the washout time, whereas the N-TIMS analysis takes approximately 20-30 minutes per sample.

Discharge and Water Sources of Southern Alberta Rivers

4.1 Discharge Data

4.1.1 Introduction

The discharge or flow data of the studied rivers were compiled using the Water Survey of

Canada’s archived hydrometric online database (Environment Canada, 2016b). The purpose of compiling this discharge data was to observe changes in flow with downstream distance and between recent years in each watershed. The available flow data from 2013-2015 was also used to calculate nutrient and boron loads during the 2014/2015 sampling campaign (Chapter Six).

The Little Bow River was the only natural tributary with archived hydrometric data in the OMR,

Haney and Battersea Drain being irrigation canals are not monitored for flow.

4.1.2 River Discharge Data

The monthly average discharge (m3/s) of the Bow River from Lake Louise to the mouth of the BRB for 2013 and 2014 is plotted in Figure 4.1 (calculated from daily discharge rates). Peak- flow at all sampling locations occurs during June and base-flow from October through March.

The maximum average discharge rate measured in 2013 was 664.4 m3/s in June at the mouth of the Bow River and the minimum rate was 4.3 m3/s in October at the Lake Louise water gauge station (Lake Louise station only has flow data recorded from May through October each year).

In 2014 the maximum monthly average discharge rate was 349.5 m3/s in June at the mouth and

the minimum rate was 6.9 m3/s in October at Lake Louise. The lowest flow rates were observed at Lake Louise and flow rates increase as tributaries progressively contribute to riverine flow in the Bow River further downstream to maximum rates at the eastern end of the Bow River, either at Bassano Dam or the mouth of the BRB depending on the month. During the growing season

(May to September) the flow at the mouth is lower than at Bassano Dam due to increased water allocation for irrigation of crops. In 2013 there were overall greater discharge rates than in 2014 and extreme discharge rates in June caused a major flooding event. This was due to late snowmelt and higher than average precipitation rates (Pomeroy et al., 2016). In

August/September 2013 and July/August 2014 the discharge at Calgary was higher than at the downstream sites. This is due to water diversion in the irrigation districts where water is lost due to seepage, evaporation, and evapotranspiration.

700 LAKE LOUISE a) 2013 600 BANFF COCHRANE 500

/s) CALGARY 3 BASSANO DAM 400 MOUTH 300

Dsicharge (m Dsicharge 200

100

0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

400 b) 2014 350 LAKE LOUISE BANFF 300 CALGARY /s) 3 250 BASSANO DAM 200 MOUTH 150

Dsicharge (m Dsicharge 100 50 0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Figure 4.1: The monthly average a) 2013 and b) 2014 discharge data for the Bow River at

Environment Canada flow monitoring stations (Environment Canada, 2016b). The

historical discharge data for Cochrane is only available until 2013.

In the studied area of the OMR, flow rates also increase with downstream distance with slightly higher rates at the mouth of the OMR than at Lethbridge. Figure 4.2 plots the discharge rates at Lethbridge, the mouth of the OMR, and the mouth of the Little Bow River for 2013-

2015. Downstream of Lethbridge natural tributaries and man-made irrigation return-flow drains flow into the OMR and increase the flow rates. Peak-flow for the Oldman River sampling locations occurred in June and base-flow between October through March. Flow in the Little

Bow River tributary peaked in June in 2013 and 2014, but occurred in July in 2015 while base flow data was not available. The highest flow rates in the study area of the OMR occurred in

2014, with maximum flow of 588.5 m3/s in June at the mouth of the Oldman River and minimum flow of 18.2 m3/s in February at Lethbridge. Flow rates in 2013 were intermediate with maximum flow of 371.7 m3/s in June at the mouth of the OMR and 23.4 m3/s in February at

Lethbridge. 2015 had low flow rates, with the maximum flow of 117.1 m3/s in June at the mouth and 15 m3/s in December at Lethbridge. Discharge at the mouth of the Little Bow River ranges from 2.5 to 8.7 m3/s for the months of May to October.

A) B) 700 700 Lethbridge Mouth 600 600 2013 2013 500 2014 500 2014 /s) /s) 3 2015 3 2015 400 400

300 300 Discharge (m Discharge 200 (m Discharge 200

100 100

0 0 … M Jul Jan Jun Jul Oct Feb Sep Apr Jan Dec Jun Mar Oct Aug Nov Feb Sep Apr May Dec Mar Aug Nov C) 10 Little Bow River 9 2013 8 2014 7 /s)

3 2015 6 5 4 3 Discharge (m Discharge 2 1 0 Jul Jan Jun Oct Feb Sep Apr Dec Mar Aug Nov May

Figure 4.2: Historical discharge data for the years 2013-2015 for A) the OMR at

Lethbridge, B) the mouth of the OMR, and C) mouth of the Little Bow River, collected at

Environment Canada flow monitoring stations (Environment Canada, 2016b).

4.2 δ2H and δ18O of water

4.2.1 Introduction

The Bow and Oldman Rivers originate in the Canadian Rockies and are fed throughout the year by contributions from glacial melt-water, snow, rainwater, and groundwater (AMEC,

2009; BRBC, 2005). Precipitation is the source for both surface waters and groundwater in

Alberta and is formed from condensation of water vapor moving inland driven by three major airmasses; the dominant airmass from the Northern Pacific Ocean, the continental polar air mass, and the tropical air masses from the Gulf of Mexico (Peng et al., 2004). Due to the Rayleigh distillation process this water vapor becomes increasingly depleted in the heavy isotopes of 2H and 18O during every rainfall as it moves inland and across the Rocky Mountains following an equilibrium isotope fractionation process (Appelo & Postma, 2005). The cooling of air during transport to higher latitudes or altitudes drives these rainouts and the isotope enrichment factor of the produced rain or snow can be calculated (Clark & Fritz, 1997). The Local Meteoric Water

Line (LMWL) for Calgary defines the relationship between δ2H and δ18O values of precipitation in southern Alberta (Peng et al., 2004):

δ2H = 7.67*δ18O – 0.21 (1)

This slope equation of the LMWL for Calgary is unique because the δ2H and δ18O values of precipitation in Calgary are lower than other locations at similar latitudes. This relates back to the Pacific airmasses flowing over the Rocky Mountains heavily depleting water vapor in 2H and

18O. Surface water isotopic compositions in the headwaters of the BRB and OMR are governed

by source waters with rather negative δ2H and δ18O values but are also affected by evaporation that is accompanied by kinetic isotope fractionation that enriches water in 18O more than 2H

(Appelo and Postma, 2005). This leads to δ2H and δ18O relationships that define a Local

Evaporation Line (LEL) (Katvala, 2008):

δ2H= 4.3*δ18O- 60.2 (2) and a Shallow Groundwater Line (SGWL) for Alberta (Cheung, 2009):

δ2H= 6.3*δ18O- 30.1 (3)

When these equations are plotted as lines against the water isotope data of the sampling sites along the BRB and OMR, predictions can be made on the main sources of water along these watersheds. Katvala (2008) conducted in-depth analysis of isotope hydrology of the BRB and found that the headwaters were fed mainly by winter precipitation (November to April) via subsequent snowmelt with rather negative δ2H and δ18O values. In the downstream regions there is an influx of water from foothills and prairie tributaries that causes an increase in δ2H and δ18O values. Chao (2011) also studied the hydrology and isotopic composition of BRB water and in addition to these conclusions determined that isotopic fractionation caused by evaporation during summer months (June to October) lowered the slope of the δ2H vs. δ18O regression line of surface waters. In a study of the OMR isotope hydrology by Rock & Mayer (2007), a significant increase in δ2H and δ18O values in the eastern portion of the mainstem and tributaries of the

OMR downstream of the Oldman Reservoir (part of this study) was found to be partly due to surface water and groundwater influx with higher isotopic values and partly due to evaporation.

Seasonal variation in isotopic composition of upstream river water was low, with hardly any

negative excursion in δ18O during peak flows in June, indicating the dominant source of runoff is well-mixed groundwater rather than snow melt from the previous winter. This is referred to as the ‘old-water paradox’ (Rock & Mayer, 2007).

The δ2H and δ18O values of surface water obtained from sampling sites along the BRB and OMR were measured during peak-flow and base-flow periods of 2015 and were additionally analyzed in the BRB in June 2014. The isotope composition of WWTP effluents along the BRB were also measured in February and October of 2014, and the water isotopic composition of

Lethbridge WWTP effluent discharging into the OMR was measured only in October 2014. The isotope hydrology of these watersheds has been monitored extensively within the last 10 years and the sources of water have been well determined. For the purposes of this study only a few sample sets were analyzed to confirm that past conclusions on the major surface water inputs are also valid for 2014/2015 and to obtain information on the H and O isotopic composition of water that may influence the isotopic compositions of nutrients such as nitrate.

4.2.2 Results

4.2.2.1 Bow River

The δ2H and δ18O values for each water-sampling site are compiled in Table 4.1. In the

Bow River δ2H and δ18O values of surface waters ranged from -151 to -130‰ and -20.4 to

-16.6‰, respectively. The lowest δ2H and δ18O values occurred at the upstream site near Lake

Louise, AB. Surface waters displayed a trend of slightly increasing δ2H and δ18O values with

downstream distance towards the mouth of the Bow River at Ronalane (Figure 4.3). The highest

δ2H and δ18O values were found below the Fish Creek WWTP in Calgary, with values of δ2H -

130‰ and δ18O -17.1‰ in June 2015, followed closely by δ2H of -131‰ and δ18O of -17‰ below Pine Creek in September 2015. This distinct increase in δ2H (up to 17‰ increase) and

δ18O (up to 2.3‰ increase) values at 220 and 228 km (Figure 4.3) occurred below the WWTPs but the measured δ2H and δ18O values in the river were higher than the corresponding WWTP effluent water isotope signatures, although they still fall along the Local Meteoric Water Line

(LMWL) for Calgary (Figure 4.6). This difference in δ2H and δ18O values between the river and the WWTP must have been due to the WWTP effluent samples collected and measured in

February/October while the river samples were mainly collected and measured in June. The isotopic composition of precipitation is typically less depleted in 2H and 18O during warmer months because higher temperatures result in less rainout (Friedman et al., 1992; Nativ &

Riggio, 1990). The greater change in the δ2H values compared to δ18O is due to the atomic mass of hydrogen being much lighter than oxygen resulting in greater mass difference between isotopes and thus greater change in δ-values with each shift in 2H/1H ratios compared to 18O/16O.

In Figure 4.6, indicated by the red shaded box, there were excursions of the river water water isotope values away from the LMWL towards the Local Evaporation Line (LEL), occurring during the September 2015 sampling campaign in Reach 3 including the water sampling sites at

Bow City, Scandia, and Ronalane. Many sampling locations in the downstream reaches are situated in open areas and susceptible to evaporation during periods of higher temperatures, as

well as receiving agricultural return flows from the downstream region entering the Bow River that have been affected by evaporation.

The δ2H and δ18O values of the WWTP effluents discharging into the BRB range from

-152 to -142‰ and -20.2 to -18.5‰, respectively. These values generally fall within the range of isotope ratios measured in the river around the location of each WWTP (Figure 4.5). Lake

Louise and Banff WWTP effluent were slightly more 2H and 18O depleted than the average river isotope values in the area but this difference is due to the effluent samples collected during winter months (February and October) while the river samples were dominantly measured on

June samples (Figure 4.5). When plotted against the LMWL, the LEL, and the Shallow

Groundwater Line (SGWL) in Figure 4.6, the WWTP effluents water isotopic compositions fall along the LMWL for Calgary, and the Lake Louise and Banff WWTP signatures plot closely to peak-flow (June) upstream BRB values.

2 18 Table 4.1: Summary of δ HH2O and δ OH2O values in per mil (‰) measured in river water and WWTP final effluent between June 2014 and October 2015 along the Bow and Oldman

Rivers.

Jun-14 Jun-15 Sep-15 2 18 2 18 2 18 Distance Sample δ HH2O δ OH2O δ HH2O δ OH2O δ HH2O δ OH2O Sample ID (km) Type (‰) (‰) (‰) (‰) (‰) (‰) BOW RIVER Mainstem Ronalane -146 -19.1 571 River -137 -18 -131 -16.63 Mainstem Scandia -146 -19.0 513 River -143 -19 -134 -17.15 Mainstem Bow City -145 -19.0 482 River -140 -18 -135 -17.45 Mainstem Carseland -147 -19.4 302 River -144 -20 -142 -18.66 Below Pine Mainstem -147 -19.3 Creek 228 River -146 -20 -131 -17.01 Below Fish Mainstem -144 -18.9 Creek 220 River -130 -17 -143 -18.92 Below Mainstem -149.0 -19.6 Bonnybrook 205 River -147 -19 -143 -18.7 Mainstem Cochrane 179 River -151 -19.8 -149 -20 -144 -18.82 Canmore- Mainstem above 79 River -150 -19.7 -146 -20 -144 -19.12 Mainstem Banff- below 64 River - - -151 -20 -145 -19.19 Mainstem Banff- above 56 River - - -151 -20 -145 -19.22 Lake Louise- Mainstem above 0 River - - -151 -20 -145 -19.31

Jun-15 Oct-15 2 18 2 18 Distance Sample δ HH2O δ OH2O δ HH2O δ OH2O Sample ID (km) Type (‰) (‰) (‰) (‰) OLDMAN

RIVER Mainstem Taber 140 River - - -129 -17 -127.56 -16.4 D/S Mainstem Tributaries 113 River - - -130 -17 -128.1 -16.5 Little Bow Tributary River 110 - - -124 -15.6 -128.7 -15.8 Battersea Tributary Drain 98 - - -118 -14.6 -121.92 -15.1

Haney Drain 96 Tributary - - -123 -14.6 -124.09 -15.3 Mainstem Pavan Park 52 River - - -131 -17.1 -128.14 -16.7 Below Belly Mainstem River 16 River - - -134 -17.4 -128.3 -16.6 Mainstem Monarch 0 River - - -134 -17.5 -129.89 -16.9 Feb-14 Oct-15 2 18 2 18 Distance Sample δ HH2O δ OH2O δ HH2O δ OH2O WWTP (km) Type (‰) (‰) (‰) (‰) Final Pine Creek -143 -18.6 -142 -18.5 225 Effluent - - Final Fish Creek -143 -18.6 -143 -18.6 219 Effluent - - Final Bonnybrook -145 -19.0 -145 -18.6 204 Effluent - - Final Canmore -147 -19.3 -149 -18.6 80 Effluent - - Final Banff -153 -20.2 -152 -19.5 63 Effluent - - Final Lake Louise -152 -20.2 -152 -19.8 3 Effluent - - Final Lethbridge 42 -130 -16.2 Effluent - - - -

δ2H June 2015 δ2H Sept 2015 δ2H WWTP -120 δ18O June 2015 δ18O Sept 2015 δ18O WWTP -11 -130 -13 -140 -15 (‰) ‰) -150 -17 H2O H2O( -19 O 18 H 2 δ

δ -160 -21 -170 -23 LL BF CM CR Calg. CS BC SD RL -180 -25 0 100 200 300 400 500 600

Distance from a point above Lake Louise (km)

Figure 4.3: δ2H (red) and δ18O (blue) values of river and WWTP sites with downstream distance of the Bow River. The river sites are plotted for both June (peak-flow) and

September (beginning of base-flow) 2015. The points indicate sampling sites in Lake Louise

(LL), Banff (BF), Canmore (CM), Cochrane (CR), Calgary (Calg.), Carseland (CS), Bow

City (BC), Scandia (SD), and Ronalane (RL).

4.2.2.2 Oldman River Basin

In the mainstem of the OMR, δ2H values ranged from -134 to -128‰ and δ18O values from -17.5 to -16.4‰ from Monarch to downstream near Taber (Table 4.1, Figure 4.4). The total distance sampled was not large enough to show marked large-scale isotope trends with downstream distance in the OMR basin. The δ2H and δ18O values plot along the LMWL for

Calgary (Figure 4.6). The water isotope signature of the WWTP effluent at Lethbridge was only measured once and this isotopic signature is within 2‰ of either δ2H value and within 1‰ of either δ18O value measured in the river below the WWTP (Figure 4.5).

The sampled tributaries have δ2H values ranging from -129 to -118‰ and δ18O values of

-15.8 to -14.6‰, hence the isotopic compositions of the tributary water are significantly higher than that of water from the OMR main stem sites (Figure 4.5). Furthermore, the δ2H and δ18O values of the irrigation return-flow Haney and Battersea Drains have 1 to 6‰ higher δ2H values and 0.5 to 1‰ higher δ18O values than those of the Little Bow River, a natural tributary (Figure

4.5). Prairie tributaries are partially sourced from groundwater that results in water less depleted in 2H and 18O compared to water sourced from higher altitude mountain snow melt. The

Battersea and Haney irrigation drains receive much less flow than natural streams such as the

Little Bow River and therefore the groundwater content will be higher within the irrigation canals. Also, these irrigation drains have low flow and are more susceptible to evaporation during warmer months. This trend is clearly shown in Figure 4.6, indicated by the red shaded box, where all the measured tributary δ2H and δ18O values from both June and October 2015 plot between the LMWL, Shallow Groundwater Line (SGWL), and Local Evaporation Line (LEL) indicating a mixing of groundwater with some surface evaporation in the tributaries. The increases in δ2H and δ18O values measured from up- to downstream of the tributaries in the mainstem of the OMR are within the ±1.0‰ and ±0.1‰ accuracy of the measurement technique, respectively and are therefore not significant (Figures 4.4, 4.5). This indicates these tributaries may not be a significant water source to the OMR.

δ2H June 2015 δ2H Oct 2015 δ2H WWTP δ18O June 2015 δ18O Oct 2015 δ18O WWTP -16.0

-120 -16.2 -16.4 -16.6 -125 -16.8 (‰) ‰) ( -17.0 H2O O H2O 18

H -130 2 -17.2 δ δ -17.4 -135 -17.6 -17.8 MN BL LB Tributaries DS TB -140 -18.0 0 50 100 150 Distance from Monarch sampling site (km)

Figure 4.4: δ2H and δ18O values of the mainstem river sites and the Lethbridge WWTP along the OMR. The river sites are plotted for June (peak-flow) and October (base-flow) of

2015. Points refer to sampling sites near Monarch (MN), Below Belly River (BL),

Lethbridge (LB), Downstream Tributaries (DS), and Taber (TB).

WWTP Upstream BRB Downstream BRB -136 -138 a) -140 -142 FC

(‰) -144 PC H 2 -146 BB δ -148 CM -150 -152 LL BF -154 -21 -20 -19 -18 -17 18 δ O (‰) WWTP Upstream OMR Downstream OMR Tributaries -115 -117 b) -119 Battersea Drain -121

(‰) -123 Haney Drain H

2 -125 Little Bow δ -127 -129 -131 -133 -18 -17 -16 -15 -14 δ18O (‰)

Figure 4.5: δ2H versus δ18O values comparing a) the BRB sites upstream and downstream of Calgary to the WWTP effluents, b) OMR sites upstream and downstream of the tributaries, the tributary values and the WWTP effluents.

Bow River- June 2014 Bow River- June 2015 Bow River - Sept 2015 WWTP effluent Mean PPT Calgary LMWL Calgary LEL SGWL -120 -125 a) -130 -135

H (‰) H -140 2 δ -145 LMWL: δ2H = 7.68δ18O - 0.21 -150 SGWL: δ2H = 6.3δ18O - 30.1 -155 LEL: δ2H = 4.3δ18O - 62 -160 -22 -21 -20 -19 -18 -17 -16 δ18O (‰)

Oldman River - June 2015 OMR Tributaries- June 2015 Oldman River- Oct 2015 OMR Tributaries- Oct 2015 WWTP effluent Mean PPT Calgary LMWL Calgary LEL SGWL -110 -115 b) -120 -125 -130 H (‰) H

2 -135 δ -140 -145 LMWL: δ2H = 7.68δ18O - 0.21 -150 SGWL: δ2H = 6.3δ18O - 30.1 -155 LEL: δ2H = 4.3δ18O - 62 -160 -20 -18 -16 -14 -12 -10 18 δ O (‰)

2 18 Figure 4.6: δ HH2O versus δ OH2O values of surface water sampling sites for a) the Bow

River sampling sites and from each WWTP effluent along the BRB and b) the Oldman

River, the OMR tributaries and the WWTP effluent from Lethbridge. Also shown are the

Local Meteoric Water Line for Calgary (LMWL Calgary) from Peng et al. (2004), the

Local Evaporation Line (LEL) (Katvala, 2008), and the Shallow Groundwater Line

(SGWL) (Cheung, 2009).

Ion Chemistry of Southern Alberta Rivers

5.1 Introduction

In this section, major ion chemistry and the NO3 and B concentrations of surface water samples from various sites along the Bow and Oldman Rivers are discussed in terms of spatial and temporal variability. To assess temporal variability, June 2014, September 2014, June 2015, and September 2015 sampling events are used for the BRB and October 2014, June 2015, and

October 2015 sampling events are used for the OMR. Tables 5.1-5.4 compile the major ion concentrations of surface water samples from the Bow River and OMR during the 2014-2015

- - 2- - sampling campaign. The major anions include HCO3 , Cl , SO4 and NO3 , and major cations

4+ + + 2+ 2+ include Si (as silicic acid, H4SiO4, in aquatic systems), K , Na , Ca , and Mg . Interpretation of the chemical composition provides information on geochemical, biological, and physical processes within the watersheds. Discussion on the average, maximum and minimum concentrations of the major ions gives insight into the natural and anthropogenic sources of the solutes in these watersheds. The major ion chemistry of the Bow River was previously discussed by Chao (2011) and that of the OMR by Rock (2005). Therefore, the ion chemistry of the surface waters from these rivers sampled in 2014 and 2015 will be briefly covered to classify the water types and determine distinct trends in chemical composition. NO3 and B concentrations are each reviewed separately to understand their spatial and temporal variability as nutrient tracers and begin to give insight to their fluxes into southern Alberta rivers.

