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A toxicological and chemical evaluation of agricultural runoff discharged into the Nicomekl River, throughout… McLeay, Michael James 1998

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A TOXICOLOGICAL AND CHEMICAL EVALUATION OF AGRICULTURAL RUNOFF DISCHARGED INTO THE NICOMEKL RIVER, THROUGHOUT ONE GROWING SEASON by MICHAEL JAMES MCLEAY B.Com., The University of British Columbia, 1992 B.Sc, The University of British Columbia, 1995 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF APPLIED SCIENCE In THE FACULTY OF GRADUATE STUDIES (Department of Civil Engineering) We accept this thesis as conforming to the required standard THE UNIVERSITY OF BRITISH COLUMBIA April 1998 © Michael James McLeay, 1998 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department of c i v i l K n g - i n ^ - r i n g The University of British Columbia Vancouver, Canada Date A p r i l 28, 1998 DE-6 (2/88) II ABSTRACT Agricultural pesticide and manure use on commercial vegetable and blueberry farmlands bordering the Nicomekl River, Surrey, B.C., creates the potential for toxic effects on the biota within the drainage ditches and receiving waters. To investigate this possibility, water samples were collected from six drainage ditches and four river locations every three weeks between May and November, 1997. For each of the 85 water samples collected, chronic (7 ± 1 day) survival and reproduction of the cladoceran test organism Ceriodaphnia dubia was determined and compared to that for samples of river water collected upstream of the area of concern. Throughout the 6-month monitoring period, none of the 35 riverwater samples tested exhibited statistically significant mortality, and only two of the 50 ditchwater samples obtained were lethally toxic, with 6-day LC50's of 39.9% and 36.5%. For the remaining 83 water samples, C. dubia reproduction in 5 ditchwater and 5 riverwater samples, from the region of concern, was statistically lower than that in samples of upstream river water. One of the sublethally toxic riverwater samples exhibited toxic responses (paralysis) characteristic of organophosphorous (OP) pesticide contamination, and was collected immediately downstream of a lethally toxic ditch which had discharged within the previous 24 hours. Another sublethally toxic ditch sample had a total ammonia concentration of 10.8 mg/L NH3-N which was believed to be responsible its observed C. dubia reproduction inhibition. This ditch discharged for minimal durations (1-2 hours/day) in the days prior to and following its confirmed NH 3 contamination. In the immediate vicinity of these discharges, ammonia concentrations within the river may have exceeded acute Canadian water quality guidelines for ammonia intended to protect sensitive fish and invertebrate species from acute toxic effects. Further downstream, fully-mixed ammonia concentrations should not have exceeded these acute guidelines, but may have exceeded safe chronic exposure levels. The remaining 24 ditchwater and 18 riverwater samples which were measured for total ammonia all had NH 3 concentrations well below the water quality guidelines for chronic exposure. A biological toxicity identification evaluation (TIE) using piperonyl butoxide (PBO) determined that the toxicant(s) in each of the two ditchwater samples which proved lethal to C. dubia were metabolically active OP insecticide(s). Solid phase extraction full-ion-scan gas chromatography mass spectroscopy (GC/MS) analyses performed on the lethal and some of the sublethal samples immediately following observed toxicity were unable to detect the presence of OPs . Liquid-liquid extraction select-ion-scan G C / M S of the lethal samples detected 0.02 - 0.03 ^g /L diazinon (OP) in each of the two acutely lethal samples, and 0.03 ^g /L chlorpyrifos (OP) and 3 /u.g/L prometryn (herbicide) in one of the lethal samples; even though there was evidence of O P insecticide losses during the frozen storage of these samples before their chemical analyses. C. dubia bioassays using portions of the thawed samples used for these later chemical analyses exhibited lesser toxicity relative to that for the fresh samples. Consideration of the analytical values for diazinon and chlorpyrifos together with the toxicity values for these pesticides, determined as part of this investigation and by other researchers, led to the tentative conclusion that diazinon and/or chlorpyrifos were responsible (or at least partly so) for the observed toxic effects. Prometryn is appreciably (i.e., four orders of magnitude) less toxic than either of these two O P pesticides. Diazinon's American suggested acute water quality criteria of 0.08 /^g/L was possibly exceeded in the Nicomekl River during the recorded discharge of one of the two O P contaminated ditches, in the days prior to its confirmed lethal toxicity. Five ditch sediments and three river sediments were collected from the study site in October, 1997, in order to appraise their toxicity to benthic invertebrates. Chronic (14-day) survival and growth inhibition to the amphipod test organism Hyallela azteca for each sample was compared to that for sediment collected from the upstream Nicomekl site. None of sediments collected within the region of agriculture under investigation showed statistically-lower survival relative to that for river sediment collected upstream of the region of concern. One drainage ditch sediment statistically inhibited H. azteca growth. Using a standardized (Environment Canada) laboratory test method and upstream river water as the control water, the majority of ditchwater and riverwater samples collected at 3-week intervals throughout this 6-month monitoring study were not acutely or chronically toxic to Ceriodaphnia dubia. This study did identify that there can be occasional municipal pumping of toxic agricultural runoff waters into the Nicomekl River in the summer months. However, overall, the study site's 1997 agricultural activities and drainage ditch discharges should not have had a significant toxic effect on the biota of the Nicomekl River. iv TABLE OF CONTENTS Page ABSTRACT ii LIST OF TABLES viii LIST OF FIGURES xi ACKNOWLEDGMENTS xii 1.0 INTRODUCTION 1 1.1 Study Objectives 2 1.2 The Nicomekl River and Watershed 4 1.2.1 River Modifications 4 1.2.2 River Flow 6 1.2.3 Dissolved Oxygen Concentrations 10 1.2.4 Possible Sources of Nicomekl River Contaminants 14 1.2.5 Nicomekl River Fish Resources 15 1.3 Study Area 19 2.0 AGRICULTURAL LAND USE IN THE STUDY AREA 21 2.1 Predominant Crops Grown 21 2.2 Drainage Ditch Water Management 21 3.0 POSSIBLE SOURCES OF TOXICITY IN THE STUDY AREA 24 3.1 Routes of Pesticide Entry into the Nicomekl River and Drainage Ditches . 25 3.2 Predominant Pesticides Used in the Study Area 26 3.3 Literature Review of Pesticide Toxicity 34 3.3.1 Acute Lethality to Fish and Invertebrates of Pesticides Likely Used on the Nicomekl Farmlands 34 3.3.2 Sublethal Effects of Pesticides to Fish 40 3.3.3 Sublethal Effects of Pesticides to Invertebrates 43 3.3.4 Environmental Factors Affecting Pesticide Toxicity 47 3.3.5 The Chemistry, Use, and Persistence of Diazinon 49 3.3.6 The Chemistry, Use, and Persistence of Chlorpyrifos 51 3.3.7 The Chemistry, Use, and Persistence of Prometryn 52 3.4 Ammonia Contamination 53 3.4.1 Farm Animals 53 3.4.2 Fertilizer 56 V Page 3.5 Literature Review of Ammonia Toxicity 58 3.5.1 Terrestrial Sources and Speciation of Ammonia in the Aquatic Environment 58 3.5.2 Fate of Ammonia in the Aquatic Environment 58 3.5.3 Toxicity to Fish 60 3.5.4 Toxicity to Invertebrates 65 3.6 Pesticide, Ammonia, and Metal Water Quality Guidelines for the Protection of Aquatic Life 68 3.7 The Relevance of the C. dubia Test 72 4.0 REVIEW OF PREVIOUS CHEMICAL AND BIOLOGICAL TESTING IN THE STUDY AREA 73 5.0 SAMPLING PROGRAM AND EXPERIMENTAL METHODOLOGY 83 5.1 Ditch and River Water Sampling Locations and Frequency 83 5.2 Ditch and River Water Sample Collection, Transport, and Storage 87 5.3 Water Toxicity Testing with Ceriodaphnia dubia 88 5.3.1 C. dubia Culture 88 5.3.2 Initial Full Strength (100%) Ditch and River Water Sample Chronic Testing for Inhibition of Survival and Reproduction . . . 89 5.3.3 Dilution Series Chronic Testing of Lethally Toxic Samples . . . . 92 5.3.4 Lethally Toxic Samples Biological Toxicity Identification Evaluation (TIE) 92 5.3.5 Toxicity Tests on Diazinon, Chlorpyrifos, and Prometryn 94 5.3.6 Reference Toxicant Testing.and Culture Health 96 5.3.7 Test Endpoints and Statistical Analyses 97 5.4 Chemical Analyses of Ditch and River Water Samples 98 5.4.1 Organics 99 5.4.2 Metals 102 5.4.3 Total Ammonia 103 5.4.4 Dissolved Total/Inorganic/Organic Carbon 103 5.4.5 Dissolved oxygen, pH, Conductivity, Hardness, and Colour . . 104 vi Page 5.5 Sediment Sampling Locations 104 5.6 Sediment Sample Collection, Transport, and Storage 106 5.7 Hyallela azteca Chronic Sediment Toxicity Testing 106 5.7.1 Test Method 107 5.7.2 Test Endpoints 108 5.8 Chemical Analyses of Sediment Samples 108 5.8.1 Percentage Organic Matter 108 5.8.2 Metals 108 6.0 RESULTS AND DISCUSSION 109 6.1 Water Samples' Ceriodaphnia dubia Chronic Toxicity Test Results . . . . 109 6.1.1 Lethally Toxic Samples 114 6.1.2 Sublethally Toxic Samples 115 6.1.3 C. dubia Culture Health and Reference Toxicant Evaluation . . 117 6.2 Lethally Toxic Samples' Biological Toxicity Identification Evaluation . . . 119 6.3 Ditch and River Water Samples' Chemical Analyses 123 6.3.1 Organics 123 6.3.2 Metals 131 6.3.3 Total Ammonia 134 6.3.4 Dissolved Total/Organic/Inorganic Carbon, pH, Conductivity, Hardness, and Colour 136 6.4 Sensitivity of C. dubia to Detected Pesticides 136 6.5 Relationship Between Rainfall and Toxic Samples 139 6.6 Dilution Calculations for Discharge from Contaminated Ditches 140 6.7 Hyallela azteca Sediment Toxicity Tests and Sediment Chemistry 152 6.8 Summary 160 7.0 GENERAL CONCLUSIONS AND RECOMMENDATIONS 167 8.0 SUGGESTED FUTURE STUDIES 169 REFERENCES 171 VII Page APPENDIX 1 Study Site Drainage Ditches' Municipal Pump Station Records 188 APPENDIX 2 Statistical Comparison of C. dubia Chronic Survival and Reproduction, Between Samples and Controls 198 APPENDIX 3 C. dubia Successive Reference Toxicant Results 203 APPENDIX 4 C. dubia Lethally Toxic Samples' GC/MS Chromatograms, Full-lon-Scan, 500X Concentration 204 APPENDIX 5 Prometryn GC/MS, Full-lon-Scan, Standards' Curves 205 APPENDIX 6 C. dubia Lethally Toxic Samples' GC/MS Chromatograms, Select-lon-Scan, 3600X Concentration 206 APPENDIX 7 C. dubia Lethally Toxic Samples' Identified Diazinon and Chlorpyrifos Select-lon-Scan Chromatogram Peaks and Mass Spectra 208 APPENDIX 8 Diazinon and Chlorpyrifos Standards' Select-lon-Scan Chromatogram Peaks and Mass Spectra 211 APPENDIX 9 Diazinon's Select-lon-Scan GC/MS Standards' Curve 213 APPENDIX 10 Chlorpyrifos'Select-lon-Scan GC/MS Standards'Curve 214 APPENDIX 11 C. dubia Chronic Test, Ditch, River, and Control Water Samples'Chemistry 215 APPENDIX 12 South Cloverdale Ditch .& Rapid Mixing of South Cloverdale Ditch Water Into the Nicomekl River (Photos) 220 APPENDIX 13 Sediment Bioassays' Initial and Final Overlying Water Chemistry 221 VIII LIST OF TABLES Page Table 1. Historically Reported Dissolved Oxygen Concentrations (mg/L and % saturation) in the Nicomekl River and Two of its Large Drainage Ditches 13 Table 2. Quantities of Pesticides Used in the Lower Mainland in 1991 for Commercial Agricultural Purposes 27 Table 3. Most Probable Pesticides Used on the Nicomekl Farmlands During the 1997 Growing Season 29 Table 4. Increased Pesticide Usage in The Lower Mainland between 1991 and 1995 for Those Pesticides Likely Used on the Nicomekl Farmlands 32 Table 5. Acute Lethality of Probable Pesticides Used on the Nicomekl Farmlands During the 1997 Growing Season 35 Table 6. Distribution of Farm Animals in the Lower Fraser Valley 54 Table 7. Acute lethality of unionized ammonia to 31 invertebrate species 65 Table 8. Current Canadian Maximum Acceptable Concentrations (MAC's) for Pesticides 68 Table 9. U.S. Suggested Water Quality Criteria for Various Insecticides 69 Table 10a. Maximum 1 -Hour Average Total Ammonia Concentration for the Protection of Salmonids and Other Cold Water Species 70 Table 10b. Maximum 4-Day Average Total Ammonia.Concentration for the Protection of Salmonids and Other Cold Water Species 70 Table 11. Maximum Acceptable Aqueous Total Metal Concentrations 71 Table 12. Historical total ammonia measurements in the Nicomekl River 82 Table 13. Specific Locations of Water Sampling Sites 85 Table 14. Dates and Respective Sites of Water Sampling 86 Table 15. Ditch and River Water Samples' C. dubia Chronic Survival and Reproduction Test Results 109-113 ix LIST OF TABLES Cont. Page Table 16. Lethally Toxic Samples' C. dubia Dilution Series Test Endpoints . . . . 114 Table 17. Samples Exhibiting Sublethal Toxicity, where Reproduction was Less than Upstream Control 116 Table 18. Results of C18 C. dubia Biological Toxicity Identification 119 Table 19. Results of 96-h C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at -10 °C for Two Months in Plastic Bottles) 120 Table 20. Results of 96-h C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at 4°C in glass for 5 months) 121 Table 21. Results of 7-day C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at 4°C in glass for 5 months) 122 Table 22. ICP Metal Scan Results for Toxic, Non-Toxic, and Control Samples 132 Table 23. Total Ammonia of All Water Samples, on Dates Exhibiting Toxicity in Select Samples 135 Table 24. C. dubia Lethality to Detected Pesticides 137 Table 25. Diazinon and Chlorpyrifos C. dubia Chronic Test (6-day) Reproduction Inhibition Test Endpoints 138 Table 26. Sediment Samples Appearance, Odour, and Visible Indigenous Invertebrates, Prior to Toxicity Testing 152 Table 27. Sediment Bioassays % Survival and Growth 153 Table 28. T-test Results for Hyallela azteca 14 Day Survival and Growth in Ditch and River Sediment Samples Versus the Upstream Sediment Sample, Plus Sediments' Organic Matter (%) 154 Table 29. Sediment Sample's Measured Dry Weight Metal Content 156 Table 30. MESS-2 Reference Sediment's Dry Weight Metal Content, Author's Results Compared to NRC's 157 X LIST OF TABLES Cont. Page Table 31. Sediment Sample's Hypothetical True Dry Weight Metal Content . . . 158 Table 32. Fraser River Sediment Background Trace Metal Concentrations, Canadian Sediment Quality Guidelines Trace Metal TEL, and Other Studies Nicomekl Sediment Metal Concentrations 159 xi LIST OF FIGURES Page Figure 1. Nicomekl River Location 5 Figure 2. Rainfall Near the Study Area, May to November, 1997 7-8 3 rcl Figure 3. Daily Nicomekl River Flow (m Is) at 203 St., 3 km Upstream of the Study Area From May to October, 1997 9 Figure 4. Free-Flowing Artesian Wells in the Study Area 11-12 Figure 5. Salmonid Spawning Areas in the Nicomekl River and its Tributaries . . 16 Figure 6. Study Site 20 Figure 7. Predominant Crops Grown in the Study Site, July, 1997 22 Figure 8. Fur Farms and Henneries Located in the Study Site and its Uplands . 55 Figure 9. Excess Nitrogen Application in the Nicomekl-Serpentine Basin 57 Figure 10. Study Site Water Sampling Locations and Site Numbers 84 Figure 11. Study Site Sediment Sampling Locations and Site Numbers 105 Figure 12a. Ammonia Contaminated Old Logging Ditch, Pumped Discharge (hours/day) Following Heavy June rainfall 141 Figure 12b. OP Insecticide Contaminated South Cloverdale Ditch, Pumped Discharge (hours/day) Following Heavy July Rainfall 142 Figure 13a. Minimum and Estimated Maximum Nicomekl River Flows at the Old Logging Ditch, Between Rainfall and Observed Ammonia Contamination 144 Figure 13b. Minimum and Estimated Maximum Nicomekl River Flows at South Cloverdale Ditch, Between Rainfall and Observed OP Contamination 144 Figure 14. Estimated Peak Daily Total Ammonia Concentrations in the Nicomekl River During Discharge of the Old Logging Ditch at 10.8 mg/L Total Ammonia, in June, 1997 146 Figure 15. Estimated Peak Daily Concentrations of Toxic South Cloverdale Ditch Water in the Nicomekl River During July 1997 149 XII ACKNOWLEDGMENTS I would like to express my utmost gratitude to Dr. Ken Hall for his support and advice throughout the research and reporting phases of this study; Dr. Howard Bailey of EVS Consultants for providing information regarding insecticide toxicity identification and critiquing this document; and Professor Jim Atwater for also reviewing this thesis. The financial support for this work was provided by the Natural Sciences and Engineering Research Council of Canada. The Civil Engineering Environmental Laboratory staff were also instrumental. Paula Parkinson provided invaluable laboratory assistance and expertise for the GC/MS, ammonia, and sediment chemical analyses performed. Susan Harper procured all necessary supplies. B.C. Research graciously provided the initial C. dubia culture, Hyallela azteca test organisms, and food for the cultures and bioassays. I wish to thank Janet Pickard for facilitating this contribution, and Patricia Keen for occasionally collecting the White Rock pond culture/control water. Special thanks go to Karen Kinnee, for assisting with the water sampling, calculating the LC50's, and for locating several key references. This thesis is dedicated to my Father, Dr. Donald J. McLeay, whose encouragement, advice, and opinions were greatly appreciated. 1 1.0 INTRODUCTION The Nicomekl River is economically and recreationally valued by commercial and sport anglers for its fisheries resources. The use of pesticides (herbicides, fungicides, and insecticides) and manure (ammonia source) on the farmlands immediately adjacent to the Nicomekl River creates the potential for aquatic toxicity in its runoff ditches and receiving waters. Prior to this study, the frequency and severity of pesticide- and ammonia-contaminated agricultural discharges in this region had not been adequately investigated. Efforts to identify sporadic contamination events of these degradable compounds are very labour intensive and require sampling at numerous locations, at regular intervals, over a lengthy period. Toxicity tests with samples of runoff can be used to assess potential ecological impacts of non-point-source agricultural pollution on aquatic communities, because the organism response integrates the bioavailability and combined effects of multiple water quality parameters. Insecticides pose the greatest threat of toxicity due to the extremely low concentrations of this group of pesticides which produce adverse biological effect. The decreased use of organochlorine insecticides due to their persistence in the environment and tendency to bioaccumulate in organisms, and bioconcentrate in food chains, has increased the quantities of organophosphorous (OP) and carbamate insecticides being applied. Compared to organochlorines, these latter compounds generally have higher water solubilities and a lesser affinity to bind to soils and sediments. Research has shown that insecticide concentrations in receiving waters can be elevated to levels causing acute sublethal and lethal effects following rainfall events and runoff from farmland to which these types of insecticides have been applied (Matthiessen et al., 1995; Wood, 1997). 2 Numerous pesticides have been identified in water and sediment samples collected from the Nicomekl system (Wan, 1989; Coastline, 1989; EVS, 1993; Wan et al., 1994; Wan et al., 1995). EVS (1993) identified a downstream change in benthic invertebrate species diversity within the region of the Nicomekl River most inundated with agricultural runoff ditches. Agricultural pesticide and manure use on the farm lands in this region may be adversely impacting ditch and river aquatic biota. Environment Canada's standardized biological test methods for testing the toxicity of water (EC, 1992) and sediment samples (EC, 1996), using the organisms Ceriodaphnia dubia and Hyallela azteca, respectively, were utilized to determine the significance of agricultural contamination of the Nicomekl system. 1.1 Study Objectives The goal of this study was to determine whether the runoff waters and sediments collected by the drainage ditches transecting vegetable farmlands, during and following one growing season were contaminated due to agricultural activities, and if so, to determine the degree and frequency of contamination, source of toxicity, and whether polluted waters were being discharged into the Nicomekl River. This study enabled the development of overall conclusions on the toxic impact of agricultural pesticide and manure use on the Nicomekl River's sensitive aquatic biota to be made, and has relevance to other rivers and streams transecting vegetable farmlands in the Lower Fraser Valley. Given that the majority of the agricultural discharges into the Nicomekl are flow controlled, this study hoped to provide insight to whether measures should be undertaken to treat this runoff or restrict its discharge to periods of appreciable river flow to ensure adequate mixing and flushing from the system. The specific study objectives are summarized on the following page: To determine if agricultural runoff in the drainage ditches discharging to the Nicomekl River (or being retained for future discharge), and the river itself in the vicinity of these outflows, exhibits lethal or sublethal effects to Ceriodaphnia dubia using the 7 ± 1 day chronic test. To pinpoint the cause of observed C. dubia toxicity (pesticides, ammonia, or metals) by performing chemical analyses and biological toxicity identification on toxic samples. To compare the concentrations of detected toxic contaminants to the concentrations reported in the literature that directly or indirectly adversely impact salmonid and non-salmonid fish and other sensitive aquatic life. To comprehensively monitor toxicity in the.Nicomekl system with C. dubia from Spring to Fall, in an attempt to determine the months and rainfall patterns of greatest potential adverse impacts in future growing seasons. To attempt to ascertain whether observed pesticide toxicity is likely due to runoff from crop lands following rainfall, or pesticide over-spray, dumping, or washing of spray equipment directly into the Nicomekl River and/or its drainage ditches. To determine C. dubia test toxicity endpoints for laboratory prepared solutions of any identified pesticides found in the system, using upstream Nicomekl River water, and compare these test endpoints to those of toxic samples, as well as the concentrations of pesticide(s) in toxic samples determined by gas chromatography mass spectroscopy. To evaluate the overall impact of agricultural pesticide and manure use, and runoff ditch water management practices, on the aquatic biota in the Nicomekl River, based on dilution calculations of discharged toxic waters. To determine if sediments from the drainage ditches discharging to the Nicomekl River, and the Nicomekl River sediments themselves, are toxic to the sediment organism Hyallela azteca, i.e., are pesticides concentrating in the river and ditch sediments during the growing season to a degree negatively impacting aquatic sediment biota? To collect water and sediment chemistry data for the Nicomekl system and its drainage ditches, for reference in future studies. 4 1.2 The Nicomekl River and Watershed 2 The Nicomekl River and Serpentine River are located in the 322 km Nicomekl-Serpentine drainage basin (Halstead, 1978) 24 km southeast of Vancouver, in the municipalities of Langley and Surrey, B.C. (Figure 1). In their lower reaches, both rivers share a main valley, a former embayment of the ocean, which extends 11 km eastward from Mud Bay to Cloverdale and varies in width from 4 km to 5 km (Halstead, 1978). Mud Bay is the northeasterly extension of Boundary Bay, which faces the southern portion of the Strait of Georgia. The Nicomekl River originates 4 km east of Langley near 232 n d St. and 52 n d Ave., and flows for approximately 34 km before entering Mud Bay (Swain and Holms, 1988a). Including its two major tributaries, Murray and Anderson Creeks, 12 and 15 km 2 2 in length, each with drainage areas of 27.1 km and 24.7 km , respectively, the Nicomekl 2 River has a total drainage area of 149 km (Swain & Holms, 1988b). 1.2.1 River Modifications Agricultural land use in the Nicomekl Basin lowlands has altered the river from its natural state in many ways. To prevent salt water intrusion at high tide, and guarantee fresh water for crop irrigation, tidal gates were built near the mouth of the Nicomekl River in 1912, and rebuilt between 1972 and 1975 (Halstead, 1978). The tidal gates are located where highway 99A crosses the river. The gates are opened passively by water pressure, when water-levels on the river side exceed water-levels on the ocean side (i.e. at low tide, or following periods of heavy rainfall). Tidal gates are typically open from 1 - 9 hours per day, but may be closed for up to 8 days at a time (Town, 1986). To prevent farm lands from flooding while the tidal gates are closed and river water-levels rise, the river's banks were dyked, and natural canopy riparian vegetation removed and replaced with grasses. 