5.2 Major Ion Chemistry

5.2.1 The Bow River

The major cations in surface water samples obtained from the Bow River occurred in order from highest to lowest concentrations Ca2+>Mg2+>Na+>K+>Si4+ and major anion

- 2- - - concentrations in the order HCO3 >SO4 >Cl >NO3 . A Piper plot was used to determine the major water type along the Bow River (Figure 5.1) revealing a Ca-Mg-HCO3 dominant water type for all three reaches of the BRB. This is a typical water type for freshwater systems. The three reaches of the river had slightly different compositions from upstream in the mountains and foothills (Reach 1), through Calgary (Reach 2), to downstream in the prairies (Reach 3). The

2+ - proportion of major cations and anions made up of Ca and HCO3 , respectively, were greatest in Reach 1 where the greatest influence on surface water composition is the weathering of local carbonate-rich limestones and shales (from 78-82% of TDS in Reach 1 compared to 68-76% in

Reach 2 and 67-77% in Reach 3). From Reach 1 (147.6 to 281.9 mg/L), there was an overall increase in TDS in Reach 2 and Reach 3 to 252.1-561.8 mg/L (Table 5.1 & 5.2). This is due to input of urban runoff and agricultural return-flow waters, in turn changing water composition

+ 2- - (Figure 5.2). There were increases in the concentrations of Na , SO4 , and Cl with downstream distance, resulting in a shift towards greater percent contributions of these major ions in the Piper plot (Table 5.1 & 5.2, Fig. 5.2). The increase in the non-dominant Na and Cl ions may in part be due to anthropogenic inputs (further investigated in Chapter 7), but increases in Na and SO4 are also likely due to water-rock interaction through input by shallow groundwater. The thick

surficial deposits of till and lacustrine material in the eastern part of Southern Alberta are salt- rich and contribute to the higher Na+ content of the water. The Na+ is released through ion exchange on montmorillinite clay with the Ca2+ and Mg2+ (Nielson, 1971; Rock, 2005). Na-rich groundwater (of either Na-HCO3 or Na-SO4 water type) has also been characterized throughout

Alberta (Humez et al., 2016). The increase in Na+ concentration from Reach 1 to Reach 3 suggests that interaction with surficial deposits and influx of groundwater may be influencing

2- this trend (Fig. 5.2). The trend of increasing SO4 content in Reach 3 can also be attributed to the properties of till, when oxidation of reduced S-species, such as pyrite, occurs (Rock, 2005).

There are three identified S-species in the glacial till of Southern Alberta: sulphate, pyrite or marcassite, and organo-sulfur compounds (Hendry et al., 1986). Cl- markedly increased in concentration in Reach 2 through Calgary and remained elevated in concentration into Reach 3 and thus anthropogenic input was assumed to be the major source (Table 5.1 & 5.2). The influence of atmospheric deposition on major ion concentrations was investigated by Grasby &

+ 2- - Hutcheon (2000) and it was found that approximately 50% of K , 17% of SO4 , and 16% of Cl can be attributed to atmospheric loading at the headwaters of the Bow River. There were overlapping water compositions in Figure 5.1 during peak-flow and base-flow sampling periods on the Bow River and thus the seasonal variation of the sampling sites was low.

The TDS values ranged from a minimum of 147.6 mg/L above Lake Louise in September

2015 to a maximum of 561.8 mg/L below Fish Creek in June 2015 (Figure 5.2). The overall mean TDS value of the Bow River is 268.4 ± 44.9 mg/L in 2014 and 287.5 ± 100.4 mg/L in

2015. As seen in Figure 5.2, the chemical composition at each of the Bow River sites during each

sampling period between 2014 and 2015 was quite similar with the exception of some significantly higher TDS concentrations below Fish Creek in 2015, a TDSmax/TDSmin of 202% at this site. In both 2014 and 2015 over 75% of surface water sites had a 2 to 2.5 times increase in

TDS values during the baseflow period in September compared to peaklow in June. This was due to lower flow rates causing less dilution of solutes during baseflow (see Section 4.2.1 “Discharge

Data”).

Table 5.1: Major cation, anion, and TDS data for the surface water samples from the Bow

River 2014 sampling period. Bracketed numbers denote the river “reach”. “NA” indicates the site was not sampled during that specific campaign.

2014 Anion Concentrations (mg/L) - - 2- - Location HCO3 Cl SO4 NO3 TDS (mg/L) June Sept June Sept June Sept June Sept June Sept [3] Ronalane 173.2 167.4 6.9 9.2 46.5 54.7 1.7 2.3 305.3 311.9 [3] Scandia 174.5 168.4 6.9 9.4 46.7 58.6 1.5 2.4 303.6 319.7 [3] Bow City 130.0 164.8 7.0 9.2 42.8 53.8 1.6 1.7 237.7 305.1 [3] Carseland 162.2 166.4 5.5 8.8 32.9 51.4 1.5 3.2 268.9 304.2 [2] Below Pine Creek 164.9 155.2 8.3 13.4 34.3 42.6 1.8 7.9 278.2 295.2 [2] Below Fish Creek 195.3 157.8 18.9 7.6 28.6 39.9 1.3 2.8 330.6 277.3 [2] Below Bonnybrook 164.4 165.6 6.1 10.6 33.8 41.4 2.1 3.9 273.0 296.3 [1] Cochrane 166.5 150.6 2.3 1.9 28.3 31.3 0.5 1.0 257.6 246.2 [1] Canmore- above 140.0 137.5 0.4 0.4 27.4 33.3 0.2 0.8 219.3 227.1 [1] Banff- below - 130.6 - 1.3 - 30.6 - 0.9 - 216.2 [1] Banff- above - 130.6 - 1.4 - 26.8 - 0.9 - 209.0 Lake Louise- above[1] - 100.3 - 0.4 - 15.1 - 0.9 - 154.2 Cation Concentrations (mg/L) Location Si4+ K+ Na+ Ca2+ Mg2+

June Sept June Sept June Sept June Sept June Sept [3] Ronalane 1.4 0.4 1.5 1.9 12.3 11.9 47.8 48.1 14.0 16.0 [3] Scandia 1.1 0.5 1.5 2.0 13.8 14.6 44.4 47.1 13.2 16.6 Bow City[3] 0.9 0.5 1.1 1.8 8.7 12.3 35.4 45.2 10.1 15.9 [3] Carseland 1.3 1.2 1.1 1.7 6.9 8.3 45.5 48.0 12.1 15.1 [2] Below Pine Creek 1.0 1.7 1.3 2.6 7.8 10.2 45.7 46.2 13.2 15.2 [2] Below Fish Creek 1.4 1.5 1.9 1.3 15.4 6.0 52.2 45.6 15.5 14.8 [2] Below Bonnybrook 1.1 1.4 1.3 1.7 6.6 8.5 45.1 47.4 12.6 15.6 [1] Cochrane 1.4 1.7 0.6 0.8 3.0 3.3 43.3 42.3 11.6 13.2 [1] Canmore- above 1.1 1.4 0.4 0.4 0.9 0.8 39.5 42.0 9.4 10.7 [1] Banff- below - 1.6 - 0.4 - 1.4 - 37.1 - 12.3 [1] Banff- above - 1.5 - 0.4 - 1.4 - 33.9 - 12.1 Lake Louise- above[1] - 1.2 - 0.3 - 0.7 - 25.3 - 9.8

Table 5.2: Major cation, anion and TDS data for the surface water samples from the Bow

River for 2015 sampling period. Bracketed numbers denote the river “reach”.

2015 Anion Concentrations (mg/L) - - 2- - Location HCO3 Cl SO4 NO3 TDS (mg/L) June Sept June Sept June Sept June Sept June Sept [3] Ronalane 159.7 177.6 5.8 15.1 44.4 66.0 0.8 1.9 275.7 346.9 [3] Scandia 149.2 185.0 6.6 15.8 44.3 104.2 1.3 2.1 267.5 414.1 [3] Bow City 184.1 160.1 5.9 15.8 63.4 104.7 0.2 1.0 338.1 382.5 [3] Carseland 170.6 178.4 7.1 32.3 36.4 26.2 2.4 2.1 286.6 319.5 [2] Below Pine Creek 179.1 177.4 7.4 13.7 19.0 43.5 2.8 3.5 277.0 314.9 [2] Below Fish Creek 342.2 317.3 47.6 31.9 30.5 25.6 1.0 2.1 561.8 502.1 [2] Below Bonnybrook 140.3 132.9 8.1 14.3 37.3 47.3 3.0 4.2 252.1 270.1 [1] Cochrane 160.9 160.8 0.7 12.9 14.7 36.6 0.4 3.2 229.2 281.9 [1] Canmore- above 158.6 131.8 0.3 0.6 26.2 35.6 0.3 0.7 243.7 222.9 [1] Banff- below 130.9 131.7 0.9 1.0 24.7 29.8 0.5 0.8 206.7 213.1 [1] Banff- above 120.1 126.8 1.0 1.1 20.8 27.2 0.7 0.8 185.6 202.8 [1] Lake Louise- above 106.1 99.9 0.9 0.6 13.2 10.8 0.7 0.7 157.7 147.6 Cation Concentrations (mg/L) Location Si4+ K+ Na+ Ca2+ Mg2+

June Sept June Sept June Sept June Sept June Sept [3] Ronalane 0.8 0.4 1.4 2.5 10.2 21.6 37.6 43.8 14.8 17.7 [3] Scandia 0.3 0.2 1.2 2.8 12.0 39.2 37.9 43.0 13.0 21.4 [3] Bow City 1.8 0.2 2.8 2.6 17.7 36.1 45.5 42.3 15.2 19.5 [3] Carseland 0.6 0.1 1.4 2.2 8.1 15.5 43.6 45.7 13.3 16.7 [2] Below Pine Creek 1.2 0.6 1.7 1.8 10.5 13.1 40.9 45.0 13.2 16.0 [2] Below Fish Creek 1.8 1.7 3.2 3.0 44.5 30.8 58.3 61.7 31.5 27.6 [2] Below Bonnybrook 1.0 0.6 1.3 2.1 6.9 13.7 38.0 39.7 12.3 15.0 [1] Cochrane 1.4 1.6 0.5 0.9 1.5 7.3 36.9 43.1 11.4 15.3 [1] Canmore- above 1.4 1.5 0.5 0.4 0.5 0.8 44.7 40.1 11.1 11.3 [1] Banff- below 1.6 1.6 0.7 0.4 1.3 1.3 34.4 34.5 11.7 12.0 [1] Banff- above 1.5 1.6 0.4 0.4 1.2 1.3 29.1 32.1 10.7 11.5 [1] Lake Louise- above 1.2 1.2 0.3 0.3 0.8 0.9 25.0 24.1 9.4 9.1

Figure 5.1: Piper plot conveying the major water types of Reach 1, 2, and 3 of the Bow

River Basin based on % meq/L.

Jun-14 Sep-14 Jun-15 Sep-15 600

500

400

(mg/L) 300 TDS 200

100

LL BF CM CR Calg. CS BC SD RL 0 0 100 200 300 400 500 600 Distance from a point above Lake Louise (km)

Figure 5.2: TDS of surface water samples versus distance along the Bow River for each sampling event.

5.2.2 The Oldman River & Tributaries

5.2.2.1 The Oldman Mainstem

The major cations in surface water samples obtained from the OMR mainstem sampling sites occur from highest to lowest concentrations in the order Ca2+>Mg2+>Na+>K+>Si4+ and

- 2- - - anion concentrations in the order HCO3 >SO4 >Cl >NO3 . A Piper plot was used to determine the major water types of Reach 1, Reach 2 and of the tributaries of the OMR (Figure 5.3). Reach

1 and Reach 2 had a Ca-Mg-HCO3 dominant water type and this water type was consistent during both the peakflow and baseflow sampling periods. However, there was an approximately

- 2- - 10% increase in the Cl + SO4 contribution from Reach 1 to Reach 2. An increase in Cl in

Reach 2 may be due to anthropogenic inputs, such as run-off containing road salts, or chlorinated

2- drinking water, downstream of Lethbridge. The presence of SO4 was due to the same processes occurring in the eastern part of the BRB, the till and lacustrine sediment geochemistry and groundwater chemistry influencing water composition through water-rock interactions.

Atmospheric loading of major ions into the Oldman River was assumed to have similar contributions as to the Bow River (Grasby & Hutcheon, 2000).

The TDS in the OMR ranged from 270.2 mg/L below the Belly River confluence in June

2015 to 347.5 mg/L at Taber in October 2014. The mean TDS of the OMR was 332.3 ± 13.8 mg/L in 2014 and 309.7 ± 16.5 mg/L in 2015. There were minor fluctuations in TDS between upstream and downstream locations but no clear trend was observed (Figure 5.4). At each mainstem sampling site there were minor variations in TDS of seasonally sampled surface water, the greatest being a difference of TDSmax/TDSmin of 116 % between 2015 peakflow and 2014 baseflow below Belly River.

5.2.2.2 Tributaries

As shown in Fig. 5.2 and further detailed in Tables 5.3 and 5.4, surface water from the

Little Bow River followed the same order of major ion concentrations as the OMR sites, but the

Haney and Battersea Drain had distinctly different concentration orders and water compositions

throughout the peak- and base-flow periods. On the Piper plot, the Little Bow River plotted as the same water type as the OMR, a Ca-Mg- HCO3 dominated system, but the water type is

+ + 2- shifted towards greater proportions of Na + K and SO4 . For the same reasons as in Reach 3 of the Bow River, this indicates there was more groundwater influx and till and lacustrine geochemistry influence than in the OMR mainstem (Figure 5.2). Increased K+ concentrations are sourced from silicate dissolution from geological units of the prairies (Hamilton et al., 1999).

The TDS of the Little Bow River had a mean value of 428 ± 45.9 mg/L, slightly increased compared to the TDS ranges of the OMR. The TDS values did not vary much between peakflow and baseflow as seen at km 110 in Figure 5.4. The water chemistry of the Battersea Drain overlapped with the water composition of the OMR during peakflow in June but during baseflow diverts towards a more Ca-SO4 dominated system, caused by groundwater influenced by pyrite oxidation (Figure 5.2). Irrigation canals are fed by the mainstem of the OMR and therefore will have similar chemical composition. The TDS values ranged from 306.2 mg/L in June 2015 to

2413.5 mg/L in October 2014. The large range in TDS between June and October indicates that during peakflow at the beginning of irrigation season, the increased discharge originating from mountain snowpack melt dilutes solutes in surface water. During baseflow, once the irrigation canals are drained, there is a major reduction in the volume of water flowing through the canals and a much stronger influence by groundwater and agricultural return flows. The water in the

Haney Drain had consistently a Ca-SO4 dominant water type and a mean TDS value of 1650.5 ±

460.7 mg/L, consistently the highest TDS of the sampled tributaries (Figure 5.4). There was slight variation in TDS between peakflow and baseflow and a presence of dissolved organic

carbon was measured with >10 mg/L. This could indicate an influx of agricultural runoff potentially affected by organic-rich cow manure that will be explored in Chapter Seven.

Table 5.3: 2014 major cation, anion, and TDS data for the surface waters from the 2014

Oldman River and tributary sampling sites. Tributary sites are highlighted in grey.

Bracketed numbers denote the river “reach”.

2014 Anion Concentrations (mg/L) Cation Concentrations (mg/L) TDS Location HCO - Cl- SO 2- NO - Si4+ K+ Na+ Ca2+ Mg2+ 3 4 3 (mg/L)

Oct Oct Oct Oct Oct Oct Oct Oct Oct Oct

Taber [2] 200.6 3.7 55.2 0.8 0.7 1.9 20.1 46.3 18.2 347.5 D/S Tributaries[2] ------Little Bow River [2] 221.4 11.1 117.9 0.5 1.2 4.2 45.8 51.6 24.2 477.9 Battersea Drain [2] 370.7 73.8 1249.4 74.5 1.2 11.1 173.1 255.9 202.1 2413.5 Haney Drain [2] 457.5 83.4 889.0 90.3 5.7 11.0 141.2 205.1 182.1 2065.4

Pavan Park [1] 196.0 3.6 50.4 1.1 1.3 2.1 19.0 46.4 17.7 337.6 Below Belly River [1] 194.3 1.7 38.8 0.9 1.6 1.5 14.9 44.4 16.7 314.8 Monarch [1] 206.7 1.9 35.4 1.0 1.9 1.6 15.9 48.3 16.5 329.4

Table 5.4: 2015 major cation, anion, and TDS data for the surface waters from the 2015

Oldman River sampling sites. Tributary sites are highlighted in grey. Bracketed numbers denote the river “reach”.

2015 Anion Concentrations (mg/L) - - 2- - Location HCO3 Cl SO4 NO3 TDS (mg/L) June Oct June Oct June Oct June Oct June Oct

Taber [2] 181.3 180.9 2.8 3.1 53.8 52.5 0.0 0.7 313.9 315.5

D/S Tributaries[2] 180.4 180.6 3.1 4.0 58.1 55.5 0.5 0.8 321.8 321.5

Little Bow River [2] 204.6 203.8 6.4 11.0 78.7 98.1 0.6 0.6 387.0 421.9

Battersea Drain [2] 172.2 274.9 2.4 44.8 56.1 581.8 0.7 46.4 306.2 1255.3

Haney Drain [2] 202.3 462.7 113.6 88.6 1319.3 972.8 60.0 70.3 2130.1 2519.4

Pavan Park [1] 182.2 181.0 3.6 4.0 51.7 50.7 0.6 1.0 318.9 316.2

Below Belly River [1] 169.5 181.3 1.2 2.3 33.2 35.7 1.0 1.0 270.2 293.7

Monarch [1] 181.1 188.3 1.7 4.0 43.3 45.1 1.5 0.9 303.5 321.5 Cation Concentrations (mg/L) 4+ + + 2+ 2+ Location Si K Na Ca Mg June Oct June Oct June Oct June Oct June Oct

Taber [2] 0.3 0.2 1.6 1.7 16.9 17.9 38.9 41.5 16.6 16.8

D/S Tributaries[2] 0.3 0.3 1.7 1.8 18.0 19.1 40.7 41.8 16.9 17.2

Little Bow River [2] 0.8 1.0 2.5 3.4 27.7 35.2 42.5 46.8 19.2 21.7

Battersea Drain [2] 1.0 1.9 1.3 5.5 19.2 71.3 31.8 136.8 17.6 90.7

Haney Drain [2] 2.7 5.8 10.8 10.8 187.5 310.7 83.2 217.5 135.8 378.2

Pavan Park [1] 0.8 1.0 2.0 1.8 17.0 18.0 42.0 41.6 17.0 16.9

Below Belly River [1] 1.1 1.2 1.5 1.1 10.4 12.9 35.6 42.0 13.1 15.9

Monarch [1] 1.3 1.4 1.3 1.4 14.5 17.4 42.4 45.3 15.3 17.3

Figure 5.3: Piper plot conveying the major water types of Reach 1, Reach 2, and the sampled tributaries of the Oldman River Basin based on % meq/L.

Oct-14 OMR Jun-15 OMR Oct-15 OMR 600 3000 Oct-14 Trib. Jun-15 Trib. Oct-15 Trib. 550 2500 500 2000 450 400 1500 OMR (mg/L)OMR

- 350 1000 Tributary (mg/L) Tributary

300 - TDS 500 250 TDS 200 MN BL LB Tribs. DS TB 0 0 50 100 150 Distance from Monarch (km)

Figure 5.4: TDS of surface water samples versus distance along the Oldman River (left axis) and tributaries (right axis) for each sampling event.

5.3 NO3 Concentrations

5.3.1 The Bow River

The NO3 concentrations of surface water samples obtained from each Bow River sampling site are summarized in Table 5.5. The concentration of NO3 ranged from 0.2 mg/L above Canmore in June 2014 to 7.9 mg/L below Pine Creek in September 2014. Figure 5.5 plots the NO3 concentrations with distance from Lake Louise for each sampling campaign. In Reach 1 of the Bow River above Calgary the NO3 concentrations were typically below 1.0 mg/L. There

was an increase in NO3 concentrations as the Bow River flows in Reach 2 through Calgary, to between 1.0 and 7.9 mg/L, due to anthropogenic inputs from storm runoff and wastewater. In

Reach 3 the nitrate concentrations decreased to an average of 1.5-2.5 mg/L downstream of

Calgary. This indicates the large influx of NO3 through Calgary was either being partially removed or diluted downstream. In Figure 5.6 the NO3/Cl ratio (using units of meq/L) is plotted against distance in Reach 2 and 3 during baseflow periods when solutes are more concentrated within the watershed. Chloride (Cl) is often used as a tracer of anthropogenic inputs due to its conservative properties in aquatic systems. The NO3/Cl ratios are plotted beginning downstream of Bonnybrook WWTP since this was the location where the first significant increase in NO3 concentrations was observed. Figure 5.6 displays a similar trend to Figure 5.5, where the NO3/Cl ratio was highest through Calgary and subsequently decreasing and leveled out in Reach 3. This suggests NO3 concentrations were decreasing downstream of Calgary due to N-removal processes rather than dilution. If dilution were occurring both the NO3 and Cl concentrations would be decreasing and the NO3/Cl ratio would be constant with distance. In the following chapter concentration and discharge data will be integrated to determine NO3 fluxes in the BRB.