5 6 1.2.2 River Flow With the exception of the area around Anderson Creek, the majority of the Nicomekl River's drainage basin is underlain with stoney marine clays and thought to be relatively impervious to groundwater intrusion. Consequently, drainage in the region is primarily by surface water runoff (Swain & Holms, 1988b), resulting in a "flashy" Nicomekl River flow regime following periods of heavy rainfall. For 1987, the Greater Vancouver Regional District (GVRD, 1988) estimated that runoff into the Nicomekl River from agricultural areas 3 3 alone was on average 112,300 m /day or 41.0 million m /year. Historical average daily 3 river flows, measured 0.5 km downstream of Anderson Creek have ranged from 0.13 m /s 3 to 35.4 m Is (Swain and Holms, 1988b). Two and ten year 7-day average low flows in the 3 3 same river location were 0.24 m /s and 0.13 m /s, respectively (Swain and Holms, 1988b). Figure 2 shows the amount of rainfall in East Cloverdale, a region just North of the Nicomekl River, between May and November, 1997 (EC, 1997). Figure 3 shows the daily rd river flow at 203 St. (the only remaining Environment Canada Water Survey flow station) 3 from May to October, 1997. River flows during this period ranged from 0.4 m /s to 15.1 m 3/s(EC, 1998). Hydrogeological investigations have shown that there is a major groundwater flow beneath the Nicomekl-Serpentine basin (Halstead, 1978). The groundwater recharge areas are composite, and include distant recharge in the central and eastern Fraser Valley, and local recharge from the Clayton and Langley uplands, including a 6.4 km portion of Anderson Creek (Halstead, 1978). The stratigraphy of the Nicomekl-Serpentine Basin is that of silty clay, silty sand, sandy silts and sand lenses of fluvial, glaciofluvial and glaciomarine origin, which provides leaky conditions in the discharge zones of a major groundwater flow 7 Figure 2. Rainfall Near the Study Area, May to November, 1997 (EC, 1997). 40 l l s o c o 120 f i o Precipitation Record Cloverda le East , B C X* • Sampling j | | | M ! i ! j j - j , i I • -M • '•I j I I I i i 1 t r -—4-! ! | j - 1 I i I i I i ; i • i | ! ' ! f i ! 1 n • : 1 —1 -U M l . • h I n | | 1 • 1 3 S 7 9 11 13 15 17 1£ 21 23 25 27 29 31 May, 1997 40 E E.30 c o I IS 20 f i o Precipitation Record Cloverda le East , B C X = Sampling | , i JL.i ! ! ; i T" I i 1 j . | 1 | i (1 | X .... t 1 3 5 7 9 11 13 IS .17 IS 231 23 2i 27 29 31 July, 1997 40 |I30 c O 1120 I10 Precipitation Record Cloverdale East , B C X = Sampling 1 1 i ! ; I ! 'l | 1 — j \ | : , - | •I iji J j | • i 1 1 I , j ~ n j | i f — s I j i 1 ; ' I •t~" 1 I  n n •—j— n It 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 September, 1997 40 E" E,30 c o is 20 f i o o. Precipitation Record Cloverdale East , B C X = Sampling ] | I I l i i i i i I • I n I i T I i n i l n 1 o_ n n n 1 3 5 7 9 11 13 15 17 19 21 23 2S 27 2S June, 1997 Precipitation Record Cloverdale East , B C X = Sampling An I I I I I E E,30 B j O •520 a. I10 j V III tm a. j | _ —, n n I j | e 1 3 5 7 9 11 13 15 17 19 21 t August, 1997 23 25 27 29 31 t Precipitation Record Cloverdale East , B C X ~ Sampling l J n f l l L 1 1 1 I 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 31 t October, 1997 t 8 Figure 2 Cont. Precipitation Record Cloverdale East , B C f = Sampling An E E^n • .... i i L_ j i | I | I ! I I ! • ( i i ! i e i I I 1 i J i | o i _ i ! i i , 1 i | £ f 10 0. I i -rr ml r " i I dJ I J _ i i uk - 1 L i | •U El i •j f - Ji . J 1 ul .t I t 11 1 I i • 1.1 n -1 3 5 7 9 11 13 15 17 19 November, 199 21 't 23 25 27 29 Figure 3. Daily Nicomekl River Flow (m3/s) at 203 r d St., 3 km Upstream of the Study Area From May to October, 1997 (EC, 1998). Nicomekl River Flow May, 1997 Y = Sampling 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 31 t May 16 14 -3T12 <|10 i l 4 2 0 Nicomekl River Flow June, 1997 y = Sampling i j I • r" ... \ r ' : < \ 1 \ 4 -1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 June f Nicomekl River Flow July, 1997 y' = Sampling 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 31 t July 16 14 1»12 |10 il 4 2 0 Nicomekl River Flow August, 1997 J = Sampling U - i . 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 31 f August f 16 14 i s 2 0 Nicomekl River Flow September, 1997 X = Sampling 44-4 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 September f Nicomekl River Flow October (Incomplete), 1997 J = Sampling 1 3 5 7 9 11 13 15 17 19 21 23 25 27 29 31 f October f 10 system (Halstead, 1978). Halstead (1978) stated that during the summer months Anderson Creek is supplied largely by groundwater. Halstead (1978) also reported the presence of numerous uncontrolled free-flowing wells in the Nicomekl's lowlands, discharging groundwater into the regions drainage ditches for irrigation purposes. Halstead (1986) presented a map of these wells, which is shown in Figure 4. The author personally observed during this study that Ericson Ditch (identified in Figures 5 and 7) had a continuously high flow throughout summer months, likely due to such artesian well discharge and/or groundwater intrusion. Consequently, groundwater is a partial component of riverwater flow, perhaps a dominant component during periods of no rainfall, and is injected into both the Nicomekl River's headland tributaries and lowland ditches. 1.2.3 Dissolved Oxygen Concentrations There are various stresses on the Nicomekl system which create the potential to reduce river dissolved oxygen (DO) values. Removal of the Nicomekl River's riparian vegetation has likely increased water temperatures from spring to fall through lack of tree canopy cover, thus lowering DO saturation concentrations. Construction of the tidal gates at the river's mouth has created periods of low flow and deep water in the Nicomekl River's lower reaches. Aeration and DO concentrations of deeper low flowing waterways may be significantly less than that of shallow fast flowing waterways. Nutrient, TOC, and ammonia rich runoff from agricultural land and septic field sewage leachates entering the Nicomekl River consume oxygen through biochemical oxygen demand (BOD) and eutrophication. 11 Figure 4. Free Flowing Artesian Wells in the Study Area (Legend Follows) 12 Figure 4 Cont. Map Legend (Halstead, 1986). L E G E N D E WATER WELLS DEPTH IN FEET DEPTH IN METRES PUITS D'EAU o O 50-75 75-125 125-175 175-225 15.3-22.9 12.9-38.1 3*.1-53.3 53.3-68,6 129.5-1*4.8 A -Roads: Routes: hard surface, all weather pavle, toute salson... ami highway mora than 2 linw jj'chamriaa Upartm piui da 2 vain hard surface, all weather pave*, toute saison SgBS»-ffMlj 2voJti molrn de 2 VOMM loose or stabilized surface, afl weather gravier agglom«re\ toute saison 2itn-»ormof8 tathanaianu 2 rain ou pi in mdrnt da 2 volas loose surface, dry weather degravier.piriodeseche cart track de terra... trail or portage senUer ou portage i ldini stop itaOoji Railway, normal gauge, singta track Chemm de ter, voie unique{ecartement normal)..-^m-mJLam» i f f^a Horizontal control point, with elevation Point geodesique, avec cote 454 & Bench mark, with elevation Repire de nh/ellement avec cote BM 157-* Spot elevation, precise Point cot*; precis 450 Mine or open cut Mine ou fosse a del ouvert W Building Bltiment • Church £glise 1 School £cole f Post Office Bureau de poste p. Cemetery Cimetfire i£j Navigation Light Feu de navigation * Rock; bare or awash Roche, nue ou 1 fleur d'eau • Power transmission line Ugne de transport d'Jnergie River with bridge Rivilreavecpont " " Rapids, Falls; largejmall Rapides, Chutes; grands, petits Lake intermittent indefinite Lac intermittent rive imprecise Marsh or Swamp Maralsoumarecage Depression contours Courbesde cuvette Trees Arbres Woods Bols CONTOUR INTERVAL 10 FEET AUXILIARY CONTOUR INTERVAL 5 FEET Elevations in Feet above Mean Sea Level EQUIDISTANCE OES GOURDES IO PI EDS EQUIDISTANCE DES COURSES INTERMFJIIAJRES 5 PIEOS EMvations en pieaj lu-desus du niveau moven de Ii mar 13 Table 1 reviews historically reported dissolved oxygen concentrations in the Nicomekl River and two of its large drainage ditches. Table 1. Historically Reported Dissolved Oxygen Concentrations (mg/L and % saturation) in the Nicomekl River and Two of Its Large Drainage Ditches. Location Dates # Samples DO Range DO Mean Nicomekl at 99A Dam (Downstream) 1972-1979 52 1 2.5 - 22.5 mg/L 28.5%-253.1% sat. 9.6 mg/L 99.6% sat. Nicomekl at 168th St. 1974-1979 28 1 5.4 -16.5 mg/L 69.8%-185.8% sat. 10.3 mg/L 106.1% sat. Nicomekl at 64th Ave. (Upstream) 1972-1983 49 1 7.8-14.1 mg/L 71.4%-128.1% sat. 10.8 mg/L 96.9% sat. Nicomekl at 184th St. 11/1989 to 11/1990 9 2 8.6-11.7 mg/L Burrows Ditch 11/1989 to 11/1990 9 2 4.2-12.3 mg/L Nicomekl at Burrows Ditch 11/1989 to 11/1990 9 2 7.6-11.4 mg/L Old Log. Ditch 11/1989 to 11/1990 9 2 6.5-12.5 mg/L Nicomekl at Old Log. Ditch 11/1989 to 11/1990 9 2 7.7-12.8 mg/L (Swain and Holms, 1988b) 2 (EVS, 1993) 14 Consequently, while historical measurements of DO for the surface waters of the Nicomekl River do occasionally drop below the Canadian Water Quality Guideline of 6.0 mg/L DO for cold water fish species such as salmonids (CCME, 1986), to date the majority of river samples have met this criteria. The B.C. Ministry of Environment has reported that water quality objectives for dissolved oxygen in the Nicomekl River have been met for the last 10 years. However, bottom water and sediments in the Nicomekl's lower reaches may have much lower DO concentrations than those reported historically for surface water sampling. The author observed that the river's bottom sediments between 152 St. and 184 St. released large amounts of sediment trapped gases. These gases were very low in odour, as were the sediments. The author believes that the sediments were hypoxic to anoxic due to a combination of low river flows and the microbial degradation of riverbank grasses and possibly organic loading from runoff ditches. The gases present were likely N 2 (nitrate reduction), methane (methanogenesis), or C 0 2 (O2 reduction) and to a much lesser extent H 2S. If low bottom water and sediment DO concentrations exist in the river's lower reaches this could deleteriously impact cold water fish and oxygen demanding sediment invertebrates, or render them more susceptible to ammonia and pesticide toxicity. 1.2.4 Possible Sources of Nicomekl River Contaminants The following is a summary of some of the potential sources of contaminants in the Nicomekl River identified by Swain and Holms (1988b). • Langley operated a municipal landfill until 1978, 1 km south of the Nicomekl River equidistant from Murray and Anderson Creeks. Historical water quality 15 measurements on the Nicomekl tributary closest to the landfill revealed elevated levels of ammonia, COD, TOC, pH, conductivity and Leptomitaceas fungal blooms due to the discharge of landfill leachate. Fungal growths in the Nicomekl River have been observed at least 1 km downstream of this tributary. Old Yale Investments Ltd. operated (may still be operating?) a poultry processing plant near the headwaters of Anderson Creek. The operation had various treatment mechanisms, including a facultative lagoon, and spray irrigated its effluent on hay crops on adjacent land. Swain and Holms (1988b) recommended upstream and downstream monitoring on Anderson Creek to determine if this operation was adversely affecting this tributary. Langley had various schools near the Nicomekl on septic system. Swain and Holms (1988b) did not believe that there would be significant seepage to the Nicomekl River. The Surrey Cooperative Association operated (may still be operating?) a bulk petroleum storage plant four km downstream from Anderson Creek, near 176th St., on the Nicomekl River. This operation was shown to discharge an average 19.3 m3/day and maximum 430 m3/day to an un-named ditch which flows directly to the Nicomekl River. The mean oil and grease discharge to the ditch was 37 mg/L and an in place oil separator was not functioning adequately (Swain and Holms, 1988b). However, due to the low flow volumes of this discharge and the large dilution in the Nicomekl River, the authors questioned whether or not there would be noticeable effects on the river's biota. Major feedlots within the Nicomekl System could be a source of ammonia contamination (discussed later). Storm-water runoff from the road surfaces, residential areas, and commercial operations in Langley and to a lesser extent the study area itself are potential sources of metals, organics. 1.2.5 Nicomekl River Fish Resources As shown in Figure 5, Swain and Holms (1988b) identified numerous natural salmonid spawning habitats both in the upper Nicomekl River, and the river's lower and upper tributaries of Chantcell Creek, Elgin Creek, Ericson Ditch, Anderson Creek, and Murray Creek. Both steelhead and cutthroat trout also naturally utilize the Nicomekl system. Coho spawn in the Nicomekl River specifically between 21 to 23 km and 26 to 30 km upstream from Mud Bay (Swain & Holms, 1988b). Salmonid spawning occurs in Anderson 16 i O) 17 and Murray Creeks from 0.5 to 10.6 km and 1 to 8.6 km upstream of their conf luence with the N icomek l , respect ively (Swain & Holms, 1988b). Many of the di tches in the region follow old temporary or permanent stream beds. F i s h can move into or out of the 5 large municipally controlled dra inage di tches only w h e n their f lood boxes are open (low tide or low river f lows coinc id ing high ditch water- levels) . In addit ion to S w a i n and Ho lms ' (1988b) identification of Erickson Ditch as a salmonid spawning region, E V S (1993) stated that B. Clark, of the B.C. Ministry of the Environment, reported to them that cutthroat trout have been observed spawning in Er icson Ditch. S ince Er icson Ditch is a known sa lmonid s p a w n i n g corridor, its pumped d ischarge is by w a y of sc rew pumps, to offer further protect ion to sa lmonid fry/smolts moving downst ream into the Nicomekl River. Er i cson ditch offers suitable salmon habitat due to its consistent f low and proliferation into severa l branches in its headlands. The other 4 municipal ly control led dra inage di tches are likely not su i tab le sa lmon spawning habitats due to their shorter lengths, lower and irregular f lows, and fine gra ined sediments ( E V S , 1993). Wh i l e sa lmonids may not be spawn ing in these di tches, other less sensi t ive bottom spec ies wh ich do not demand as r igorous f lows a n d high D.O. concentrat ions, and can spawn in silty sediments or on ditch macrophytes, may be spawn ing here. E V S (1993) observed the p resence of carp, Cyprinus carpio in the Nicomekl River. Bourque and Hebert (1982) found the three sp ined st ick leback (Gasterosteus aculeatus), prickly sculp in {Cottus asper), redside shiner (Richardsonius balteatus), lamprey {Lampetra richardsoni), brown bul lhead (Ictalurus nebulosus), peamouth chub (Mylocheilus caurinum), crayf ish, and frogs in the Serpent ine sys tem. T h e s e spec ies are likely found in the Nicomekl R iver as wel l , cons ider ing its proximity and similar geography and water chemistry. Th is author observed a catf ish (appeared to be the brown bullhead, Ictalurus nebulosus) caught in the Nicomekl R iver in 18 the study site, and st ick leback (not three-spine) in both the 1 6 8 t h St. N E Ditch and O ld Logging Ditch. nd nd The Nicomekl River has a salmon hatchery located at 232 St. and 52 A v e . In the fall, the hatchery spawns its returned coho, red Chinook, white Chinook, pink, and chum sa lmon (Rhidine, 1997). In April of each year the Nicomekl Hatchery re leases all its f ish at roughly 6 months of age. In April of 1997 the hatchery released approximately 50,000 red Chinook, 29,000 white Chinook, 50,000 chum, 30,000 pink, and 29 ,000 coho (Rhidine, 1997). The hatchery only rears its fish from fall to spring. Hatchery re leased pink and chum are at the fingerl ing s tage and should migrate to the ocean immediately fol lowing re lease (Iwama, 1991). Hatchery re leased Chinook are c lose to the smolt s tage (Rhidine, 1997), and should migrate to the o c e a n immediately or dur ing the summer months. T h e hatchery 's coho are a lways released at the fingerling stage. The 1997 coho were re leased at 2 - 2.5 g rams weight (Rhidine, 1997). T h e s e f ish should spend between 1 and 2 years in the N icomek l R iver and/or its tributaries prior to smolt i f ication and migration to the o c e a n (Iwama, 1991). Typical coho returns are 600-800 fish (Rhidine, 1997). The river's natural cutthroat and s tee lhead trout populat ions should have river rearing.t imes of 5 and 1 - 2 years, respect ively (Iwama, 1991). There exists the potential that juveni le sa lmonids seek ing refuge and rearing in the lower reaches of the Nicomekl River or its dra inage ditches, or smolt ing f ish migrating through the lower reaches of the river to the ocean , cou ld be exposed to chemical ly contaminated agricultural runoff. Whi le Er icson Ditch is the only known sa lmonid spawn ing ditch, some of the other large drainage ditches in the river's lower reaches may serve as refuge a reas 19 for rearing juvenile sa lmon as wel l . A s a l ready ment ioned, st ick leback were observed in the O ld Logg ing Ditch and the 1 6 8 t h St. N E ditch. Furthermore, f ish foraging in contaminated ditches or river zones may be reduced due to the agricultural runof fs impact on sens i t i ve invertebrate organisms. There a lso exists the possibi l i ty that in the fall spawning adult f ish cou ld come in contact with toxic river water within or downstream of the study site. S ince the depletion of natural coho and Chinook s tocks is currently of paramount concern to the Department of F isher ies and O c e a n s (DFO) , further emphas is is being p laced on both the protection and hatchery product ion of this f ish. Consequent ly , whether the agricultural dra inage di tches in the Nicomekl River are d ischarg ing contaminants in concent ra t ions affecting sa lmonids or their food source organ isms w a s one of the object ives promoting this study. 1.3 Study Area nd th T h e chosen a rea for this f ield investigation is located between 152 and 184 St., and nd th be tween 32 and 48 A v e . (Figure 6). Th is region w a s chosen due to its extensive agricultural use, abundance of dra inage di tches, lack of permeab le soi l and high runoff, e a s e of site accessibil ity, location of the known salmonid spawning area at the headwaters of Er ickson Ditch, and previous water and sediment sampl ing by Coast l ine (1989), W a n (1989), E V S (1993), and W a n et. al (1994, 1995) wh ich revealed the p resence of pest ic ides in this region. 20 21 2.0 AGRICULTURAL LAND USE IN THE STUDY AREA 2.1 Predominant Crops Grown T h e study area is predominantly used for commercial vegetable production, and to a lesser extent, commerc ia l berry product ion, f lower nurser ies, and hobby vegetable farms. The vegetab les observed to be commercia l ly g rown during the study per iod inc luded c o m , potatoes, lettuce, onions, carrots, beets, pumpkins and zucch in i . Berry c rops were exclusively blueberr ies. F igure 7 shows the locat ions of the predominant vegetable and blueberry c rops grown in the study a rea near the dra inage di tches in Ju ly of 1997. 2.2 Drainage Ditch Water Management nd th There are 5 large runoff/irrigation d i tches in the study area , between 152 St. and 184 St., wh i ch e a c h have their d ischarge and irrigation control led by municipal pumping stations located at the junction between the ditches and the Nicomekl River. T h e s e 5 main di tches are speci f ical ly "South C loverda le Ditch", "Er icson Ditch", "Burrows Ditch", "O ld Logg ing Ditch", and "Ha l ls Prair ie Ditch" (Figure 7). If a d i tches water- levels become h igher than the river's water- level , such as after signif icant rainfall and runoff events, and/or at per iods of low tide w h e n the r iver 's tidal gates are opened and the river water-level drops, then d ischarge from the di tches can occurs by hydraul ic head through f lood b o x e s at the pumping stations. N o munic ipal records are kept for ditch d ischarge by gravity through the f lood boxes. If ditch water- levels are high enough that d ischarge is necessa ry to prevent farm land from be ing f looded, but the river water- level is a lso h igh due to runoff upstream of the study site and/or the c losure of the river's tidal gates at high tide, then gravity d ischarge is not poss ib le and these 5 d i tches are mechanica l ly d ischarged by pumps. 22 23 The municipal pumping stations can also be used to transfer water from the river into the ditches during the summer for irrigation of crops. The City of Surrey's Engineering Department keeps an ongoing daily log of these pumps' hours of operation. Discharge by way of pumping alone was significant throughout the summer months, following the periods of rainfall shown in Figure 1. From these municipal records, the hours of pumping at each of the study site's 5 municipal ditches was determined on a daily basis for the study period (Appendix 1). Appendix 1 displays discharge simply in terms of pump-hours per day (the cumulative number of hours of pumping for that day), assuming only a single pump was operating. As summer drainage requirements are lower than those of winter, often only one of the multiple pumps per station was in operation. The maximum discharge flows at "South Cloverdale Ditch", "Ericson Ditch", "Burrows Ditch", "The Old Logging Ditch" and "Halls Prairie Ditch" are 2.4 m3/s, 3.98 m3/s, 1.1 m3/s, 2.26 m3/s, and 3 1.9 m Is, respectively (Lalond, 1998). This is assuming all of the 2-3 pumps at each station 3 3 3 are in operation. Single pump discharge flows are 1.2 m /s, 2.0 m /s, 0.55 m Is, 0.75 3 3 m Is, and 0.95 m /s, respectively. For the study sites 5 main ditches, there appears to be one exception to the summertime use of generally only one pump per station. South 3 Cloverdale Ditch always discharged with both pumps in operation, at 2.4 m Is. Consequently, Appendix 1 shows discharge from South Cloverdale Ditch in terms of pump 3 hours per day, with both pumps operating at their combined flow of 2.4 m /s. Discharge/irrigation of the smaller ditches in the study area is controlled by individual farmers. By way of removal of "stop-logs", these ditches can drain through flap-gates when ditches' water-levels are higher than the river water-level. When the river water-level is higher than the ditches' water-levels, the flap gates prevent river inflow into the 24 ditches. These small ditches can also be irrigated either by small irrigation pumps or the opening of the flap gates during periods of high river water-level. 3.0 POSSIBLE SOURCES OF TOXICITY IN THE STUDY AREA In the municipalities of Langley and Surrey, the 1991 expenditures on all agricultural chemicals (such as fertilizers, herbicides, fungicides, and insecticides) for non-livestock farms were $ 88,000 ($ 1,063/ha), and $ 1,011,000 ($ 747/ha), respectively (FREMP, 1996). Between 1991 and 1995, the total quantity of reportable pesticides sold for commercial use (excluding domestic sales and use) in the Lower Mainland increased from 317, 000 kg to 475,000 kg (BCMELP, 1993, 1997). There was not a substantial increase in the quantities purchased/used in the Lower Mainland for landscaping services (BCMELP, 1997). Unfortunately, the increases in pesticide usage solely by commercial agricultural operations can not be determined, due to limitations in the BCMELP (1997) survey. Therefore the presented increased sales are attributed to higher usage collectively by agriculture, aquatic weed control, forestry, forest nurseries, predator control, industrial vegetation control, industrial vegetation-pavers, landscape services, mosquito and biting fly control, noxious weed control, product.fumigation, and structural-wood preservation uses (BCMELP, 1997). Extensive amounts of manure and fertilizer are being applied to the Nicomekl-Serpentine farmlands, as evidenced by FREMP's (1996) excess nitrogen calculations for the region (presented later). The heavy reliance on chemical-assisted farming increases the potential for contamination of the aquatic habitats in these regions. 25 3.1 Routes of Pesticide Entry into the Nicomekl River and Drainage Ditches. Contamination of the Nicomekl River by runoff containing agricultural pesticides from the lands in the study area is only likely to occur via the discharge of polluted ditch water into the river. Diffuse runoff from agricultural lands in the study area directly into the river is extremely unlikely due to the river's high clay-diked banks. Consequently, by monitoring the drainage ditches in the study area, one can evaluate whether pesticides are reaching the river via runoff from the agricultural lands. Transport of contaminants by ditch water may be as dissolved constituents or via their adsorption to suspended sediments washed off of the fields. Mulkey and Donigan (1984) identified three conditions which create circumstances in which pesticide residues in surface waters may adversely affect aquatic biota: (1) the applied pesticide persists on the land surface or in the soil profile long enough for subsequent transport in runoff, in subsurface flow, and on eroded sediments; (2) rainfall, infiltration and storm runoff occurs during the time period when pesticides reside on or in the soil; (3) the resulting pesticide runoff reaching surface waters persists long enough in the aquatic system to affect potential exposure to aquatic organisms. The same can likely be said for ammonia and metals. Pesticide contamination of the Nicomekl River or its feeder ditches may also occur due to the over-spray of pesticides into ditches and/or the river directly. The lack of adequately sized buffer strips alongside waterways, spraying in windy conditions, or the use of pesticide application equipment not producing the appropriate spray droplet size, can contribute to pesticide drift into ditches and/or the river. Illegal disposal of pesticides into 26 drainage ditches or the washing of spray equipment near ditches are two other possible means of pesticide contamination. 