Table 5.5: NO3 concentrations of the Bow River surface water samples during the 2014-

2015 sampling campaign. Sites with concentrations “NA” were not sampled during the specific sampling campaign.

Distance June 2014 Sept 2014 June 2015 Sept 2015 Sampling Site (km) NO3 (mg/L) NO3 (mg/L) NO3 (mg/L) NO3 (mg/L) Ronalane 571 1.7 2.3 0.8 1.9 Scandia 513 1.5 2.4 1.3 2.1 Bow City 482 1.6 1.7 0.2 1.0 Carseland 302 1.5 3.2 2.4 2.1 Below Pine Creek 228 1.8 7.9 2.8 3.5 Below Fish Creek 220 1.3 2.8 1.0 2.1 Below Bonnybrook 205 2.1 3.9 3.0 4.2 Cochrane 179 0.5 1.0 0.4 3.2 Canmore- above 79 0.2 0.8 0.3 0.7 Banff- below 64 NA 0.9 0.5 0.8 Banff- above 56 NA 0.9 0.7 0.8 Lake Louise- above 0 NA 0.9 0.7 0.7

Jun-14 Sep-14 Jun-15 Sep-15 9 LL BF CM CR Calg. CS BC SD RL 8 7 6 5 (mg/L)

3 4

NO 3 2 1 0 0 100 200 300 400 500 600 Distance from a point above Lake Louise (km)

Figure 5.5: NO3 concentrations (mg/L) in surface water samples along the Bow River with distance from the sampling point above Lake Louise.

0.40 2014 2015

0.35

0.30

0.25 /Cl 3 0.20 NO 0.15

0.10

0.05 Calg. CS BC SD RL 0.00 180 230 280 330 380 430 480 530 580 Distance from a point above Lake Louise (km)

Figure 5.6: NO3/Cl ratios (using meq/L) along the Bow River with downstream distance during baseflow (September) sampling periods of 2014 and 2015. Plot begins below

Bonnybrook WWTP (205 km) where NO3 concentration begins to increase above background concentrations.

5.3.2 The Oldman River

In the OMR, surface water NO3 concentration data were collected from five mainstem sites and three tributaries and the results are summarized in Table 5.6. The NO3 concentrations of the surface water from the mainstem sites ranged from below detection limit (NO3 detection limit

= 0.02 mg/L) at Taber in June 2015 to 1.5 mg/L at Monarch in June 2015. The NO3 concentrations along the OMR were plotted with downstream distance, including the tributary

sites, in Figure 5.7. There is a slight decrease in NO3 concentrations downstream of Monarch, from 0.9-1.5 mg/L at Monarch to n.d.-0.7 mg/L at Taber.

The three sampled tributaries: Little Bow River, Haney Drain and Battersea Drain had a large range in NO3 concentrations (Table 5.6). The Little Bow River, a natural tributary, had

NO3 levels comparable to those in the OMR, ranging from 0.5-0.6 mg/L NO3 during all three sampling campaigns. The highest NO3 concentration measured was 90.3 mg/L in the Haney

Drain in October 2014, a large contrast to 1.5 mg/L, the highest NO3 concentration measured in the mainstem of the OMR (Fig. 5.7). This tributary consistently had the highest measured NO3 concentrations during all sampling campaigns (Table 5.6). The Battersea Drain ranged from 0.7 mg/L in June 2015 during irrigation season to 74.5 mg/L NO3 in October 2014 after irrigation season. The NO3 concentrations of these man-made irrigation canals were highest after the return-flow water was released back into the canals in the fall after the irrigation season was complete (October). Although these man-made agricultural return-flow canals had extremely high NO3 concentrations this did not seem to affect the NO3 concentrations within the OMR in

Reach 2, downstream of the measured tributaries (Figure 5.7). The discharge volume from these canals was very minor in comparison to the discharge of the mainstem of the river, resulting in diluted NO3 concentrations once mixed into the OMR. The major inputs of NO3 into the mainstem of the OMR will be further explored in Chapter Six discussing NO3 fluxes.

Oct 2014 June 2015 Oct 2015 Sampling Site Km NO3 (mg/L) NO3 (mg/L) NO3 (mg/L) Taber 140 0.8 n.d. 0.7 D/S Tributaries 113 NA 0.5 0.8 Little Bow River 110 0.5 0.6 0.6 Battersea Drain 98 74.5 0.7 46.4 Haney Drain 96 90.3 55.3 69.6 Pavan Park 52 1.1 0.6 1.0 Below Belly River 16 0.9 1.0 1.0 Monarch 0 1.0 1.5 0.9

Oct-14 OMR Jun-15 OMR Oct-15 OMR Oct-14 Trib. Jun-15 Trib. Oct-15 Trib. 1.6 100 1.4 90 80 1.2 70 1.0 60 (mg/L) 0.8 50 40

OMR OMR 0.6 Tributaries (mg/L) Tributaries 3 30 - -

0.4 3 NO 20

0.2 NO MN BL LB Tribs. DS TB 10 0.0 0 0 20 40 60 80 100 120 140 160 Distance from Monarch (km)

5.4 Boron Concentrations

5.4.1 The Bow River

Boron concentrations in surface waters, with an accuracy of ± 8%, obtained from the

Bow River ranged from 1.3 µg/L above Lake Louise in September 2014 to 29.0 µg/L at Bow

City in September 2015, with data from the sampling campaign compiled in Table 5.7. Boron concentrations were between 1-6 µg/L in Reach 1, increased to between 8.4-21.5 µg/L through

Reach 2, and remain elevated downstream in Reach 3 to Bow City where B reached maximum concentrations and then decreased slightly to between 8.7-25.8 µg/L at Scandia and Ronalane

(Figure 5.8). In the September 2014 sampling campaign the boron concentrations remained most constant in Reach 3 varying by only 5 µg/L. By plotting the B/Cl ratio for surface waters of

Reach 2 and 3 in the Bow River during baseflow (Figure 5.9), it is evident that B/Cl ratios remained rather constant with distance, with most ratios between 0.012-0.020. This indicates that boron behaves similarly to chloride in the river system, which is known to remain conservative with transport in surface water systems.

June 2014 Sept 2014 June 2015 Sept 2015 Sampling Site Distance (km) B (µg/L) B (µg/L) B (µg/L) B (µg/L) Ronalane 571 8.7 17.0 12.4 20.1 Scandia 513 13.5 18.8 12.4 25.8 Bow City 482 16.7 17.9 22.6 29.0 Carseland 302 10.8 13.6 11.1 18.7 Below Pine Creek 228 3.2 18.8 11.0 21.5 Below Fish Creek 220 13.0 20.0 23.5 16.1 Below Bonnybrook 205 9.7 14.8 8.4 17.6 Cochrane 179 3.8 5.6 3.6 5.9 Canmore- above 79 4.1 3.6 3.2 3.1 Banff- below 64 NA 3.4 2.8 2.8 Banff- above 56 NA 3.3 2.5 2.6 Lake Louise- above 0 NA 1.3 1.4 1.5

Jun-14 Sep-14 Jun-15 Sep-15 40

35 LL BF CM CR Calg. CS BC SD RL 30 25 20

µ( Bg/L) 15 10 5 0 0 100 200 300 400 500 600 Distance from a point above Lake Louise (km)

Figure 5.8: B concentrations (µg/L) for surface waters obtained along the Bow River with distance from the sampling point above Lake Louise.

2014 2015 0.030

0.025

0.020

0.015 B/Cl

0.010

0.005 Calg. CS BC SD RL 0.000 200 300 400 500 600 Distance from a point above Lake Louise (km)

Figure 5.9: B/Cl ratios (using meq/L) with downstream distance during baseflow

(September) sampling periods of 2014 and 2015. Plot begins below Bonnybrook WWTP

(205 km) where B concentration begins to increase above background concentrations.

5.4.2 The Oldman River

Boron concentrations of surface water samples obtained from the OMR and the sampled tributaries are compiled in Table 5.8. Surface water samples from mainstem OMR sites ranged from 11.2 µg/L below the confluence with the Belly River in June 2015 to 17.2 µg/L at Taber in

October 2014. In Figure 5.10, the B concentrations of the samples are plotted with distance from the Monarch sampling site. Concentrations varied slightly along the mainstem sites but there was no significant trend of increasing or decreasing B concentrations with distance. In the tributary samples, B concentrations ranged from 17.4 µg/L in the Battersea Drain in June 2015 to 122.9

µg/L in the Haney Drain in October 2014. Overall, the man-made irrigation return-canal waters had much higher B concentrations than the natural Little Bow River. The elevated boron concentrations of the irrigaton return-canals did not appear to influence the B concentrations of the mainstem OMR samples. Flux calculations in Chapter Six will determine where the sources of major boron input to the OMR.

Distance Oct 2014 June 2015 Oct 2015 Sampling Site (km) B (µg/L) B (µg/L) B (µg/L) Taber 140 17.2 14.9 14.6 D/S Tributaries 113 NA 15.5 15.3 Little Bow River 110 27.3 19.2 25.2 Battersea Drain 98 121.3 17.4 67.4 Haney Drain 96 122.9 114.5 108.7 Pavan Park 52 16.4 15.6 15.4 Below Belly River 16 12.6 11.2 11.6 Monarch 0 14.2 12.8 14.5

Oct-14 OMR Jun-15 OMR Oct-15 OMR 50 Oct-14 Trib. Jun-15 Trib. Oct-15 Trib. 150 45 MN BL LB Tribs. DS TB 135 120 40 105 35 90 30 75 60 OMR µ ( g/L)OMR 25 - 45 µ( g/L) Tributary - B 20

30 B 15 15 10 0 0 20 40 60 80 100 120 140 160 Distance from Monarch (km)

NO3 and Boron Budgets for Southern Alberta Rivers

6.1 Introduction

To properly estimate the sources and sinks of NO3 and B in southern Alberta rivers, mass balance flux calculations are a useful tool because they indicate where significant masses of nutrients are input or removed from the surface water system. Mass balance estimation of nutrients in these river systems entails calculating flux rates from discharge and concentration data from surface water sampling sites along the river, comparing this to the flux rates from point sources such as WWTPs, and finally estimating the contribution by non-point sources of NO3 and B within agricultural or other areas. Based on a mass balance relationship, the flux of nutrients in a river system is equal to the sum of the flux input minus the flux output in the system. For NO3 and B, inputs are derived from point-sources such as WWTP effluent outfalls, or non-point sources brought in by tributaries, run-off, groundwater inflow, and atmospheric deposition. NO3 outputs are in the form of denitrification or N-assimilation by plants and microorganisms removing NO3 from the aquatic system and in the case of B output may occur by adsorption onto mineral surfaces. If there is conservative mixing of NO3 or B in the river system this means there are no removal processes occurring, therefore the sum of the inputs of nutrients along the river should equal the total flux at the mouth of the river.

The flow data for the BRB and OMR used were based on monthly averages from

Environment Canada’s Water Survey of Canada database (see Section 4.1). The Environment

Canada flow monitoring stations were not the same locations as the water quality sampling sites from this study. Therefore, for each of the surface water sampling sites the flow data from the

nearest Environment Canada flow monitoring site was used or a mid-point method between two flow stations was used to determine flow at the sampling site. The mid-point method entails taking the average flow value between a flow station up- and down-stream of a respective surface water sampling site. For both the Bow River and OMR monthly discharge averages

(kg/d) that are based on samples taken from baseflow months (Sept/Oct) were used for mass balance estimations to maintain consistency with the WWTP flux calculations that are also based on samples taken during baseflow. Flux from each WWTP sampled was calculated using the concentrations measured in WWTP effluent sampled in February and October 2014 combined with the daily outflow water volumes from the annual reports provided by the City of Calgary,

City of Lethbridge, Epcor and Parks Canada. For determining flux for each of these locations the following equation was used:

Nutrient concentration (kg/m3) x Discharge rate (m3/d) = Flux (kg/d) (4)

where the nutrient is either NO3 or B and the discharge represents a river or WWTP outflow rate.

There was uncertainty associated with each of the parameters in these flux calculations. To estimate the error propagation, the confidence interval is calculated using the equation:

Confidence Interval (%) = √ (a2 + b2 + c2 +...) (5)

where the total relative uncertainty is based on the square root of each relative uncertainty (a, b, c) of the parameters of the flux equation, squared and summed. The relative uncertainty for NO3

concentration measurement was ± 5% and for B concentration measurement was ± 10% (see

Chapter Three). There was no relative uncertainty value reported for the WWTP or river discharge datasets, so an uncertainty of ± 10% for the WWTP discharge rate and ± 20% for the river discharge was assumed. For the WWTP effluent the 10% uncertainty refers to the 2014 average daily outflow during baseflow (Jan-Mar and Oct-Dec) used in the flux calculations along with the average of two NO3 and B concentration measurements. A 20% uncertainty for river discharge was chosen to account for the river sampling sites for this study and the location of the

Environment Canada hydrometric stations being different, or for the interpolation of discharge rates at the river sampling sites that did not have a discharge station nearby. In most cases the baseflow river NO3 and B concentration data and discharge data were taken from the same year to calculate flux. However for certain data points, where concentrations were anomalously high or low, results from a different year were used. Another factor to consider is the time of day each sampling site was accessed and the relative uncertainty associated with diurnal cycling of NO3 in the river. NO3 can vary by as much as 30% of the mean concentration within a diurnal cycle, twice within a 24-hr period (Gammons et al., 2011; Nimick et al., 2011; Scholefield et al., 2005).

In the Bow River specifically, a fluctuation in NO3 concentrations between 8 and 40% from the mean was observed during daylight hours (10:30 to 17:30) through Calgary during the months of

June to September (Taube, unpublished data). Based on the average Bow River diurnal fluctuation, a relative uncertainty of ±20% is assigned to this cycling. Boron does not participate in biological reactions and therefore will not undergo diurnal concentration fluctuations. This results in NO3 flux relative uncertainties of approximately ±28% in the river, and ±11 % for the

100

WWTP effluents. For B flux there is a ±22 % relative uncertainty for river flux and ±14 % for the WWTP effluents.

6.2 Nutrient Fluxes in WWTP Effluents

The concentrations of NO3, NH4, and B measured in effluents of WWTPs in February and October 2014 are listed in Table 6.1. The greatest NO3 concentrations were measured in effluents from Bonnybrook WWTP (BB) with NO3 ranging from 54.4-59.9 mg/L, followed closely by effluents from Canmore WWTP (CM) with NO3 concentrations ranging from 48.8-

60.8 mg/L. For the WWTPs that use the nitrification/denitrification treatment process, the NO3 concentrations of the final effluent were between 8.5-60.8 mg/L. In the case of Fish Creek

WWTP (FC), where there is no nitrogen removal and it is discharged into the river as NH4, NO3 concentrations were consistently low, <1 mg/L. NH4 concentrations were also measured in effluent samples from each plant and were found to be low at most plants using N-removal, 0-

3.3 mg/L. NH4 concentrations were highest at FC with values ranging from 18.7-24.5 mg/L and followed by CM with concentrations from 5.0-11.4 mg/L. B concentrations in the WWTP effluents ranged from 180 µg/L at Pine Creek WWTP (PC) to 30.8 µg/L at LL. The highest boron concentrations were found in effluents from the WWTP of Calgary and Lethbridge (LB)

WWTPs: BB, FC, PC and LB ranging from 111-180 µg/L while the WWTP effluent from the smaller municipalities: Lake Louise WWTP (LL), Banff WWTP (BF), and CM had boron concentrations from 30.8-92.6 µg/L.

101

Since measurable NH4 concentrations in the river were only detected directly downstream of the FC WWTP outfall within the effluent plume, NH4 was not considered in the flux calculations. To calculate the nutrient flux of WWTP effluent into the rivers, the

3 concentration of nutrients (NO3 or B) in units of kg/m were multiplied by the daily outflow rate of the WWTP in m3/d, using equation (1). Since WWTP effluents were sampled during baseflow months, the average daily outflow rates of the WWTPs of October through March were used for flux calculations. These values are listed in Table 6.2.

Table 6.1: Nutrient concentrations of the effluents from the WWTPs. “NA” indicates the site was not sampled during that campaign and “n.d.” indicates below detection limit.

- + Location River NO3 (mg/L) NH4 (mg/L) B (µg/L) Feb Oct Feb Oct Feb Oct Lake Louise BRB 23.3 (LL) 13.9 0.1 n.d. 30.8 69.4 Banff (BF) BRB 11.7 13.4 3.3 n.d. 71.1 63.3 Canmore BRB 60.8 (CM) 48.8 11.4 5.0 92.6 61.3 Bonnybrook BRB 59.9 (BB) 54.4 0.3 0.7 131 111 Fish Creek BRB 0.83 (FC) 0.30 24.5 18.7 160 140 Pine Creek BRB 20.3 (PC) 8.50 3.3 n.d. 180 136 Lethbridge OMR 5.70 (LB) NA NA n.d. NA 130

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Table 6.2: The average baseflow daily flow rate (October-March) for the sampled WWTPs.

Baseflow Location River Daily Flow (m3/d) Lake Louise BRB (LL) 3,000 Banff (BF) BRB 8,203 Canmore BRB (CM) 5,448 Bonnybrook BRB (BB) 337,850 Fish Creek BRB (FC) 26,450 Pine Creek BRB (PC) 83,925 Lethbridge OMR (LB) 36,900

Table 6.3: The calculated average baseflow flux values of NO3 and B from each WWTP.

NO3 Baseflow B Baseflow Location Avg. NO3 NO3 Flux Avg. B B Flux mg/L kg/m3 kg/d µg/L kg/m3 kg/d Lake Louise (LL) 18.6 1.86E-02 55.9 47.0 4.70E-05 0.1 Banff (BF) 12.6 1.26E-02 103 66.5 6.65E-05 0.6 Canmore (CM) 54.8 5.48E-02 299 75.5 7.55E-05 0.4 Bonnybrook (BB) 57.1 5.71E-02 19291 121 1.21E-04 40.9 Fish Creek (FC) 0.57 5.65E-04 14.9 150 1.50E-04 4.0 Pine Creek (PC) 14.4 1.44E-02 1209 158 1.58E-04 13.3 Lethbridge (LB) 5.70 5.70E-03 210 130 1.30E-04 4.8

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The NO3 fluxes from WWTP effluents ranged from 14.9 ± 1.6 kg/d from FC to 19,291 ±

2122 kg/d from BB (Table 6.3). BB was by far the largest contributor, accounting for >90% of the total NO3 WWTP effluent flux. This was followed by PC (1209 ± 133 kg/d), and CM (299 ±

33 kg/d), contributing 6 ± % and 1% of the NO3 WWTP effluent flux to the Bow River, respectively. The effluent from LB WWTP contributed 210 ± 23 kg/d NO3 flux to the OMR and was the only WWTP effluent discharged into the river.

In terms of B flux with WWTP effluents there was a range of 0.1 ± 0 kg/d from LL to

40.9 ± 5.7 kg/d at BB. Again, the BB WWTP had the largest discharge of B, contributing approximately 70% of the total B WWTP effluent flux into the Bow River. This flux was followed by PC (13.3 ± 1.9 kg/d) and FC (4.0 ± 0.4 kg/d) contributing over 20% and 6% of B

WWTP effluent flux to the Bow River, respectively. The Calgary WWTPs combined accounted for 98% of the total B flux by WWTP effluent in the Bow River. LB WWTP contributed 4.8 ±

0.7 kg/d B as the sole contributor of WWTP B flux to the OMR.

The BB WWTP had the highest average daily flow rate out of each of the WWTPs discharging into the Bow and Oldman Rivers. With effluent containing significant NO3 and B concentrations, this resulted in the highest NO3 and B fluxes from WWTP effluent to these southern Alberta riverine systems.

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6.3 Riverine NO3 Budgets

6.3.1 The Bow River

For NO3 flux estimations the September 2014 surface water concentration and discharge values were typically used, as listed in Table 6.4. The initial NO3 flux in the BRB was 1033 ±

289 kg/d above Lake Louise and increased to a final flux of 25,523 ± 7146 kg/d at Ronalane, a

25-fold increase. The overall trend of NO3 flux was a 3- to 4-fold increase from Lake Louise to

Canmore (Reach 1), a 2-fold increase at Cochrane, up to a 10-fold increase by the time the Bow

River flows through Calgary (Reach 2) compared to Reach 1, and no further increase in NO3 flux downstream of Calgary to Ronalane (Reach 3). A conceptual model of the NO3 flux along the

Bow River can be found in Figure 6.1.

Reach 1

There was a significant increase in NO3 flux from the headwater site above Lake Louise to above Banff during baseflow, from 1033 ± 289 to 3810 ± 1067 kg/d, respectively. The 3- to 4- fold increase in NO3 flux in this reach was caused by the increase in Bow River flow fed by

3 3 tributaries, from 13 m /s above Lake Louise to 49 m /s above Banff while the NO3 concentrations in this reach remain consistent around 0.90 mg/L (Table 6.4). The NO3 flux remained at 3810 kg/d until Cochrane where there was an approximate 2-fold increase to 6247 ±

1749 kg/d. This was due to a minor increase in NO3 concentration from 0.90 up to 1.0 mg/L but most importantly another increase in riverine discharge, from 49 m3/s above Canmore to 72.3

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m3/s at Cochrane. This increase in flow was due to tributaries such as the Kananaskis and the

Ghost Rivers feeding into the Bow between Canmore and Cochrane.

Reach 2

NO3 flux increased from 6247 ± 1749 kg/d at Cochrane to 32,296 ± 9043 kg/d below

Bonnybrook WWTP, a significant 5-fold increase, or a 10-fold increase compared to NO3 flux at

Banff and Canmore. The discharge of the Bow River increased from 72.3 up to 95.6 m3/s through Calgary due to the Elbow River feeding into the Bow, but the driver of this increase in

NO3 flux are the NO3 concentrations. NO3 increased from 1.0 mg/L at Cochrane to up to 3.9 mg/L below the Bonnybrook WWTP in the Bow River. The NO3 flux remained elevated through

Calgary, ranging from 23,293 ± 6522 kg/d below Fishcreek WWTP to 29,240 ± 8187 kg/d below

Pine Creek. The largest increase in NO3 flux along the entire length of the Bow River occurred below the Bonnybrook WWTP (Section 6.2) (Figure 6.1).

Reach 3

Downstream of Calgary, the Bow River NO3 flux was 33,517 ± 9385 kg/d at Carseland, the same amount of NO3 flux measured through Calgary within the measurement uncertainty.

Once downstream of the Bassano Dam, the NO3 flux decreased to 18,409 ± 5155 kg/d at Bow

City, a 50% decrease in NO3 flux. The dam causes water –level drawdown through diversion to the agricultural irrigation canals, so NO3 flux was expected to decrease to some degree downstream of the dam. Whether there are biological processes contributing to the decrease in

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NO3 load, such as denitrification, will be discussed in Section 6.5. The NO3 flux was measured at

25,729 ± 7204 kg/d at Scandia, an approximate 40% (or 1.5-fold) increase in NO3 flux compared to Carseland, and remained constant through Ronalane at 25,523 ± 7146 kg/d (Figure 6.1).

15 18 11 Isotope data (δ NNO3, δ ONO3, and δ B) will be used in Chapter Seven to investigate what the driving factors may be causing these NO3 flux variations.

Table 6.4: NO3 flux during peakflow (June) and baseflow (Sept) periods along the BRB.

NO3 NO3 June 2014 Q NO3 Flux Sept 2014 Q NO3 Flux Km mg/L m3/s kg/d mg/L m3/s kg/d Ronalane 571 1.7 349.5 50429 2.3 129.0 25523 Scandia 513 1.5 340.5 43540 2.4 124.6 25729 Bow City 482 1.6 340.5 46777 1.7 124.6 18409 Carseland 302 1.5 288.8 37678 3.2 120.1 33517 Below Pine Creek 228 1.8 246.2 37651 3.5 95.6 29240 Below Fish Creek 220 1.3 246.2 27866 2.8 95.6 23293 Below Bonnybrook 205 2.1 246.2 44883 3.9 95.6 32296 Cochrane 179 0.5 191.3 8760 1.0 72.3 6247 Canmore- above 79 0.2 136.4 2475 0.8 49.0 3302 Banff- below 64 n.d. - - 0.9 49.0 3810 Banff- above 56 n.d. - - 0.9 49.0 3810 Lake Louise- above 0 n.d. - - 0.9 13.0 1033

6.3.2 The Oldman River

The flux rates calculated from baseflow (Oct 2015) were used for NO3 load analysis in the Oldman River because there was available flow data and the greatest number of samples with

NO3 concentration data were collected during this campaign (Table 6.5). In the studied area of

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the OMR, both Reach 1 and Reach 2 flow through areas with similar land use, dominantly agricultural. Effluents from agriculture, feedlots, and grazing all are non-point source inputs of nutrients to river systems when they infiltrate into the river through groundwater interaction.

However, irrigation canals that collect agricultural return-flow waters, such as the Haney and

Battersea Drain, can be viewed as point-source conduits for these nutrient inputs back to the

OMR.

In October 2015, the baseflow NO3 flux at Monarch was 1702 ± 477 kg/d and increased by 30% to 2187 ± 612 kg/d below the confluence with the Belly River, and remained around this

NO3 load downstream for a final flux of 2096 ± 587 kg/d at Taber (Figure 6.2). These relatively constant NO3 flux values suggest there were not any significant N-removal processes occurring.

The only significant change in NO3 flux was measured below the confluence with the Belly

River and this is likely attributed to the increase in discharge from the Belly River bringing in greater NO3 loading from its drainage area.

There was a calculated NO3 flux of 116 ± 33 kg/d from the Little Bow River, approximately 5% of the NO3 flux in the mainstem of the Oldman River above and below the mouth of the tributary. The Battersea and Haney Drain do not have recorded flow data, however if an estimated baseflow discharge rate of 0.01 m3/s is assigned to the irrigation drains, 250 times less than the Little Bow River, this would result in associated NO3 flux values of 40.1 ± 11.2 and

60.2 ± 16.7 kg/d, respectively. If this estimation is correct, the irrigation canals combined would contribute almost 10% of the upstream Pavan Park NO3 load to the mainstem, but there were no significant changes in NO3 flux downstream of the sampled tributaries (Figure 6.2).

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Table 6.5: NO3 flux during peakflow (June) and baseflow (Oct) periods along the OMR.

Data for the measured tributaries are highlighted in grey.

NO3 NO3 NO3 Oct NO3 Jun NO3 Oct NO3 2014 Q Flux 2015 Q Flux 2015 Q Flux Km mg/L m3/s kg/d mg/L m3/s kg/d (mg/L) m3/s kg/d Taber 140 0.8 68.7 4689 0.0 117.1 0 0.7 36.2 2096 D/S Tributaries 113 - - - 0.5 112.6 4572 0.8 32.8 2267 Little Bow River 110 0.5 2.5 102 0.6 3.8 181 0.6 2.4 116 Battersea Drain 98 74.5 0.1 644 0.7 0.2 12 46.4 0.1 40.1 Haney Drain 96 90.3 0.1 780 55.3 0.2 955 69.6 0.1 60.2 Pavan Park 52 1.1 58.1 5672 0.6 108.1 5417 1.0 29.4 2439 Below Belly River 16 0.9 40.5 3044 1.0 89.7 7518 1.0 26.1 2187 Monarch 0 1.0 40.5 3429 1.5 89.7 11393 0.9 22.9 1702

6.3.3 WWTP influence on Riverine NO3

Along the Bow River, below the Bonnybrook (BB) WWTP the observed NO3 flux increased 5-fold, significantly beyond the 28% uncertainty margin. This increase in NO3 flux was driven by an increase in riverine NO3 concentration from 1.0 to 3.9 mg/L and little change to riverine flow rate. The NO3 flux increase observed below BB WWTP was equal to what is input by the WWTP effluent within the margin of error. This was the most significant change in NO3 flux, along the Bow River but also within the entire study area, caused by WWTP effluent. The other WWTPs within Calgary, Fish Creek (FC) and Pine Creek (PC), did not affect the riverine

NO3 load outside of the 28% relative uncertainty margin. There was also a 3-fold increase in

NO3 flux between above Lake Louise (LL) and Banff, where the LL WWTP discharges in

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between. However, with only 56 ± 6 kg/d NO3 flux from the WWTP effluent and no change in riverine NO3 concentration there was no evidence that the effluent was contributing markedly to riverine NO3 flux. There was also a 2-fold increase in NO3 flux between Canmore and Cochrane, from 3302 ± 925 to 6247 ± 1749 kg/d, and the Canmore (CM) WWTP that discharges 299 ± 33 kg/d NO3 into the Bow River between these two sampling locations. The NO3 flux from the CM

WWTP contributed approximately 10% of the NO3 flux increase, however a large contributor to this change in NO3 flux was due to the large increase in discharge in the Bow River due to tributaries feeding into the river.

In the Oldman River, the input of NO3 flux from the Lethbridge WWTP was 210 ± 23 kg/d, approximately 10% of the NO3 flux both above and below the WWTP. This input was not reflected in the NO3 flux at the Pavan Park site directly downstream of the WWTP, or was undetectable within the uncertainty of the calculations (Figure 6.2). Since there was no obvious increase in NO3 flux below the known point-source, it was difficult to conclude whether there was any influence of NO3 loading due to WWTP effluent based on the flux data alone. Further investigation using isotopic tracing tools will be described in Chapter Seven.

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uncertainty andWWTPauncertainty. effluents 11% uncertainty return sitesTheWWTP agricultural effluent represents sampling and (squares). rectangle the (circles) waterthe 6.1: Figure LL 1033 - 56 LL flow flow non

3810 NO BF - A 3 103 - BF

pointsour flux (kg/d) in the Bow River.Bow (kg/d) The valuesflux onin based the basefloware values flux fromthe surface BF 3810 - B CM 3302 - ce of NO of ce A CM 299 6247 CR 3

downstreamofNOCalgary.riverineEach 19291 BB BB 32296 - B 15 FC 23293 FC - B 1209

PC 29240 PC - B CS 33517 Bassano Dam Bassano Agriculture Return Agriculture 3

flux valuefluxhas 28% a BC 18409 - Flow 25729 SD 25523 RL

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and WWTP effluent auncertainty.WWTP 11% effluent and agriculturalreturn representedwithin r non of Lethbridge the downstream sitesandWWTPeffluent(circles)Both (square).fromtheupstream Lethbridge and sampling NO 6.2: Figure epresented by rectangles. The of rectangles.sampled Lethbridge tributaries threeare downstream epresentedby 3

fluxinestimatedvalues theOMR baseflow (kg/d) of using surfacewater - point sources of nutrients, agricultural nutrients, sourcespoint return of - flow. Eachflow. NO riverine

flux value has a 28% value fluxa uncertainty has - flow, areflow,

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6.4 Riverine B Budget

6.4.1 The Bow River

The overall trend in B load in the Bow River was increasing with downstream distance from 1.0 ± 0.3 kg/d above Lake Louise to 189 ± 42 kg/d at Ronalane, based on baseflow B flux values (Sept 2014) (Table 6.6). Since the boron loads did not significantly decrease at any point with downstream distance, there was no evidence of any major boron removal processes occurring in the Bow River.

Reach 1

At the headwaters above Lake Louise boron flux was 1.0 ± 0.3 kg/d during baseflow. The flux increased by a factor of 14 in the Bow River downstream of the sampling sites above Banff.

This change in B flux was associated with both an approximate 3-fold increase in the river discharge and in B concentration (Table 6.6). The increase in discharge was due to the Spray

River and Cascade River tributaries, among others, flowing into the Bow River near Banff.

These tributaries increased the B loading to the Bow River, but without additional tracers it cannot be determined if there are other B sources causing the increase in concentration at this site. The boron flux doubled between Canmore and Cochrane from 15.1 ± 3.3 kg/d to 35.2 ± 7.7 kg/d. The addition of tributary flow caused this increase in B loading and there was also an increase in B concentration from 3.6 to 5.6 µg/L in the river at Cochrane. Whether there were

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point sources of B such as WWTP effluents or non-point sources causing an increase in B flux will be investigated in Chapter Seven.

Reach 2

As the Bow River flows through Calgary there was a 3.5-fold increase in boron flux from

35.2 ± 7.7 kg/d at Cochrane to 122 ± 27 kg/d below the BB WWTP. The B flux remained constant in the remainder of Reach 2 within the 22% relative uncertainty associated with riverine

B fluxes. The most significant change in B flux was the overall 4.5-fold increase from Reach 1, at Cochrane, to Reach 2, downstream of the Calgary WWTPs. The calculated B input from the

Calgary WWTPs effluents combined was 58.2 ± 8.0 kg/d and this input, taking the relative uncertainty into account, was not large enough to account for the 4.5-fold increase in B loading in the river (Figure 6.3). Some additional B flux may have also been supplied by the confluence of the Elbow River into the Bow River through Calgary, since tributaries bring in boron and nutrient loads from further reaches of the BRB. The major sources influencing this increase in riverine B flux will be further investigated using the isotopic compositions of B in Chapter

Seven.

Reach 3

The B load in the Bow River remained constant within uncertainty downstream of the

Calgary WWTPs to Carseland (Table 6.6). Boron flux increased from 127 ± 28 kg/d at Carseland to 193 ± 42 kg/d at Bow City, a 50% increase, then remained constant to a final B flux value of

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189 ± 42 kg/d at Ronalane (Figure 6.3). There was no apparent decrease in B flux associated with the water drawdown at the Bassano Dam between Carseland and Bow City. The source of the significant increase in B flux will be investigated further in Chapter Seven.

Table 6.6: B flux during peakflow (June) and baseflow (Sept) periods along the BRB.

B B B B June 2014 Q Flux Sept 2014 Q Flux Km µg/L m3/s kg/d µg/L m3/s kg/d Ronalane 571 8.7 349.5 261 17.0 129 189 Scandia 513 13.5 340.5 398 18.8 124.6 202 Bow City 482 16.7 340.5 491 17.9 124.6 193 Carseland 302 10.8 288.8 270 13.6 107.9 127 Below Pine Creek 228 3.2 246.2 68.1 18.8 95.6 156 Below Fish Creek 220 13.0 246.2 277 20.0 95.6 165 Below Bonnybrook 205 9.7 246.2 206 14.8 95.6 122 Cochrane 179 3.8 191.3 62.5 5.6 72.3 35.2 Canmore- above 79 4.1 136.4 48.6 3.6 49.0 15.1 Banff- below 64 n.d. - - 3.4 49.0 14.5 Banff- above 56 n.d. - - 3.3 49.0 13.9 Lake Louise- above 0 n.d. - - 1.3 13.0 1.5

6.4.2 The Oldman River

During baseflow (Oct 2015) in the OMR, there was an overall increasing trend in B flux from the upstream site at Monarch with 28.7 ± 6.3 kg/d moving downstream to Taber with 45.6

± 10.0 kg/d (Table 6.5). The largest change in B flux occurred between below the confluence of the Belly River, 26.1 ± 5.7 kg/d to Pavan Park in Lethbridge, 39.2 ± 8.6 kg/d, with a 50% increase in flux that falls outside the 22% relative uncertainty of the B flux calculations. Between these two sampling locations there is the confluence with the St. Mary River and Six Mile

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Coulee, as well as the Lethbridge WWTP, all of which could contribute to the increase in boron flux in the mainstem of the OMR. There is no available flow data for these tributaries but the St.

Mary River is a substantial tributary and hence flow, and B flux, in the OMR is expected to increase downstream of the confluence.

The same discharge values were assigned to the Battersea and Haney Drains as with the

3 NO3 flux calculations, 0.01 m /s. This resulted in both irrigations canals having B flux values of

0.1 ± 0 kg/d. The Little Bow River had a calculated B flux of 5.2 ± 1.1 kg/d. The combined B flux of the tributaries was 5.4 ± 1.2 kg/d, approximately 12-14% of the B load of the OMR both up- and downstream of the tributaries. This added B flux from the tributaries is within the 22% relative uncertainty of the mainstem B load and therefore a definite increase in B fluxes due to input by the Little Bow River and irrigation canals cannot be identified. The gradual increase in

B flux with increasing downstream distance indicated boron could be conservatively mixing in the OMR. To determine the sources that affect the B load in the OMR (e.g. natural, WWTP effluent, or agricultural), isotope data will be utilized as described in Chapter Seven.

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Table 6.7: B flux during peakflow (June) and baseflow (Oct) periods along the OMR. The measured tributaries are highlighted in grey.

Oct June Oct B 2014 B Q B Flux 2015 B Q B Flux 2015 B Q Flux Km µg/L m3/s kg/d µg/L m3/s kg/d µg/L m3/s kg/d Taber 140 17.2 68.7 102 14.9 117.1 151 14.6 36.2 45.6 D/S Tributaries 113 - - - 15.5 112.6 151 15.3 32.8 43.4 Little Bow River 110 27.3 2.5 5.9 19.2 3.8 6.3 25.2 2.4 5.2 Battersea Drain 98 121.3 0.1 1.0 17.4 0.2 0.3 67.4 0.1 0.1 Haney Drain 96 122.9 0.1 1.1 114.5 0.2 2.0 108.7 0.1 0.1 Pavan Park 52 16.4 58.1 82.4 15.6 108.1 145 15.4 29.4 39.2 Below Belly River 16 12.6 40.5 44.1 11.2 89.7 86.6 11.6 26.1 26.1 Monarch 0 14.2 40.5 49.8 12.8 89.7 99.3 14.5 22.9 28.7

6.4.3 WWTP influence on B riverine flux

Like NO3 riverine flux, the Bonnybrook (BB) WWTP effluent had the highest B load out of the WWTPs discharging into the Bow River, 40.9 ± 5.7 kg/d, and was associated with one of the highest increases in riverine B flux, a 3.5-fold increase from Cochrane to below Bonnybrook

(Table 6.6). As mentioned above a portion of this B flux increase may also have been in part due to the Elbow River loading B into the Bow. However, the increase in B concentration from 5.6

µg/L to 14.8 µg/L below Bonnybrook was a 3-fold increase that was likely due to input by BB

WWTP effluent. Although the other two WWTPs through Calgary, Fish Creek and Pine Creek, had the next highest B fluxes discharging into the Bow River, the riverine B flux did not clearly reflect their input within the relative uncertainty of the B flux estimates.

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Another significant increase in B flux along the Bow River with associated WWTP effluent discharge was found between above Lake Louise (LL) and above Banff sampling sites shifting from 1.0 ± 0.3 to 13.9 ± 3.1 kg/d. The LL WWTP discharged its effluent between these two sampling points with a measured B flux of 0.1± 0.0 kg/d. This represented less than 1% of the downstream B load and therefore the LL WWTP effluent was not a significant influence on boron loading in the river. The combined boron flux from the Banff and Canmore WWTP was approximately 1 kg/d and therefore was also not the major contributor to increased B fluxes in

Reach 1 of the river. The 3-fold increase of B fluxes observed in the Bow River between

Canmore and Cochrane was likely influenced by tributary flow rather than WWTP effluent.

The Lethbridge (LB) WWTP effluent had a B flux input of 4.8 ± 1.0 kg/d to the OMR.

This B flux accounted for half of the 25% increase in B load in the OMR from 29.4 ± 6.5 kg/d below the Belly River confluence to 39.2 ± 8.6 kg/d directly below the WWTP at Pavan Park

(Figure 6.4). This increase in riverine B loading fell just outside the margin of the 22% uncertainty so it is difficult to confirm that the amount of B flux discharged by the LB WWTP significantly contributed to the increase of B flux in the river within the margin of error. As mentioned previously there are also tributaries in this reach that contribute some B flux to the

OMR. Chapter Seven will introduce the B isotopic signal of the OMR to further define the changes in the riverine B flux.

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14% uncertainty. 14% 22%Eacha relative downstreama Bhasand WWTP fluxof Buncertainty Calgary. riverine effluent of WWTP effluent sitesand (circles) 6.3: Figure LL 1.0 0.14 LL

B flux (kg/d) along the Bow(kg/d)flux periodBalong the duringRiver baseflowfromthe surfacewater sampling BF 14 - A 0.55 BF

BF 14 - B CM 15 - A CM 0.41 (squares). The agricultural returnThe (squares). agricultural CR 35 41 BB BB 122 - B 4.0 FC 165 FC - B 13 PC PC 156 - B CS 127 - flow (rectangle) is aflowis(rectangle) non Bassano Dam Bassano Agriculture Return Agriculture BC 193 - Flow SD 202 - point sourcepoint 189 RL

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flux valuearelativeahasand 22% uncertainty 14% effluent flux uncertainty. WWTP tributa agriculturalsourceof nutrients, return point effluenttheBothLethbridge(square).up WWTP and (circles) ofBOMR6.4: (kg/d) values fluxduringthethe basef Figure ries downstream ofLethbridgeriesrepresented aredownstream theagriculturalreturn within - flow, are representedareTheby flow, rectangles.sampled three low periodsitesfromthelow surfacewater samping -

anddownstreamnon of Lethbridge the

- flow. Each riverine BriverineEach flow. -

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6.5 Riverine NO3/B Ratios

Because the entire length of the Bow River was used in this study, it was possible to compare the NO3/B loading ratios along the length of the river to evaluate how differently the two calculated parameters behave as a large-scale spatial trend. In the Oldman River, the 140 km distance sampled focuses on changes downstream of Lethbridge and some agriculturally influenced tributaries but does not capture overall changes between NO3 and B flux in the OMR over long distances. Therefore, only the Bow River flux data was used for NO3/B flux ratio comparison. The NO3/B flux ratios for the Bow River during baseflow (Sept 2014) were plotted against distance in Figure 6.5. The ± 28% uncertainty for riverine flux was applied to these ratios.

In Reach 1 of the Bow River from above Lake Louise to Cochrane (0 to 179 km), the

NO3/B flux ratio declined by a factor of 7 as boron flux increased at a greater magnitude than

NO3 flux with downstream distance. In Reach 2 through Calgary (180 to 228 km), the NO3/B flux ratio remained constant within the margin of relative uncertainty. This indicated that B and

NO3 were loaded at similar rates and both sufficiently represented the WWTP effluent input to the Bow River in this reach. In Reach 3 (229 to 571 km), the NO3/B flux ratio at Carseland (302 km) remained at the same ratio as those observed in Reach 2 within the error of the flux determinations, indicating little NO3 removal. Downstream of the Bassano Dam (390 km), the

Bow City (482 km), Scandia (513 km), and Ronalane (571 km) NO3/B flux ratios expressed a 2 to 2.7-fold decrease from the Carseland site, a significant shift. In this final reach of the Bow

River the boron flux increased from Carseland to Bow City and then remained constant

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downstream while the NO3 flux decreased from Carseland to Bow City then remained constant.

This suggests there were some NO3 removal processes occurring in the Bow River between

Carseland and Bow City while boron was added to the system along this reach. This supports the hypothesis of boron acting as a conservative tracer of potential nutrient inputs in riverine systems.

800

700

600

500 Flux 400 /B 3

NO 300

200

100 LL BF CM CR Calg. CS BC SD RL 0 0 100 200 300 400 500 600 Distance from a point above Lake Louise (km)

Figure 6.5: The NO3/B flux ratio during baseflow 2014 in the Bow River with distance from above Lake Louise. A 28% error is applied to this ratio.