3.2 Predominant Pesticides Used in the Study Area The use of organochlbrine pesticides has been heavily restricted due their persistence, ability to bioconcentrate in organisms and propensity to bioaccumulate in food chains, severely affecting non-target organisms. This has led to more intensive use of the comparatively easily degraded, generally non-bioaccumulating, organophosphate and carbamate insecticides. The B.C. Ministry of Environment, Lands and Parks (BCMELP) estimated agricultural pesticide sales in the Lower Mainland for over 200 different pesticides in 1991. Annual sales were assumed to approximate annual usage (BCMELP, 1993). A list of the approximate weights of the top 90 pesticides (insecticides, herbicides, and fungicides) used in the Lower Mainland in 1991 solely for commercial agricultural purposes (from BCMELP, 1993) is shown in Table 2 on the following page. More recently, the BCMELP summarized 1995 pesticide sales/usage (kg) for the Lower Mainland (BCMELP, 1997). However, this summary did not separate the agricultural sales from total sales, which includes those pesticides sold to companies holding pest service licences, and is not as useful for this study as the 1991 report which determined uses for commercial agricultural operations. The BCMELP's 1995 data does show that there were no sales of dinoseb, fensulfothion, chloramben, lindane, and phorate in B.C. in 1995. The B.C. Ministry of Agriculture Fisheries and Food's (BCMAFF) Field Crop Production Guide for 1996 does not suggest farmers use any of these five pesticides. Consequently, it is assumed they 27 had no use in the study site in 1997, and it is believed that these pesticides have been either banned or now have only extremely limited uses. Maneb is still recommended by the 1996 Field Crop Guide (BCMAFF, 1996a); however, only 55.7 kg were sold throughout B.C. in 1995, compared to the 3,400 kg sold in the Lower Mainland alone in 1991. Consequently it is assumed that maneb use in the study site in 1997 was similar to 1995 use, reducing its rank as quantity used from #14 to #88, on the list of the 90 most abundantly used pesticides. Table 2. Quantities of Pesticides Used in the Lower Mainland in 1991 for Commercial Agricultural Purposes (BCMELP, 1993). Pesticide kg Used Pesticide kg Used Pesticide kg Used Glyphosate 40,807 EPTC 882 Fluazifop-Butyl 202 Metam 23,448 Bentazon 854 Pyridate 192 Captan 15,796 Naled 847 Chlorpropham 182 Herbicidal Mineral Oil 11,790 2,4-D Amine 828 Aluminum Phosphide 176 Atrazine 9,273 Zineb 807 Dicofol 170 Malathion 7,393 Propoxur 782 Fensulfothion 166 Mancozeb 6,875 Azinphos-Methyl 732 Methomyl 160 Paraffin Base Mineral Oil 5,149 Metiram 684 Oxyflourfen 153 Metolachlor 4,391 Maleic Hydrazide 610 Metribuzin 144 Simazine 4,029 Triforine 601 Tebuthiuron 142 Diazinon 3,907 Cupric Hydroxide 589 Prometryne 141 Insecticidal Mineral Oil 3,767 Amitrole 585 Thiram 137 Maneb 3,400 Treflan 579 MCPB 133 Copper Oxychloride 3,382 Metalaxyl 543 Metobromuron 132 Formaldehyde 2,724 Endosulfan 508 Oxamyl 132 Dinoseb 2,695 Carbofuran 496 Lindane 124 Chlorpyrifos 2,432 Iprodione 451 Cycloate 115 Chlorothalonil 2,354 Chlorthal 412 Hexazinone 113 Sodium Metaborate Tetrahydrate 2,340 Sulfotep 410 Diclofop-Methyl 107 28 Benomyl 2,073 Disulfoton 387 Thiophanate-Methyl 102 Napropamide 2,003 Butylate 384 Methoprene 98 Methamidophos 1,935 Propargite 342 Sethoxydim 93 Dichlobenil 1,862 Sulphur 339 Chlormequat 92 Fonofos 1,664 Acephate 331 Thiabendazole 86 Paraquat 1,582 Oxydemeton-Methyl 319 Pendimethalin 68 Parathion 1,306 Dimethoate 290 Picloram Esters 65 Linuron 1,106 Mecoprop K Salt 281 Etridiazole 60 Dicamba 1,088 Nicotine 267 Chloramben 58 Sodium Chlorate 1,053 Diquat 257 Dodemorph-Acetate 54 MCPA Amine 1,035 Phorate 235 Dichlorvos 54 Bromoxynil 971 Carbaryl 210 Ziram 49 It was determined which of these 90 pesticides were still recommended for use in 1997, specifically for the types of crops observed in the study area, under the guidelines set by the BCMAFF in the 1996 Field Crop Guide (BCMAFF, 1996a) with its 1997 update, and BCMAFF 1996/97 Berry Crop Production Guide (BCMAFF, 1996b). Those pesticides which have suggested uses for the crops observed in the study area, as well as peas and beans (in the even these common crops were also present in the study site but not observed), are shown in Table 3. These serve as a list of the pesticides most likely used on the Nicomekl's adjacent agricultural lands during the 1997 growing season. They are listed in order of decreasing quantity sold for agricultural use in the Lower Mainland in 1991, which serves as an approximation of the likely relative abundance of each used in the study site in 1997. 29 Tab le 3. Most Probable Pest ic ides Used on the Nicomekl Farmlands during the 1997 Growing S e a s o n (Descend ing Order of Est imated Amount of E a c h Used) . Pesticide Use Class of Compound and Action When Typically Used Vegetable Crops Used On Glyphosate Herbicide Phosphonic Acid (amino-acid synthesis inhibitor) Sprayed pre-harvest for beans and peas; when necessary, once per season, blueberries beans, peas, blueberries Captan Fungicide Phthalate Seed treatment; blooms sprayed for blueberries corn, beans, peas, blueberries Atrazine Herbicide Triazine (photosynthetic electron-transport inhibitor) Sprayed pre-emergence or early post emergence corn Malathion Insecticide Organophosphate (cholinesterase inhibitor) Sprayedtwheninsects;. appear; also late winter for blueberries potatoes, corn, beets, blueberries Mancozeb Fungicide Thiocarbamate Seed treatment potatoes, corn Metolachlor Herbicide Amide (cell division inhibitor) Sprayed pre-plant or pre-emergence corn & field crops Simazine Herbicide Triazine (photosynthetic electron-transport inhibitor) Sprayed pre-emergence blueberries Diazinon insecticide Organophosphate (cholinesterase inhibitor) Seed treatment for com and potato pieces & sprayed when insects appear potatoes, com, beans Chlorpyrifos Insecticide Organophosphate (cholinesterase inhibitor) Sprayed when insects appear potatoes, com, beets Benomyl Fungicide Benzimadazole Sprayed between 50% and full bloom beans Napropamide Herbicide Aryloxyalkanamide (cell division inhibitor) Sprayed between fall and spring. blueberries Methamidophos Insecticide Organophosphate (cholinesterase inhibitor) Sprayed when insects appear potatoes Dichlobenil Herbicide Benzonitrile (cellulose biosynthesis inhibitor) Granules broadcast in early spring blueberries Paraquat Herbicide Bipiridyl (photosynthetic electron flow diverter) Sprayed pre-emergence potatoes, beets, com, blueberries 30 Parathion Insecticide Organophosphate (cholinesterase inhibitor) Sprayed when insects appear peas Linuron Herbicide Urea herbicide (photosynthetic electron-transport inhibitor) Sprayed when com 38 cm high corn Dicamba Herbicide Arenecarboxylic acid Sprayed post-emergence when corn 20 - 50 cm high corn MCPA Amine Herbicide Aryloxyalkanoic acid Sprayed when corn 15 cm high; peas 2 to 5 nodes corn, peas Bromoxynil Herbicide Hydroxybenzonitrile (photosynthetic electron-transport inhibitor) Sprayed when corn 4 to 8 leaf stage com Bentazon Herbicide Benzothiadiazinon (photosynthetic electron-transport inhibitor) Sprayed post-emergence corn, peas 2,4-D Amine Herbicide Aryloxyalkanoic acid Sprayed until com 15 cm high com Amitrole Herbicide Triazole (carotenoid synthesis inhibitor) Sprayed prior to com planting com, beans Carbofuran Insecticide Carbamate (cholinesterase inhibitor) Potatoes & beets granulated at seeding; corn sprayed when insects appear potatoes, com, beets Dimethoate Insecticide Organophosphate (cholinesterase inhibitor) Sprayed when insects appear potatoes, com, beets, peas, beans Diquat Herbicide Bipyridyl (photosynthetic electron flow diverter) Sprayed 2 weeks before potato harvest potatoes, peas Carbaryl Insecticide Carbamate (cholinesterase inhibitor) Sprayed when insects appear potatoes, com Fluazifop-butyl Herbicide Alkanoic acid (fatty-acid synthesis inhibitor) Sprayed when necessary; once per season blueberries potatoes, beets, blueberries Methomyl Insecticide Carbamate (cholinesterase inhibitor) Sprayed when insects appear potatoes, com, peas Prometryn Herbicide Triazine (photosynthetic electron-transport inhibitor) Sprayed pre-emergence peas 31 Thiram Fungicide Dimethyldithiocarba-mate Seed treatment corn, beans, peas Hexazinone Herbicide Sprayed when necessary blueberries Diclofop methyl Herbicide Propionic acid (fatty-acid synthesis inhibitor) Sprayed when necessary beets, beans, peas Pendimethalin Herbicide Dinitroanaline (cell division inhibitor) Sprayed pre-emergence or post-emergence com E V S (1993) conducted interviews with farmers in the study a rea during the 1989 growing season. The farmers indicated that they used glyphosate (applied prior to Spr ing plowing), ch lorpropham (used a s a top-killer) and chlorpyrifos ( E V S , 1993). Ch lorpropham is currently only recommended for use as spot treatment against dodder in alfalfa forage c rops ( B C M A F F , 1996a), had zero sa les in B .C . in 1995, and w a s unl ikely used in the study site in 1997. The author observed pest ic ide spraying on Ju ly 1 6 t h , 1997, of a white mist (tractor-tank-boom setup) on potato crops located adjacent to Burrows ditch, Si te 11, th on the f ield south of 40 Ave . Over -spray into Burrows ditch w a s observed at this time. S a m p l i n g of Burrows Ditch w a s performed on this date. Simi lar spray ing w a s a lso th th observed on July 29 , 1997, on the potato crops to the north of 40 Ave . , between th th between the O ld Logging Ditch and 176 St. No samp les were col lected on July 29 . Fo r the study area, the most abundant ly used insect ic ides are organophosphate and carbamate compounds. The most abundant ly used herb ic ides are more var iable in chemica l composi t ion, but are primarily ni trogenous compounds such as amides and tr iazines. 32 Many of the pesticides likely used in the study area show increased sales in the Lower Mainland between 1991 and 1995. Table 4 shows the increase/decrease of the pesticides likely used in the study area between 1991 and 1995. As 1995 data for the amount of pesticides used solely for agricultural use was not available as it was in 1991, the weights (Table 4) are those for total sales (for agriculture and other pest control uses presented in Section 3.0), excluding domestic sales, for both 1991 and 1995. Table 4. Increased Pesticide Usage in The Lower Mainland between 1991 and 1995 for Those Pesticides Likely Used on the Nicomekl Farmlands. Pesticide 1991(kg) 1995 (kg) % Change Class Glyphosate 77,575 79, 035 + 2 Herbicide Captan 16,164 20, 853 + 29 Fungicide Atrazine 15, 360 8, 611 -44 Herbicide Malathion 10, 843 5, 749 -47 OP Insect. Mancozeb 7, 589 24,999 + 229 Fungicide Metolachlor 7, 475 5, 356 -28 Herbicide Simazine 5, 667 7, 250 + 28 Herbicide Diazinon 8, 153 9, 459 + 16 OP Insect. Chlorpyrifos 3, 523 4, 633 + 32 OP Insect. Benomyl 2, 258 3, 036 + 35 Fungicide Napropamide 2,065 4, 672 + 126 Herbicide Methamidophos 1, 973 1, 877 -5 OP Insect. Dichlobenil 3, 024 4, 942 + 63 Herbicide Paraquat 2, 238 3, 852 + 72 Herbicide Parathion 1, 711 2, 051 + 20 OP Insect. Linuron 1, 106 3, 689 +234 Herbicide Dicamba 1, 908 1, 173 -39 Herbicide MCPA Amine 1,231 1, 637 + 33 Herbicide Bromoxynil 1, 096 258 -76 Herbicide Bentazon 1,050 1, 317 + 25 Herbicide 2,4-D Amine 2, 304 6, 418 + 179 Herbicide Amitrole 747 1, 150 + 54 Herbicide Carbofuran 503 925 + 84 % Carb. Insect. 33 Dimethoate 1, 201 5, 586 + 365 % OP Insect. Diquat 296 1, 213 + 310% Herbicide Carbaryl 299 1, 731 + 479 % Carb. Insect. Fluazifop-butyl 205 164 - 20 % Herbicide Methomyl 167 401 + 140% Carb. Insect. Prometryne 226 385 + 70 % Herbicide Thiram 137 630 + 360 % Fungicide Hexazinone 113 56 - 50 % Herbicide Diclofop methyl 107 0 - 100% Herbicide Pendimethalin 68 543 + 699 % Herbicide TOTAL 178, 382 213,678 + 20 % Pesticides With respect to both agricultural and pest control service pesticide usage in the Lower Mainland, for the pesticides suggested for field crop and blueberry use (BCMAFF, 1996a, 1996b), total usage was 20% higher in 1995 than in 1991. With respect to insecticide usage in the Lower Mainland, carbaryl and dimethoate usage increased dramatically, 479% and 365%, respectively. The top 9 insecticides, eligible for field crop and berry use, used in the greatest abundance in the Lower Mainland in 1995, in order of decreasing weight used, were diazinon (OP), malathion (OP), dimethoate (OP), chlorpyrifos (OP), parathion (OP), methamidophos (OP),'carbaryl (carbamate), carbofuran (carbamate), and methomyl (carbamate). Usage of these insecticides increased 14%; between 1991 and 1995, from 28, 373 kg to 32,412 kg. These increases may be solely due increased urban development and non-agricultural uses, or they may in part reflect greater farm dependence on chemical pest control. In general, organophosphate insecticides are more toxic than carbamate insecticides (Macek and McAllister, 1970). Organophosphate and carbamate insecticides are both acetycholinesterase inhibitors. Acetycholine (Ach) is the neurotransmitter which allows 34 for transmission of nerve impulses across the synapse between adjacent neurons. Acetycholinesterase is the enzyme responsible for the breakdown of Ach, thereby terminating the electrochemical connection between neurons. Organophosphorous insecticides phosphorylate acetycholinesterase, which inhibits Ach from hydrolyzing neurotransmitter at the nerve synapse (Smith, 1987). Carbamate chemicals carbamylate acetycholinesterase, with the same result. The destruction of acetycholinesterase and the accumulation of acetycholine results in continuous nerve firing and eventual failure of nerve impulse propagation (Smith, 1987). Consequently, this invterferes with neuromuscular junction, producing rapid twitching of voluntary muscles and finally paralysis (Ware, 1978). Respiratory paralysis is generally the immediate cause of death (Murphy, 1975). In addition to inhibited Ach activity, delayed neurotoxicity is believed to be caused by inhibition of another enzyme, neurotoxic esterase, which results in ataxia and paralysis caused by axonal degeneration (Davies & Richardson, 1980). Herbicides generally have less of an effect on non-target animals because these chemicals are designed to biochemically inhibit photosynthesis and cell division of plants. 3.3 Literature Review of Pesticide Toxicity 3.3.1 Acute lethality to Fish and Invertebrates of Pesticides Likely Used on the Nicomekl Farmlands. The acute lethality of the most abundantly used pesticides, to various fish and invertebrate species, is shown in Table 5, to give the reader an understanding of the relative toxicity of pesticides. These data also demonstrate how the majority of these pesticides are lethal to invertebrates at much lower concentrations than they are to fish. 35 Table 5. Acute Lethality of Probable Pesticides Used on the Nicomekl Farmlands During the 1997 Growing Season. Pesticide & Solubility Lethal Toxicity To Fish and Invertebrates (Mg/L) Reference Glyphosate 12g/L(25 °C) Rainbow trout 96-h LC50 = 86,000 Rainbow trout (0.8 g) 96-h LC50 = 130,000 * Rainbow trout (0.8 g) 96-h LC50 = 130,000 * Rainbow trout (1.0 g) 96-h LC50 = 8,300 (R) Rainbow trout 96-h LC50 = 50,000 Rainbow trout 96-h LC50 = 54,800 Channel catfish (2.2 g) 96-h LC50 - 130,000 * Daphnia 48-h LC50 > 780,000 Daphnia magna 48-h EC50 = 3,000 (R) Daphnia magna 48-h LC50 = 2,950 Midge, Chironomus plumosus 48-h LC50 = 55,000 (R) Roundup formulation, 3 -42 times more toxic than technical grade (BCPC, 1991) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Johnson & Finley, 1980) (Folmar, 1976) (Hildebrand etal., 1982) (Johnson & Finley, 1980) (BCPC, 1991) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) (Johnson & Finley, 1980) Caotan 3.3 mg/L (25 °C) Rainbow trout (1.0 g) 96-h LC50 = 73 * Coho (0.8 g) 96-h LC50 = 138 * Chinook (fingerling) 96-h LC50 = 120 Cutthroat trout (0.4 g) 96-h LC50 = 56 Brook trout 96-h LC50 = 34 Channel catfish (1.2 g) 96-h LC50 = 78 * Daphnia pulex 26-h LC50 = 1300 (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) (BCPC, 1991) (Johnson & Finley, 1980) (Frear& Boyd, 1967) Atrazine 30 mg/L (20 °C) Rainbow trout 96h-LC50 = 2,600 - 3,200 Rainbow trout 96-h LC50 = 4,500 Rainbow trout 48-h LC50 = 12,600 Daphnia magna 48-h LC50 = 3,600 (BCPC, 1991) (Bathe etal., 1975) (FWPCA, 1968) (FWPCA, 1968) Malathion 145 mg/L (room temp.) Rainbow trout 96-h LC50 = 170 Rainbow trout 96-h LC50 = 200 Cutthroat trout (1.0 g) 96-h LC50 = 280 * Cutthroat trout (2.9 g) 96-h LC50 = 230 Coho 96-h LC50 = 101 Channel catfish (1.5 g) 96-h LC50 = 8,970 Carp (0.6 g) 96-h LC50 = 6,590 * Daphnia magna 48-h EC50 = 1.0 * Daphnia pulex 48-h EC50 = 1.8 * Isopod, Asellus 96-h LC50 = 3,000 * Amphipod, Gammarus asciatus 96-h LC50 = 0.76 * Amphipod, Gammarus lacustris 48-h LC50 = 1.8 Stonefly, Simocephalus sp. 48-h LC50 = 3.0 Mayfly, Baetis sp. 48-h LC50 = 6.0 (Macek & McAllister, 1970) (Mayer & Ellersieck, 1986) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Macek & McAllister, 1970) (Macek & McAllister, 1970) (Macek & McAllister 1970) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Cope, 1966) (Cope, 1966) (Cope, 1966) Mancozeb 6-20 mg/L Rainbow trout 48-h LC50 = 2,200 Catfish 48-h LC50 = 5,200 (BCPC, 1991) (BCPC, 1991) 36 Metolachlor 530 mg/L (20 °C) Rainbow trout 96-h LC50 = 2,000 Carp 96-h LC50 = 4,900 Daphnia magna 48-h LC50 = 25,100 (BCPC, 1991) (BCPC, 1991) (Vilkas, 1976) Simazine 5 mg/L (20 °C) Rainbow trout (1.2 g) 96-h LC50 > 100,000 * Rainbow trout 96-h LC50 > 100,000 Rainbow trout 96-h LC50 = 68,000 Rainbow trout 96-h LC50 = 56,000 Carp 96-h LC50 > 100,000 Daphnia magna 48-h LC50 = 1,100 * Daphnia magna 48-h EC50 > 3,500 Amphipod, G. fasciatus 96-h LC50 > 100,000 * Stonefly, Pteronarcys sp. 96-h LC50 = 1,900 (Johnson & Finley, 1980) (BCPC, 1991) (Cope, 1965a) (Bohmont, 1967) (BCPC, 1991) (Johnson & Finley, 1980) (Marchini et al., 1988) (Johnson & Finley, 1980) (USDA, 1984) Diazinon 40 mg/L (20 °C) Rainbow trout (1.2 g) 96-h LC50 = 90 * Rainbow trout 96-h LC50 = 2,600 - 3,000 Rainbow trout 96-h LC50 = 4,300 Rainbow trout 24-h LC50 @ 22°C = 52 Brook trout 96-h LC50 = 770 Cutthroat trout (2.0 g) 96-h LC50 = 1,700 * Lake trout (3.2 g) 96-h LC50 = 602 * Carp 96-h LC50 = 7,600 - 23,400 Daphnia pulex 48-h EC50 = 0.8 * Dapnia pulex 48-h EC50 = 0.9 Ceriodaphnia dubia 48-h LC50 = 0.35 Ceriodaphnia dubia 24-h LC50 = 0.37 - 0.75 Ceriodaphnia dubia 48-h LC50 = 0.26 - 0.58 Ceriodaphnia dubia 96-h LC50 = 0.32 - 0.35 (pH 7.4 - 8.2, hardness 80- 100 mg/L as CaC03) (laboratory and natural waters) Ceriodaphnia dubia 24- or 48-h LC50 = 0.3 - 0.5 Amphipod, Gammarus fasciatus 96-h LC50 = 0.2 * Amphipod, G. pseudolimnaeus 48-h LC50 = 4 Amphipod, Gammarus lacustris 96-h LC50 = 170 Amphipod, Hyallela azteca 48-h LC50 = 22 Stonefly, Pteronarcys sp. 48-h LC50 = 60 Midge, Chironomus tentans 48-h LC50 = 0.1 Isopod, Asellus communis 96-h LC50 = 21 Mollusc, Helisoma trivolvis 7-day LC50 = 528 Crayfish, Orconectes propinquus 48-h LC50 = 537 (Johnson & Finley, 1980) (BCPC, 1991) (Murty, 1986) (Cope, 1965b) (Allison & Hermanutz, 77) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (BCPC, 1991) (Johnson & Finley, 1980) (Cope, 1966) (Amatoetal., 1992) (Bailey et al., 1997) (Bailey et al., 1997) (Bailey et al., 1997) (Bailey et al., 1997) (Katznelson & Mumley,'97) (Johnson & Finley, 1980) (Morgan, 1976) (Morgan, 1976) (Morgan, 1976) (FWPCA, 1968) (Morgan, 1976) (Morgan, 1976) (Morgan, 1976) (Morgan, 1976) 37 Chlorpvrifos 2 mg/L (20 °C) Rainbow trout 96-h LC50 = 3.0 Rainbow trout 96-h LC50 = 8.0 Rainbow trout (1.4 g) 96-h LC50 = 7.1 Rainbow trout (1.4 g) 96-h LC50 < 1.0 Rainbow trout 48-h LC50 = 20 Rainbow trout 24-h LC50 = 53 Cutthroat trout (1.4 g) 96-h LC50 = 18 * Channel catfish (0.8 g) 96-h LC50 = 280 * Daphnia pulex 48-h EC50 = 1.78 Ceriodaphnia dubia 24-h LC50 = 0.063 - 0.101 Ceriodaphnia dubia 48-h LC50 = 0.058 - 0.059 Ceriodaphnia dubia 96-h LC50 = 0.055 (pH 7.4 - 8.2, hardness 80- 100 mg/L as CaC03) (laboratory and natural waters) Ceriodaphnia dubia 48-h LC50 = 0.117 (storm-water sample) Amphipod, Gammarus lacustris 96-h LC50 = 0.11 * Amphipod, Gammarus lacusths 24-h LC50 = 0.76 Amphipod, Hyallela azteca 48-h LC50 = 0.1 Amphipod, Hyallela azteca 10-day LC50 = 0.086 Stonefly, Pteronarcella badia 24-h LC50 = 4.2 Midge, Chironomus tentans 48-h LC50 = 0.3 Midge, Chironomus tentans 10-day LC50 = 0.07 (BCPC, 1991) (Holcombe, et al., 1982) (Maceketal., 1969) (Mayer & Ellersieck, 1986) (FWPCA, 1968) (Maceketal., 1969) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Wood, 1997) (Bailey etal., 1997) (Bailey et al., 1997) (Bailey et al., 1997) (Bailey et al., 1997) (Johnson & Finley, 1980) (Sanders, 1969) (Moore et al., 1998) (Phipps et al., 1995) (Sanders & Cope, 1968) (Moore et al., 1998) (Ankley et al., 1994a) Benomvl 4 mg/L (25 °C) Rainbow trout (1.2 g) 96-h LC50 = 170 * Rainbow trout 96-h LC50 = 170 Rainbow trout (1.0 g) 96-h LC50 = 310 Channel catfish (1.2 g) 96-h LC50 = 28 * Daphnia magna 48-h LC50 = 2,800 Amphipod, G. Pseudolimnaeus 96-h LC50 = 750 Midge, Chironomus plumosus 48-h EC50 = 7,000 (Johnson & Finley, 1980) (BCPC, 1991) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) Napropamide 73 mg/L (20 °C) Rainbow trout 96-h LC50 = 16,600 (BCPC, 1991) Methamidophos > 200 g/L (20 °C) Rainbow trout 96-h LC50 = 51,000 Carp 96-h LC50 = 68,000 (BCPC, 1991) (Chin & Sudderuddin,1979) Dichlobenil 18 mg/L (20 °C) Rainbow trout (1.0 g) 96-h LC50 = 6,300 * Daphnia 48-h LC50 = 9,800 Daphnia pulex 96-h LC50 = 3,700 Amphipod, G. lacustris 96-h LC50 = 11,000 Isopod, Asellus 96-h LC50 = 35,000 (Johnson & Finley, 1980) (BCPC, 1991) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) Paraquat very soluble Rainbow trout (0.5 g) 96-h LC50 = 15,000 Rainbow trout 96-h LC50 = 32,000 Channel catfish (1.4 g) 96-h LC50 > 100,000 * Daphnia pulex 48-h EC50 = 4,000 Daphnia pulex 48-h LC50 = 3,700 Amphipod, G. fasciatus 96-h LC50 = 11,000 Amphipod, G. lacustris 24-h LC50 = 38,000 (Johnson & Finley, 1980) (BCPC, 1991) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (FWPCA, 1968) (Johnson & Finley, 1980) (Sanders, 1969) 38 Parathion (Methvl) 11 mg/L (20 °C) Rainbow trout (1.1 g) 96-h LC50 = 3,700 * Rainbow trout 96-h LC50 = 2,750 Rainbow trout 96-h LC50 = 2,700 Rainbow trout 96-h LC50 = 2,800 Cutthroat trout (0.2 g) 96-h LC50 = 1,850 * Coho (1.0 g) 96-h LC50 = 5,300 Channel Catfish (1.4 g) 96-h LC50 = 5,240 Channel Catfish 96-h LC50 = 5,710 Carp (0.6 g) 96-h LC50 = 7,130 Daphnia magna 48-h LC50 = 0.14 * Daphnia pulex 48-h EC50 = 0.60 Amphipod, Gammarus fasciatus 96-h LC50 = 3.8 (Johnson & Finley, 1980) (Macek & McAllister, 1970) (BCPC, 1991) (Palawski et al., 1983) (Johnson & Finley, 1980) (Macek & McAllister, 1970) (Johnson & Finley, 1980) (Macek & McAllister, 1970) (Macek & McAllister, 1970) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Johnson & Finley, 1980) Linuron 81 mg/L (24 °C) Rainbow trout 96-h LC50 = 16,000 Daphnia magna 48-h EC50 = 270 Midge, Chironomus plumosus 48-h EC50 = 2,900 (BCPC, 1991) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) Dicamba 6.5 g/L (25 °C) Rainbow trout (0.8 g) 96-h LC50 = 28,000 * Rainbow trout 96-h LC50 = 135,000 Rainbow trout 48-h LC50 = 35,000 Coho 24-h LC50 = 151,000 Daphnia magna 48-h LC50 > 100,000 * Daphnia pulex 48-h LC50 = 11,000 Amphipod, G. fasciatus 96-h LC50 > 100,000 * Amphipod, Gammarus lacustris 48-h LC50 = 5,800 Isopod, Asellus 96-h LC50 > 100,000 * (Johnson & Finley, 1980) (BCPC, 1991) (Bohmont, 1967) (Bond etal., 1965) (Johnson & Finley, 1980) (Sanders and Cope, 1966) (Johnson & Finley, 1980) (FWPCA, 1968) (Johnson & Finley, 1980) MCPA Amine 825 mg/L (room temp.) Rainbow trout 96h-LC50 = 117,000 Daphnia magna EC50 = 100,000 Daphnia magna 48-h EC50 > 230,000 (MCPA Task Force, 1987) (Crosby & Tucker, 1966) (Rhone-Poulenc, 1992) Bromoxvnil 130 mg/L (20 °C) Rainbow trout 48-h LC50 = 150 Catfish 48-h LC50 = 63 (BCPC, 1991) (BCPC, 1991) Bentazon Could not find available toxicity data 2.4-D Amine 620 mg/L (25 °C) Rainbow trout (1.0 g) 96-h-LC50 = 24,000 Rainbow trout 24-h LC50 = 250,000 Channel catfish (1.5 g) 96-h LC50 > 100,000 Cutthroat trout 96-h LC50 = 150- 1,200 Daphnia lumholtzi 38-h LC50 - 10,000 (Johnson & Finley, 1980) (Alabaster, 1969) (Johnson & Finley, 1980) (BCPC, 1991) (George & Hingorani,1982) Amitrole 280 g/L (25 °C) Salmon 48-h LC50 = 3,250,000 Daphnia magna EC = 23,000 (Bohmont, 1967) (Crosby & Tucker, 1966) Carbofuran 320 mg/L (25 °C) Rainbow trout (1.5 g) 96-h LC50 = 380 * Coho (0.6 g) 96-h LC50 = 530 * Channel catfish (1.0 g) 96-h LC50 = 248 * Trout 96-h LC50 = 280 Daphnia pulex 48-h LC50 = 35 (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (BCPC, 1991) (Hartman & Martin, 1985) 39 Dimethoate 25 g/L (21 °C) Rainbow trout (1.5 g) 96-h LC50 = 6,200 * Rainbow trout (1.5 g) 96-h LC50 = 8,600 Rainbow trout 24-h LC50 = 19,000 Amphipod, Gammarus lacustris 96-h LC50 = 200 * Amphipod, Gammarus lacustris 24-h LC50 = 900 Daphnia magna 48-h LC50 = 3,320 Daphnia magna 48-h LC50 = 2,500 Stonefly, Pteronarcys sp. 48-h LC50 = 140 (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Alabaster, 1969) (Johnson & Finley, 1980) (Sanders, 1969) (Song, et al., 1997) (FWPCA, 1968) (Cope, 1965b) Diquat 700 g/L (20 °C) Rainbow trout 96-h LC50 = 21,000 Rainbow trout 96-h LC50 = 12,000 Rainbow trout 48-h LC50 = 12,300 Chinook 48-h LC50 = 28,500 Mirror Carp 96-h LC50 = 67,000 Daphnia magna EC = 7,100 Amphipod, G. fasciatus 96-h LC50 > 100,000 * (BCPC, 1991) (Folmar, 1976) (FWPCA, 1968) (Bohmont, 1967) (BCPC, 1991) (Crosby & Tucker, 1966) (Johnson & Finley, 1980) Carbaryl 40 mg/L (30 °C) Rainbow trout (1.5 g) 96-h LC50 = 4,340 Rainbow trout 96-h LC50 = 1,300 Rainbow trout 96-h LC50 = 1,100 Cutthroat trout (0.5 g) 96-h LC50 = 7,100 * Coho (1.0 g) 96-h LC50 = 764 Channel catfish (1.5 g) 96-h LC50 = 15,800 Channel catfish (1.5 g) 96-h LC50 = 7,790 Carp (0.6 g) 96-h LC50 = 5,280 * Carp 96-h LC50 = 1,700 Daphnia pulex 48-h EC50 = 6.4 * Amphipod, Gammarus fasciatus 96-h LC50 = 26 * Amphipod, Gammarus lacustris 96-h LC50 = 22 * Stonefly, Pteronarcella badia 24-h LC50 = 5.0 Isopod, Asellus 96-h LC50 = 280 * (Macek & McAllister, 1970) (BCPC, 1991) (Mayer & Ellersieck, 1986) (Johnson & Finley, 1980) (Macek & McAllister, 1970) (Macek & McAllister, 1970) (Mayer & Ellersieck, 1986) (Macek & McAllister, 1970) (Chin & Sudderuddin, 79) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Sanders & Cope, 1968) (Johnson & Finley, 1980) FluazifoD-butvl 1 mg/L Rainbow trout 96-h LC50 = 1,370 Mirror Carp 96-h LC50 =1,310 (BCPC, 1991) (BCPC, 1991) Methomyl 58 g/L (25 °C) Rainbow trout (1.1 g) 96-h LC50 = 1,600 * Rainbow trout (0.8 g) 96-h LC50 = 860 Rainbow trout 96-h LC50 = 3,400 Cutthroat trout (1.0 g) 96-h LC50 = 6,800 * Channel catfish (1.0 g) 96-h LC50 = 530 * Daphnia magna 48-h LC50 = 8.8 * Amphipod, G. pseudolimnaeus 96-h LC50 = 920 Midge, Chironomus plumosus 48-h EC50 = 88 (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (BCPC, 1991) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) (Mayer & Ellersieck, 1986) Prometryn 33 mg/L (20 °C) Rainbow trout 96-h LC50 = 2,500 Rainbow trout LC50 = 2,900 Daphnia magna EC50 = 18,590 (BCPC, 1991) (U.S. EPA, 1996) (U.S. EPA, 1996) Thiram 30 mg/L (room temp.) Rainbow trout 48-h LC50 = 130 Carp 48-h LC50 = 400 Channel catfish 72-h LC50 = 790 (BCPC, 1991) (BCPC, 1991) (Clemens & Sneed, 1959) Hexazinone 33 g/L (25 °C) Rainbow trout 96-h LC50 > 180,000 Dapfcn/a 48-h LC50 = 151 (Mayer & Ellersieck, 1986) (BCPC, 1991) 40 Diclofop methyl 3 mg/L (22 °C) Rainbow trout 96-h LC50 = 350 Rainbow trout 96-h LC50 = 250 Daphnia magna 48-h LC50 = 317 (BCPC, 1991) (Mayer & Ellersieck, 1986) (Lintott, 1992) Pendimethalin 0.3 mg/L (20 °C) Channel Catfish 96-h LC50 = 420 (BCPC, 1991) * Technical grade pesticide used for bioassay. Some EC50s were included, which show combined sublethal/lethal effect concentrations. In some cases the toxicity endpoints exceed solubility. The cause of which is unknown. 