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Nutrient Sources and Cycling Based on δ15N, δ18O, and δ11B Values

7.1 Introduction

Stable isotope tracing is an additional tool used to identify natural and anthropogenic sources of NO3 and boron in this southern Alberta riverine nutrient study. As previously mentioned, stable isotope fingerprinting has been applied in numerous studies worldwide to distinguish between different natural and anthropogenic sources of N-containing compounds in riverine systems (Battaglin et al., 2001; Chang et al., 2002; Kendall et al., 2003; Kratzer et al.,

2004; Mayer et al., 2002; Panno et al., 2006; Rock & Mayer, 2004; Sebilo et al., 2006). One of

15 18 the main objectives of this study was to use δ N and δ O of NO3 along with the co-tracing

11 isotope ratios of boron, δ B, to better identify the major NO3 contributors within southern

Alberta river systems. To define the anthropogenic end-members, the δ15N, δ18O (where applicable), and δ11B values of mineral fertilizers used in Alberta, cow manure, and WWTP effluents were measured. The isotopic compositions measured in this study are compared against previously reported isotope values found in the literature. The control site for this study was above Lake Louise on the Bow River. This site was considered having a “background” or natural isotopic composition for surface water with only minor anthropogenic influences. The NO3 isotope ratios for surface water samples from the locations along the Bow River and OMR were measured during baseflow 2014 and both peakflow and baseflow 2015. The available δ15N

+ values of NH4 (NH4) in the rivers were analyzed during baseflow 2014 but concentrations were

15 rarely above the detection limit therefore this δ N -NH4 data do not give much insight into

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nutrient sources and are only briefly discussed. The δ11B values of the river water samples during baseflow 2015 were measured and used as a fingerprinting tool on their own as well as in

15 combination with the δ NNO3 data to determine natural and anthropogenic nutrient sources and mixing scenarios in both rivers, and compare the differences in nutrient sources between the two major rivers. Based on the results of Chapter Six, along the Bow River the Bonnybrook (BB)

WWTP effluent creates the largest NO3 flux increase and therefore is the greatest anthropogenic

NO3 contributor to the river. The B flux along the Bow River reveals a large flux increase below the BB WWTP that continues to increase downstream of Calgary suggesting there could be additional agricultural sources contributing to NO3 and B loading. Along the OMR the only significant increase in NO3 flux occurs below the confluence with the Belly River and B fluxes suggest there could be an influence on load from the Lethbridge WWTP. The Little Bow River and two irrigation canals sampled do not change either the downstream B or NO3 load in the

OMR river beyond the measurement uncertainty. These preliminary findings will be tested using

15 18 11 a combination of δ NNO3, δ ONO3, and δ B values in an attempt to identify less dominant NO3 and B inputs in the rivers, such as other WWTP effluent signatures or specific agricultural

15 18 11 signals such as nutrients derived from fertilizer or manure. The δ NNO3, δ ONO3, and δ B values of the river will also with assist with determining if anthropogenic inputs, such as the

Bonnybrook WWTP and agricultural return-flow, persist with downstream distance in the river or if they are reduced by biological reactions.

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7.2 Results

7.2.1 Isotopic Composition of Anthropogenic Nutrient Sources

The following compiles the isotopic data measured for local samples of the major anthropogenic nutrient sources in southern Alberta: WWTP effluent, mineral fertilizers and manure. Since cattle are the main livestock, with over 5 million head in the province of Alberta

(StatsCan, 2016a), isotopic analyses were focused on cattle manure. The results of NO3 and B concentrations and isotope ratios for these end-members are compiled in Table 7.1.

WWTP Effluent

The isotope analyses of WWTP effluent discussed in this chapter are based on the

October 2014 dataset. The NO3, NH4, and B concentrations for the obtained WWTP effluent

15 samples are found in Chapter Six. The relationship between NO3 concentrations and δ N and

18 15 δ O values are shown in Figure 7.1. The range of δ NNO3 values spanned from 15.3‰ at Banff

(BF) to -0.8‰ at Fish Creek (FC) and δ18O values vary from -11.6‰ at Lake Louise (LL) to

15 18 -1.8‰ at Lethbridge (LB). With decreasing NO3 concentration both the δ NNO3 and δ ONO3 values increased, indicating that during the treatment process the lighter isotopes, 14N and 16O, were being reduced from NO3 to N2 gas and progressively more complete denitrification results

15 18 in N and O enrichment in the remaining nitrate. Only the FC NO3 isotope data showed deviation from this trend since there is no nitrification/denitrification treatment at FC WWTP.

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The boron isotope data from the WWTP typically range between +0.9 to +2.9‰ except for the LB WWTP that had a δ11B signature of +7.2‰. Since the δ11B values of sewage effluent were similar to the sources of the boron, namely Ca/Na-borate detergents and cosmetics, boron isotopes are assumed to remain conservative during the WWTP process (Komor, 1997;

Leenhouts et al., 1998; Petelet-Giraud et al., 2009).

15 The δ N-NH4 values were not able to be measured from each WWTP due to NH4 concentrations typically below the detection limit. However, in effluent from FC, the NH4

15 concentration was 24.5 mg/L and a δ N value of NH4 of 7.0‰ was observed. Most WWTP

15 15 effluents had a δ NNO3 range of 8-15‰. Since the δ NNH4 value was close to this isotopic range this likely suggests that all the other WWTP that undergo nitrification/denitrification processes

15 for N-removal had near complete NH4 conversion to NO3, causing their δ NNO3 values to be

15 similar to the δ NNH4 value of the FC effluent.

Mineral Fertilizer

Nitrogen isotope analyses for all major fertilizers applied in Alberta were completed by the Applied Geochemistry Group at the University of Calgary (unpublished data). The fertilizers that were further analyzed for this study are listed in Table 7.1. These were chosen to cover the most widely used fertilizers in the province that had measurable boron contents (see Section 2.3).

There were attempts to measure boron contents in ammonium nitrate fertilizer however only trace boron concentrations were detected (0.005 µg/g) and therefore these fertiizers were excluded from this study. The majority of the fertilizers chosen for boron analysis are NH4-

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15 based with one urea-based fertilizer ((NH2)2CO). The δ N values measured in the fertilizer samples range from -1.3 to 1.7‰, which concurs with what is stated in the literature (Section

1.2.1). The mineral fertilizers were measured for δ15N values directly from the fertilizer samples

18 18 before nitrification so there was no δ O-NO3 value measured. Typical δ O values associated with fertilizer nitrification can be found in Section 1.2.1.

The boron content in these fertilizers ranged from 4.8 µg/g in urea fertilizer to 1011 µg/g in ammonium phosphate (+other) lawn fertilizer. The δ11B values of the fertilizers ranged from

1.0‰ in ammonium phosphate (+other) lawn fertilizer to 8.7‰ in agricultural ammonium phosphate fertilizer. The δ11B value measured in the urea fertilizer sample from this study was

1.5‰, which falls outside the range of previously reported values (-2 to 0.7‰) (Komor, 1997).

The δ11B values of fertilizers are based on the source from which the fertilizer was mined and therefore must be defined depending on the source of the fertilizer. Komor reported a δ11B value of 14.8‰ for phosphate fertilizers, higher than the δ11B values from this study. Other studies found it difficult to measure δ11B values in fertilizers because of their low boron concentrations

(see Section 1.2.2).

Cow Manure

There were two sample types of cow manure provided for analysis: stockpiled and composted manure. The concentrations and isotopic values of stockpiled versus composted manure samples are compiled in Table 7.1. The measured NO3 concentration for the analyzed samples was 51.4 mg/g for stockpiled and 21.4 mg/g for composted manure, measured in a

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15 filtered manure solution (preparation described in Section 3.2.2). The δ NNO3 values for

18 stockpiled and composted manure were +22.8‰ and +14.1‰ and the δ ONO3 values are +1.6‰ and -10.6‰, respectively. The NO3 in manure forms from nitrification of urea. Because the NO3 concentration of the composted manure was half that of the stockpiled manure NO3 concentration, this potentially indicates that some of the nitrogen content may have been reduced

15 to N2 gas during denitrification in the decomposition process. The δ NNO3 values fell within the

15 18 higher of the previously reported range (δ N +8 to +25‰) for manure and the δ ONO3 values were well within the previously reported range (δ18O -20‰ to +10‰) as described in Section

1.2.1 (Kendall et al., 2007). During the decomposition process from “stockpiled” to “composted” manure, there is a 20-25% decrease in N-content in a manure sample (Dr. Benjamin Ellert,

14 private communication). One would expect N to be lost during volatilization of NH3, causing

15 15 an increase in the remaining δ N-NH3 value. However, there were lower δ N-NO3 values observed in the composted manure compared to stockpiled manure. This may be related to the decomposition of the straw or woodchip bedding the manure is typically mixed with, but further investigations would be required to test this.

The boron content in these samples was 96.6 µg/g for stockpiled and 105.3 µg/g for composted manure. The average of two analyses of δ11B values for the manure samples were

+25.4 ± 3.3‰ and +25.7 ± 0.4‰ with their standard deviations, for stockpiled and composted manure respectively. The boron concentrations and isotopic compositions are similar between the two samples indicating there was no boron lost during decomposition. These δ11B values are within the previously reported values for cattle manure of 25 to 29‰ (Widory et al., 2004).

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Boron content in manure is mainly sourced from the type of feed the animal is given (Komor,

1997). Thus, the boron isotopic ratio may vary slightly depending on the diet of a given animal species.

Table 7.1: δ15N, δ18O, and δ11B values of the major nutrient end-members within southern

Alberta.

[NO -] [B] 3 δ15N (‰) δ18O (‰) δ11B (‰) Source (mg/L) NO3 (µg/L) WWTP

Lake Louise (LL) 23.3 3.2 -11.6 69.4 0.9 Banff (BF) 13.4 15.3 -5.2 63.3 2.9 Canmore (CM) 60.8 7.2 -11.3 61.3 2.0 Bonnybrook (BB) 59.9 9.6 -8.2 111.0 1.2 Fish Creek (FC) 0.83 -0.8 -3.7 140.1 1.2 Pine Creek (PC) 20.3 14.9 -2.0 136.0 1.9 Lethbridge (LB) 5.7 13.2 -1.8 129.9 7.2 [B] [NO -] δ15N (‰) δ18O (‰) δ11B (‰) Fertilizer 3 NO3 (µg/g)

(NH4)2SO4 n/a 1.7 n/a 246.2 3.6

(NH4)2SO4 n/a 0.1 n/a 168.4 3.6

NH4*H2PO4 n/a -0.3 n/a 76.4 7.5

NH4*H2PO4 n/a 0.1 n/a 319.6 8.7

NH4*H2PO4 n/a 0.4 n/a 92.8 6.1

(NH2)2CO n/a -1.3 n/a 4.8 1.5

NH4*H2PO4+other (lawn) n/a -0.5 n/a 209.2 1.0

NH4*H2PO4+other (lawn) n/a -0.7 n/a 1011 4.3 [NO -] [B] δ11B (‰) 3 δ15N (‰) δ18O (‰) Cow Manure (mg/g) NO3 (µg/g) (avg) Stockpiled 51.4 22.8 1.6 96.6 25.4 Composted 21.4 14.1 -10.6 105.3 25.7

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18 -1 16 14 -3 12 -5 10 -7 (‰) (‰) 8 NO3 O NO3 18 N 6 -9 δ 15 δ 4 -11 2 -13 0 -2 -15 0 10 20 30 40 50 60 70

NO3 (mg/L) 15 18 Figure 7.1: δ N (blue) and δ O (red) values of NO3 versus NO3 concentration for effluents from the WWTPs.

The δ15N and δ18O values of these anthropogenic end-members allowed the distinction between nitrate in mineral fertilizer and sewage effluent but there was overlap between the isotopic signals of nitrate in sewage effluent and cow manure. The δ11B values allowed for a distinction between sewage effluent and cow manure, but there were overlapping δ11B values for mineral fertilizer and sewage effluent. Using a combination of δ15N, δ18O and δ11B values there appeared to be strong potential for these end-members to be distinguished since they were isotopically distinct. This will be investigated in the discussion in Section 7.3.

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7.2.2 The Bow River

15 18 11 The δ NNO3, δ ONO3, and δ B values for surface water samples obtained from the Bow

River sampling sites are plotted against distance starting from 0 km above Lake Louise in Figure

15 18 7.2. The δ N and δ O values of NO3 were measured for samples obtained during both peakflow and baseflow periods between 2014 and 2015 and boron isotope ratios were measured for the

15 18 11 15 September 2015 sample set. Table 7.2 compiles the δ NNO3, δ ONO3 and δ B values. δ NNH4 data from September 2014 samples are also discussed.

15 The average δ NNO3 values from all sampling campaigns increased with downstream distance ranging from -0.1‰ above Lake Louise to +10.4‰ at Ronalane. The majority of the increase occurred between Cochrane (179km) to below Fish Creek WWTP (220km). Thereafter the δ15N values increased at a lower rate (up to +4‰ shift) or remained constant through the final

18 reaches to the Ronalane sampling sites (Figure 7.2). The average δ O-NO3 values over all sampling campaigns displayed an overall decreasing trend with a range from +8.2‰ above Lake

18 Louise to -8.4‰ below Bonnybrook WWTP. Thereafter δ O-NO3 values increased to Bow City and then remain constant until a final average δ18O value of -4.3‰ at Ronalane. The decreasing trend in δ18O values from Lake Louise (0km) to below Bonnybrook WWTP (205km) was the largest change in δ18O values. The increase downstream of Bonnybrook towards Ronalane had a gentle upward slope. The trends in NO3 isotope ratios with distance were similar between sampling campaigns, indicating the changes observed in the NO3 isotope ratios were not constrained by seasonal discharge fluctuations.

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The δ11B values of the Bow River, measured in samples obtained in September 2015, were rather constant with an average δ11B value of +8 ± 4‰ from above Lake Louise (0 km) to

Carseland (302km). At Bow City (482km) the δ11B value increased to +17.1‰ and decreased to a final δ11B value of +13.4‰ at Ronalane (571km). δ11B values remain constant for the majority of river until below Bassano Dam, an agriculturally dominant reach of the Bow River.

The only NH4 concentrations along the Bow River that were above the detection limit were within the Fish Creek WWTP plume in a surface water sample collected by a canoe transect during the September 2014 sampling campaign. Three sampling sites within the plume

15 yielded water samples with NH4 concentrations high enough to determine δ NNH4 values, compiled in Table 7.3. The NH4-containing Bow River water was sampled directly downstream of the WWTP outfall and the three samples were collected approximately 10m apart from one another. The samples yielded NH4 concentrations of 0.39, 0.27, and 0.15 mg/L with downstream

15 distance from the FC WWTP outfall. The corresponding δ NNH4 values were +8.2‰, +7.5‰,

15 and +5.4‰ respectively. The δ NNH4 value from the FC WWTP was +7.0‰, measured in

October 2014, one month later than the BRB 2014 baseflow sampling. The effluent and river

15 samples were collected during different times, however all the δ NNH4 values measured in the

15 river downstream of the FC WWTP are within ± 2‰ of the δ NNH4 value measured in the final effluent, and hence it appears that these values are representative of the isotopic signature from

15 the effluent. If nitrification were occurring below the WWTP outfall one would expect δ NNH4

14 values to increase with decreasing NH4 concentration, as N is being preferentially converted to

NO3 (Wetzel, 2001).

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15 18 11 Table 7.2: δ NNO3, δ ONO3 and δ B values of surface water samples collected along the

Bow River.

Distance Site ID (km) Sep-14 Jun-15 Sep-15 Average δ15N δ18O δ15N δ18O δ15N δ18O δ15N δ18O (‰) (‰) (‰) (‰) (‰) (‰) δ11B (‰) (‰) (‰) Ronalane 571 11.5 -3.8 9.6 -3.9 10.0 -5.0 13.4 10.4 -4.3 Scandia 513 10.2 -4.1 11.5 -2.1 11.2 -4.2 15.7 11.0 -3.4 Bow City 482 9.4 -3.6 2.1 n.d. 11.0 -0.9 17.1 7.5 -2.2 Carseland 302 9.2 -5.4 8.7 -8.2 9.2 -8.2 6.7 9.1 -7.3 Below Pine Creek 228 8.0 -8.4 7.8 -9.3 10.3 -3.8 5.0 8.7 -7.2 Below Fish Creek 220 8.2 -6.1 9.0 0.1 9.3 -8.0 10.2 8.8 -4.7 Below Bonnybrook 205 8.3 -7.7 7.6 -8.5 7.5 -8.9 6.0 7.8 -8.4 Cochrane 179 4.0 -2.6 3.2 -1.5 8.0 -5.9 11.2 5.0 -3.3 Canmore- above 79 3.0 -0.7 1.7 1.1 2.8 2.4 5.6 2.5 0.9 Banff- below 64 0.7 -1.3 1.2 -0.5 0.7 2.5 9.2 0.9 0.3 Banff- above 56 2.3 0.6 1.0 0.5 1.0 1.8 9.7 1.4 1.0 Lake Louise- above 0 -0.7 5.2 0.5 10.6 -0.1 8.8 8.4 -0.1 8.2

15 + Table 7.3: Concentrations and δ N values of NH4 in surface water samples collected within the FC WWTP plume of the Bow River.

15 δ N-NH4 Location NH4 (mg/L) (‰) Below Fish Creek 1 0.39 8.2 Below Fish Creek 2 0.27 7.5 Below Fish Creek 3 0.15 5.4

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Sep-14 Jun-15 Sep-15 15

(‰) 10 NO3 N

15 5 δ

10 5

(‰) (‰) 0 NO3

O -5 18

δ -10 -15

20 15 (‰) (‰)

B 10 11 δ 5 0 LL BF CM CR Calg. CS BC SD RL 0 100 200 300 400 500 600 Distance from a point above Lake Louise (km)

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7.2.3 The Oldman River

15 18 11 The δ NNO3, δ ONO3 and δ B values for surface water samples taken along the mainstem of the OMR from Monarch (0 km) to Taber (140 km) are compiled in Table 7.4. The

δ15N and δ18O values were measured on samples taken during October 2014 and 2015 (baseflow) and June 2015 (peakflow) while δ11B values were obtained on samples collected during October

2015. Figure 7.3 plots the δ15N, δ18O and δ11B values of surface water samples with distance along the OMR.

15 The average δ N values of NO3 along the mainstem of the OMR ranged from +6.9‰ below the confluence with the Belly River (16km) to +12.8‰ at Pavan Park (52km),

15 downstream of the Lethbridge WWTP. The majority of the δ NNO3 values ranged from +10.8‰

15 to +12.8‰ with the exception of the lower δ NNO3 values measured below the confluence with

15 the Belly River. Taking into account the ±0.5‰ uncertainty associated with the δ NNO3 measurements of the water samples, samples from below the confluence of the Belly River were the only occasion with a significant decrease in δ15N values. Along the studied section of the

15 OMR the same trend in δ N-NO3 values was observed between sampling campaigns, indicating the changes in δ15N values were independent of season.

18 The average δ ONO3 values of samples from the mainstem sampling sites ranged from

+1.9‰ at Monarch (0km) to +7.6‰ downstream of the tributaries (113km). There was no

18 apparent trend in δ ONO3 values along this reach of the OMR spatially or between sampling campaigns. There was a significantly elevated δ18O value of +17.5‰ during peak run-off below

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the Belly River confluence in June 2015. The downstream tributaries and Taber sampling sites

18 did not have adequate NO3 concentrations for δ O measurements in June 2015 due to high discharge rates diluting nutrient concentrations.

The δ11B values along the mainstem of the OMR ranged from +7.3‰ at Monarch (0km) to +12.1‰ downstream of the tributaries (52km). There were increasing δ11B values from

Monarch to downstream of the tributaries, then the δ11B values remained relatively constant to a final value of +11.2‰ at Taber (140km).

15 18 11 Table 7.4: δ NNO3, δ ONO3 and δ B values of surface water samples obtained along the

Oldman River.

Location Km Oct-14 Jun-15 Oct-15 Average δ15N δ18O δ15N δ18O δ15N δ18O δ11B δ15N δ18O (‰) (‰) (‰) (‰) (‰) (‰) (‰) (‰) (‰) Taber 140 11.5 3.4 - - 10.3 8.6 11.2 10.9 6.0 D/S Tributaries 113 - - - - 10.8 7.6 12.1 10.8 7.6 Pavan Park 52 13.0 2.7 12.3 5.6 13.2 1.9 9.6 12.8 2.8 Below Belly River 16 9.6 2.1 3.2 17.5 7.8 0.1 8.9 6.9 7.1 Monarch 0 10.6 1.3 14.4 2.6 11.1 1.5 7.3 12.0 1.9

Tributaries

15 18 11 The δ NNO3, δ ONO3, and δ B values of surface water samples obtained from the Little

Bow River, Battersea Drain, and Haney Drain tributaries are summarized in Table 7.5. The

15 18 δ NNO3 and δ ONO3 values of NO3 measured at the three tributary sites are plotted for each sampling campaign of October 2014, June 2015, and October 2015 in Figure 7.4. The Little Bow

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15 18 River, the natural tributary, had the most variation in both δ NNO3 and δ ONO3 values between

15 18 each sampling campaign, with δ NNO3 values ranged from +0.4 to +9.1‰ and δ ONO3 values

15 18 from +6.1 to +25.1‰. During baseflow, the δ NNO3 values were higher and δ ONO3 values lower than during peakflow. The surface water NO3 from the Little Bow River sampling site also

15 had much lower δ NNO3 values than the other two man-made irrigation canal tributaries, an average of +4.9‰ in the Little Bow compared to +19.7‰ and +20.1‰ in the Battersea and

Haney Drains.