3.3.2 Sublethal effects of Pesticides to Fish Table 5 showed various acutely lethal concentrations of the pesticides most likely used in the study area. However, what is the likelihood of the study areas' drainage ditches, or the Nicomekl River itself experiencing these lethal concentration? In addition, if lethal concentrations do arise, could fish seek refuge and avoid regions of the drainage ditches and the Nicomekl River which are contaminated to this degree. Folmar (1976) found that fish showed avoidance to the herbicides diquat and glyphosate once concentrations reached 10,000 //g/L for each, a concentration close to those producing lethality 12,000 fig/L and 50,000 //g/L, respectively. Hildebrand et al. (1982) found that rainbow trout avoided diquat concentrations of 40,000 //g/L, and should be able to avoid acutely toxic concentrations. Although, concentrations: of this magnitude are very unlikely. "Mass mortality of fish due to pesticide exposure is rare, and results only from accidents or direct spraying of the water bodies. More commonly, fish are subjected to long-term stress arising from exposure to sublethal concentrations. In the long run, these sublethal concentrations may prove more deleterious than lethal concentrations, because subtle and small effects on the fish may alter their behavior, feeding habits, position in the school, reproductive success, etc. Behavioral or morphological changes may make the fish more conspicuous in the environment and more susceptible to predation or parasitisation, thereby reducing the ability of the population to survive and reproduce. Likewise, subtle effects at the organ or cellular level may alter the metabolism of the fish and hence its ability to withstand stress. Even if the fish is not directly affected, any effect on fish-food organisms may result in a starved population offish." (Murty, 1986) 41 Sublethal effects to fish inhabiting the study a reas dra inage di tches or the Nicomekl River itself are more feas ib le than lethal effects, g iven that much lower concentrat ions of pest ic ides are required. W e i s s (1961) reported that f ish exposed to 100 / /g/L O P insecticides showed inhibition of brain acetychol inesterase, and that regenerat ion of this e n z y m e to normal levels may take up to 30 days. Wi th respect to d iaz inon, W e i s s demonstrated that largemouth b a s s and fathead minnows exposed to 500 / /g /L d iaz inon that d id not dye had brain acetychol inesterase activity reduced by 82%. W e i s & W e i s (1975) found that regeneration of removed caudal f ins of the killifish Fundulus heteroclitus was retarded after two weeks of exposure to only 1 0 / / g / L o f the O P insect ic ides malathion and parathion, and the carbamate insecticide carbaryl. Dodson & Mayf ie ld (1979) showed that ra inbow trout exposed to 500 / /g/L of the herbic ide diquat for 24 hours showed a significant dec rease in swimming speeds , and modif icat ions in rheotaxis, result ing in an increased inc idence of downst ream drift. Hatf ield and Ande rson (1972) found that after 24 hours of exposure to 1,000 / /g /L of the O P insect ic ide fenitrothion, At lant ic sa lmon (Salmo salar) parr were more vulnerable than unexposed f ish to predation by large brook trout (Salvelinus fontinalis). Symons (1973) reported that At lant ic sa lmon exposed to 100 / /g /L and 1000 / /g/L fenitrothion for 15 hours showed a 2 0 % and 5 0 % dec rease in the number of f ish holding territories 6 days fol lowing the treatment, respect ively. S y m o n s (1973) a lso showed that young At lant ic sa lmon regurgitated food that w a s contaminated with fenitrothion. Bul l and Mc lnerney (1974) found that coho (Oncorhynchus kisutch) e x p o s e d to fenitrothion for 2 hours at concentrat ions of 100 to 230 / i g / L exhibi ted decreased feeding and those exposed to 480 to 750 / /g/L exhibited dec reased locomotion and inability to maintain posit ion in a current. 42 Lorz et al. (1979) studied the downstream migration of yearling coho salmon smolts held in sublethal concentrations of three herbicides for periods of 4 or 15 days, and then marked and released into a nearby stream. He found no significant effect on downstream migration of the groups of 50-100 fish exposed to Tordan (contains 2,4-D and picloram) or dinoseb (no longer in use). However, diquat inhibited downstream migration offish groups exposed for 4 or 15 days at concentrations as low as 2% of the 96-h LC50. Using Folmar's (1976) rainbow trout 96-h LC50 for diquat of 12,000 //g/L as an approximation of the 96-h LC50 for coho, 2% would be roughly 200 //g/L. This is only a brief review of the numerous sublethal effects that insecticides and herbicides have on fish. With the exception of reduced brain acetycholinesterase activity, those shown above are primarily behavioral effects. Numerous studies referenced by Murty (1986) show various physiological effects produced by sublethal concentrations of pesticides; however the implications of such effects to overall organism health are often unclear. The point this thesis wishes to stress is that adverse sublethal behavioral effects such as disturbance to locomotion, feeding, the maintenance of channel position, and reduction in smolting migration have-beenrdemonstrated at concentrations much lower than concentrations producing acute lethality often in relatively short exposures. With respect to OP pesticides, in general much lower concentrations are required to affect invertebrates and hence potentially impact the fishes' food source. Invertebrates may be killed, or chemically induced to drift out of a ditch or a reach of the river at very low concentrations of this class of compounds. 43 3.3.3 Sublethal Effects of Pesticides to Invertebrates With respect to sublethal effects of diazinon on invertebrates, Morgan (1976) found the following: C. tentans reared in a concentration of only 0.003 //g/L (0.11X the 7-day LC50) for 7 days significantly delayed egg hatch, increased the duration of the larva stage, slightly depressed, pupation and emergence of adults from puparia, and overall lengthened the time of development from egg to larvae by 33.6%. In addition, fewer adult flies emerged from the 0.003 //g/L treatment. Morgan concluded that these sublethal effects could result in a shift in C. tentans usual life cycle such that organisms would emerge when environmental conditions were less favorable, and that the number of generations produced per year could be reduced. The crayfish O. propinquus exposed to 3 //g/L diazinon (0.2X the 7-day LC50) in flow-through testing increased the locomotion of males and females, and the number and duration of fights between males and females. The amphipod G. lacustris reared in 3 //g/L diazinon (0.02X the 7-day LC50) showed a 1.7X increase in its level of activity. Increased locomotion and activity is both a waste of energy reserves and could expose the invertebrates increasing their vulnerability to predators. It is believed that these results were all for purely aqueous tests conducted in the laboratory, not tests using organisms inhabiting:sediment in the natural environment. Morgan also performed field testing by dosing a stream with a concentration of 3 //g/L diazinon for 20 minutes, at 3-day intervals for 11 weeks. Morgan observed that there was increased drift of chironomid larvae, Ephemerella sp., Chloroterpes sp., Hydropsyche sp., and Cheumatopsyche sp. within the first 3 hours of stream treatment (Morgan, 1976). Drift of chironomid larvae, Hydropsyche sp., and Cheumatopsyche sp. was observed for greater than three hours after every dosing, in the first 5 weeks of the total 11 week study period. 44 Fewer animals and fewer species were collected at the sampling station closest to the source of diazinon than 300 m and 600 m downstream during the diazinon treatment. However, while Morgan found that there were changes in the benthos' species diversity while diazinon dosing was underway, these changes were temporary, and the communities recovered within 4 weeks after the diazinon treatment stopped (Morgan, 1976). On the basis of his study results, Morgan stated "that in order to protect aquatic invertebrates, a maximum permissible concentration of diazinon in receiving water should not exceed 3 //g/L" at any time". However, this was.based on field work in which this was the lowest concentration tested. Other authors have performed similar studies using lower diazinon dosings. Arthur et al. (1983) found that total macroinvertebrate density in a channel which had been dosed continuously with diazinon at 0.3 //g/L for 12 weeks (May to August), followed by 5.0 //g/L for 4 weeks (August to September) was only 5% lower than their un-dosed control channel. Likewise, they also found that total macroinvertebrate density in a channel which had been dosed continuously with 3.0 //g/L for 12 weeks (May to August), followed by 8.0 //g/L for 4 weeks (August to September) and 20.0 /^ g/L for 2 weeks (September) was only 25% lower than their un-dosed channel, Arthur et al. (1983) did not observe increased drift of benthic invertebrates during the initial 12 hour period of the 0.3 ug/L or 3.0 ug/L dosing. Sustained drift rates were 5 to 7.8 times higher in the treated channels (0.3 and 3.0 //g/L) than the control, but only 3 weeks after dosing began. Increasing the low dose channel from 0.3 to 5.0 //g/L slightly increased drift. Increasing the high dose channel from 3 to 8 //g/L sharply increased drift. Amphipods were the most abundantly drifting organisms (Arthur et al., 1983). 45 In terms of overall effect, for their low dosing regime, 3 months of exposure to 0.3 //g/L diazinon had no effect on Hyallela azteca abundance; however after 11 days of exposure to the increased concentration of 5 //g/L, the experimental channels Hyallela azteca population was eliminated (Arthur et al., 1983). Chironomid abundance did not appear to be impacted by either the low or high dosing regime. Mayflies, caddisflies, damselflies were not found in either low or high dose channels in the latter part of the study; however, there does not appear to have been an affect on the abundance of these organisms during the first 12 weeks of continuous exposure to either 0.3 or 3.0 //g/L (Arthur et al., 1983). Arthur et al. categorized benthic macroinvertebrates in terms of their sensitivity to diazinon as: tolerant - flatworms, physid snails, isopods and chironomids; moderately sensitive -leeches and the amphipod Crangonyx; sensitive - the amphipod Hyallela, mayflies, caddisflies and damselflies. Arthur et al. (1983) concluded by stating "it is unknown how much lower than 0.3 //g/L of diazinon would be fully needed to protect the macroinvertebrate communities in outdoor experimental channels". However, while Arthur et al. (1983) did observe an effect on benthic invertebrates at 0.3 //g/L, effect was not observed for short term (days) exposure. Macek et al. (1972) investigated the effects of chlorpyrifos application to ponds for mosquito control on non-target invertebrates and fish. Their experiment involved dosing experimental ponds twice over 63 days with the recommended highest dosing of 0.05 lb Dursban per acre of pond. Pond concentrations 24 hours after the first and second dosing were 2.39 ug/L and 2.03 uo/L. respectively. The concentrations prior to the second dosing and at the end of their 49 day the test were only 0.20 //g/L and 0.05 //g/L, respectively. The chlorpyrifos application severely impacted invertebrate plate colonization, relative to 46 that of the control pond. The number of insects colonizing plate samplers w a s reduced by 75%, mayfly colonization was severely reduced, and caddisfly colonization w a s completely e l iminated. L ikewise, the 2.39 / /g/L concentrat ion of chlorpyri fos kil led 5 5 % of the 275 bluegills and 4 6 % of the 275 largemouth bass wh ich had been introduced to the pond for the test. Muirhead-Thomson (1978) found the amphipod Bammarus pulex to be the most prone to channe l drift at sublethal concentrat ions of chlorpyrifos. Pusey et al. (1994) reported that an artificial stream dosed for 6 hours with chlorpyrifos at 0.1 / /g/L had no signif icant effect on stream macroinvertebrate communit ies: however, 6 hour dos ing at 5.0 / /g/L did reduce the abundance of severa l spec ies . For subsequent testing, Ward et al. (1995) reported that an artificial s t ream exposed to chlorpyri fos for 21 d a y s at the same concentrat ions exhibited a dec reased number of taxa and total inver tebrate abundance for both the high (5.0 / /g/L) and low doses (0.1 / /g/L). The individual abundances of 9 of the 36 non-chironomid taxa and 13 of the chironomid taxa coun ted were signif icantly reduced by the study's insect ic ide appl icat ion (Ward et a l . , 1995). Eaton et a l . (1985) reported reduced taxon r ichness in an artificial stream for repeated acute exposures to 3.1 //g/L chlorpyrifos. and a 100 day exposure to 0.22 / /g/L. W i t h respec t to O P pest ic ides, C. dubia does not appear to demonstrate signif icant inhibition of reproduction to diazinon in the 7-day chronic test, other than that due to death of the organ isms. Katzne lson and Mumley (1997) observed that C. dubia exposed to urban runoff samples from residential and mixed land use catchments where d iaz inon w a s shown to be a major c a u s e of toxicity produced numerous offspring before they d ied, in s a m p l e s with d iaz inon concentrat ions low enough to a l low time for reproduct ion, those 47 causing mortality after 5 to 7 days. Statistical analyses showed that reproduction was not inhibited prior to death. Hansen (1994) and WCC (1996) were reported to have observed minor inhibition of reproduction to C. dubia by diazinon, in laboratory water experiments (Katznelson and Mumley, 1997). It is probable that other acetycholinesterase inhibiting OP insecticides may also not significantly inhibit C. dubia reproduction irrespective of mortality. With respect to herbicides, Folmar et al. (1979) found significant increases in stream drift for Chironomus plumosus after 12 hours of exposure to 2,000 //g/L of the herbicide Roundup (Glyphosate). Inhibition C. dubia reproduction was reported by Ort et al. (1994) for Bicep (atrazine and metolachlor) and Extrazine (atrazine and cyanazine); however acute-to-chronic ratios comparing the 48-h LC50's (shown previously) with the 7-day NOEC-LOEC values of 8,840 //g/L and 17,680 //g/L were only 1.80 and 1.86, respectively (Ort et al., 1994). Kersting and van der Honing (1981) observed a decrease in feeding and filtering rates for Daphnia magna after 4 hours of exposure to dichlobenil, with a lowest observed effect concentration of 10% of the 48-h LC50. Schober and Lampert (1977) determined the inhibition of:reproduction. LOEC for Daphnia pulex exposed to atrazine for 28 days as 1,000 //g/L. Presing (1981) determined the inhibition of reproduction LOEC for Daphnia magna exposed to 2,4-D for 21 days as 25,000 //g/L (8% its 48-h LC50). Fitzmayer et al. (1982) determined the inhibition of reproduction LOEC for Daphnia pulex exposed to simazine for 26 days as 4,000 //g/L (4% its 48-h LC50). 3.3.4 Environmental Factors Affecting Pesticide Toxicity As can be observed in Table 5, multiple toxicity tests on a single compound using the 48 same organism often produce different toxicity endpoints due to variat ions in the chemica l formulation, age of organism, test apparatus, etc. A brief descr ipt ion of the environmental (and test) parameters which affect pesticide toxicity follows. Many pest ic ides are avai lab le for purchase by farmers in different forms. D iaz inon is avai lab le as a wettable powder (in which diazinon is 5 0 % of the powder by weight) or as an emulsifiable concentrate (500 g/L d iaz inon) . In genera l , emulsi f iable concentrates are more toxic than technical g rade material, whereas dust and powder formulations are less toxic (Murty, 1986), likely due to their reduced solubility. Commercial pesticide products are often more toxic than technical grade material as the act ive ingredient is often combined with synergists to increase the products ef fect iveness. For example, Johnson and F in ley (1980) reported that the herb ic ide Roundup w a s 3 - 4 2 t imes more toxic than the technical g rade of its act ive ingredient, g lyphosate. Most toxicologists use technical grade material for their testing; however , even technical grade material will vary in its percent purity. The age of the chemical (or stock solut ions) used by the farmer or toxicologist may affect its toxicity due to hydro lys is , biological degradat ion, adsorpt ion to storage vesse ls , and the poss ib le formation of more toxic degradation byproducts. The toxicity of aged solut ions of technical grade d iaz inon in water has actual ly been shown to increase due to the formation of the more toxic breakdown compound d iazoxon (Murty, 1986). Toxicity tests conducted in p las t ic vesse l s may produce higher L C 5 0 ' s than those performed in g lass vesse l s depend ing on the propensity of the pest ic ides to bind to plast ics. Wi th the except ion of malathion, the activity of most O P insect ic ide compounds is not inf luenced by p H , and pH only af fects toxicity with respect to its control on the rate of hydrolysis (Murty, 1986; H e n d e r s o n & P icker ing, 1957). Mos t pest ic ides are more toxic to f ish at higher temperatures , eg chlorpyrifos (Macek et a l . , 1969) and d iaz inon (Murty, 1986). Wa te r 49 hardness does not influence the toxicity of most pesticides (Henderson & Pickering, 1957), with the exception of diquat, which was less toxic in hard water (Murty, 1986). Turbidity can affect the toxicity of organic compounds in solution, depending on the organism being tested. Toxicity to fish should decrease in more turbid waters due to increased adsorption of the more non-polar pesticides to suspended sediments. However, with respect to filter feeding invertebrates the situation could be the opposite, depending on the bioavailability of the adsorbed chemical. Fredeen (1953) found that DDT applied to a region of the Saskatchewan River with a TSS concentration of 551 mg/L eliminated blackfly larvae for a 98 mile stretch, whereas similar dosing of the same river in a region with low TSS only eliminated blackfly larvae for a 9 mile stretch. Fredeen attributed the greater toxicity in the areas of higher turbidity to the fact that the blackfly larvae are filter feeders and their exposure to the insecticide was actually increased due to its adsorption to the suspended solids it was feeding on. Two insecticides and one herbicide were identified in this investigation. More detailed information is now provided on the chemistry, use, and persistence of these three pesticides. 3.3.5 The Chemistry, Use, and Persistence of Diazinon Diazinon (0,0-diethyl 0-2-isopropyl-6-methylpyrimidin-4-yl phosphorothioate) (^12^21^2^3^^)' i s a c ' e a r colourless liquid, which acts on animals by way of acetycholinesterase inhibition. It has a vapour pressure of 0.097 mPa (20°C), with a solubility in water of 40 mg/L (20°C) (BCPC, 1991). Diazinon is currently sold in B.C. for agriculture under the brand names of Diazinon and Basudin by United Agri Products and 50 C i b a C r o p Protect ion a s a wettable powder ( W P ) (50% diazinon) or a n emulsi f iable concentrate (EC) (500 g/L diazinon) ( B C M A F F , 1996a). It has a lso been so ld under the b rand n a m e s of Neoc ida l , Nuc ido l , and Knox Out ( B C P C , 1991), and D a z z e l , D iagran, D ianon , Diaterrfos, Diazajet, D iaz ide, Didazi ta l , D iazo l , D iz inon, Dyzo l , Gardentox, Kayazol , Nipsan, Sarolex, and Spectracide (Smith, 1987). The B C M A F F recommends the p o w d e r e d d iaz inon be used as a s e e d treatment for corn, mixed with a fungic ide (75% captan or 7 5 % thiram), and that a solut ion of d iaz inon made from wettable powder (WP) or emulsifiable concentrate (EC) be used as a pre-planting emers ion treatment for potato pieces or sprayed with ground equipment on potato and corn crops when insects appear ( B C M A F F , 1996a). D iaz inon is used to protect corn from root maggots and protect potatoes from aphids, Co lo rado potato beet les, f lea beet les, leafhoppers, and leafminers ( B C M A F F , 1996a). Diazinon's recommended B.C. usage rate is 1.11 kg/ha and 1.11 L/ha (445 mL/acre) , W P and E C , respect ively ( B C M A F F , 1996a). D iaz inon is reported to have a soi l half-life of 32 to 48 days (Rao and Dav idson, 1980). In aqueous solutions, diazinon degrades by hydrolysis. At 20°C it has a hydrolysis half-life of 11.8 hours (pH 3.1), 185 days (pH 7.4) and 6.0 days (pH 10.4) ( B C P C , 1991). Morgan (1976) reported a hydrolysis half-life of 43 days at 16°C and pH 7.6. In addit ion to hydro lys is , d iaz inon can be biological ly degraded. Katzne lson and Mumley (1997) reported a half-life of 7 to 40 days in sur face water samples . Sethunathan and Pathek (1972) found that for prepared solut ions of d iaz inon in water from rice f ie lds wh ich had previously been exposed to repeated d iaz inon dos ing, bacterial degradat ion of d iaz inon by Arthrobacter and Flavabactehum sp . w a s rapid (23 mg/L to 0 mg/L in 5 days)(23 °C), whereas incubation of d iaz inon in water from untreated rice f ie lds showed little bacterial 51 degradat ion of d iaz inon (23 mg/L to 20 mg/L in 12 days). Further, they found that water f rom f ields previously treated with chlorpyri fos or carbofuran did not acce lerate the degradat ion of d iaz inon. Hence , the development of bacter ia speci f ical ly ab le to rapidly deg rade d iaz inon may occur in aquat ic environments wh ich have rece ived numerous previous diazinon applications. Sethunathan and Pathak (1972) a lso found that d iaz inon persisted longer in solut ions wh ich had been incubated with soi l than those without soi l , stat ing that adsorpt ion to soi l part ic les dec reased biological degradat ion/hydrolysis. D iaz i non has been reported to accumulate in creek sediments and persist for w e e k s ( W C C , 1996). 3.3.6 The Chemistry, Use, and Persistence of Chlorpyrifos Chlorpyrifos (0,0-diethyl 0-3,5,6-tr ichloro-2-pyridyl phosphorothioate) ( C g H ^ C ^ N O s P S ) is a colourless crystal and a lso acts on an imals by way of acetychol inesterase inhibition. It has a vapour pressure of 2.5 m P a (25°C) and a solubil i ty in water of 2 mg/L (25°C) ( B C P C , 1991). Chlorpyrifos is currently sold in B.C. for agriculture under the brand names of L o r s b a n 4 E , Pyr inex 480 E C by D o w E lanco and United Agr i Products, sole ly as an emuls i f iab le concentrate (EC) (480 g/L, 10L jugs) ( B C M A F F , 1996a). It has a lso been sold under the brand names of Dursban, Spannit, and S i l f r i fos . (BCPC, 1991) and Brodan, D o w c o 179, Eradex, and Ki l lmaster (Smith, 1987). The B C M A F F recommends that chlorpyr i fos be sprayed with ground equipment when insect damage appears on corn, potatoes, and sugar beets for the control of cutworms and Co lo rado potato beet les ( B C M A F F , 1996a). Chlorpyr i fos ' recommended B .C . dosage rate is 1.0 to 2.4 L/ha (405 to 970 mL/acre) for vegetable crops ( B C M A F F , 1996a). 52 Chlorpyrifos persists in soils for 60 to 120 days (BCPC, 1991). In solution, its rate of hydrolysis increases with pH, the presence of copper, and possibly other metals. It has a hydrolysis half-life of 1.5 days (pH 8 and 25°C) to 100 days (phosphate buffer at pH 7 and 15°C)(BCPC, 1991). Hughes et al. (1980) dosed a natural pond lined with leaf litter with 10 //g/L chlorpyrifos, and found that in 18 days chlorpyrifos concentrations were reduced to 0.01 //g/L. Compared with diazinon, chlorpyrifos has a higher tendency to adsorb to solid surfaces, resulting in reduced migration in water runoff (Katznelson and Mumley, 1997), but perhaps greater transport on suspended sediments.. 3.3.7 The Chemistry, Use, and Persistence of Prometryn Prometryn (N2,N4-di-isopropyl-6-methylthio-1,3,5-triazine-2,4-diamine) (C 1oH 1gN 5S) is a triazine herbicide, appears as a white powder, and acts on plants by inhibiting photosynthetic electron transport. It has a vapour pressure of 0.133 mPa (20°C), with a solubility in water of 33 mg/L (20°C) (U.S. EPA, 1996), 48 mg/L (20°C) Humburg et al., 1989). Prometryn is currently sold in B.C. for agriculture under the brand name of Gesagard as a wettable powder (WP) (80% Prometryn) (BCMAFF, 1996a). It has also been sold as Caparol, Primatol, Prometrex (U.S. EPA, 1996), and in mixtures as Codal, Cotogard, Gesatene, and Peaweed (BCPC, 1991). BCMAFF recommends that prometryn be wetted and sprayed with ground equipment only on peas after planting, prior to emergence, to control the weeds such as lamb's quarters, redroot pigweed, corn spurry, wild mustard, lady's thumb, hemp-nettle, common chickweed, green foxtail, and purslane (BCMAFF, 1996a). Prometryn's recommended B.C. dosage rate is 2.3 to 2.8 kg/ha (0.93 to 1.13 kg/acre) (BCMAFF, 1996a). BCPC states that prometryn can also be used on 53 broad beans, carrots, celery, cotton, leeks, lentils, parsley, potatoes, and sunflowers (BCPC, 1991). Humburg et al. (1989) states that prometryn is used on cotton and celery. Prometryn has a half-life in soils of 40 to 70 days, and in slightly acid, neutral, or slightly alkaline solutions at 20°C it is relatively stable to hydrolysis (BCPC, 1991). Humburg et al. (1989) states that prometryn is most readily adsorbed to soils containing clays and organic matter, and adsorbs more readily than most other commercial triazine herbicides. Prometryn undergoes microbial breakdown by several microbial organism which utilize it as a source of energy, nitrogen, and sulphur (Humburg et al., 1989). Prometryn resists hydrolysis and photolysis, and has a half-life in solution in excess of 270 days (U.S. EPA, 1996). 3.4 Ammonia Contamination 3.4.1 Farm Animals Table 6 shows the abundance of farm animals in different regions of the Fraser River Estuary Management Plan (FREMP) study area in the Lower Fraser Valley (LFV) for 1991 (FREMP, 1996). Langley and Surrey, the two regions the Nicomekl traverses, have the greatest abundance of farm animals of any region of the Fraser River Estuary Management Plan. Consequently, the potential exists for manure-runoff ammonia contamination of the Nicomekl River, and/or its tributaries and drainage ditches due to manure produced in the region; whether it be dilute runoff from the lands rearing the animals, concentrated runoff from manure piles, or runoff from farm lands to which the manure from the region's animals has been applied. 54 Table 6. Distribution of Farm Animals in the Lower Fraser Valley (FREMP, 1996). Chickens Turkeys Cattle & Calves Pigs Sheep Horses Mink Langley 1585054 161328 14303 17428 6715 3860 100137 Surrey 1133045 117805 19101 4442 1709 1225 25752 Richmond NA NA 1632 NA 333 325 0 Delta 1088 78 5147 81 637 591 0 Maple Ridge 45173 NA 2022 209 780 587 0 Pitt Meadows 1341 NA 6475 58 207 134 0 Total FREMP 2765701 279211 48680 22218 10381 6722 125889 Total LFV 8796715 718786 131472 154726 14233 8635 189233 NA = data not available It does not appear that there are many farm animals on the farmland immediately adjacent to the river in the study area between Sites 1 and 12, with the exception of a few cows, chickens, and horses on small hobby farms. However, the B.C. Ministry of Environment Water Management Branch did identify numerous fur farms and henneries located on the nd th th lands south of 32 Ave. on a 1982 map of the area between 160 and 180 St., shown in Figure 8 (BCME, 1982). The Old Logging Ditch, Burrows Ditch, and Ericson Ditch nd th extend into this area. Further, this region between 32 and 24 Ave. is elevated and likely drains into these ditches. Consequently, ammonia contamination of the drainage ditches in this area from animal waste storage, or runoff from the lands where animals are being reared appears to be feasible. 55 Figure 8. Fur farms and Henneries Located In the Study Site and its Uplands (BCME, 1982). 56 While, ammonia contamination of the river in the study area may occur due to this regions discharge of ditches contaminated with animal waste ammonia, elevated river ammonia levels may also be caused by animal waste ammonia contamination originating upstream of the study site, in the river's regions of higher farm animal abundance. Swain & Holms (1988b) identified the lands bordering the north side of the Nicomekl River adjacent to 192 n d St., and the headlands of Anderson Creek at 216 t n St. and 24 t h Ave. as locations of significant cattle use. 3.4.2 Fertilizer The Nicomekl River and its drainage ditches risk possible ammonia contamination due to runoff from farmlands where manure and urea based fertilizers or pesticides have been applied. The Nicomekl-Serpentine basin is one of three regions in the Lower Fraser Valley where nitrogen application to the land is 100 - 130 kg N/ha in excess of that which is actually necessary for the agriculture being performed (FREMP, 1996) as shown in Figure 9. The nitrogen excess was calculated by adding manure and fertilizer inputs and then subtracting plant uptake by crop type, and denitrification. The herbicide Linuron is a urea based chemical, recommended for use on corn crops 38 cm in height, with relatively abundant use for agriculture in the Lower Mainland (Tables 2 & 4), and may be another possible source of runoff ditch ammonia contamination. 57 58 3.5 Literature Review of Ammonia Toxicity 3.5.1 Terrestrial Sources and Speciation of Ammonia In the Aquatic Environment Unionized ammonia (NH3), is a cellular metabolic waste, produced by both carnivorous and herbivorous animals during the amino acid deamination step of protein digestion. Ammonia is excreted by animals as urea [NH2-(C=0)-NH2] within their urine. Once urea enters aquatic environments (and likely moist soils), it undergoes hydrolysis to form unionized ammonia (NH3). Feces contains proteins, as the animals' unassimilated consumed protein and the animals' own dead blood cell and intestinal tissue. These proteins are converted to NH 3 through protein deamination described above. Runoff from lands which have been dosed with urine or feces is a potential aquatic ammonia source. An overabundance of farm animals, the use of manure and urea based fertilizers, and possibly urea herbicides such as Linuron, creates the potential for elevated ammonia concentrations in the receiving waters bordering these lands. 3.5.2 Fate of Ammonia in the Aquatic Environment Once in an aqueous medium, unionized ammonia (NH3) is converted to ammonium (NH 4 +) in the following reaction: NH3(g) + nH20(l) NH3(nH20)(aq.) N H 4 + + OH" + (n-1)H20(l) Water pH, and to a lesser extent temperature, are the predominant water quality parameters which control this reaction and the relative concentrations of NH 3 and NH 4 + . At higher pH's and temperatures more total ammonia is in the form of unionized ammonia (NH3). Very roughly, for a one unit pH increase and a 1 °C temperature increase, the unionized ammonia (NH3) concentrations increase 10 fold and 1.1 fold, respectively. 59 Under aerobic conditions, unionized ammonia may be converted to nitrite (N0 2 ) by Nitrosomonas sp. bacteria, and the nitrite may then be simultaneously converted to nitrate ( N 0 3 ) by Nitrobacter sp. bacteria. Nitrite is very toxic to fish [96-h LC50 rainbow trout Onchorhynchus mykiss = 0.19 - 0.39 mg/L N0 2 (CCME, 1986)]; however, the nitrite/nitrate conversions generally happen simultaneously, such that NH 3 is essentially converted directly to the relatively non-toxic N0 3 " [96-h LC50 rainbow trout Onchorhynchus mykiss = 6.0 g/L (CCME, 1986)] without elevating aquatic N0 2 " concentrations. Ammonia (and nitrite) which is not converted to nitrate has an appreciable potential to cause toxicity. Unionized ammonia, being uncharged and hydrophobic, easily diffuses through phospholipid cell membranes. In contrast, the ammonium ion's charge makes it hydrophilic and much more restricted in diffusing through cell membrane (Haywood, 1983). Consequently, ammonia toxicity to aquatic life is predominantly caused by unionized ammonia (NH3). Thurston & Russo (1981a) found that N H 4 + may be toxic to salmonids as well as NH 3 ; however, they concluded that NH 3 is 300 to 400 times more toxic than N H 4 + Laboratory measurements of ammonia often measure total ammonia (NH 3 and NH 4 +) using the colorimetric "Phenate Method" and typically report the concentration as NH3-N (total ammonia as N). This nomenclature should not be confused with the concentration of unionized ammonia (NH3) or unionized ammonia as N (NH3-N). Ammonia toxicity test endpoints are commonly reported as NH3-N or NH 3 (mg/L). To aid in the comparison of this studies ammonia measurements with the toxicity endpoints from the literature and the Canadian Water Quality Guidelines for ammonia, from herein all toxicity endpoints 60 r ev iewed wil l be in terms of the concentrat ion of un ion ized ammon ia a s nitrogen mg/L N H 3 - N . The necessary convers ions have a l ready been made from the original literature where required. 3.5.3 Toxicity to Fish Lethal Toxicity Arthur et a l . (1987) ranked the sensit ivity of 5 different f reshwater f ish to un ion ized ammonia, from most to least sensitive, with their respective mean 96-h L C 5 0 ' s a s ra inbow trout {Onchorynchus mykiss) (0.44 mg/L N H 3 _ N ) > wa l leye {Stizostedion vitreum) (0.54 mg /L N H 3 - N ) > channe l catf ish (Ictalurus punctatus) (0.71 mg/L N H 3 - N ) > white sucker (Catastomus commersoni) (1.26 mg/L N H 3 - N ) > fathead minnow (Pimephales promelas) (1.78 mg/L N H 3 - N ) , all tests at pH 7.7 to 8.3). Sa lmon ids are extremely sensi t ive to un ion ized ammonia. Thurston and R u s s o (1983) found 96-h LC50 's for rainbow trout ranging from 0.17 to 0.37 mg/L N H 3 - N ( @ 10 - 1 3 °C and pH 7.7 - 7.9). Observab le acute responses for sa lmonids to N H 3 are typical ly hyperexcitabil i ty, hyperventi lat ion, violent erratic movements, convuls ion and coma, lead ing to death. T h e s e effects appear to be . caused by neurological disorder. Acu te lethal i ty wi l l usual ly occur within 96-h of testing, if it is to occur at al l , a s shown by Thurston and R u s s o (1983) wh ich s h o w e d little di f ference between 35-day L C 5 0 va lues and 96-h L C 5 0 va lues. 61 Life Stage Sensitivity A salmonids' developmental stage affects its sensitivity to unionized ammonia. Rice and Stokes (1975) found that the 24-h median tolerance limit (TLm) for both fertilized eggs and alevins of rainbow trout was > 2.95 mg/L NH3-N (the highest concentration tested), but til that at the end of the alevins' yolk absorption (50 day of development) the 24-h TLm was only 0.059 mg/L NH3-N, the same value they reported for adult trout. Rice and Stokes also found that egg fertilization was not prevented in concentrations up to 1.47 mg/L NH 3 -N (again the highest concentration tested). The:resistance of eggs.and alevins to acute ammonia exposure is similar to that observed for other compounds such as endrin (Wenger, 1973), DDT (Burdick et al., 1964), and zinc (Skidmore, 1965). Rice and Stokes (1975) did not believe that the eggs' membranes were offering protection from ammonia absorption since newly hatched alevins did not demonstrate a greater susceptibility to ammonia than the eggs. They hypothesized that increased sensitivity to NH 3 upon the transition to the fry stage is likely due to physiological changes following yolk absorption. The sensitivity of rainbow trout fry and adults to NH 3 is similar; however, there is some variation in sensitivity at different life stages,beyond the alevin stage (Rice and Stokes, 1975). For post-alevin rainbow trout (fry and beyond), the sensitivity to NH 3 decreased as the fish developed through the larval stages, was the lowest at the juvenile and yearling stages, and increased with age thereafter (Thurston and Russo, 1983). Thurston and Russo found 96-h LC50s of approximately 0.3, 0.6, 0.3, 0.2 mg/L NH3-N for fish weights of 0.4, 4.0, 40, and 400 g, respectively (the actual paper showed results for much smaller increments of weight increase). Regression of their data suggested that the greatest tolerance to NH 3 occurred at roughly 1-2 g fish weight (96-h LC50 approximately 0.5 - 0.7 62 mg/L NH 3_N) (Thurston and Russo, 1983). This weight is the approximate size of the Nicomekl hatchery (Rhidine, 1997) and natural coho in the spring commencing these fishes' first year of rearing in the Nicomekl River and its tributaries. Fish 10 - 20 g in weight are less tolerant to NH 3 with 96-h LC50's of approximately 0.2 - 0.3 mg/L NH3-N (Thurston and Russo, 1983). Consequently, two year old smolting salmonids migrating through, or completing their rearing, in the study area would likely be the most sensitive to ammonia contamination. Sublethal Toxicity Chronic exposure to much lower ammonia concentrations may not produce the neurological disorders observed at higher ammonia concentrations, but can produce other sublethal biological effects. Carbalo and Munoz (1991) found that juvenile rainbow trout exposed to 0.06 mg/L NH3-N for 10 days (@ 15 °C, pH 7.3) exhibited a decreased resistance to fungal infection. Smith and Piper (1975) exposed rainbow trout to 0.013 mg/L NH3-N for 9.5 months. There was no mortality in test fish at 9.5 months; however, these fish became emaciated and lethargic. Examination of gill tissue from the lethargic fish at 9.5 months showed extensive proliferation of gill epithelial tissue and severe fusion of gill lamella, preventing normal respiration. Smith and Piper continued the exposure to 12 months. By the 12 t h month, fish exposed to NH 3 incurred bacterial gill disease infection, resulting in severe mortalities among the population. Thurston et al. (1984) found that adult rainbow trout exposed to > 0.06 mg/L NH3-N for 4 months (@ 9 °C, pH 7.7) had altered hematocrit (white blood cell count, related to disease resistance) and hemoglobin. 63 The Influence of Water Quality Parameters on Ammonia Toxicity Temperature and pH regulate the concentration of unionized ammonia in an aquatic system and are the dominant factors in determining whether enough NH 3 is available to cause toxicity. However, these parameters play an additional role than just controlling the NH 3 <-*• N H 4 + equilibrium. They also influence the toxicity of unionized ammonia. pH While lower pH's reduce the percentage of unionized ammonia, the unionized ammonia present is actually more toxic at lower pH's. Szumski et al. (1982) reported a 96-h LC50 for coho fingerling of 0.27 mg/L NH3-N (pH 7.0), 0.54 mg/L NH3-N (pH 7.3), 0.71 mg/L NH3-N (pH 7.6) and 0.88 mg/L NH3-N (pH 7.9), @ 15 °C. Temperature Higher temperatures increase the percentage of unionized ammonia present, although the change is not as drastic as that observed for pH variations (Section 3.5.2). While higher temperatures increase the percentage of unionized ammonia, the unionized ammonia present is actually less toxic at higher temperatures. Arthur et al. (1987) reported a 96-h LC50 for rainbow trout fingerling of 0.21 mg/L NH3-N (4 °C), 0.50 mg/L NH3-N (10 °C), and 0.86 mg/L NH3-N (19 °C), @ pH 7.7 - 8.3. Dissolved Oxygen Unionized ammonia has a greater toxicity to salmonids at low D.O. levels. Haywood (1983) proposed two possible explanations for this phenomenon based on increased ventilation by the fish to meet oxygen requirements. The first is that increased ventilation 64 simply causes greater NH 3 exposure and uptake. The second is that increased ventilation flushes the gill chamber of excreted C0 2 , which raises gill chamber pH and causes a shift of more total ammonia to unionized ammonia than that found in the gill chamber under regular ventilation rates. Both are conceivable explanations. Thurston et al. (1981b) reported a 96-h LC50 for rainbow trout fingerling of 0.40 mg/L NH3-N (D.O. = 3.6 mg/L), 0.48 mg/L NH3-N (D.O. = 6.6 mg/L), and 0.62 mg/L NH3-N (DO. = 8.6 mg/L), (@ 12 - 13 °C, pH 7.8 - 7.9). Given that low dissolved oxygen concentrations in the Nicomekl River and its drainage ditches may coincide with elevated ammonia concentrations the Nicomekl system should be considered relatively sensitive to ammonia contamination. Water Hardness Water hardness does not appear to affect the toxicity of NH 3 (Haywood, 1983). N H o Acclimation and Avoidance Fish are capable of gradually acclimating to a uniform elevated total ammonia in their environment. Some teleosts have been found to contain ornithine-urea enzymes capable of converting ammonia to urea (Huggins et al., 1969; Read, 1971). Fish may also produce glutamine to reduce ammonia toxicity (Levi et al., 1974). Lloyd and Orr (1969) found that exposure of fish to sublethal levels of NH 3 increased the tolerance of these fish to lethal levels of NH 3 , but this resistance only lasted 3 days (perhaps by activating the above possible means of detoxification). Under conditions where waters are polluted by point source discharges, such as drainage ditch outflows, which produce ammonia concentration gradients, fish may avoid the 65 regions of high ammonia. Stickleback (Jones, 1948) and green sunfish (Summerfelt and Lewis, 1967) showed avoidance to lethal concentrations of NH 3 in a gradient tank. Consequently, if the drainage ditches are discharging lethal concentrations of ammonia into the Nicomekl River, and these discharges are not adequately diltuted, fish migration up or down the river may be impeded if fish are actively avoiding zones of the River contaminated with ammonia. Avoidance/attraction to sublethal NH 3 concentrations is less clear. Green sunfish were not repelled by sublethal NH 3 concentrations and stickleback were attracted to sublethal NH 3 concentrations. Hepner (1959) showed that carp avoided ammonia fertilizer added to ponds. 3.5.4 Toxicity to Invertebrates There is less concern over ammonia toxicity to invertebrates than salmonids since the majority of invertebrates are generally more resistant to elevated ammonia concentrations. Three authors' 96-h LC50s for unionized ammonia to 31 invertebrate species are shown in Table 7. Table 7. Acute lethality of unionized ammonia (mg/L NH3-N) to 31 invertebrate species. Organism 96-h LC50 (mg/L NH3-N) Reference Cladoceran, Simocephalus vetulus 0.5 (U.S. EPA, 1983) Flatworm, P. tenuis 0.58 (Kendall etal., 1986) Cladoceran, Ceriodaphnia acanthina 0.6 (U.S. EPA, 1983) Snail, L. stagnalis 0.8 (Kendall etal., 1986) Fingernail clam, Musculium transversum 0.91 (Arthur etal., 1987) (pH 7.7 - 8.3) 66 Cladoceran, Daphnia pulicaria 1.0 (U.S. EPA, 1983) Flatworm, Dendrocoelum lacteum 1.2 (U.S. EPA, 1983) Insect, L. inermis 1.3 (Kendall etal., 1986) Midge, Chironomus ripan'us 1.36 (Kendall etal., 1986) Insect, B. rodani 1.4 (Kendall etal., 1986) Snail, P. fontinalis 1.4 (Kendall etal., 1986) Mayfly, Callibaetis sp. 1.5 (U.S. EPA, 1983) Insect, E. ignita 1.52 (Kendall etal., 1986) Oligochaete, L. hoffmeisteri 1.58 (Kendall etal., 1986) Stonefly, Ancynopteryx parallels 1.6 (U.S. EPA, 1983) Snail, Physa gyrina 1.61 (Arthur etal., 1987) (pH 7.7 - 8.3) Amphipod, Gammarus pulex 1.69 (Kendall etal., 1986) Crustacean, A. aquaticus 1.9 (Kendall etal., 1986) Snail, Physa trivolvis 1.95 (Arthur et al., 1987) (pH 7.7 - 8.3) Tubificid worm, Tubifex tubifex 2.2 (U.S. EPA, 1983) Isopod, Asellus racovitzai 2.4 (U.S. EPA, 1983) Caddisfly, H. angustipennis 2.43 (Kendall etal., 1986) Amphipod, Cragonyx pseudogracilis 2.57 (Arthur etal., 1987) (pH 7.7 - 8.3) Crayfish, Orconectes nais 2.8 (U.S. EPA, 1983) Mayfly, Callibaetis skokianus 3.21 (Arthur etal., 1987) (pH 7.7 - 8.3) Mayfly, Ephemerella grandis 4.0 (U.S. EPA, 1983) Isopod, Asellus racovitzai 4.13 (Arthur etal., 1987) (pH 7.7 - 8.3) Aquatic beetle, Stenelmis sexlineata 6.6 (U.S. EPA, 1983) Caddisfly, Philartcus quaeris 8.32 (Arthur etal., 1987) (pH 7.7 - 8.3) Crayfish, Orconectes immumis 15.07 (Arthur etal., 1987) (pH 7.7 - 8.3) 67 However, note the high sensitivity to NH 3 of the cladocerans. In addition to the above values, Arthur et al. (1987) reported a 48-h LC50 for Simocephalus vetulus of 1.41 mg/L NH3-N. The 48-h LC50 for Daphnia magna was reported as 2.94 mg/L NH3-N (pH 8.3 -8.6 @ 20°C) (Gersich and Hopkins, 1986). Arthur et al. (1987) reported a 48-h LC50 for adult Ceriodaphnia vetulus of 1.1 mg/L NH3-N (pH 8.1 @ 20.4°C). Nimmo et al. (1989) reported a Ceriodaphnia dubia 48-h LC50 value of 1.06 mg/L NH3-N (pH 7.8 @ 25°C). With respect to sublethal toxicity to Ceriodaphnia dubia, Nimmo et al. (1989) reported that the 7-day reproduction inhibition concentration of unionized ammonia using river water was 0.68 to 0.88 mg/L NH3-N (pH 8 @ 25°C). Nimmo et al. (1989) pointed out that there is excellent agreement between the range of "acute" toxicity NH 3 to fishes (their fish toxicity endpoints were for johnny darters, larval fathead minnows, and juvenile fathead minnows) and Ceriodaphnia "chronic" limits under both cold and warm test conditions. Nimmo et al. (1989) concluded that Ceriodaphnia dubia chronic testing appears to be a useful surrogate for testing wastes or receiving waters for acute NH 3 toxicity to some fish. Although, the author points out that the standard 7-day sublethal Ceriodaphnia test at 25 °C may show no toxicity for a sample which could be lethal to rainbow trout. However, if a sample demonstrates sublethal effects in the Ceriodaphnia chronic test due to ammonia, the sample would most likely be acutely lethal to salmonids. 68 3.6 Pesticide, Ammonia, and Metal Water Quality Guidelines For the Protection of Aquatic Life. The Canadian Water Quality Guidlines for the Protection of Aquatic Life (CCME, 1986, 1989, 1991, 1993a, 1993b, 1994, 1995) do not stipulate maximum acceptable concentrations (MACs) for all of the pesticides likely used in the study area. Those for which Canadian guidelines currently exist are shown in Table 8. The development of Canadian water quality guidelines predominantly for herbicides and fungicides is likely based on these compounds higher solubility, and slower degradation, in comparison to the OP and carbamate insecticides. Table 8. Current Canadian Maximum Acceptable Concentrations (MACs) (//g/L) for Pesticides (CCME, 1986 to 1995, inclusive). Pesticide Use (MAC) Mg/L Glyphosate Herbicide 65 Captan Fungicide 2.8 Atrazine Herbicide 2 Metolachlor Herbicide 8 Simazine Herbicide 10 Dicamba Herbicide 10 MCPA Amine Herbicide 2.6 Bromoxynil Herbicide 5 Carbofuran Insecticide-Carbamate 1.75 Dimethoate Insecticide-OP 6.2 Diclofop methyl Herbicide 6.1 Davis et al. (1997) summarized various suggested U.S. water quality criteria for several of the insecticides likely used on the Nicomekl farmlands, not covered by the CCME water quality guidelines. These are shown in Table 9. 69 Table 9. U.S. Suggested Water Quality Criteria (fxg/L) for Various Insecticides (Davis etal., 1997). Pesticide Acute Chronic RMC Diazinon (OP) 0.08 1 0.04 1 Chlorpyrifos (OP) 0.083 2 0.041 2 Malathion (OP) 0.1 3 Parathion (OP) 0.065 2 0.013 2 Carbaryl (Carbamate) 0.02 4 Azinphos-methyl (OP) 0.01 3 Acute = Short Term Exposure Chronic = Long Term Exposure RMC = Recommended Maximum Concentration 1 (Menconi and Cox, 1994) for California Department of Fish and Game 2 Washington State Water Quality Standards, WAC 173-201A (reference unavailable) 3 (US EPA, 1986) 4 (NAS, 1973) The current Canadian Water Quality Guidelines (CCME, 1986) for ammonia are those adopted from U.S. EPA (1985a). The maximum acute (1-hour) and chronic (4-day) average total ammonia concentrations for the protection of salmonids and other sensitive cold water species are shown in Tables 10a and 10b. The concentrations have been converted to mg/L total ammonia as N (NH3-N), for comparison with the ammonia measurements performed on this study's water samples. 70 Table 10a. Maximum 1-Hour Average Total Ammonia Concentration (mg/L NH3-N) for the Protection of Salmonids and Other Cold Water Species (U.S. EPA, 1985a) PH 10°C 15°C 20°C 25°C 6.5 25.48 24.66 23.84 . 