15 18 In the Battersea Drain δ N values of NO3 ranged from +18.2‰ to +20.9‰ and the δ O values from +1.7 to +6.5‰. This irrigation canal had the same trend as the Little Bow River,

15 18 with higher δ NNO3 and lower δ ONO3 values in surface water samples obtained during baseflow than during peakflow, but the range of values observed was not as large as in the Little Bow.

15 18 The Haney Drain δ NNO3 values of NO3 ranged from +14.9 to +20.9‰ and δ ONO3 from

-1.5 to +5.3‰. This drain displayed a different trend between seasons than the other two

15 18 15 tributaries; the highest δ NNO3 and δ ONO3 values occurred during peakflow. The δ NNO3 and

18 δ ONO3 values of NO3 from all three tributaries appeared to be affected by seasonal variations.

15 The δ NNO3 values of the tributaries were usually isotopically distinct from the surrounding

18 mainstem OMR sampling sites and the δ ONO3 values tended to overlap more with the mainstem

NO3 isotope values.

The δ11B value of surface waters from the three tributaries ranged from +16.7‰ at Little

Bow River, +18.1‰ at Battersea Drain to +22.8‰ at Haney Drain in October 2015. Further analysis is needed to determine any temporal trends in δ11B values.

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15 18 11 Table 7.5: δ NNO3, δ ONO3 and δ B values of surface water samples collected from tributaries of the Oldman River.

Location Km Oct-14 Jun-15 Oct-15 Average δ15N δ18O δ15N δ18O δ15N δ18O δ11B δ15N δ18O (‰) (‰) (‰) (‰) (‰) (‰) (‰) (‰) (‰) Little Bow River 110 5.1 12.0 9.1 6.1 0.4 25.1 16.7 4.9 14.4 Battersea Drain 98 20.0 2.0 18.2 6.5 20.9 1.7 18.1 19.7 3.4 Haney Drain 96 14.9 -1.5 24.1 5.3 21.2 3.2 22.8 20.1 2.3

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OMR Oct-14 OMR Jun-15 OMR Oct-15 Trib Oct-14 Trib Jun-15 Trib Oct-15 30 25 20 (‰) (‰) 15 NO3 N 5 1 10 δ 5 0

30 25 20 (‰) (‰) 15 NO3 10 O 8 1

δ 5 0 -5

15 (‰)

B 1

1 10 δ 5

0 0 20 40 60 80 100 120 140 160 Distance from Monarch (km)

15 18 11 Figure 7.3: δ NNO3, δ ONO3 and δ B values of surface water versus downstream distance from the Monarch sampling site along the OMR.

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7.3 Discussion

This section is a synthesis of data collected from the BRB, the OMR, some of its tributaries, and their nutrient sources. The reasons for the changes in NO3 and B fluxes discussed

15 18 in Chapter Six will be identified using the isotopic composition of nitrate (δ NNO3 and δ ONO3) and boron (δ11B) and the knowledge of the isotopic fingerprints of natural and anthropogenic nutrient end-members. The two southern Alberta river systems were chosen for isotopic analysis of NO3 and boron sources because of their contrasting land use with major urban impacts (BRB) and agricultural land use (OMR).

15 18 7.3.1 δ NNO3 and δ ONO3 Values Revealing Sources and Processes Affecting Riverine NO3

18 15 In Figure 7.4, the δ O versus δ N values of NO3 are plotted for Bow River sampling sites during baseflow and the local NO3 sources (WWTP effluent, mineral fertilizer, cow manure) are also displayed; these points are overlain by the range of isotope ratios of major NO3 sources found in this study and previously reported in the literature (Kendall et al., 2007). It must be noted that according to the δ15N and δ18O values measured in the surface water samples, no nitrate from atmospheric deposition was detected in the Bow River. In the case of the Bow River, the sites along Reach 1 fell within the values of natural soil nitrification. This supports the NO3 flux calculations from Chapter Six that suggested the major increases in NO3 load in the Bow

River were sourced from the incoming NO3 flux from tributaries. The isotopic compositions of nitrate in Reach 2 fell within the δ15N and δ18O values for sewage effluent and manure, close to the isotopic signature of the BB WWTP effluent, the largest point-source NO3-input in the Bow

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River. In combination with the NO3 flux values that had the greatest increase in Calgary below the Bonnybrook WWTP, this showed wastewater effluent was the major source of NO3 in Reach

2. However, the CM WWTP, with the third-largest NO3 flux contribution of sewage effluents, had an isotopic signal that was similar to the isotopic signatures in Reach 2 of the Bow River.

Below the CM WWTP, between the Canmore and Cochrane surface water sampling sites, there is a 2–fold increase in NO3 flux as described in Chapter Six. This change in NO3 load was for the most part due to nitrate contributions from tributaries such as the Kananaskis and Ghost River flowing into the Bow. However, since the δ15N and δ18O values in Reach 2 fall between the CM and BB WWTP effluent signals, this could suggest an NO3 load in the river influenced in part by the CM WWTP effluent. This will be further discussed in Section 7.3.3.

In Reach 3, δ15N and δ18O values of nitrate also fell in the range of isotopic values for manure and wastewater effluent close to the BB WWTP effluent isotope signal. The isotope signature of the second largest sewage effluent contributor, the Pine Creek WWTP, did not appear to affect the isotopic composition of the river markedly. This confirmed the NO3 flux trends in the river that did not increase significantly downstream of either the Fish Creek or Pine

Creek WWTPs. Reach 3 is dominated by agricultural land use. Whether the isotopic signature of nutrients in Reach 3 was influenced by agricultural nitrate sources such as livestock manure will

11 15 be investigated by incorporating δ B values since this distinction is not possible with δ NNO3

18 and δ ONO3 values alone. NO3 and B flux calculations in Reach 3 revealed by comparing NO3/B flux ratios that there is a 2 to 2.7-fold decrease in NO3/B ratio from Carseland to Bow City. Since boron is considered to be a conservative tracer, there may have been NO3-removal processes

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occurring within this 180 km distance. The δ15N values of nitrate along this reach increased slightly or remained constant (+1 to +2‰ shift) until Ronalane while the δ18O values also increased until Bow City or Scandia (+1 to +5‰ shift) and then remained constant until

Ronalane. Referring back to isotopic fractionation during N-transformations (Section 1.2.1), one

15 18 would expect δ N values to gradually increase in the remaining NO3 along with the δ O values at a rate of between 1:1 to 2:1 and an enrichment factor of -10 to -30‰ in the case of riparian denitrification, or a small -1.5 to -3.6‰ isotopic fractionation in the case of benthic denitrification or assimilation by plants or algae (Kendall et al., 2007; Sebilo et al., 2003).

15 18 Considering the increase in δ NNO3 and δ ONO3 values between Carseland and Ronalane was so small, it is not clear whether N-transformation processes, such as benthic denitrification, were removing NO3 from the surface water and more evidence is needed. An alternate explanation for the decrease in NO3 flux seen downstream of Carseland may be due to water drawdown at the

Bassano Dam, diverting water to the agricultural irrigation canals.

In Figure 7.5 the δ15N and δ18O values of the surface water nitrate and local nutrient sources along the OMR are plotted in the same fashion as in Figure 7.4. There was no evidence

15 18 for atmospheric deposition of NO3 detected by δ NNO3 and δ ONO3 values in the Oldman River or its tributaries. The isotopic composition of nitrate in Reach 1 and Reach 2 of the mainstem

15 18 OMR sites were within the range of δ N and δ O values of manure and sewage. From NO3 flux data from Chapter Six it is known that the LB WWTP effluent, in Reach 1, does not significantly change the NO3 flux in the river. The Pavan Park and Monarch mainstem sites had nitrate with

δ15N and δ18O values closest to the LB WWTP. Monarch is upstream of the LB WWTP but

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Pavan Park is the sampling site directly downstream of the plant, so similar isotopic signatures of nitrate suggested some influence by sewage effluent in this section of the river. Along the OMR basin livestock farming accounts for >50% of land-use, therefore it has also been suggested that the δ15N and δ18O values of nitrate within the river reflect a dominant manure signal rather than sewage effluent (Rock, 2005). This hypothesis will be further tested with the use of δ11B values.

Rock (2005) found that synthetic fertilizers constitute a significant nitrogen flux into the OMR

15 18 (accounting for 15-35% of N-input) but their δ NNO3 and δ ONO3 values were not clearly expressed in the mainstem of the river and the findings of this study confirm this. The reasons may be that farmers and ranchers in the region are sticking to the Alberta Agriculture guidelines for fertilizer application (AAFRD, 2004) or that there is no direct runoff of mineral fertilizer nitrate into the river (Rock, 2005). N-transformation processes do occur after ammonium fertilizer is applied to soil such as rapid N-assimilation by plants or ammonia volatilization after

15 application, both of which will enrich the remaining NH4 in N before nitrification occurs

(Freyer & Aly, 1974; Letolle, 1980). If this is the case, it is not expected to find a clear isotopic signal from synthetic fertilizers in agricultural return-flow due to its rapid uptake, cycling and associated N isotope effects impacting the nitrate in the seepage water.

The OMR tributaries had baseflow δ15N and δ18O values of nitrate that varied to different degrees between baseflow 2014 and 2015 sampling (see Section 7.2). The man-made irrigation return-flow drains downstream of Lethbridge receive irrigation return water and contain nutrient and pesticide concentrations that are much higher than elsewhere in the basin (AMEC, 2009;

Saffran, 2005). The natural streams, such as the Little Bow River, are not directly fed by

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irrigation return-flow water but also receive significant amounts of agricultural runoff. The isotope signatures of NO3 in the Haney and Battersea Drains fell within the range of values for manure and sewage effluent stated in the literature and were very close to the δ15N and δ18O values measured in the local cow manure samples (Figure 7.5). Knowing that the irrigation canals are fed by agricultural return-flow and that there is no major sewage effluent source, this is strong evidence of a manure-derived nitrate in the canals. The Little Bow River also had δ15N

18 15 and δ O values of nitrate indicative of a manure-based signal, but measurement of δ NNO3 and

18 δ ONO3 values on a sample taken in October 2015 fell within the range of NO3-based synthetic fertilizer. Urea ammonium nitrate fertilizer is the only NO3-containing fertilizer applied in the

Canadian prairies, but it is far less common than pure urea and ammonium phosphate fertilizer

(Section 2.3). Also, referring to the earlier statements of nitrogen in synthetic fertilizers being very reactive once applied to soil, this was likely not the source of the isotopic composition. The

3 Little Bow River had low water flow during baseflow (2.4 m /s), low NO3 concentrations for an agricultural tributary (0.6 mg/L), and has many riparian zones intact. This may suggest that much of the agricultural NO3 was removed by riparian N-assimilation and denitrification before it reached the surface waters of the Little Bow and thus altered the nitrate isotopic composition

(Kendall et al., 2007). Therefore, the October 2015 NO3 isotopic signal of the Little Bow River may not reflect the nitrate sources but instead the N-transformation processes in this tributary.

11 The δ B values will be used to further assess the sources of NO3 and B in this tributary.

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Figure 7.4: δ18O versus δ15N values of nitrate measured in the Bow River during baseflow, the WWTP effluents that discharge into the Bow River, local cow manure and fertilizer samples. These isotope values are overlain by previously reported values for NO3 sources

(Kendall et al., 2007).

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7.3.2 δ11B Values Revealing Sources and Processes Affecting Riverine Boron

Due to the experimental nature of working with a novel isotope tracer, δ11B analysis on various sample materials was time-consuming and only samples from the October 2015 sampling event were analyzed, chosen as the most representative of the conditions of the watersheds.

Using δ11B values as a nutrient tracer in this study enabled the distinction between sewage effluent, with δ11B values of +0.9 to +7.2‰, from cow manure with δ11B values of +23.1

15 18 11 to +27.8‰. This was not possible using the δ N and δ O values of NO3. The measured δ B values for Alberta synthetic fertilizers were between +1.0 to +8.7‰, overlapping with sewage effluent values, but distinction of these sources was possible using NO3 isotopic compositions.

The δ11B values for these nutrient end-members in the study area are compared against δ11B values previously reported in the literature and against the values measured in the OMR, its selected tributaries, and the Bow River in Figure 7.6. The δ11B values for sewage effluent from local WWTPs fell within the previously stated range of -5 to +10‰ (Barth, 1993; Tirez et al.,

2010; Vengosh et al., 1994). The values for cow manure also fell within the previously reported range of δ11B values from +23‰ (Komor, 1997) to +25 to +29‰ (Widory et al. 2004). The mineral fertilizers in this study had measured δ11B values that were depleted in 11B compared to the δ11B value of +14.8‰ for phosphate synthetic fertilizers and higher δ11B values compared to the range measured for ammonium nitrate and urea fertilizer, -2.0 to +0.7‰ (Komor, 1997). The

δ11B values for atmospheric deposition (precipitation) range in the literature between -10 to

+50‰ largely dependent on the study area. There was no δ11B value of precipitation determined

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for southern Alberta. In this study the naturally occurring boron in surface water was determined to have a δ11B value of approximately +8‰, measured at a point along the Bow River above

Lake Louise. This value is within the range of δ11B values for sewage effluent and mineral fertilizers, however there are no major anthropogenic sources believed to be impacting this site.

Natural boron is likely sourced in part from water interaction with soil, the local argillaceous limestone, or clayey till (Pawluk & Bayrock, 1969). However, δ11B measurements of the soil, local geologic material and in precipitation are required to determine how much of each are contributing to the natural δ11B value above Lake Louise.

Applying these boron isotope values to tracing nutrient sources in the Bow River, the

δ11B values with downstream distance remained rather constant in the range of approximately +8

± 4‰ from above Lake Louise to Carseland, increased to a maximum value of +17.1‰ at Bow

City, and then decreased to a final value of +13.4‰ at Ronalane (Figure 7.2). This increase in

δ11B values between Carseland and Bow City was also associated with a 1.5-fold increase in B flux as discussed in Chapter Six. This significant increase in δ11B values was likely caused by influx of 11B enriched cow manure-derived B as the land use shifts to being agriculture-dominant in Reach 3 and agricultural return-flow waters enter the Bow River (Figure 7.6). Hence, the novel B isotope tracing technique revealed the influence of agricultural return flows in this portion of the Bow River by an increase in δ11B value in Reach 3 towards a manure-influenced isotopic signature. Using NO3 isotopes, any agricultural influence in Reach 3 was masked by the

Bonnybrook WWTP effluent signal and hence was previously undetected.

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There are other significant changes in B flux in the Bow River, a 14-fold increase in B load upstream within Reach 1 between Lake Louise and Banff, a 2-fold increase between

Canmore and Cochrane (Chapter Six), and a 4.5-fold increase in B flux from Cochrane to below the Calgary WWTPs. The δ11B values measured in the river at each of these points of increased

B flux remained rather constant in the δ11B +8 ± 4‰ range. This may indicate that the major source of additional B in Reach 1 and Reach 2 was naturally occurring B from soil and bedrock weathering entering the surface water through runoff and groundwater. The δ11B values provided no clear evidence of major anthropogenic influence on the boron isotopic composition in Reach

1. However, the δ11B isotopic signatures of WWTP effluent discharging into the Bow River ranged from +0.9‰ to +2.9‰ and may have been responsible for the lower δ11B values measured below the BB and PC WWTP in Calgary, decreasing from +11.2 to +6.0‰ below BB

WWTP and from +10.2‰ to +5‰ below the PC WWTP. The change in the Bow River B flux measured between Cochrane and below the BB WWTP is larger than the ± 22% uncertainty assigned to river B load values and is a significant increase. Combined, the B flux contribution from WWTPs in Calgary is 58.2 ± 8.0 kg/d which is over 35% of the B flux measured downstream of the WWTPs of 156 ±34 kg/d. Decreased δ11B values of +5.0 to +6.7‰ were observed from below the WWTPs in Calgary to the sampling site at Carseland downstream of

Calgary. These decreased δ11B values suggest the WWTP effluent does influence the B flux in

Reach 2 and downstream to Carseland in the Bow River. The main advantage of using δ11B values as a nutrient tracer in the Bow River is that there is a distinct enrichment in 11B in Reach 3 paired with a significant increase in B load that appears to be caused by influence from animal

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manure-derived nutrient sources, unique information that can only be derived from δ11B values.

In reaches where both natural and anthropogenic sources, such as sewage effluent, may be present in the surface water it is difficult to determine the amount of influence from sewage effluent since the range of δ11B values considered to be naturally sourced in this study area is close to the δ11B signature of the WWTP effluent. Another drawback to using δ11B values as a stand-alone tracer is that there is a large overlap in isotopic values between mineral fertilizers

11 15 and sewage effluent (Fig. 7.6). Using δ B and δ NNO3 values in combination may help to resolve some of these issues.

In the case of the OMR and the selected tributaries, there is substantial agricultural activity as well as urban influence from the city of Lethbridge, such that all the δ11B values measured in surface water are considered to have some degree of anthropogenic influence. The

δ11B data from the Bow River suggested a naturally sourced boron isotope signal to be around

+8‰. The δ11B values increased from +7.3‰ at the Monarch site (0 km) in Reach 1 to +11.2‰ at the Taber site (140 km) in Reach 2. This was a significant shift in δ11B values, accompanied by a 1.5-fold increase in B flux that occurs gradually from upstream to downstream. The δ11B value measured at Monarch could indicate influence from naturally-sourced boron, WWTP effluent, synthetic fertilizer, or a mixture of these sources. The increasing δ11B values with downstream distance in the OMR indicates a shift towards a more manure-derived isotopic signal (Fig 7.6).

The Haney Drain, Battersea Drain and Little Bow River all had δ11B values more enriched in 11B than the OMR mainstem (δ11B +16.7‰ to +22.8‰), and the greatest increase in δ11B values in

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the mainstem was found between up- to downstream of the tributaries. This strongly suggests that the enrichment in 11B in OMR surface water was due to the tributary flow bringing in manure-derived nutrients with elevated δ11B values. However, the change in B load up- and downstream of the tributaries was not significant since it does not increase outside of the 22% relative uncertainty for flux values. Since the δ11B values of the tributaries feeding the mainstem of the OMR with manure-rich return-flow waters are known the % contribution of B from the tributaries can be estimated using a mass and isotope balance equation:

11 11 11 δ BDownstream= ƒ(δ BUpstream) + (1-ƒ)(δ BTributaries) (6) where ƒ is the fraction of B contribution from upstream of the tributaries, and 1- ƒ is the fraction of B contribution from the tributaries, to the downstream tributaries site. Using an average δ11B value of 19‰ for the tributaries, this suggests that 74 ± 3% of the boron was coming from upstream of the tributaries and 26 ± 1% was sourced from manure-derived nutrient sources through the tributaries. The error for these contributions is calculated using Equation (5) in

Chapter Six. To confirm this with the B flux data, more precise discharge data would be required since these estimates are close to measurement uncertainties of the B fluxes in this study.

Another significant area of B input along the OMR was between below the confluence with the Belly River (16 km) and Pavan Park (52 km) sampling sites. As discussed in Chapter

Six, there was a 50% increase in B load between the two sites, from 26.1 ± 5.7 kg/d to 39.2 ± 8.6 kg/d, that have the St. Mary River and Six Mile Coulee tributaries as well as the LB WWTP effluent discharging into the river between them. Both tributaries flow through agricultural land with irrigation and hence it is likely that manure and synthetic fertilizer may have influenced the

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boron load and δ11B value. The δ11B of the LB WWTP effluent was +7.2‰ and had a 4.8 ± 1.0 kg/d B flux into the river, accounting for approximately half of the increase in B flux calculated from up- to downstream of the WWTP. The δ11B value of OMR surface water at Pavan Park was

+9.6‰, identical to the +8.9‰ value measured upstream at the Belly River site (a ± 2‰ error is associated with the δ11B values). If there was influence by the LB WWTP effluent on boron load in the river this was not conclusive based on the δ11B values, due to similar δ11B values of different sources.

The advantage to using δ11B as an isotopic tracer of nutrient sources in the OMR is that it gives a clear indication of manure-derived nutrient sources coming from the tributaries, especially the Haney and Battersea irrigation canals. Knowing that the OMR flows through dominantly agricultural land it is possible that the δ11B values are influenced by both manure and synthetic fertilizer signals. However, the δ11B values of synthetic fertilizer, the LB WWTP effluent, and naturally occurring boron all overlap and therefore it is difficult to distinguish these sources using only δ11B values. In contrast, manure-derived nutrient sources were clearly identified by elevated δ11B values in the Haney and Battersea Drains. A surprising finding is that the δ11B values in the mainstem of the OMR, especially below the confluence with these irrigation canals, are not closer to the δ11B values of cow manure, given that it is thought to

15 11 contribute the majority of anthropogenic NO3 to the river (Rock 2004). Using δ NNO3 and δ B in combination may help to draw some final conclusions about NO3 versus boron sources in the

OMR.

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7.3.3 Combining δ15N and δ11B Values for Assessing Nutrient Sources

One of the main objectives of this study was to explore the use of δ11B values in combination with the isotopic composition of NO3 to more conclusively identify nutrient sources

15 18 compared to using NO3 isotopes on their own. The investigation of δ NNO3 and δ ONO3 values

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and δ11B values as independent tracers revealed there are strengths associated with each tracer but both also have sources that cannot be distinguished due to overlapping isotope ratios.

Therefore, the next step is to combine δ11B and δ15N values to determine if any sources can be identified more conclusively and mixing patterns between nutrient sources are revealed within the watersheds. In Figures 7.7 and 7.8 the δ11B versus δ15N values are plotted for the nutrient isotopic end-members and the surface water samples from sites along the Bow River and OMR,

11 15 respectively. Using the δ B versus δ NNO3 plot there is a clear distinction between manure- derived nutrients and sewage effluent, and mineral fertilizers have only a slight overlap in δ15N values with sewage effluent. Manure-derived nutrients and sewage effluent cannot be

15 18 distinguished using δ NNO3 and δ ONO3 while synthetic fertilizers and sewage effluent cannot

11 be distinguished using δ B values. The range of NO3 and B isotope signatures for atmospheric deposition vary widely and encompass most of the range of isotope values for both δ11B and

δ15N observed in natural systems. However, atmospheric deposition was determined not to be a major source of NO3 based on nitrate isotope values, and there was not enough evidence to determine how much B, if any, in surface water is due to atmospheric deposition.

In the Bow River sources and mixing trends are revealed more conclusively when δ11B

15 11 15 versus δ NNO3 are plotted. In Reach 1, both the δ B and δ N values measured at Cochrane

(CR) shift away from the isotopic composition of uncontaminated background or sewage effluent values slightly toward the animal manure isotopic signature (Figure 7.7). Both the B and NO3 flux data show a 2-fold increase from Canmore to Cochrane. There are some livestock operations around Cochrane (BRBC, 2005) and their effluents may contribute to a shift in the isotopic

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signal of boron and nitrate. This unique isotopic signature appears to indicate that a mix of nutrients from Canmore WWTP effluent, naturally occurring NO3 and B from soil and bedrock weathering, and some cow manure-derived nutrients influence the isotopic composition of NO3 and B of Bow River surface water at Cochrane. More analysis is required to fully quantify the mixing ratios of nutrients at this site.

In Reach 2, the δ15N versus δ18O values of nitrate in the Bow River surface waters revealed that the major input of nutrients is from the sewage effluent, mainly the BB WWTP

(Section 7.3.1). The δ15N and δ11B values concur that BB WWTP effluent is the major contributor of nutrients to the Bow River in Reach 2, paired with a 4-fold increase in B flux and a 5-fold increase in NO3 flux compared to flux data from Cochrane. The sewage effluent in

Canmore has a similar NO3 and B isotopic composition to that of the Bonnybrook WWTP that can make it appear as though Canmore WWTP effluent could also influence the nutrient load in

Reach 2 of the Bow River. However, the NO3 and B isotopic composition of Canmore WWTP effluent is not reflected at the Cochrane sampling site, in between Canmore and Reach 2.

Furthermore, the nutrient flux coming from the Canmore WWTP is only 1-2% of the flux coming from Bonnybrook WWTP and is not likely to have much influence. The second largest

WWTP along the Bow River, Pine Creek, has NO3 and B isotopic compositions very different than those measured in the Bow River, so it is not considered to currently have a significant influence on nutrient loads in the river.

In Reach 3 of the Bow River, a clear trend was found from a sewage effluent signal towards the isotope signature of nutrients from animal manure, in particular the local cow

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manure. This is supported by a 50% increase in boron load between Carseland and Bow City sampling sites in Reach 3 and then a B load that remains constant until Ronalane. NO3 flux decreases in this reach but NO3/B flux ratios revealed that there is significant N-removal either by denitrification or N-assimilation in riparian and riverine zones, or that there is NO3 removed from the river during diversion of water for irrigation at the Bassano Dam. Therefore, it appears that nitrate removal within the Bow River in concert with influx of nutrients from agricultural

11 15 return flows best explains the observed data trends. Using a combination of δ B and δ NNO3 values, there is a more clearly defined mixing trend in the river between sewage effluent and

11 15 manure-derived nutrients. The δ B and δ NNO3 values also confirms that BB WWTP is the most influential sewage effluent signal in the Bow River.

In the OMR, Reach 1 sampling points yielded samples within the range of δ11B and

15 δ NNO3 values for sewage effluent. The only sampling point in Reach 1 that lies downstream of the Lethbridge WWTP plant was Pavan Park, the point in Figure 7.8 that lies closest to the

WWTP isotopic signature. Therefore, the other two OMR sampling sites upstream must have been influenced by other nutrient sources. There is a WWTP upstream of these sites in Fort

Macleod that discharges effluent into the OMR, with a capacity for up to 10,000 L/d of effluent,

15 11 and may be influencing the δ NNO3 and δ B values of OMR surface water. It is also possible that this isotope signal may be influenced by some septic tank contamination leaching into the river through groundwater, or that there is a mixing relationship between naturally occurring

11 15 NO3 and B, synthetic fertilizer and livestock manure that results in δ B and δ N values in the same range as sewage effluent. Further investigation is required to better define these sources.

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NO3 fluxes did not significantly increase in Reach 1, and there was a significant 50% increase in

B flux from below the Belly River confluence (BL) to Pavan Park (LB) in Reach 1. This significant shift in B flux paired with the δ15N and δ11B values of Pavan Park suggested that there was influence by the LB WWTP effluent on the isotopic signature of the OMR, thus it was a significant source of nutrients. This is a conclusion that could only be derived using the δ11B

15 and δ NNO3 co-tracing technique.

In Reach 2 of the OMR, there was a shift towards the manure-derived nutrient isotope composition of NO3 and B. The mainstem of the OMR in Reach 2 is downstream of the sampled tributaries: Haney Drain, Battersea Drain, and Little Bow River, and these tributaries have

15 11 δ NNO3 and δ B values strongly indicative of an agricultural, manure-derived nutrient source.

15 11 Therefore, it is surprising to find that the δ NNO3 and δ B values of the OMR in Reach 2 did not fall clearly within the range of values typical for animal manure, such as in the tributaries.

Rather, the mainstem OMR sites in Reach 2 appeared to express mixing of sewage effluent δ15N and δ11B values with the manure isotopic compositions. There were also no significant shifts in either NO3 or B flux outside of the measurement uncertainty downstream of the tributaries in

Reach 2. The man-made irrigation return canals, Haney Drain and Battersea Drain, derived nutrients that were very close to the isotopic signature measured for stockpiled cow manure while the natural tributary, Little Bow River, appeared to have a signature that is a mixture of cow manure and fertilizer isotopic signals (Figure 7.8). This tributary also had the highest

15 variability in δ NNO3 values of nitrate between sampling campaigns. This indicates the source of nutrients in the tributary may shift depending on whether it is receiving agricultural surface

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runoff during peakflow or is dominated by groundwater influx during baseflow (Chapter Five).

Using δ11B values it was estimated that 26% of the boron load in Reach 2 of the OMR mainstem is contributed by the tributaries (Section 7.3.2). The same isotope and mass balance calculation

15 was not made for δ NNO3 because NO3 is not conservative and may be taken up or be denitrified in the riparian zones in the tributary catchments.

15 11 In summary, combining δ NNO3 and δ B values was effective in determining the significance of the LB WWTP effluent as a significant nutrient source in Reach 1 of the OMR. It also suggests that in Reach 2 there is a mixing relationship between manure-dominated nutrients from the tributaries and other nutrient sources.

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7.3.4 Summary

15 11 In summary, the major advantage of applying the δ NNO3 and δ B tracing technique to the Bow River was the ability to identify the impact of agricultural return flows in Reach 3 and determine a mixing relationship between agricultural manure-derived nutrient inputs and the BB

WWTP effluent also in this reach. In the OMR, this dual isotope tracing technique revealed some nutrient input by the LB WWTP effluent in Reach 1. The sampling sites upstream of

15 11 Lethbridge in Reach 1 appear to have the δ NNO3 and δ B values of another sewage effluent source, perhaps from upstream at Fort Macleod. It also suggests the mixing of manure-derived nutrients in Reach 2 through the irrigation canals, Haney and Battersea Drain, with other nutrient sources in the OMR. However, there are still some uncertainties that remain as to what nutrient sources are influencing some of mainstem sites along the OMR. Further analysis is needed and will be discussed in Chapter Eight.

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Chapter Eight: Conclusions and Future Work

8.1 Conclusions

15 18 Although in the past stable isotope techniques using δ NNO3 and δ ONO3 values have been applied to tracing nitrate sources and N cycling processes in the Bow and Oldman Rivers, two main watersheds that make up part of the South Saskatchewan River Basin, questions remained as identified in Chapter One. These questions are based on the difficulty in using

15 18 δ NNO3 and δ ONO3 values to identify NO3 sourced from agricultural return flows downstream of NO3 input by wastewater effluent in a watershed with mixed land use, such as the Bow River.

There are also questions about how to identify and determine the rates of N transformations in watersheds such as denitrification and N-uptake by plants and microorganisms. It has become a priority of government bodies to identify sources of nutrients, such as NO3, in freshwater systems in order to maintain integrity of aquatic ecosystems as well as protect the quality of

15 18 drinking water supplies. The main issues in using the δ N and δ O values of NO3 as an isotopic tracer of nutrient sources the overlapping δ15N and δ18O values for wastewater effluent and animal manure, as well as the N-transformations that occur in surface water and groundwater systems. These transformations, such as denitrification or N-uptake by plants and

15 18 microorganisms, alter the original δ N and δ O values of the NO3 sources and make fingerprinting of nitrate sources more challenging. The objective of this study was to address these issues by incorporating δ11B values of surface water samples as a co-tracer of nutrient sources on their own as well as in combination with δ15N and δ18O values of nitrate. In combination with other isotopic, geochemical, and hydrologic data the obtained results allowed

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to draw conclusions about major nutrient sources and transformation processes occurring in the

Bow and Oldman Rivers. The key findings of Chapters Four through Seven are summarized in the following paragraphs.

2 18 In Chapter Four discharge and water isotope values (δ H and δ O of H2O) for the Bow and Oldman rivers were discussed. It was found that for both the Bow and Oldman Rivers, peakflow occurs typically in June and baseflow from October through March. In terms of the

δ2H and δ18O values between peakflow and baseflow, the isotopic values tended to be lower during peakflow since there is a greater contribution of water from snowpack melt in the mountains in June. The higher δ2H and δ18O values during baseflow for both rivers indicate an increased groundwater influence. Along the Bow River in September 2015, downstream of

Calgary, the water isotope values deviated away from the Local Meteoric Water Line (LMWL) towards the Local Evaporation Line (LEL). This may indicate that the water discharged by tributaries into the Bow River undergo higher amounts of evaporation in summer and fall and in turn affect the δ2H and δ18O values of the mainstem. In the OMR the water isotopic compositions of samples from mainstem sites all fell along the LMWL. The prairie tributaries deviate away from the LMWL towards both the LEL and Shallow Groundwater Line (SGWL), indicating the influence of evaporation and that their flow is groundwater dominated. However, these tributaries do not significantly impact the water isotope values of the OMR.

In Chapter Five, the ion chemistry of the Bow and Oldman Rivers were investigated. It was found that the Bow River has a Ca-Mg-HCO3 dominated water type, influenced by the carbonate-rich limestone and shale geology of the region. In Reach 2 and 3, through and

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downstream of Calgary, there was an increase in TDS and an increase in Na, SO4, and Cl concentrations. The increases in Na and Cl can be attributed to anthropogenic inputs from

Calgary and increased Na and SO4 is due to groundwater influx for instance from the prairie tributaries. The mainstem of the OMR had the same Ca-Mg-HCO3 dominated surface water

chemistry as the Bow River due to similar geologic influence. There were increased Cl and SO4 concentrations in Reach 2, caused by influence from anthropogenic inputs within and downstream of Lethbridge and increased groundwater influx throughout the prairies. TDS remains constant in the mainstem of the river from Reach 1 to Reach 2 (west to east). In the tributaries, the major water types ranged from Ca-Mg-HCO3 in the Little Bow River with TDS similar to that of the OMR, to Ca-SO4 dominated water types in the Haney and Battersea Drains with TDS values up to 8 times greater than that of the OMR. The two man-made irrigation canals, Haney and Battersea Drain, had the lowest flow and greatest influence from groundwater affected by agricultural contamination, evidenced by surface water NO3 concentrations typically

>45 mg/L, that caused this distinct difference in water type. Understanding the ion chemistry of surface waters gives better insight into where the major sources of solutes may be sourced from.

In Chapter Five and Six, NO3 and B concentrations were discussed and then combined with flow data from the Bow and OMR to determine where the major NO3 and B loading was occurring. It was found that the most significant increase in NO3 loading along the Bow River occurred below the Bonnybrook WWTP, causing a 10-fold increase in NO3 flux from upstream in the Rocky Mountains to downstream of the WWTPs in Calgary. The NO3 load decreased downstream of the Bassano Dam by circa 50% and then remained constant to the mouth of the

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Bow River. The Bow River progressively accumulated B load with downstream distance. NO3/B flux ratios along the Bow River decreased by a factor of 7 from Reach 1 in the Rocky Mountains to Reach 2 through Calgary and then decreased below the Bassano Dam by a factor of 2.0 to 2.7.

This indicates that while B remained conservative in the Bow River, NO3 was not accumulating load as quickly or was decreasing due to nitrate removal processes.

The NO3 load in the Oldman River mainly remained constant throughout the sampled area besides a 1.3-fold increase below the confluence with the Belly River. This consistency suggests there are no major N-removal processes occurring in this portion of the OMR, or alternatively that the rates of NO3 input and denitrification are essentially equal in much of the river. NO3 flux additions from the Lethbridge WWTP effluent, Little Bow River, Battersea Drain and Haney Drain did not increase the riverine NO3 flux of the OMR outside of the error associated with the flux calculations. The greatest measured B flux in the OMR occurred between below Belly River to below the Lethbridge WWTP with a 1.5-fold increase, otherwise the B load remained constant within the measurement of uncertainty. The Lethbridge WWTP effluent was found to impact the B load in the river between these sites, investigated using isotopic tracers in Chapter Seven. A significant increase in B load downstream of the tributaries was not identified.

The load calculations in Chapter Six revealed where major changes in NO3 and B fluxes

15 18 11 occurred in the investigated rivers and by integrating the δ NNO3, δ ONO3, and δ B values in

Chapter Seven the sources of these load changes were determined. The δ15N and δ18O values of

NO3 were useful in determining the NO3 sources in Reach 1 of the Bow River, namely natural

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nitrification of soil N, and confirmed that the major increase in NO3 load below Bonnybrook

WWTP was due to wastewater effluents, which constitute the dominant NO3 source in the Bow

River in Reach 2 and Reach 3. δ11B values identified an impact on nutrient load of the Bow

River by agricultural return-flows in Reach 3, which remained previously undetected based on the isotopic composition of nitrate alone. Hence, nitrate in Reach 3 of the Bow River was sourced from WWTP effluents and manure-derived agricultural inputs while nitrate was also affected by nitrate removal processes as revealed by decreasing NO3/B ratios. In the OMR, the

15 18 δ NNO3 and δ ONO3 values indicated that a manure/sewage effluent-based nitrate sources were dominant. The tributaries are fed by groundwater influx and are impacted by agricultural return- flow as evidenced by δ15N and δ18O values of nitrate that mainly reflect a manure-based source.

11 15 Using a combination of δ B and δ NNO3 values revealed that in Reach 1, upstream of the sampled tributaries, nutrient input from the Lethbridge WWTP effluent affected the isotope composition of the river directly downstream of the plant. Upstream of the LB WWTP there was evidence of mixing of nutrients from another sewage effluent, perhaps from Fort Macleod, possibly other rural sources such as synthetic fertilizer or septic tank leachate, or naturally- derived nutrients. In Reach 2, the Little Bow River, Battersea Drain and Haney Drain tributaries had a strong nutrient signal indicating manure-impacted agricultural return-flow water. The δ11B values of irrigation canal water were very close to the B isotopic composition of cow manure. An estimated 26% of boron load was contributed from the tributaries to the mainstem of the OMR based on an isotope mass balance calculations using δ11B values. The remaining 74% of the

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boron load in Reach 2 is from an undefined mix of nutrients from sewage effluent and other anthropogenic or natural sources.

8.2 Implications and Future Work

11 15 18 This study was the first in western Canada to employ δ B, δ NNO3 and δ ONO3 values as co- tracers for identifying sources of nutrients such as NO3 in watersheds. The integration of these three isotope tracers allowed for the distinction between nutrient inputs from wastewater effluents and animal manure, not previously possible in other NO3 isotope tracing studies in this area. It was also determined that B appears to remain conservative in watersheds and compliments the use of NO3 isotope tracing by revealing through NO3/B flux ratios where NO3 is being lost from the surface water system. This multi-isotope approach to tracing nutrients such as NO3 in watersheds could be further developed and used to better constrain nutrient sources in other watersheds in western Canada with multiple types of anthropogenic land use, both municipal and agricultural. From the conclusions of this study several ideas for future research arose:

• In future studies, a record of δ11B values of surface water samples in both watersheds

should be established over periods of peakflow and baseflow over the span of several

years. Analyzing monthly surface water samples for δ11B values may give more insight

into potential seasonal variations in the isotopic composition. In the Oldman River,

sampling should be extended upstream of the area chosen for this study to capture any

other anthropogenic upstream sources of nutrients that may have affected NO3 and B

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loads, such as sewage effluent from other WWTPs. This may resolve the mixing

relationship that remained undefined in Reach 1 and 2 of the OMR. It may also allow for

a background δ11B value for the Oldman River to be determined. There would also be

merit in analyzing the nitrogen, oxygen, and boron isotopic compositions of WWTP

effluent of any other plants that discharge in the Oldman River, other livestock species

such as pig, poultry, and horse, and more samples of urea fertilizer.

15 18 11 • The scope of δ NNO3, δ ONO3, and δ B values of groundwater throughout Alberta

should be fully constrained by expanding the data already available through the

15 Groundwater Observation Well Network (GOWN) that has measured δ NNO3 and

18 δ ONO3 values of 186 groundwater monitoring wells throughout Alberta (Humez et al.,

2016). This would help to better understand groundwater/surface water interactions and

how this may affect the isotopic composition of surface water.

• There is still no direct evidence of whether there is denitrification occurring in the

sediments in the benthic zone of the Bow and Oldman Rivers. Further isotopic studies

15 18 investigating the δ N and δ O values of the dissolved NO3 in the water compared to the

15 18 δ N and δ O composition of the nitrous oxide (N2O) and N2 gas released during

denitrification may enable a better quantification of potential denitrification in the

watersheds.

15 18 • Another novel isotope tracing technique could be combined with δ NNO3, δ ONO3 and

δ11B values to help to fully address all the nutrient sources and transformation processes

occurring in the watershed. This could involve employing 87Sr/86Sr ratios, used in surface

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water and groundwater studies to determine groundwater contributions, water-rock interactions and mixing processes (Petelet-Giraud et al., 2009; Widory et al., 2004). By using Δ17O, the extent of oxygen isotope fractionation may be used to determine contribution by atmospheric-N. This approach has been used in recent studies to also quantify nitrification in a water body with seasonal variation (Tsunogai et al., 2016).

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References

AAFRD. (2000). Agriculture Statistics yearbook, Statistics and Data development unit,

Economics and Competitiveness division, Alberta Agriculture, Food and Rural

Development.

AAFRD. (2004). Alberta Fertilizer Guide.

Accoe, F., Berglund, M., Duta, S., Hennessey, C., Taylor, P., Van Hoof, K., & De Smedt, S.

(2008). Source apportionment of nitrate pollution in surface water using stable isotopes of N

and O in nitrate and B: a case study in Flanders (Belgium). JRC Scientific and Technical

Reports. Belgium.

Aggarwal, J. K., Sheppard, D., Mezger, K., & Pernicka, E. (2003). Precise and accurate

determination of boron isotope ratios by multiple collector ICP-MS: origin of boron in the

Ngawha geothermal system, New Zealand. Chemical Geology, 199(3), 331–342.

Aggarwal, S. K., & You, C.-F. (2016). A Review on the Determination of Isotope Ratios of

Boron with Mass Spectrometry. Mass Spectrometry Reviews, 9999, 1–21.

Agriculture and Agri-Food Canada. (1998). The Canadian System of Soil Classification.

Agriculture Canada. (1991). Soils of the Pincher Creek Crowsnest Pass Area, Alberta.

Alberta Environment. (2003). Water for Life Strategy.

Alberta Environment. (2006). Approved Water Management Plan for the South Saskatchewan

River Basin ( Alberta ). Medicine.

Alberta Environment. (2016). South Saskatchewan River Basin Water Quality Data. Retrieved

May 1, 2016, from http://www.environment.alberta.ca/apps/basins/default.aspx

170

AMEC. (2009). South Saskatchewan River Basin in Alberta: Water Supply Study. Lethbridge,

Alberta.

Appelo, C. A. J., & Postma, D. (2005). Geochemistry, groundwater and pollution (2nd ed.). A.A.

Balkema Publishers.

Aravena, R., Evans, M. L., & Cherry, J. A. (1993). Stable isotopes of oxygen and nitrogen in

source identification of nitrate from septic systems. Groundwater, 31, 180–186.

Aravena, R., & Robertson, W. D. (1998). Use of multiple isotope tracers to evaluate

denitrification in ground water: study of nitrate from a large-flux septic system plume.

Groundwater, 36(6), 975–982.

Barth, S. (1993). Boron isotope variations in nature: a synthesis. Geologische Rundschau, 82(4),

640–651.

Barth, S. R. (2000). Utilization of boron as a critical parameter in water quality evaluation:

Implications for thermal and mineral water resources in SW Germany and N Switzerland.

Environmental Geology, 40(1–2), 73–89.

Bassett, R. L., Buszka, P. M., Davidson, G. R., & Chong-Diaz, D. (1995). Identification of

groundwater solute sources using boron isotopic composition. Environmental Science &

Technology, 29(12), 2915–22.

Battaglin, W. A., Kendall, C., Chang, C. C. Y., Silva, S. R., & Campbell, D. H. (2001). Chemical

and isotopic evidence of nitrogen transformation in the Mississippi River, 1997-98.