16.44 6.75 23.02 22.19 22.19 15.28 7.00 20.55 19.73 18.91 13.48 7.25 16.40 16.19 15.78 11.01 7.50 12.7 12.25 12.00 8.38 7.75 13.3 8.63 8.47 5.92 8.00 5.83 5.67 5.59 3.95 8.25 3.37 3.29 3.21 2.30 8.50 1.89 1.89 1.89 1.41 8.75 1.13 1.13 1.17 0.88 9.00 0.68 0.71 0.75 0.59 Table 10b. Maximum 4-Day Average Total Ammonia Concentration (mg/L NH3-N) for the Protection of Salmonids and Other Cold Water Species (U.S. EPA, 1985a) PH 10°C 15°C 20°C 25°C 6.5 1.81 1.81 1.22 0.85 6.75 1.81 1.81 1.22 0.85 7.00 1.81 1.81 1.22 0.85 7.25 1.81 1.81 1.22 0.85 7.50 1.81 1.81 1.23 0.86 7.75 1.73 1.81 1.15 0.81 8.00 1.13 1.09 0.76 0.54 8.25 0.64 0.62 0.44 0.32 8.50 0.37 0.36 0.26 0.19 8.75 0.21 0.22 0.16 0.12 9.00 0.13 0.13 0.11 0.08 71 The maximum acceptable aqueous total metal concentrations as suggested by the Canadian Water Quality Guidelines for Freshwater Aquatic Life (CCME, 1986) and BC Ministry of the Environment Approved and Working Criteria for Water Quality (BCME, 1989) are shown in Table 11. Table 11. Maximum Acceptable Aqueous Total Metal Concentrations (CCME, 1986; BCME, 1989). Metal Maximum Acceptable Concentration Source Ag o.i Mg/L CCME (1986) Al 0.1 mg/L, pH > 6.5, C a 2 + > 4.0 mg/L CCME (1986) As 50 //g/L CCME (1986) Ba 1,000 /ug/L BCME (1989) Be 1,100 /ug/L for hardness > 75 CCME (1986) Ca < 8,000 fxg/L BCME (1989) Cd 0.2 /ug/L for hardness 0 - 60 0.8 /ug/L for hardness 60-120 1.3 /ug/L for hardness 120-180 1.8 /ug/L for hardness > 180 CCME (1986) Co 50 ,ug/L BCME (1989) Cr 2.0 ,ug/L all aquatic life 20 /ug/L fish CCME (1986) Cu 2 /ug/L for hardness 0-120 3 //g/L for hardness 120-180 4 /ug/L for hardness > 180 CCME (1986) Fe 0.3 mg/L CCME (1986) Hg 0.1 ng(L CCME (1986) Mn 100- 1,000 Mg/L BCME (1989) Mo Acute: 2 mg/L Chronic (30 days): < 1 mg/L BCME (1989) Ni 25 z^ g/L for hardness 0 - 60 65 /ug/L for hardness 60-120 110 /ug/L for hardness 120-180 150 /ug/L for hardness > 180 CCME (1986) Pb 1 /ug/L for hardness 0 - 60 2 /ug/L for hardness 60 -120 4 /ug/L for hardness 120-180 7 /ug/L for hardness > 180 CCME (1986) Se 1 /ug/L CCME (1986) Zn 30 /ug/L CCME (1986) 72 3.7 The Relevance of the C. dubia test Daphnids are freshwater microcrustaceans belonging to the Order Cladocera. Cladocerans including Daphnia sp. and Ceriodaphnia sp., are ubiquitous in temperate and fresh waters, and abundant in lakes, ponds, and quiescent section of streams and rivers throughout North America (Pennak, 1978). They are ecologically important as they convert phytoplankton into protein and are a significant portion of the diet of numerous fishes including young salmonids. The C. dubia chronic toxicity test has been widely applied as a tool in screening for the presence of low levels of contaminants, often saving time and money versus conventional chemical testing (EC, 1992). Chemical testing is limited in its ability to fully measure water quality since the scientist has to choose specific compounds he/she wishes to test for, and may not select the contaminants present at a site that actually are having a significant adverse impact on aquatic life. In contrast, toxicity tests reflect the actual bioavailability of contaminants in a sample, as well as additive, antagonistic, and synergistic effects of multiple contaminants in the context of the basic water chemisty of the sample. With respect to pesticides, the C. dubia test will reveal the presence of chlorpyrifos and diazinon, as lethality to test organisms, at concentrations below the detection limits of conventional organic chemistry techniques. With the 7 ± 1 day C. dubia chronic test, inhibition-of reproduction may be a sensitive indicator of toxic concentrations of ammonia, metals, and herbicides. This test is particularly useful as an indicator of OP insecticides since the lethality endpoints for OP insecticides are much lower than the concentrations required to produce sublethal/lethal effects to fish and the majority of other invertebrates. 73 4.0 REVIEW OF PREVIOUS CHEMICAL AND BIOLOGICAL TESTING IN THE STUDY AREA To date, a number of studies have investigated the contamination of ditches, streams, and rivers by agricultural pesticides in the Lower Fraser Valley. Wan (1989) reported the results of chemical monitoring for diazinon, azinphosmethyl, fensulfothion, dinoseb and endosulfan at various locations in the Lower Fraser Valley, including the Nicomekl River's th 168 St. drainage ditch. Sampling was conducted for both 1985 and 1986 in the first week of May, in July, and during the rainy season of October, December, and February. In 1985 diazinon, azinphosmethyl, endosulfan, and fensulfothion, were not found at a detection limit of 1 //g/L in the ditch water at any of the study sites. Unfortunately this detection limit was higher than the toxicity level of diazinon (Table 5), and azinphosmethyl [Daphnia magna 48-h LC50 = 0.2 //g/L (FWPCA, 1968)]. The herbicide dinoseb was consistently found in ditch water at numerous sites in the Lower Fraser Valley (although, th not consistently at the Nicomekl 168 St. ditch). Dinoseb was not detected in early May water samples, but was detected in July, October, December, and February at concentrations ranging from 0.3 -18.6 //g/L (Wan, 1989). Wan also observed an instance where pesticide overspray created ditch water with pesticide concentrations that were likely highly toxic to fish ("catastrophic event"). The ditch was located on Westham Island, and ditchwater endosulfan concentrations went from non detectable levels to 500-2700 //g/L (mean 1530 //g/L) one-half hour following endosulfan spraying on the adjacent field, despite a crop setback of 3 m from the ditch. This level of endosulfan exceeded the 96-h LC50 for rainbow trout, Daphnia sp., and stonefly by 1100, 6, and 670 times respectively (Wan, 1989). 74 In light of the restrictions on use of organochlorine pesticides such as endosulfan, water residues of the organophosphate and carbamate pesticides during the growing season have possibly increased since Wan's study, as these products are being used as replacements. As shown previously, the quantity of these types of compounds sold in the lower mainland, rose 14% between 1991 and 1995. In addition, the predominantly used insecticides of malathion, diazinon, chlorpyrifos, parathion, carbofuran, dimethoate, carbaryl, and methomyl have maximum water solubilities of 145 mg/L, 40 mg/L, 2 mg/L, 11 mg/L, 320 mg/L, 25 mg/L, 40 mg/L and 58 g/L, respectively. Organochlorine products such as endosulfan and DDT have much lower water solubilities. Endosulfan's solubility is 0.32 mg/L and DDT is practically insoluble in water (BCPC, 1991). Hence if solubility is a crude indication of dissolution, then organophosphate and carbamate insecticides may offer greater potential for field runoff contamination of the Nicomekl system than the organochlorines used in the past. In 1985, Wan found low levels of diazinon, azinphosmethyl, fensulfothion, endosulfan, and dinoseb in ditch sediments. Azinphosmethyl was found in sediments from two sites in July 1985 (Westham Island and near the Fraser River in Ladner) with a mean level of 2.7 fig/kg. Diazinon and fensulfothion were detected in sediments from two sites (near the Fraser River in Ladner and near the Sumas River) in July 1985, with respective mean concentrations of 4.0 and 10.3 ,ug/kg. Dinoseb was found in sediment from one ditch (near Sumas River) in July 1985 at 22.9 //g/kg. In 1987, in December and February, both th dinoseb and endosulfan residues were found in the sediments of the Nicomekl's 168 St. ditch, at average concentrations of 81.2 //g/kg and 652 //g/kg, respectively. 75 Environment Canada contracted Coastline Environmental Services Ltd. to test for dinoseb and endosulfan in water and sediments collected from six Nicomekl River locations and six of its drainage ditches. Coastline performed its sampling in March 1989. Dinoseb and endosulfan were only detected in two of the 12 water samples. The 168 t h St NE ditch water contained dinoseb at 5 //g/L, and the 176 th St. inner SE ditch sample exhibited trace quantities of endosulfan (< 0.1 //g/L) (Coastline, 1989). With respect to the sediments, dinoseb was measured in trace quantities (< 10 //g/kg) in most of the Nicomekl River sites th th th th measured (at 168 St., 176 St., and 184 St., and at the upstream site 10m below 64 Ave) (Coastline, 1989). The only significant sediment dinoseb concentrations were 49 //g/kg (NW ditch at 176 t h St.), 36 //g/kg (inner SE ditch at 176 t h St.), 37 //g/kg (Nicomekl River, 50m above 176 t h St.), and 29 //g/kg (NE ditch at 184 t h St.) (Coastline, 1989). Endosulfan was not detected in any of the river sediments, but was found at 428 //g/kg in sediments from the inner 176 t h St. ditch and at trace concentrations (< 10 //g/kg) in the sediments from the SE 176 th St., NW 176 th St., and inner SE 176 t h St. ditches (Coastline, 1989). Coastlines winter sampling regime may hve accounted for the low concentrations detected. Wan et al. (1995) reported concentrations of endosulfan for ditches on Westham Island, Ladner, South Burnaby, Cloverdale, Sumas Prairie, and the Sumas Canal for sampling performed between July and December, 1991. Of 21 samples of ditch water taken from the SE and SW ditches alongside 168 th St. over these 6 months, 4 contained endosulfan; however, concentrations were less than 0.02 //g/L. The highest endosulfan concentration was 13.4 //g/L, and was found in a sample collected from South Burnaby (Wan et al., 1995). In terms of effect levels, for endosulfan, rainbow trout's 24-h LC50 was 3.2 //g/L 76 (Macek, Hutchinson and Cope, 1969) and the amphipod Gammarus lacustris's 24-h LC50 was 9.2 //g/L (Sanders, 1969). Wan (1995) noted that in 1991 endosulfan accounted for only 1.9% of the insecticides sold in the Lower Fraser Valley and that organophosphate and carbamate insecticides represented 91.8% and 5% of total insecticide sales. This may have in part accounted for Wan's relatively low concentrations found in the Nicomekl's ditches. Wan (1994) also reported the results of chemical testing Tor OP insecticides diazinon, chlorfenvinphos, parathion, and dimethoate, at the same time (July to December, 1991) and same Lower Fraser Valley sites as stated above. Malathion and azinphos-methyl were not detected in any of the 84 water samples collected during Wan's 6 month study throughout the Fraser Valley. These two pesticides both break down rapidly. Wan (1994) found diazinon in 15 of the 21 samples collected from the 168 t h St. Nicomekl River ditch (0.01 - 0.22 //g/L, mean 0.1.1 //g/L), dimethoate in 8 of the 21 samples (0.02 - 1.27 //g/L, mean 0.26 //g/L), and parathion in 3 of the 21 samples (0.06 - 0.40 //g/L). Wan's study clearly shows that the acute suggested water quality criteria for diazinon (0.08 //g/L) and parathion (0.065 //g/L) (Table 9) were violated with respect to the drainage ditch alonside th 168 St. Wan did not determine whether this.ditch discharged. Nicomekl River water concentrations were not measured. The impact of this OP contamination on the Nicomekl River was undetermined. Coastline (1989) recommended toxicity tests be performed on ditchwater and riverwater samples in the region of their preliminary investigation. As a follow up study, EVS Consultants, under contract to Environment Canada, performed minor toxicity testing on 77 ditchwater and riverwater samples within same study site as used by this author. Unfortunately, the majority of EVS's toxicity testing was performed during the winter months, with an extremely limited number of test sites and sampling dates. Specifically, EVS performed water sampling, and chemical/toxicity testing on samples collected from the Old Logging Ditch, Burrows Ditch, Hall's Prairie Ditch, as well as the Nicomekl River in the vicinity of these three ditches. EVS performed 7-day C. dubia chronic tests on samples collected from Burrows Ditch, The th Old Logging Ditch, and the Nicomekl River upstream of 184 St. on samples collected twice in December 1989 and once in February 1990. For all three sample dates, survival and reproduction in the 100% ditch samples was not statistically-less than that in the 100% riverwater sample, or EVS's perrier-water control. Reproduction was actually slightly higher in the ditches than in the river water, and slightly higher in the river water than the Perrier-control water. The latter is of no surprise. Reproduction in EVS's Perrier-control water never met the test requirements of an average 15 neonates produced per brood (EC, 1992), and hence their Perrier-water control was likely unsuitable. It would have been very unlikely for EVS to find toxic water samples with their limited number of samples and winter collection time. EVS also performed invertebrate colonization studies using artificial substrates at 4 river locations (184 t h St. to downstream of the Old Logging Ditch). Their colonization studies were conducted from Dec. 14 t h to Jan 29 t h (1989), Feb. 23 r d to April 12 t h (1990), April 26 to July 4 (1990), and Nov. 2 to 27 (1990). Oligochaeta and chironomidae were the dominant river taxa and accounted for > 80% of the total number of invertebrates 78 throughout the year (EVS, 1993). The only other taxa to account for > 5% of the total number of invertebrates were Amphipoda in February/April and Plecoptera (stoneflies) in December/January. The other macroinvertebrates EVS (1993) found were estuarine and marine amphipods Paramoera nr. carlottensis, Allorchestes sp., and Ramellogammarus ramellus, and the isopods Gnorimosphaeroma oregonense (estuarine) and Asellus occidentalis (freshwater) (EVS, 1993). EVS (1993) reported that taxa belonging to the Orders Ephemeroptera (mayflies), Plecoptera (stoneflies), and Trichoptera (caddisflies) were generally in low abundance. Goodnight and Whitley (1961) concluded that in streams having the highest degree of organic enrichment, oligochaete tubificids constituted 97% of the macroinvertebrate bottom fauna, whereas at the cleanest stations, they represented only 13%. Their general conclusion was that whenever the population of these oligochaetes constituted more than 80% of the total population of macroinvertebrates, a high degree of either organic enrichment or industrial pollution was indicated. A percentage between 60% and 80% indicated doubtful conditions, and below 60% indicated good conditions (Goodnight & Whitley, 1961). Oligochaetes and chironomids are generally less sensitive to nutrient enrichment and its associated low dissolved oxygen levels than the more sensitive Ephemeroptera, Plecoptera, and Trichoptera (EVS, 1993) aquatic larval insects. Likewise, in the natural environment, tube/burrow organisms such as oligochaetes and chironomids may be isolated from recent sediment contaminants since these species have a head-down anus-up orientation and feeding is primarily on older deeper sediments (Lee, 1991). In addition, 79 sediment-bound organic pollutants (such as pesticides) may bind to the organisms polysaccharide tubes or may be degraded more rapidly as a result of the enhanced microbial activity in their tube/burrow walls (Lee, 1991). Metals may bind with acid volatile sulphides present in high concentrations in certain burrow walls (Aller and Yingst, 1983). Consequently, in the natural environment sediment-bound contaminants may be less bioavailable to tube dwelling organisms such.as chironomids and oligochaetes than to non tube dwelling organisms like amphipods and larval insects, accounting for the difference in sensitivity. th EVS's study found that moving down the mere 6 km reach of the Nicomekl from 184 St. to the Old Logging ditch, the ratio of the less tolerant Ephemeroptera and Plecoptera species relative to the abundance of the more tolerant Chironomidae species (EP/Chironomidae ratio) markedly decreased (EVS, 1993). This phenomenon was consistent throughout the year, but was most severe during the summer months. EP species were completely absent from the April to July colonization substrate located just downstream of the Old Logging Ditch (EVS, 1993). EVS found that ammonia, phosphorous, and total organic carbon concentrations were significantly higher in Halls Prairie Ditch, Burrows Ditch and the Old Logging Ditch than in the river throughout the year. EVS also found that the River concentration of nutrients increased from upstream to downstream (EVS, 1993) in 6 km distance covered by their study. Interpreting EVS's nutrient measurements, it appears that the upstream to downstream nutrient enrichment within the study site is most profound for Total Organic Carbon in the spring (April) and the fall (November to December). Ammonia enrichment occurred primarily in the summer 80 and fall (July to December). The TOC and ammonia enrichment of the river due to runoff from the drainage ditches support the potential of pesticide enrichment/contamination in this reach of the Nicomekl River. The ammonia enrichment is itself a separate issue of concern. Based on EVS's observations on relative species abundance, Goodnight and Whitley's (1961) conclusions relating the magnitude of pollution to the percentage of total invertebrates being chironomids and oligochates, the study area likely exhibited high organic enrichment and/or chemical contamination when EVS performed their 1989/1990 study. The downstream nutrient and TOC enrichment supports the declining EP/Chironomidae ratio. Since EVS (1993) reported that summer amphipod abundance showed no variation between upstream and downstream sites, this may suggest that the contamination was predominantly organics/nutrients hypoxia based. Amphipods such as Hyallela azteca are extremely sensitive to many metal and organic contaminants, such as pesticides (Arthur et al., 1983), yet they are capable of surviving extremely low dissolved oxygen concentrations (30-day LC50 < 0.3 mg/L 0 2 ) (EC, 1996). Nevertheless, Athur et al. (1983), Morgan (1976), Ward et al. (1995), and Eaton et al.'s (1985) demonstrated that chronic exposure to low doses, or acute exposure to higher doses, of OP insecticides, can also severely affect species diversity due to invertebrate drift. The possibility that pesticides caused or contributed to EVS's observations cannot be ruled out. Sediment toxicity testing was performed on 5 samples of sediment collected from the mouth of the Nicomekl River in June of 1993 (Swain and Walton, 1994). Amphipod 81 {Rhepoxynius abronius) survival in the 5 sediment samples was 88 - 97 % (mean 93 %). Microtox solid phase EC50's ranged from 0.1910 to 0.7655 % (mean 0.43 %). A Microtox EC50 of > 2% is non-toxic, 1-2% slightly toxic, 0.1-1% moderately toxic, and 0.0-0.1% extremely toxic (FREMP, 1996; from the B.C. Ministry of Environment Toxicity Laboratory) Metal concentrations in these sediments did not exceed threshold effect levels (TEL). FREMP pointed out that the observed moderate toxcity in the solid phase microtox test could have been caused by PAH's which were found in concentrations exceeding TEL. Numerous historical ammonia measurements on Nicomekl River samples are shown in Table 12. There is much less information on ammonia concentrations of the drainage ditches in the study area. The river ammonia measurements reviewed by this author were all well below the suggested acute (1-hour) average ammonia concentrations for the protection of aquatic life (Table 10a). The suggested chronic (4-day) average total ammonia concentrations (Table 10b) may have occasionally been exceeded. 82 Table 12. Historical total ammonia measurements (mg/L NH3-N) in the Nicomekl River. Location Dates # Samples Total NH3-N Range Total NH3-N Mean Nicomekl at 99A Dam (Downstream) 1976- 1979 20 1 <0.01 -0.13 < 0.01 Nicomekl at 99A Dam (Downstream) 1988- 1992 28 2 Primarily Fall 0.012 - 0.446 0.153 Nicomekl at 168th St. 1976-1980 16 1 0.027 - 0.352 0.141 Nicomekl at 64th Ave. (Upstream) 1974- 1983 40 1 < 0.005 - 0.525 0.092 Nicomekl at 64th Ave. (Upstream) 1988- 1992 28 2 Primarily Fall 0.005-0.177 0.049 Nicomekl at 184th St. 12/1989 to 11/1990 6 3 0.03-0.13 Burrows Ditch 12/1989 to 11/1990 6 3 0.08 - 0.27 Nicomekl (below Burrows Ditch) 12/1989 to 11/1990 6 3 0.02-0.12 Old Logging Ditch 12/1989 to 11/1990 6 3 0.02 - 0.3 Nicomekl (below Old Logging Ditch) 12/1989 to 11/1990 6 3 0.01 - 0.25 (Swain and Holms, 1988b) 2 (BCME, 1997) 3 (EVS, 1993) 83 5.0 SAMPLING PROGRAM AND EXPERIMENTAL METHODOLOGY 5.1 Ditch and River Water Sampling Locations and Frequency Several criteria were used in choosing sampling locations. The ditches chosen for sampling were large in size and remained wet throughout the summer, are possible fish habitats/refuge areas, discharge large volumes of water into the Nicomekl, are accessible without trespassing, and/or have revealed the presence of pesticides in previous chemical testing. The river sampling sites chosen were located downstream or upstream/downstream of the accessible ditch discharge points. In addition, river sampling was also performed upstream of the study area in the Nicomekl's headwaters. The locations of sampling sites 1-12 are shown in Figure 10. The upstream river control water collection point, Site 13, and White Rock pond culture/control water collection locale were previously labeled on Figure 1. A brief description of the specific location of each of these sites is provided in Table 13. Site 1 served as a river location for mixed ditch and riverwater leaving the study site. The named ditches were chosen due to their size and high municipally controlled and recorded discharge flows. Sites 10, 11, and 12 were not located at the convergence of the ditches with the Nicomekl, due to lack of access to these locations. Sites 3, and 8, the smaller un-named ditches were chosen for sampling since previous studies had detected pesticides at these locations (Section 4.0). Site 13 served as a river upstream control site to which toxicity and chemistry results from the study site ditch and riverwater samples were compared. Site 13, should theoretically be the most uncontaminated region of the Nicomekl River. 84 85 Table 13. Specific Locations of Water Sampling Sites. Site Description of Location 1 th Nicomekl River, 50 m upstream of the 40 Ave. bridge 2 th Nicomekl River, north side of river beneath 168 St. bridge 3 th Un-named drainage ditch, NE 168 St. 4 Nicomekl River, 5 m downstream of South Cloverdale Ditch discharge point 5 South Cloverdale Ditch, a t 4 0 t h Ave. 6 Nicomekl River, 40 m upstream of South Cloverdale Ditch discharge point 7 Nicomekl River, south side of river 50 m upstream of the 176th St. bridge, downstream of Site 8 ditch discharge point. 8 th Un-named drainage ditch, SE 176 St; inner ditch (furthest from road) 9 Nicomekl River, 50 m upstream of Site 8 ditch discharge point 10 Ericson Ditch, at 40 Ave. And 180 St. 11 Burrows Ditch, at 40 t h Ave. 12 Old Logging Ditch, at 40 t h Ave. 13 Nicomekl River, at 64 t h Ave. east o f 2 1 6 t h St. (upstream control site) W.R.P. th White Rock pond, Southmere Park, 16 Ave. and Oxford St. (culture/control site) Water sampling commenced in the late.Spring and concluded in the late Fall of the 1997 growing season. Sampling was performed at three week intervals. The specific dates of water sampling, along with the sites sampled on these dates are shown in Table 14. When possible, water samples were collected within a few days of rainfall (Figure 1). 86 Table 14. Dates and Respective Sites of Water Sampling. Dates of Water Sampling (1997) Sites Sampled (Figure 10) May 6 Pumped Discharge from Site 8 June 6 1,2, 3, 4, 5, 7, 8, 10, 11, 12, 13 June 26 1,2, 3, 4, 5, 7, 8, 10, 11, 12, 13 July 16 1,2, 3, 4, 5, 7, 8, 10, 11, 12, 13 August 7 1,2, 3, 4, 5, 7, 8, 10, 11, 12, 13 August 29 1,2, 3, 4, 5, 7, 8, 10, 11, 12, 13 September 17 1,2,3,4,5,6,7,8,10,11,12,13 October 3 1,2, 3, 4, 5, 6, 7, 8, 10, 11, 12, 13 October 16 5, 10-Nearby Field Puddle November 21 1,2, 3, 4, 5, 6, 7, 8, 10, 11, 12, 13 If no discharge was occurring from the Site 5 South Cloverdale ditch at the times of sampling, then upstream sampling at Site 6 was not undertaken, and only one river sample was taken downstream of the South Cloverdale Ditch, at Site 4. Significant discharge was only observed from the Site 8, 176 t h St. ditch at one time, on May 6. A sample of this discharge was collected; however the sample was collected in a plastic bottle as opposed to glass, and upstream and downstream sampling was not performed around the discharge point, since the May 6 site visit was a preliminary site investigation and full sample . collection had not been intended on this date. No significant discharges were observed from this site during subsequent sampling events. Slight discharge through the Site 8 176 th St. ditch's flap gates was observed on the August 7 t h and September 17 t h sampling dates. However, since discharge volume at these times was extremely low, river sampling both upstream and downstream of the discharge was not felt to be warranted, and only the 87 downstream location, Site 7, was sampled. Consequently, sampling at the upstream location, Site 9 was never performed throughout the study. 