Hydrological Processes, 15, 1285–1300.

Benkovitz, C. M., Scholtz, M. T., Pacyna, J., Tarrasón, L., Dignon, J., Voldner, E. C., …

171

Graedel, T. E. (1996). Global gridded inventories of anthropogenic emissions of sulfur and

nitrogen. Journal of Geophysical Research, 101(D22), 29239.

BRBC. (2005). A report on the state of the Bow River Basin.

Bronders, J., Tirez, K., Desmet, N., Widory, D., Petelet-Giraud, E., Bregnot, A., & Boeckx, P.

(2012). Use of compound-specific nitrogen (δ15N), oxygen (δ18O), and bulk boron (δ11B)

isotope ratios to identify sources of nitrate-contaminated waters: A guideline to identify

polluters. Environmental Forensics, 13(1), 32–38.

Casciotti, K. L., Sigman, D. M., Hastings, M. G., Bohlke, J. K., & Hilkert, a. (2002).

Measurement of the oxygen isotopic composition of nitrate seawater and freshwater using

the dentirifier method. Anal. Chem., 74(19), 4905–4912.

Chang, C. C. Y., Kendall, C., Silva, S. R., Battaglin, W. A., & Campbell, D. H. (2002). Nitrate

stable isotopes: tools for determining nitrate sources among different land uses in the

Mississippi River Basin. Canadian Journal of Fisheries and Aquatic Sciences, 59, 1874–

1885.

Chao, J. (2011). Major ion and stable isotope geochemistry of the Bow River, Alberta, Canada.

University of Calgary. M.Sc. Thesis, Department of Geoscience, University of Calgary.

Cheung, K. (2009). The Geochemistry of Shallow Groundwater and Produced Fluids Associated

with Coalbed Methane (CBM) Exploration in Alberta, Canada. M.Sc. Thesis, Department

of Geoscience, University of Calgary.

City of Calgary. (n.d.). Wastewater Treatment Process. Retrieved June 1, 2016, from

http://www.calgary.ca/UEP/Water/Pages/Water-and-wastewater-systems/Wastewater-

172

system/Wastewater-treatment-tour.aspx

Clark, I. D., & Fritz, P. (1997). Environmental isotopes in hydrogeology. CRC Press.

Duchateau, N. L., & De Bievre, P. (1983). Boron isotopic measurments by thermal ionization

mass spectrometry by using the negative BO2- ion. International Journal of Mass

Spectrometry and Ion Processes Elsevier Science Publishers B.V. Joint Research Centre,

54, 289–297.

Elliot, E. M., & Brush, G. S. (2006). Sedimented organic nitrogen isotopes in freshwater

wetlands record long-term changes in watershed nitrogen source and land use.

Environmental Science and Technology, 26(40), 2910–2916.

Environment Canada. (2016a). Historical Climate Data. Retrieved May 14, 2016, from

http://climate.weather.gc.ca/historical_data/search_historic_data_e.html

Environment Canada. (2016b). Historical Hydrometric Database. Retrieved June 7, 2016, from

https://wateroffice.ec.gc.ca/search/historical_e.html

Eppich, G. R., Singleton, M. J., Roberts, S. K., Wimpenny, J. B., Derubeis, E., Moran, J. E., …

Yin, Q.-Z. (2012). Source determination of anthropogenic NO3 in groundwater by analysis

of δ15N,δ18O, and δ11B: A case study from San Diego County, California. Groundwater

Resources Association of California, Fresno, CA., june 13(3), 11p.

Eppich, G. R., Wimpenny, J. B., Yin, Q., & Esser, B. K. (2011). Stable isotopic composition of

groundwater boron – analytical method development LLNL, 30.

Fogg, G. E., Rolston, D. E., Decker, D. L., Louie, D. T., & Grismer, M. E. (1998). Spatial

Variation in Nitrogen Isotope Values Beneath Nitrate Contamination Sources. Ground

173

Water, 36(3), 418–426.

Freyer, H. D., & Aly, A. I. M. (1974). Nitrogen-15 variations in fertilizer nitrogen. Journal of

Environment Quality, 3(4), 405–406.

Friedman, I., Smith, G. I., Gleason, J. D., Warden, A., & Harris, J. M. (1992). Stable isotope

composition of waters in southeastern California 1. Modern precipitation. Journal of

Geophysical Research, 97(D5), 5795.

Gadd, B. (1995). Handbook of the Canadian Rockies. Corax Press.

Gammons, C. H., Babcock, J. N., Parker, S. R., & Poulson, S. R. (2011). Diel cycling and stable

isotopes of dissolved oxygen, dissolved inorganic carbon, and nitrogenous species in a

stream receiving treated municipal sewage. Chemical Geology, 283(1–2), 44–55.

Gonfiantini, R., & Pennisi, M. (2005). The behaviour of boron isotopes in natural waters and in

water–rock interactions.

Grasby, S. E., & Hutcheon, I. (2000). Chemical dynamics and weathering rates of a carbonate

basin Bow River, southern Alberta. Applied Geochemistry, 15(1), 67–77.

Guerrot, C., Millot, R., Robert, M., & Négrel, P. (2011). Accurate and High-Precision

Determination of Boron Isotopic Ratios at Low Concentration by MC-ICP-MS (Neptune).

Geostandards and Geoanalytical Research, 35(2), 275–284.

Hamilton, W. N., Price, M. C., & Langenberg, C. W. (1999). Geological Map of Alberta.

Harrington, R. R., Kennedy, B. P., Chamberlain, C. P., Blum, J. D., & Folt, C. L. (1998). 15N

enrichment in agricultural catchments: field patterns and applications to tracking Atlantic

salmon (Salmo salar). Chemical Geology, 147(3), 281–294.

174

HealthCan. (2003). Guidelines for Canadian Drinking Water Quality, Health Canada.

Heaton, T. (1986). Isotopic studies of nitrogen pollution in the hydrosphere and atmosphere: a

review. Chemical Geology: Isotope Geoscience Section, 59, 87–102.

Hemming, N. G., & Hanson, G. N. (1994). A Procedure for the Isotopic Analysis of Boron by

Negative Thermal Ionization Mass-Spectrometry. Chemical Geology, 114(1–2), 147–156.

Hendry, M. J., Cherry, J. A., & Wallick, E. I. (1986). Origin and Distribution of Sulfate in a

Fractured Till in Southern Alberta, Canada. Water Resources Research, 22(1), 45–61.

Hoefs, J. (2009). Stable Isotope Geochemistry, 6th Edition (6th ed.). Springer.

Holcomb, M., DeCarlo, T. M., Schoepf, V., Dissard, D., Tanaka, K., & McCulloch, M. (2015).

Cleaning and pre-treatment procedures for biogenic and synthetic calcium carbonate

powders for determination of elemental and boron isotopic compositions. Chemical

Geology, 398, 11–21.

Holland, W. D., & Coen, G. M. (1983). Ecological ( Biophysical ) Land Classificatio n Of Banff

and Jasper National Park s Ecological ( Biophysical ) Land Classification Of Banff and

Jasper National Park s Volume I : Summary. Alberta Institute of Pedology M-83-2.

Howarth, R. W., Billen, G., Swaney, D., Townsend, A., Jaworski, N., Lajtha, K., … Zhu, Z. L.

(1996). Regional nitrogen budgets and riverine N&P fluxes for the drainages to the North

Atlantic Ocean: Natural and human influences. Biogeochemistry, 35(1), 75–139.

Hubner, H. (1981). Isotope effects of nitrogen in soils and the biosphere. Akadamie der

Wissenschaften der DDR, Leipzig, Germany.

Humez, P., Mayer, B., Ing, J., Nightingale, M., Becker, V., Kingston, A., … Taylor, S. (2016).

175

Occurrence and origin of methane in groundwater in Alberta (Canada): Gas geochemical

and isotopic approaches. Science of the Total Environment, 541, 1253–1268.

http://doi.org/10.1016/j.scitotenv.2015.09.055

Katvala, S. M. (2008). Isotope hydrology of the Upper Bow River Basin, Alberta, Canada. M.Sc.

Thesis, Department of Geoscience, University of Calgary.

Kendall, C. (1998). Tracing nitrogen sources and cycling in catchments. In Isotope tracers in

catchment hydrology (pp. 519–576).

Kendall, C., Elliott, E. M., & Wankel, S. D. (2007). The citation for the chapter is : Tracing

anthropogenic inputs of nitrogen to ecosystems, 375–449.

Kendall, C., Silva, S. R., Chang, C. C. Y., Wankel, S. D., Hooper, R. P., Frey, J. W., … Delong,

M. D. (2003). Tracing nutrient sources in the Mississippi River Basin, USA. International

Symposium on Isotope Hydrology and Integrated Water Resources Management, IAEA, 69–

70.

Komor, S. C. (1997). Boron contents and isotopic compositions of hog manure, selected

fertilizers, and water in Minnesota. Journal of Environmental Quality1, 26(5), 1212–1222.

Kratzer, C. R., Dileanis, P. D., Zamora, C., Silva, S. R., Kendall, C., Bergamaschi, B. A., &

Dahlgren, R. A. (2004). Sources and transport of nutrients, organic carbon, and

chlorophyll-a in the San Joaquin River upstream of Vernalis, California, during summer

and fall, 2000 and 2001.

Leenhouts, J. M., Bassett, R. L., & Maddock III, T. (1998). Utilization of intrinsic boron isotopes

as co-migrating tracers for identifying potential nitrate contamination sources.

176

Groundwater, 36(2), 240–250.

Lemarchand, D., Gaillardet, J., Gopel, C., & Manhes, G. (2002). An optimized procedure for

boron separation and mass spectrometry analysis for river samples. Chemical Geology,

182(2–4), 323–334.

Letolle, R. (1980). Nitrogen-15 in the natural environment. In P. Fritz & J. C. Fontes (Eds.),

Handbook of Environmental Isotope Geochemistry, Vol. 1 (pp. 407–433). Elsevier Ltd.

Levson, V. M., & Rutter, N. W. (1996). Evidence of Cordilleran Late Wisconsinan Glaciers in

the "Ice-Free Corridor"*. Quaternary International, 32(95).

Macko, S. A., & Ostrom, N. E. (1994). Pollution studies using stable isotopes. In Stable Isotopes

in Ecology and Environmental Science (pp. 45–62). Blackwell Scientific Publication.

Mayer, B., Bollwerk, S. M., Mansfeldt, T., Hütter, B., & Veizer, J. (2001). The oxygen isotope

composition of nitrate generated by nitrification in acid forest floors. Geochimica et

Cosmochimica Acta, 65(16), 2743–2756.

Mayer, B., Boyer, E. W., Goodale, C., Jaworski, N. A., Breemen, N. Van, Howarth, R. W., …

Paustian, K. (2002). Sources of nitrate in rivers draining sixteen watersheds in the

northeastern U.S.: Isotopic constraints. Biogeochemistry, 5758, 171–197.

Nadelhoffer, K., & Fry, B. (1994). Nitrogen isotope studies in forest ecosystems. In Stable

Isotopes in Ecology and Environmental Science (pp. 22–44). Blackwell Scientific

Publication.

Nativ, R., & Riggio, R. (1990). Precipitation events in the southern high plains: Meteorologic

and isotopic features. Journal of Geophysical Research, 95(D13), 22,559-22,564.

177

Nielson, G. L. (1971). Hydrogeology of the irrigation study basin, Oldman river drainage,

Alberta, Canada. In Brigham Young University Research Studies (pp. 3–98).

Nimick, D. A., Gammons, C. H., & Parker, S. R. (2011). Diel biogeochemical processes and

their effect on the aqueous chemistry of streams: A review. Chemical Geology, 283(1–2),

3–17.

OWC. (2010). Oldman River State of the Watershed Report 2010. Lethbridge, Alberta.

Palmer, M. R., Spivack, A. J., & Edmond, J. M. (1987). Temperature and pH controls over

isotopic fractionation during adsorption of boron on marine clay. Geochimica et

Cosmochimica Acta, 51(9), 2319–2323.

Panno, S. V., Hackley, K. C., Kelly, W. R., & Hwang, H.-H. (2006). Isotopic Evidence of Nitrate

Sources and Denitrification in the Mississippi River, Illinois. Journal of Environment

Quality, 35(2), 495.

Pawluk, S., & Bayrock, A. (1969). Some characteristics and physical properties of Alberta tills.

Research Council of Canada, Edmonton, Canada.

Peierls, B. L., Caraco, N. F., Pace, M. L., & Cole, J. J. (1991). Human influence on river

nitrogen. Nature, 350(6317), 386–387.

Peng, H., Mayer, B., Harris, S., & Krouse, H. R. (2004). A 10-year record of stable isotope ratios

of hydrogen and oxygen in precipitation at Calgary, Alberta, Canada. Tellus, Series B:

Chemical and Physical Meteorology, 56(2), 147–159.

Petelet-Giraud, E., Klaver, G., & Negrel, P. (2009). Natural versus anthropogenic sources in the

surface- and groundwater dissolved load of the Dommel river (Meuse basin): Constraints by

178

boron and strontium isotopes and gadolinium anomaly. Journal of Hydrology, 369(3–4),

336–349.

Pomeroy, J. W., Stewart, R. E., & Whitfield, P. H. (2016). The 2013 flood event in the South

Saskatchewan and Elk River basins: Causes, assessment and damages. Canadian Water

Resources Journal / Revue Canadienne Des Ressources Hydriques, 41(1–2), 105–117.

Powlson, D. S., Addiscott, T. M., Benjamin, N., Cassman, K. G., de Kok, T. M., van Grinsven,

H., … van Kessel, C. (2008). When does nitrate become a risk for humans? Journal of

Environmental Quality, 37, 291–295.

Rabalais, N. N., Turner, R. E., & Wiseman, W. J. (2001). Hypoxia in the Gulf of Mexico.

Journal of Environment Quality, 30(2), 320.

Robertson, G., & Groffman, P. M. (2007). Nitrogen transformations. In Soil Microbiology,

Biogeochemistry, and Ecology (pp. 341–364). Springer.

Rock, L. (2005). Assessment of the usefulness of the isotopic composition of surface water

nitrate ( δ 15 N and δ 18 O ) for identifying its sources in a watershed with differing land

use. Ph.D. Thesis, Department of Geoscience, University of Calgary.

Rock, L., & Mayer, B. (2004). Isotopic assessment of sources of surface water nitrate within the

Oldman River Basin, Southern Alberta, Canada. Water, Air, and Soil Pollution: Focus, 4(2–

3), 545–562.

Rock, L., & Mayer, B. (2006). Nitrogen budget for the Oldman River Basin, southern Alberta,

Canada. Nutrient Cycling in Agroecosystems, 75(1–3), 147–162.

Rock, L., & Mayer, B. (2007). Isotope hydrology of the Oldman River basin, southern Alberta,

179

Canada. Hydrological Processes, 21(24), 3301–3315.

Saccon, P., Leis, A., Marca, A., Kaiser, J., Campisi, L., Bottcher, M. E., … Erbland, J. (2013).

Multi-isotope approach for the identification and characterisation of nitrate pollution

sources in the Marano lagoon (Italy) and parts of its catchment area. Applied Geochemistry,

34(3), 75–89.

Saffran, K. A. (2005). Oldman River Basin Water Quality Initiative Surface Water Quality

Summary Report. Lethbridge, Alberta.

Schindler, D. W., Hecky, R. E., & McCullough, G. K. (2012). The rapid eutrophication of Lake

Winnipeg: Greening under global change. Journal of Great Lakes Research, 38, 6–13.

http://doi.org/10.1016/j.jglr.2012.04.003

Scholefield, D., Le Goff, T., Braven, J., Ebdon, L., Long, T., & Butler, M. (2005). Concerted

diurnal patterns in riverine nutrient concentrations and physical conditions. Science of the

Total Environment, 344(1–3 SPEC. ISS.), 201–210.

Sebilo, M., Billen, G., Grably, M., & Mariotti, A. (2003). Isotopic Composition of Nitrate-

Nitrogen as a Marker of Riparian and Benthic Denitrification at the Scale of the Whole

Seine River System. Source: Biogeochemistry, 63(1), 35–51.

Sebilo, M., Billen, G., Mayer, B., Billiou, D., Grably, M., Garnier, J., & Mariotti, A. (2006).

Assessing nitrification and denitrification in the seine river and estuary using chemical and

isotopic techniques. Ecosystems, 9(4), 564–577.

Sebilo, M., Mayer, B., Grably, M., Billion, D., & Mariotti, A. (2004). The use of the

“ammonium diffusion” method for δ15N-NH4+ and δ15N-NO3- Measurements:

180

Comparison with other techniques. Environmental Chemistry, 1(2), 99–103.

Seiler, R. L. (2005). Combined use of 15N and 18O of nitrate and 11B to evaluate nitrate

contamination in groundwater. Applied Geochemistry, 20(9), 1626–1636.

Sigman, D. M., Casciotti, K. L., Andreani, M., Barford, C., Galanter, M., & Böhlke, J. K. (2001).

A bacterial method for the nitrogen isotopic analysis of nitrate in seawater and freshwater.

Analytical Chemistry, 73(17), 4145–4153.

Smil, V. (1999). Nitrogen in crop production : An account of global flows adds recycled in

organic up by harvested and Quantification of N losses from crop to 26-60. Global

Biogeochemical Cycles, 13(2), 647–662.

Sosiak, A. (2002). Long-term response of periphyton and macrophytes to reduced municipal

nutrient loading to the Bow River (Alberta, Canada). Canadian Journal of Fisheries and

Aquatic Sciences, 59, 987–1001.

Stalker, A. M. (1977). The probable extent of Classical Wisconsin ice in southern and central

Alberta. Canadian Journal of Earth Sciences, 14(11), 2614–2619.

StatsCan. (2016a). Cattle inventories, by province. Retrieved July 20, 2016, from

http://statcan.gc.ca/tables-tableaux/sum-som/101/cst01/prim50a-eng.htm

StatsCan. (2016b). Service bulletin Fertilizer Shipments Survey, (21), 1–15.

Stednick, J. D. (1991). Wildland water quality sampling and analysis. Elsevier Ltd.

Sturm, P., & Knohl, a. (2009). Water vapor δ H and δ O measurements using o ff -axis

integrated cavity output spectroscopy, 2055–2085.

Tirez, K., Brusten, W., Widory, D., Petelet, E., Bregnot, A., Xue, D., … Bronders, J. (2010).

181

Boron isotope ratio (δ11B) measurements in Water Framework Directive monitoring

programs: comparison between double focusing sector field ICP and thermal ionization

mass spectrometry. Journal of Analytical Atomic Spectrometry, 25(7), 964.

Tsunogai, U., Miyauchi, T., Ohyama, T., Komatsu, D. D., Obata, Y., & Nakagawa, F. (2016).

Quantifying nitrate dynamics in lakes using triple oxygen isotopes as tracers. In

Goldschmidt Conference Abstracts (p. 3197).

US EPA. (1994). Method 200.7, Revision 4.4: Determination of Metals and Trace Elements in

Water and Wastes by Inductively Coupled Plasma-Atomic Emission Spectrometry.

USGS. (2006). National field manual for the collection of water-quality data. In Handbook for

Water-Resources Investigations. Reston, VA.

Vanderpool, R. A., Hoff, D., & Johnson, P. E. (1994). Use of Inductively Coupled Plasma-Mass

Spectrometry in Boron-lO Stable Isotope Experiments with Plants, Rats, and Humans.

Environmental Health Perspectives, 102, 13–20.

Vengosh, A., Barth, S., Heumann, K. G., & Eisenhut, S. (1999). Boron isotopic composition of

freshwater lakes from central Europe and possible contamination sources. Acta

Hydrochimica et Hydrobiologica, 27(6), 416–421.

Vengosh, A., Heumann, K. G., Juraske, S., & Kasher, R. (1994). Boron Isotope Application for

Tracing Sources of Contamination in Groundwater. Environmental Science & Technology,

28(11), 1968–1974.

Wassenaar, L. I. (1995). Evaluation of the Origin and Fate of Nitrate in the Abbotsford Aquifer

Using the Isotopes of 15N and 18O in NO3. Applied Geochemistry, 10, 391–405.

182

Wetzel, R. G. (2001). Limnology : lake and river ecosystems (3rd Ed.). Academic Press.

WHO. (1996). Guidelines for drinking water quality, 2nd edition, Volume 2 - Health criteria nad

other supporting information.

Widory, D., Kloppmann, W., Chery, L., Bonnin, J., Rochdi, H., & Guinamant, J. L. (2004).

Nitrate in groundwater: An isotopic multi-tracer approach. Journal of Contaminant

Hydrology, 72(1–4), 165–188.

Widory, D., Petelet-Giraud, E., Brenot, A., Bronders, J., Tirez, K., & Boeckx, P. (2013).

Improving the management of nitrate pollution in water by the use of isotope monitoring:

the δ15N, δ18O and δ11B triptych. Isotopes in Environmental and Health Studies, 49(1),

29–47.

Widory, D., Petelet-Giraud, E., Negrel, P., & Ladouche, B. (2005). Tracking the sources of

nitrate in groundwater using coupled nitrogen and boron isotopes: A synthesis.

Environmental Science and Technology, 39(2), 539–548.

Xiao, J., Jin, Z. D., Xiao, Y. K., & He, M. Y. (2014). Controlling factors of the δ11B-pH proxy

and its research direction. Environmental Earth Sciences, 71(4), 1641–1650.

Xiao, Y. K., Vocke, R. D., Swihart, G. H., & Xiao, Y. (1997). Boron volatilization and its

isotope fractionation during evaporation of boron solution. Analytical Chemistry, 69(24),

5203–5207.

183