5.2 Ditch and River Water Sample Collection, Transport, and Storage Water samples were collected several feet from the bank of the river or a ditch using a 1 L polymethylpentene plastic container affixed to a 3 m steel pole. Replicate samples (15-20) at each site were collected to homogenate possible temporal and spatial variances in contaminant concentration at each location. Each small sample collected with the dip-pole was poured through 60 im\ silk plankton netting, to remove detritus and indigenous organisms, and collected in a 25 L high density polyethylene bucket. This step was repeated numerous times until 10-12 L of sample had been collected in the polyethylene bucket. The pooled water sample was then immediately poured from the polyethylene bucket through a plastic funnel into two separate solvent rinsed, acid washed 4 L amber glass bottles. The amber glass bottles were filled completely to eliminate air space and sealed with teflon-lined screw caps. In addition to the 8 L of sample collected for toxicity testing from each site, smaller volumes of sample were collected in nalgene and polyethylene sample bottles, respectively, for ammonia and metal analyses to be performed in the event the sample exhibited toxicity. Immediately following collection, the two 4 L glass bottles were placed in coolers with frozen gel packs to prevent the water samples from warming above their temperature at the time of collection. Once all water samples were collected from the study site, they were immediately transported to the laboratory for storage at 4 °C in the dark. Upon arrival at the laboratory, samples collected for metal and ammonia analysis were 88 preserved by a reduction to pH <2 and 2-3, respectively. Metal samples were preserved using 500 //I of concentrated nitric acid, whereas ammonia samples were preserved using 2 -3 drops of concentrated sulphuric acid. Preserved metal and ammonia samples were also stored at 4 °C in the dark until testing. 5.3 Water Toxicity Testing with Ceriodaphnia dubia The Ceriodaphnia dubia chronic and acute tests performed generally followed the test guidelines of EC (1992) and U.S. EPA (1993a), respectively. 5.3.1 C. dubia Culture The initial C. dubia culture was provided by a local commercial toxicity laboratory, B.C. Research Inc. The culture/control water used was the same water used by BCRI, hardened water (hardness raised to 80 mg/L as CaC0 3) from a small pond located in Southmere Village Park, White Rock (Figure 1), at the north side of 16 t h Ave. just east of Oxford St.. The 22 L of collected culture/control water was filtered using 60 //m silk plankton netting and stored in a large nalgene container at 4 °C for up to 3 weeks of use. th The White Rock pond is bordered on three sides by condominiums and one side by 16 Ave. However, there is a minimum of 20 m grass parkland bordering the pond on all sides, and the only runoff the pond appears to receive is that from approximately 4 acres of perimeter park grassland. The edges of the pond have extensive stands of cat-tails, which should help in adsorbing any organic/metal contaminants the pond may receive. Fish, frogs, and turtles were personally observed in the pond, suggesting it is a reasonably healthy ecosystem. BCRI has good success using this water as culture/control water in 89 the C. dubia tests they have performed for several years. The pond is susceptible to occasional algae and fungus blooms in the summer and fall months, which can influence C. dubia culture health. Individual C. cubia brood culture organisms were reared in separate plastic Phoenix Biomedical medicine cups. Throughout the study period, thirty to sixty brood organisms were continuously maintained until two weeks in age, at which time their neonates were no longer eligible to be used for toxicity testing (EC, 1992), andthe brood organisms were replaced with a subsequent set. Brood organisms were fed 100 IA. YCT (yeast, Cerophyll, fermented trout chow) and Selenastrum provided by BCRI, either daily or every second day. Brood temperature and lighting conditions were identical to test conditions, and followed EC (1992). 5.3.2 Initial Full Strength (100%) Ditch and River Water Sample Chronic Testing for Inhibition of Survival and Reproduction. Ditchwater and riverwater toxicity testing was initiated usually within 24 hours of sample collection, and never exceeding 96 hours. Initially, water samples were tested at full-strength (100%). One of the two 4 L samples from each site was selected as a sample source for the entire full-strength toxicity test. From this 4 L vessel, 250 mL subsamples were poured into soap/acid-washed glass beakers initially to begin the test and subsequently daily thereafter, when performing the necessary 24-h water changeovers. Before taking a 250 mL subsample, the 4 L bottle was agitated to resuspend any of the solids which had settled during storage, and homogenize the solution. The 250 mL subsamples from the various sites were warmed to 25 °C in a hot-water bath. Once at 25 90 °C, the subsamples' initial dissolved oxygen concentrations, pH, and conductivity were measured and recorded. Most subsamples had dissolved oxygen concentrations typically near saturation after warming. However, as testing progressed, the water-level in each 4 L sample bottle decreased, increasing the bottles air space. This air space, along with agitation of the 4 L sample bottles prior to subsampling, caused the samples to become saturated with D.O. at the 4 °C storage temperature, yielding supersaturation of the subsamples once they were warmed to 25 °C (D.O. saturation at 25°C is 8.4 mg/L). Supersaturated subsamples often exhibited D.O. concentrations of > 1.0.0 mg/L. There is a trade-off between subjecting C. dubia to unaerated D.O. supersaturated subsamples and detoxification of a sample by volatilization and oxidation of toxic organics or ammonia during pre-aeration to rid the D.O. supersaturation. Consequently, aeration to rid supersaturation was kept to a minimal rate and duration. Supersaturated subsamples were pre-aerated with compressed air using a glass pipette, typically for 10 to 20 minutes at 500-1000 bubbles/min, until DO was within 1.0 mg/L of saturation (under 9.4 mg/L). There is the possibility that this aeration could have reduced some concentrations of some contaminants. However, most subsamples did not require pre-aeration until the start of th th th the 4 ,5 , or 6 day of the test, and organisms should have received exposure to any volatile toxic compounds present in the sample, such as ammonia, in the first 72 hours of testing, had these compounds been present. Each 100% subsample (at 25 °C and pre-aerated, if necessary) was poured into 10 separate 30 mL Phoenix Biomedical plastic medicine cups, sitting in cradles of holes cut in styrofoam boards. The styrofoam board medicine cup holders provided an excellent means of holding, and moving tests, and also insulated tests from any slight temperature 91 variations in the 27 °C climate controlled test room. The test room was kept at 27 °C in order to maintain solution temperatures of 25 ± 1 °C (EC, 1992), which tended to be lower than room temperature likely due to the room's air conditioning fan and slight sample evaporation despite covering. Each medicine cup containing full strength sample was then fed 100 /A. of a yeast, Cerophyll, fermented trout-chow (YCT) solution and 100 (A. of a concentrated Selenastrum solution (EC, 1992). C. dubia neonates were collected from the 30 - 60 brood organisms the morning the test was to commence and pooled in one medicine cup. Neonates were less than 24 hours old (EC, 1992). Test neonates were transferred by glass medicine dropper into each medicine cup containing sample. The complete set of full strength tests for all sample sites were placed beneath a cool-white light source with a photoperiod of 16 hours of light (at 400 - 900 Lux) and 8 hours of darkness (EC, 1992). Tests were covered with plexiglass to prevent evaporation of solutions. Daily, all test neonates were transferred by glass medicine dropper to fresh test solutions (which had been warmed, aerated as necessary, and fed with 100 jA. YCT and Selenastrum) (EC, 1992). The DO, pH, and.temperature of both the new test solution as well as the test solution being replaced were all measured and recorded at this time (EC, 1992). The survival and neonate production of each of the 10 test organisms per full strength sample was recorded during and after each 24-hour transfer, respectively. As required by EC (1992) tests were concluded once 60% of the ten test organisms in the White Rock pond culture/control water had produced 3 broods of neonates. Test duration was 7 ± 1 days (EC, 1992). 92 5.3.3 Dilution Series Chronic Testing of Lethally Toxic Samples If significant mortality was observed for a water sample at full strength (100%), the second stored 4 L bottle of sample was used to perform a toxicity test using the dilution series of 100%, 56%, 32%, 18%, and 10% sample in White Rock pond culture/control water. The second 4L bottle was used as a sample source since this bottle was stored with no air space since collection. Dilution-series testing commenced as soon as lethality was observed in the original full strength test. 5.3.4 Lethally Toxic Samples Biological Toxicity Identification Evaluation. Two approaches were used to attempt to determine the cause of toxicity in the lethally and sublethally toxic samples. If a sample demonstrated acute toxicity, along with retesting the sample using the above dilution series, a 1.5 L portion of the lethally toxic sample was poured from the second 4 L storage container into a large glass Erlenmeyer flask. Then silica-bound C18 chromatography gel was added to this aliquot at 1 g/L and the mixture was rapidly stirred for 20 minutes. The sample was then separated from the C18 gel by filtration through a Whatman 934H 1.5//m borosilicate micro-fibre filter, and returned to the 4 °C storage room. The 1.5 L of C18 treated sample was then used to conduct a separate 7-day C. dubia survival and reproduction inhibition test, run concurrently with the toxic sample's dilution series test. The hypothesis behind this testing procedure was that the Q 1 8 gel would adsorb organic contaminants, and remove the toxicity if it was predominantly due to non-polar organic contaminants. Removal of OP insecticide toxicity by C 1 8 SPE is a component of the U.S. EPA's phase I TIE procedure (Mount and Anderson-Carnahan, 1988), and was observed for diazinon toxicity removal by Amato (1992). 93 Lethally toxic samples were also retested solely for 96-h survival at 100% concentration with and without the addition of a proven metabolically activated organophosphate pesticide detoxifying agent, piperonyl butoxide (PBO), based on the methods of Bailey et al. (1996) and Ankley et al. (1991). PBO was added to the samples at 200 ppb in a methanol carrier not exceeding 1.5% concentration in the bioassay. Bailey et al. (1996) found that 200 ppb of piperonyl butoxide in methanol carrier added to solutions of 1.5 //g/L diazinon and 0.75 //g/L chlorpyrifos completely eliminated the mortality of these two OP insecticides in 48-h testing. They also found no mortality in control tests which contained methanol (up to 1.5% concentration) and 200 ppb PBO (in the absence of diazinon and chlorpyrifos). Elimination of toxicity using PBO is strong evidence that toxicity is due to . metabolically activated OP insecticides (Ankley et. al., 1991), those requiring metabolic activation to an oxon derivative capable of efficiently inhibiting acetycholinesterase (Matsumura, 1975). Reported metabolically activated OPs include malathion, diazinon, chlorpyrifos (Bailey et a l , 1997), methyl parathion, dimethoate, and azinphos methyl (Matsumura, 1975), in decreasing order of the amount of each likely used for commercial agriculture in the Lower Mainland in 1991. Unfortunately, this studies' piperonyl butoxide tests were not conducted immediately following the original bioassays which showed lethality. Early in the study it was expected that the solid phase extraction full-ion-scan gas chromatography/mass spectroscopy (GC/MS) analyses performed would adequately identify organic toxicants, and the PBO analyses would not be necessary. Failed detection of any suspect toxicants by the SPE GC/MS analysis prompted the PBO analyses. Consequently, this meant that the PBO analyses were performed on portions of the lethally toxic samples which had been stored. 94 Initial PBO testing used subsamples which had been frozen in nalgene bottles at -10 °C for 2 months, and then thawed overnight at 4 °C. One of the two thawed samples failed to produce significant mortality, making its PBO/non-PBO comparison meaningless. PBO testing was repeated using portions of the lethally toxic samples subjected to a different means of storage, 4 °C in glass for 5 months. The PBO tests were 96-h tests and used the test methods set out by US EPA (1993a) which called for feeding of C. dubia prior to the start of the test, and 2 hours prior to the only solution changeover at 48 hours. PBO testing utilized 4 replicates of each test solution with 5 C. dubia neonates per cup, for a total of 20 neonates per test solution. In the second round of PBO testing, using the samples which had been stored 5 months in glass, the lethality testing was continued to 7-days duration to allow time for toxicity to manifest and better show the effect of PBO. Lack of changeover or feeding beyond that performed at 48 hours did not appear to affect the survival in the food-rich ditchwater samples treated with PBO as it did the controls. 5.3.5 Toxicity Tests on Diazinon, Chlorpyrifos, and Prometryn The C. dubia 7-day chronic test was performed using diazinon at the concentrations of 0.8, 0.4, 0.2, 0.1, 0.05, 0.025 //g/L. Diazinon was purchased as a 100 //g/mL (methanol) chromatography standard from Accustandard Inc. 80 //L of this solution were pipetted into 10 mL of methanol, which was used as the test's 800 //g/L diazinon stock solution. The stock solution was refrigerated in the dark as recommended by Accustandard and used for the duration of the 7-day test. For each daily water changeover in the 7-day test, 0.5 mL of the stock solution was transferred by volumetric pipette into a 500 mL volumetric 95 flask of upstream river control water. This was used as the test's highest diazinon concentration (0.8 //g/L); consequently, the highest methanol concentration was 0.1%. The lower concentrations were prepared by performing successive 50% dilutions of the 0.8 //g/L solution. The testing was performed in the same plastic medicine cups used for testing of the field samples. The 7-day chronic test was performed using chlorpyrifos at the concentrations of 0.132, 0.066, 0.033, 0.016, 0.008, 0.004 //g/L. .Chlorpyrifos was purchased as a 100 //g/mL (methanol) chromatography standard from Accustandard Inc. 132 //L of this solution were pipetted into 100 mL of methanol which was used as the test's 132 //g/L chlorpyrifos stock solution. The stock solution was stored in the dark at ambient temperatures as recommended by Accustandard for the duration of the 7-day test. For each daily water changeover in the 7-day test, 0.5 mL of the stock solution were transferred by volumetric pipette into a 500 mL volumetric flask of upstream control water. This was used as the test's highest chlorpyrifos concentration (0.132 //g/L); consequently, the highest methanol concentration was 0.1%. The lower concentrations were prepared by performing successive 50% dilutions of the 0.132 //g/L solution. The testing was again performed in plastic medicine cups. Unfortunately, the 7-day test on chlorpyrifos failed to produce any lethality at the highest concentration (0.132 //g/L). The lack of lethality was attributed to adsorption of these low concentrations to the plastic cups and/or the food added to the test solutions. Time did not allow for re-testing for 7-days at higher concentrations. Therefore a 48-hour test was performed using the concentrations of 1.32, 0.66, and 0.33 //g/L chlorpyrifos in upstream 96 river water with a maximum methanol concentration of 1 %. This test was again performed in plastic medicine cups with 10 cups/organisms per concentration. Test solutions were initially fed with 100 [A. of YCT and Selenastrum, as the purpose of this test was to find out what concentration of chlorpyrifos could be toxic using the plastic cups and feeding regime of the 7-day test. A 96-hour lethality test was performed on prometryn using the concentrations of 20, 10, 5, 2.5 and 1.25 mg/L. Prometryn was purchased as 99% pure technical grade powder from Supelco Inc. The highest concentration tested was prepared by weighing out 0.0101 g of prometryn, and adding this to 500 mL of upstream control water in a volumetric Erlenmeyer flask. The lower concentrations were prepared by performing successive 50% dilutions of the 20 mg/L solution. This test followed EPA (1993a) with a water changeover at 48 hours and feeding of neonates prior to test initiation and the 48-h water changeover. 5.3.6 Reference Toxicant Testing and Culture Health Reference toxicant tests were performed using sodium chloride at concentrations of 0, 320, 560,1000,1800, and 3200>g/L in White Rock pond culture/control water using the chronic survival and reproduction inhibition test (EC, 1992) described above. The first two reference toxicants were performed prior to testing the initial June 6 t h set of samples; the remaining 6 reference toxicants were conducted once a month throughout the study period. An IC50 (50% reproduction inhibition) ± 2 S.D. endpoint was determined for each reference toxicant, and the mean IC50 ±2 S.D. was determined collectively for all the reference toxicants. IC5o's were calculated despite the EC (1992) recommendation for IC25s, in order that this authors reference toxicant results could be compared with that of 97 the C. dubia supplier BCRI (which reports its reference toxicants as IC50s). Individual IC50s were compared to the collective mean ± 2 SD. as required by EC (1992). IC50s were determined using the ICPIN 20 computer program (U.S. EPA, 1993b). In addition to the reference toxicant tests, culture health was continuously evaluated by tracking the neonate production of 4 brood organisms from each maximum 2 week old brood set. 5.3.7 Test Endpoints and Statistical Analyses Survival and reproduction in each full strength sample were separately statistically compared with that in the control solutions (both the upstream, Site 13 control as well as the White Rock pond culture control) using student t-tests. T-testing was performed using Corel's Quatro Pro 7.0 t-test data function, assuming normality of data for each sample and heterogeneous variances in data between samples. Both one-tailed and two-tailed t-testing was performed, using a=0.05. Qualitative statements were made as to whether samples exhibited higher or lower reproduction than controls, based on the one-tailed results. With respect to the biological toxicity identifications performed: survival and reproduction in Q 1 8 treated samples was compared to that in the controls using student t-tests (a = 0.05); survival in the untreated toxic samples was compared with survival in the controls and treated samples using student t-tests (a = 0.05). If a sample's data is not normally distributed, simple t-testing without normalization of data should provide reasonably accurate results. Zar (1984 ) states: "The theoretical basis of t-testing-assumes that the sample data came from a normal population, assuring that the mean at hand came from a normal distribution of means. Fortunately, the t-test is robust, meaning that its validity is not seriously affected by moderate deviations from the underlying assumption." 98 Recent, 1997, updates to EC (1992) also state: "Ideally, the data should conform to normal distribution and homogeneity of variance, but the t-test is robust in the face of non-conforming data." For dilution series testing of lethally toxic samples and the pesticides identified in the study site, LC50s (median lethal concentrations) were determined using BCRI's in-house LC50 program followed the methods of Stephan (1977). The endpoints of NOEC (no-observable-effect concentration) and LOEC (lowest-observable-effect concentration) were determined for reproduction using the Toxstat 3.2 computer program (Gulley et al., 1989) and the guidelines of the U.S. EPA (1994a), which requires exclusion of the test concentrations for which there was complete lethality from the computer analysis. An IC25 (25% inhibition concentration) for reproduction was also determined for these dilution series tests. These calculations were performed using the ICPIN 2.0 computer program (U.S. EPA, 1993b), and included using the neonate production of test organisms which died during the test EC (1992). Thus, the IC25 value may not be purely a sublethal reproductive effect, but a combination of reduced reproduction and mortality, or simply mortality. Therefore, the calculated IC25s were reported alongside the LC50s to indicate the influence of lethality on the IC25 endpoint. 5.4 Chemical Ana lyses of Ditch and River Water Samples Analyses were performed for organics, metals, and ammonia on toxic samples in an attempt to aid in determining the cause(s) of toxicity. Total dissolved carbon (TDC), total dissolved inorganic carbon (TDIC), total dissolved organic carbon (TDOC), and colour, conductivity, and hardness were measured on every water sample collected, in order to characterize the water chemistry of the region and possibly help explain causes/variations in toxicity. 99 5.4.1 Organics Various methods were used for the organic analyses, both with respect to the extraction/concentration of the samples' contaminants into solvent and the setup of the gas chromatograph/mass sprectrophotometer (GC/MS) in an attempt to chemically determine if pesticides were present in the toxic samples. A complete discussion of all the techniques used, including a detailed justification for their use are provided in the results and discussion section of this thesis. The following simply outlines the details of the materials and methods used for the solid phase extractions (SPE's), liquid-liquid extractions, and gas chromatography/mass spectroscopy (GC/MS) analyses performed. Solid phase extractions were the main technique used to remove and concentrate the organics of both lethally toxic and sublethally toxic samples. SPE's were used since they allowed for stable storage of the contaminants on the SPE tubes until a later date when elution, concentration, and chemical analysis of a multiple number of collected toxic samples could be performed. Lacorte et al. (1995) demonstrated that C18 SPE tubules used to filter groundwater samples spiked at 10 //g/L with 19 different OP pesticides showed complete recovery for 16 of 19 OP:pesticides tested when SPE columns were stored at -20 °C for 8 months, and complete recovery of-the same 16 OP pesticides when SPE columns were stored at only 4 °C for up to 3 months. Once a sample was determined to be lethally or sublethally toxic and selected for organics analysis, a subsample from the second 4 L glass sample bottle was collected. This bottle had no, or minimal air space, and was the best attainable duplicate of the sample used in the original toxicity test. Subsamples used for the SPE's were 250 mL in volume, since 100 it is recommended that the volume filtered not exceed 250ml_ for the Chromosep 1000 mg Q 1 8 columns used (Supelco, 1996). Subsamples were warmed to 25 °C before extraction, to duplicate the conditions of the solutions used in the bioassays. The SPE tubes used were Chromosep brand 1000 mg Q 1 8 (Octadecyl) 6.0 mL tubes. The SPE tubes were sequentially pre-conditioned by flushing them with 2 mL of acetonitrile, 2 mL of methanol, and 2 mL of de-ionized water prior to sample filtration. These solvents were recommended by Lacorte et al. (1995). Immediately following pre-conditioning, 250 mL of subsample was vacuum filtered through the SPE tube at a filtration rate of 5-10 mL/min. (25-50 min. filtration time) using a Baker-10 SPE System. The SPE tubes were left under suction after filtration for 20 minutes in order to dry, then wrapped in tin-foil, and stored in the freezer at -10 °C. The same procedure was performed on the non-toxic White Rock pond dilution/control water and non toxic ditchwater and riverwater samples, for comparison of results. Within two months of storage, the SPE tubes were removed from the freezer and thawed for 1.5 hours. Approximately 2 mL of acetonitrile was slowly vacuum filtered through each SPE tube into one of the Baker System's glass cuvettes, to elute any organics each SPE tube had retained. The acetonitrile elutions contained traces of water which had to be removed prior to chemical analysis. The elutions were transferred to glass test tubes and approximately 1 g of anhydrous sodium sulphate was added to each acetonitrile elution to absorb the water. Water absorption was allowed to take place overnight at 4 °C. The acetonitrile elutions were poured off of the hydrated sodium sulphate into volumetric glass test tubes. A small amount of the acetonitrile elutions was unfortunately trapped in pockets of the solidified sodium sulphate. 101 The recovered acetonitrile elutions were evaporated to 0.5 mL in volume by gently blowing nitrogen gas across the eluates surface. The acetonitrile SPE elutions, now nominally 500X the concentration of the original samples, were then transferred to Pyrex GC vials, sealed with teflon lined caps, and returned to -10 °C storage for 2-3 days. The eluates were analyzed by gas chromatography/mass spectroscopy (GC/MS), in which the mass spectrophotometer used a full-ion-scan (monitored for ions of all mass to charge ratios). Using the full-ion-scan allowed the complete mass to charge spectra of chromatogram peaks to be matched with those in the computer's database, to identify compounds. The SPE GC/MS techniques employed could not detect pesticides even in samples which were lethally toxic due to OP insecticide contamination as evidence by the initial PBO analysis (discussed in results). Liquid-liquid extraction with methylene chloride was performed on lethally toxic samples as well as the culture/control water in hopes of insecticide detection using this different technique. The toxic samples used for the liquid-liquid extraction were the same samples used in the first round of PBO testing, which had been frozen in nalgene bottles at -10 °C for 2 months. The frozen samples were thawed at 4 °C overnight, and 360 mL subsamples warmed to 25 °C. Each subsample was vigorously shaken for one minute with 30 mL of methylene chloride in a pyrex separatory funnel, and then the phases were allowed to separate for 10 minutes. This extraction procedure was repeated 3 times. The 90 mL of methylene chloride extract was poured through anhydrous sodium sulphate (1-2 g) in a glass funnel lined with a paper filter, to remove water. The sodium sulphate and filter paper apparatus was rinsed with additional methylene chloride to extract any traces of organics adsorbed to the sodium sulphate and/or filter paper. The methylene chloride was evaporated to approximately 2 mL in 102 volume using a Rotovap evaporator. The 2 mL concentrates were further reduced by nitrogen gas to 1.0 mL. The solvent extractions, now 360X the concentration of the original samples, were transferred to pyrex GC vials, sealed with teflon-lined caps, and analyzed by GC/MS, using the full-ion-scan, and a select-ion-scan looking for diazinon and chlorpyrifos at a later date. All the gas chromatography mass spectroscopy (GC/MS) analyses were performed using the Civil Engineering Department's Hewlett Packard 6890 Series GC System and 5973 Series Mass Selective Detector. A Hewlett Packard 5MS (cross-linked 5% phenyl methyl siloxane) column (30 m by 0.25 mm with 0.25 /xm film thickness) was used with helium carrier gas at a constant flow of 0.9 mL/min. The injection volume was 1.0 /A.. The temperature program used was: 40 °C for 2 minutes, increased by 5 °C per minute to a final temperature of 300 °C which was held for 8 minutes, for a total run time of 62 minutes per sample. 5.4.2 Metals Trace metal concentrations were determined for the two lethally toxic ditchwater samples. Metals were also measured for river (Site 13) and White Rock pond control samples, and ditch water from the sites where toxicity was observed, collected on a date which produced no toxicity. All samples were preserved to a pH <2 using nitric acid, and stored at 4 °C in the dark for < 4 months prior to metal analyses. The metal analyses were contracted to Chemex Laboratories, who used an inductively coupled plasma/mass spectroscopy analysis (ICP/MS). Total metals were measured since dissolved metals may have adhered to suspended solids in the samples during storage. In addition, particulate bound metals 103 in the original ditchwater bioassays could have contributed to the observed toxicity, since C. dubia are filter feeders. Chemex was instructed to thoroughly rinse the sample bottles with the same strong acid used for the samples' digestion to remove any metals which may have adhered to the plastic sample bottles. 5.4.3 Total Ammonia Total ammonia was measured in the samples from all sites for the dates which produced lethal toxicity and select dates which produced sublethal toxicity. The sulfuric acid-preserved water samples were filtered through disposable blood serum isolating filters (Iso-Filter 1071) to remove particulate material which might have interfered with the automated analysis. Total ammonia was measured using the phenate method, which measures the abundance of the blue compound indophenol formed by the reaction of ammonia, hypochlorite, and phenol byway of light absorbance at 600-630 nm as outlined in Standard Methods 4500-NH3 G (AWWA, 1995). Testing was conducted using the Civil Engineering Department's Lachat Quikchem AE 2300-000 Instrument. 5.4.4 Dissolved Total/lnorganic/Organic Carbon Carbon analyses were performed on all water samples. Samples were first filtered through 0.45 ijm cellulose membrane filters. The filtrate was analyzed for total dissolved carbon (TDC) and total dissolved inorganic carbon (TDIC). Total dissolved organic carbon (TDOC) was calculated by difference. A Shimadzu TOC-500 Total Organic Carbon Analyzer was used which measured TDC and TDIC as C 0 2 using an infrared detector, following combustion of the samples at 680°C and 150°C, respectively. The instrument was calibrated with total carbon and total inorganic carbon standards of 5 and 50 mg/L. 104 5.4.5 Dissolved oxygen, pH, Conductivity, Hardness, and Colour Dissolved oxygen concentrations and pH were measured daily in the water used for the toxicity tests solution renewal, as well as the bioassays' discarded (24-h old) solutions. Conductivity, total hardness, and colour were measured once for each sample. Dissolved oxygen was measured to 0.1 mg/L using a YSI Model 54A meter, pH to two decimal places using a Beckman 44 digital pH meter, conductivity using an analog Radiometer Copenhagen CDM3 conductivity meter, colour using a Hellige Aquatester, and total hardness using the EDTA titrimetric method outlined in Standard Methods 2340 C (AWWA, 1995). 5.5 Sediment Sampling Locations Surface sediment samples were collected on October 16, 1997 from most of the ditch and river locations where water samples were collected. Specifically, sediments were collected from water sampling Sites 1, 5, 7, 8, 11, and 13. Sediments were also collected approximately 200 m downstream of Site 2 (in the vicinity of the boom-supported irrigation system intake pipe which partially spans the river), 10 m east of Site 10 (Ericson Ditch) th from the ditch parallel to the north side of 40 Ave., and from Site 12 (Old Logging Ditch) th 20 m south of 40 Ave. See Figure 11. A Fall collection date was selected to determine if pesticides had accumulated in the sediments from the 1997 growing season. 105 106 5.6 Sediment Sample Collection, Transport, and Storage Sediment collection, transport, and storage techniques were based on the recommendations of EC (1994). Sediments were collected using the same polymethylpentene 1 -L container affixed to an iron pole that was used for the water sample collections. Sediments were scooped from the ditch and river bottom or banks (whichever could be obtained) and slowly raised to the surface to avoid loss of the fine material. The overlying ditch/river water atop the sediments was gently poured off, and sediments were transferred to polyethylene freezer bags, vacated of air space and sealed, placed in coolers with frozen gel packs, and transported back to the laboratory for storage in the dark at 4 °C. After sediment toxicity testing, sediments were frozen until percentage organic matter and total dry weight metal concentrations were determined. 5.7 Hyallela azteca Chronic Sediment Toxicity Testing Hyallela azteca was selected as the sediment test organism. Correlations between chemical testing and sediment toxicity tests using amphipods have shown that these tests can provide reliable evidence of biologically-adverse contamination of sediment in the field (Swartzetal., 1982,1985,1986,1994; Becker etal., 1990; Canfield etal., 1994; US EPA, 1994b). Hyallela azteca has been shown to be euryhaline, and can be successfully tested with estuarine sediments. This species has an extremely wide tolerance to sediment grain size. Ingersoll and Nelson (1990) found that in long term exposures to sediments ranging from > 90% silt- and clay-size particles to 100% sand- size particles, no detrimental effects on either survival or growth were observed. In uncontaminated sediments, Ankley et al. (1994b) found no correlation between amphipod survival rates and sediment particle size, organic carbon content, or mineralogical composition, providing that the test animals were 107 fed. Suedal and Rodgers (1994) found that H. azteca was tolerant to all possible sediment particle size distributions (0 to 100% sand, 0 to 100% silt, and 0 to 60% clay) and ranges of organic carbon content they examined (0.1 to 8.0%). H. azteca can survive low dissolved oxygen conditions, and has been shown to have a 48-h LC50 of 0.7 mg/L 0 2 (de March, 1981) and a 30-day LC50 of < 0.3 mg/L 0 2 (Nebeker et al., 1992). West et al. (1993) found that H. azteca was more sensitive than C. tentans and L. variegatus in 10-day whole sediment tests with field collected sediments. In a study of contaminated sediments from the Great Lakes, H. azteca was the most sensitive of 24 organisms tested (Burton and Ingersoll, 1994). 5.7.1 Test Method Growth and survival in sediments collected from 3 river and 5 drainage ditch locations were compared to sediment collected from the upstream control location (Site 13). The sediment tests followed Environment Canada's 14-day chronic Test For Growth and Survival in Sediment Using the Freshwater Amphipod Hyallela azteca (EC, 1996), with the exception that the recommended 5 replicates was reduced to 4 replicates due to test equipment, and Hyallela limitations. Each replicate contained 100 mL whole wet sediment overlain by 175 mL upstream river control water, in 400 mL glass beakers (EC, 1996). These test vessels were washed with soap and water, acid-washed, solvent-washed, and thoroughly rinsed with distilled water prior to testing (EC, 1996). Test sediments were stirred to homogenize the samples, and large detritus and visible indigenous organisms were removed with forceps (EC, 1996). Hyallela azteca, 2 to 9 days in age, were provided by B.C. Research Inc. Ten organisms were used per replicate. The test was static (no changeover of overlying water) except for the replacement of water lost due to 108 evaporation. There was continuous aeration of each test chamber at 2-3 bubbles per second (EC, 1996). The lighting regime was 16 h light (900 lux), and 8 h darkness, and the temperature was 25 °C. Test chambers were fed 1.5 mL of YCT 3 times per week (EC, 1996). Total ammonia, pH, conductivity, and hardness were measured in the overlying test waters at the start and end of the test, and dissolved oxygen and temperature were measured 3 times per week (EC, 1996). 5.7.2 Test Endpoints Mean percent survival, and mean Hyallela dry weight, were calculated for each sample and statistically compared to that of the upstream control sediment using one-tailed t-tests. 5.8 Chemical Analyses of Sediment Samples 5.8.1 Percentage Organic Matter The frozen sediment samples were thawed at room temperature. Sediments were dried at 103°C, weighed, ashed at 550 °C, and then re-weighed. The percent weight loss on ignition (LOI) was used as a measure of the sample's percent organic matter content. 5.8.2 Metals Sediment samples dried at 103 °C were ground with mortar and pestle, and sieved through a 1.5 mm screen to remove debris. Five grams of each dried sediment was ashed for 1 hour at 400 °C to destroy the organic matter. The remaining ashed sediments were dissolved with 5 ml of concentrated nitric acid for 1 hour. The acid slurries were transferred into Erlenmeyer flasks, and refluxed with approximately 25 mL of distilled water on a hot plate for a half hour. These solutions were then filtered (Whatman 541 filter), and 109 the filtrates were diluted to 50 mL volume using distilled water. These extracts were analyzed using an ICP analysis performed by UBC's Soil Science Department. The digestion/extraction procedure's accuracy was tested by comparing measured metal concentrations of a certified reference sediment (MESS-2) with those reported by its supplier, the National Research Council of Canada (NRC, Date unknown). 6.0 RESULTS AND DISCUSSION 6.1 Water Samples' Ceriodaphnia dubia Chronic Toxicity Test Results Table 15 summarizes the C. dubia chronic survival and reproduction test results. Table 15. Ditch and River Water Samples' C. dubia Chronic Survival and Reproduction Test Results. May 6 Sample Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Discharge Site 8 7d 100 24.2 +/- 6.7 > Culture Cont. Culture Control X 7d 80 14.3 +/-4.6 June 6 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 8d 100 37.2 +/- 5.3 = Upstream, > Culture Cont. Site 2 8d 100 41.0 +/- 6.5 = Upstream, > Culture Cont. Site 3 8d 100 44.8 +/- 8.4 = Upstream, > Culture Cont. Site 4 8d 90 34.4+/-13.4 = Upstream, > Culture Cont. Site 5 8d 100 33.6 +/- 6.7 = Upstream, > Culture Cont. Site 6 Not Sampled Site 7 8d 100 40.3 +/- 5.9 = Upstream, > Culture Cont. Site 8 8d 100 33.5 +/- 4.0 = Upstream, > Culture Cont. Site 9 Not Sampled Site 10 8d 90 35.8 +/- 4.0 = Upstream, > Culture Cont. Site 11 8d 100 41.1 +/-9.0 = Upstream, > Culture Cont. Site 12 8d 100 40.6 +/- 7.5 = Upstream, > Culture Cont. Culture Control 8d 89 26.2 +/-11.1 Upstream Control 8d 100 36.6 +/- 4.4 110 Table 15 Cont. June 26 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 6d 100 21.9 +/-6.7 = Upstream, = Culture Cont. Site 2 6d 90 18.3 +/-8.2 = Upstream, = Culture Cont. Site 3 6d 100 21.5 +/- 7.3 = Upstream, = Culture Cont. Site 4 6d 90 21.5 +/- 8.6 = Upstream, = Culture Cont. Site 5 * 6d 100 15.7 +/-6.3 < Upstream, < Culture Cont. Site 6 Not Sampled Site 7 6d 100 19.3 +A8.4 = Upstream, = Culture Cont. Site 8 © 6d 100 0 < Upstream, < Culture Cont. Site 9 Not Sampled Site 10 6d 100 24.6 +/- 3.5 = Upstream, = Culture Cont. Site 11 * 6d 100 16.4 +/-7.5 < Upstream, = Culture Cont. Site 12 * 6d 100 12.5 +/-5.0 < Upstream, < Culture Cont. Culture Control 6d 100 21.3 +/- 6.5 Upstream Control 6d 100 22.9 +/- 5.2 July 16 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 7d 80 23.1 +/- 11.0 = Upstream, = Culture Cont. Site 2 7d 100 25.8 +/- 6.7 > Upstream, > Culture Cont. Site 3 7d 100 18.4 +/- 12.9 = Upstream, = Culture Cont. Site 4 * E 7d 60 10.9+/-6.0 < Upstream, = Culture Cont. Site 5 © 7d 0 0 < Upstream, < Culture Cont. Site 6 Not Sampled Site 7 7d 100 20.3 +/- 11.1 = Upstream, = Culture Cont. Site 8 7d 100 13.5 +/-7.1 — Upstream, = Culture Cont. Site 9 Not Sampled Site 10 7d 70 17.0 +/-11.7 = Upstream, = Culture Cont. Site 11 7d 100 19.2 +A7.2 = Upstream, = Culture Cont. Site 12 7d 90 18.0 +/-11.1 = Upstream, = Culture Cont. Culture Control 7d 90 15.3 +/-9.3 Upstream Control 7d 100 19.7 +/-8.2 I l l Table 15 Cont. Aug. 7 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 6d 90 13.0 +A8.2 = Upstream, = Culture Cont. Site 2 6d 100 19.1 +A3.2 > Upstream, = Culture Cont. Site 3 6d 100 15.1 +/-5.1 = Upstream, = Culture Cont. Site 4 6d 100 15.7 +/-4.7 = Upstream, = Culture Cont. Site 5 6d 100 16.2 +/-6.6 = Upstream, = Culture Cont. Site 6 Not Sampled Site 7 6d 90 10.9 +/-6.0 = Upstream, < Culture Cont. Site 8 6d 100 19.2 +/-3.6 = Upstream, = Culture Cont. Site 9 Not Sampled Site 10 6d 90 12.7+/-5.0 = Upstream, = Culture Cont. Site 11 6d 80 9.9 +/- 6.7 = Upstream, < Culture Cont. Site 12 6d 100 14.3+/-7.2 = Upstream, = Culture Cont. Culture Control 6d 100 15.7 +A6.2 Upstream Control 6d 100 14.9 +/-4.7 Aug. 29 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 6d 100 19.7 +/-4.5 = Upstream, = Culture Cont. Site 2 6d 100 17.2 +/-5.3 = Upstream, = Culture Cont. Site 3 6d 100 19.8 +A5.3 = Upstream, > Culture Cont. Site 4 6d 90 18.9 +/-7.1 = Upstream, = Culture Cont. Site 5 6d 90 16.5+/-6.7 = Upstream, = Culture Cont. Site 6 Not Sampled Site 7 6d 100 16.8 +/-6.5 = Upstream, = Culture Cont. Site 8 6d 100 18.8 +/-2.4 = Upstream, > Culture Cont. Site 9 Not Sampled Site 10 6d 100 17.0 +/-5.3 = Upstream, = Culture Cont. Site 11 6d 100 22.0 +/- 3.4 = Upstream, > Culture Cont. Site 12 * 6d 100 13.2 +/-7.0 < Upstream, = Culture Cont. Culture Control 6d 100 15.1 +/-5.7 Upstream Control 6d 100 19.7 +/-5.7 112 Table 15 Cont. Sept. 17 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 * 6d 100 16.3 +/-4.5 < Upstream, = Culture Cont. Site 2 6d 100 19.1 +/-5.5 = Upstream, > Culture Cont. Site 3 * 6d 100 16.3 +/- 6.8 < Upstream, = Culture Cont. Site 4 6d 100 21.0 +/-3.7 = Upstream, > Culture Cont. Site 5 * 6d 100 17.2 +/-4.0 < Upstream, > Culture Cont. Site 6 6d 100 20.0 +/- 4.7 = Upstream, > Culture Cont. Site 7 * 6d 90 16.6 +A8.7 < Upstream, = Culture Cont. Site 8 6d 100 23.7 +/- 2.3 = Upstream, > Culture Cont. Site 9 Not Sampled Site 10 6d 100 22.6 +/- 2.9 = Upstream, > Culture Cont. Site 11 6d 90 24.3 +/- 6.4 = Upstream, > Culture Cont. Site 12 6d 100 21.4 +/- 3.9 = Upstream, > Culture Cont. Culture Control X 6d 100 13.6 +A2.2 Upstream Control 6d 100 22.0 +A 2.4 Oct. 3 Samples Test Duration Survival % Mean # Neonates +A S.D. Reproduction Comparison With Controls Site 1 7d 90 18.5 +A5.0 > Upstream, > Culture Cont. Site 2 7d 70 14.5 +A8.9 = Upstream, < Culture Cont. Site 3 7d 100 20.8 +A 6.6 > Upstream, = Culture Cont. Site 4 7d 100 16.5 +A6.8 = Upstream, < Culture Cont. Site5 7d 90 16.2 +A7.8 = Upstream, < Culture Cont. Site 6 7d 70 11.3 +A 8.5 = Upstream, < Culture Cont. Site 7 7d 100 21.1 +A9.0 > Upstream, = Culture Cont. Site 8 7d 100 16.1 +A4.4 = Upstream, < Culture Cont. Site 9 Not Sampled Site 10 7d 100 16.7 +A6.1 = Upstream, < Culture Cont. Site 11 7d 90 16.6 +A4.1 = Upstream, < Culture Cont. Site 12 7d 100 17.5 +A8.8 = Upstream, < Culture Cont. Culture Control 7d 100 23.6 +A 5.7 Upstream Cont. X 7d 100 13.2 +A 7.5 Oct. 3 Retesting (Began Oct. 21) Test Duration Survival % Mean # Neonates +A S.D. Reproduction Comparison With Controls Site 2 6d 100 21.2 +A2.5 = Upstream, = Culture Cont. Site 6 6d 100 17.7 +A4.3 = Upstream, = Culture Cont. Culture Control 6d 100 18.1 +A6.3 Upstream Control 6d 100 18.9 +A4.0 Table 15 Cont. 113 Oct. 16, Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 5 6d 100 18.8 +/-5.3 = Upstream, = Culture Cont. Site 10-Nearby * Field Puddle 6d 80 8.6 +/- 4.5 < Upstream, < Culture Cont. Culture Control 6d 100 18.1 +/-6.3 Upstream Control 6d 100 18.9+/-4.0 Nov. 21 Samples Test Duration Survival % Mean # Neonates +/- S.D. Reproduction Comparison With Controls Site 1 7d 90 21.2 +/- 9.1 = Upstream, = Culture Cont. Site 2 7d 100 25.1 +/- 9.6 - Upstream, = Culture Cont. Site 3 7d 100 28.4 +/- 7.7 > Upstream, > Culture Cont. Site 4 7d 100 23.4 +/- 6.5 = Upstream, = Culture Cont. Site 5 7d 100 20.8 +/-3.1 = Upstream, = Culture Cont. Site 6 7d 100 22.2 +/- 5.0 = Upstream, = Culture Cont. Site 7 7d 100 20.2 +/- 7.1 = Upstream, = Culture Cont. Site 8 * 7d 90 11.4 +A6.5 < Upstream, < Culture Cont. Site 9 Not Sampled Site 10 7d 100 22.4 +/- 7.5 = Upstream, = Culture Cont. Site 11 7d 100 26.6 +/- 7.5 = Upstream, > Culture Cont. Site 12 7d 100 19.7 +/-5.7 = Upstream, = Culture Cont. Culture Control 7d 100 20.7 +/- 3.6 Upstream Control 7d 100 23.4 +/- 4.5 © Lethally toxic sample. Sublethally toxic sample. E Effect observed; partial paralysis by test end of 3 of 6 surviving organisms. X Control does not meet EC (1992) requirements for a valid control (mean neonate production per surviving female < 15). Conclusions from comparison to this control are questionable. The t-test results supporting Table 15's sublethal results are shown in Appendix 2. 114 6.1.1 Lethally Toxic Samples Of the 85 ditchwater and riverwater samples collected from the study site, only two produced C. dubia survival that was significantly less than survival in the upstream river control (Site 13) and/or the White Rock pond culture/control in t-testing. These two lethally th toxic samples were Site 8, the inner ditch running parallel to the east side of 176 St., collected on June 26 t h, 1997, and Site 5, South Cloverdale Ditch at 40 t h Ave., collected th on July 16 , 1997. Immediate retesting of these toxic samples using the dilution series of 100%, 56%, 32%, 18% and 10% ditch sample in culture/control water produced the LC50s, NOECs, LOECs, and IC25s presented in Table 16. Table 16. Lethally Toxic Samples' C. dubia Dilution Series Test Endpoints. Sample 6-Day Reproduction NOEC, LOEC, and IC25 (+/- S.D.) Survival LC50 (+/- 95% C.l.) NOEC LOEC IC25 48-h LC50 96-h LC50 6-day LC50 Site 8, June 26 176th St. Ditch 32% NA 42.7 % (+/- 3.5) > 100% 74.8 %1 (56.0,100.0) 39.9 % 1 (32.0,56.0) Site 5, July 16 South Cloverd. Ditch 32% NA 19.9% (+/-12.8) 92.9 % 2 52.2 %1 (32.0,100.0) 36.5 % 3 (28.6, 46.3) NA = There was no LOEC since reproduction was not inhibited in 10%, 18%, or 32% relative to the control, and complete mortality in the 56% and 100% concentrations eliminated these concentrations from the LOEC calculations, following US EPA (1993a). 1 Computer program used Binomial method to calculate the LC50. 2 Computer program used Non-linear interpolation method to calculate the LC50. 3 ^ Computer program used Probit method to calculate the LC50. 115 The reproduction LOECs and IC25s for the Sites 5 and 8 samples are largely, if not entirely, based on lethality. For the lethally toxic samples, prior to mortality the C. dubia exhibited erratic swimming, followed by lack of movement except for rapid twitching of their antennae, and finally complete paralysis. 6.1.2 Sublethally Toxic Samples Qualitative conclusions from the statistical comparison of mean neonate production in the runoff ditch and river water samples versus the upstream control (Site 13) and culture/control were included in Table 15. The conclusions from statistical comparisons with controls which did not produce a mean 15 neonates are questionable. EC (1992) states that a control is invalid if mean young production is < 15 neonates per surviving th th female. This would invalidate these controls: May 6 , culture/control; September 17 • culture/control; and October 3 r d ' upstream control. Since these controls came close to meeting EC (1992) requirements they were still included for statistical analyses. A sample was deemed sublethally toxic if reproduction was statistically significantly less than the upstream control. Comparisons were made to the White Rock pond culture/control in order to draw conclusions in the event that the upstream control did not meet EC (1992) requirements. The t-test results supporting Table 15s sublethal effect conclusions have been included in Appendix 2. The sublethally toxic samples are summarized in Table 17. The July 16 t h riverwater sample from Site 4 had a 7-day survival of 60%. This survival was not statistically lower than the upstream control; however, only 3 of Site 4's surviving 6 test organisms appeared healthy by day 7 of the test. These 3 unhealthy organisms exhibited the same symptom of partial paralysis, except for antennae twitching, which was 116 observed in the two lethally toxic ditch samples and tests using purchased pesticides. The Site 4 C. dubia test results suggest that there was possibly OP river contamination on July th 16 , likely due to discharge from the South Cloverdale Ditch. This is later elaborated on in greater detail. Table 17. Samples Exhibiting Sublethal Toxicity, where Reproduction was Less than Upstream Control (excluding 2 lethally toxic samples). Sample Location Date Collected Mean # Neonates +/- S.D. Site 5, South Cloverdale Ditch June 26 15.7+/-6.3 Site 11, Burrows Ditch June 26 16.4 +/- 7.5 Site 12, Old Logging Ditch June 26 12.5 +/- 5.0 Site 4, Nicomekl River July 16 10.9+/-6.0 Site 12, Old Logging Ditch August 29 13.2 +/-7.0 Site 1, Nicomekl River Sept. 17 16.3+/-4.5 Site 3, Nicomekl River Sept. 17 16.3+/-6.8 Site 5, South Cloverdale Ditch Sept. 17 17.2 +/-4.0 Site 7, Nicomekl River Sept. 17 16.6 +/- 8.7 Site 10 - Nearby Field Puddle Oct. 16 8.6 +/- 4.5 Site 8, 176th St. Ditch Nov. 21 11.4 +/-6.5 Reproduction in the non-toxic ditch and river samples from the study site was usually rd higher than that in the White Rock pond culture/control samples. The October 3 test results were atypical, with neonate production in most of the Nicomekl system samples being significantly less than in the White Rock pond culture/control water. It appears as if the poor reproduction in Nicomekl system waters collected on this date may have been rd due to fungal contamination. The October 3 Site 13 upstream riverwater sample and 117 Site 6 river sample both exhibited fungal contamination. The fungus observed in the test vessels did not appear to have been introduced with the test neonates, which had been produced in fungal contaminated culture water, since the fungus observed in the river and ditchwater samples was parasitic and filamentous, whereas the fungus observed in the White Rock pond culture water was non-parasitic and non-filamentous. Swain and Holms (1988b) reported that fungal growths in the Nicomekl River have been caused by leachate from the landfill between Murray and Anderson Creeks. However, since Site 13 is upstream of the old landfill, it is more likely that the river's fungal contamination was a natural fall event. The culture/control water used initially for the October 3 test was contaminated with fungus, but was replaced with filtered water that was not contaminated th with fungus for the third water changeover (October 5 ) and for the remainder of the test. rd Hence, the majority of the October 3 culture/control's testing was performed using fungus-free water, likely accounting for this control's atypical higher reproduction relative to the Nicomekl ditch and river samples. 6.1.3 C. dubia Culture Health and Reference Toxicant Evaluation The majority of neonates used to test the samples and reference toxicants appeared to come from brood organisms in good health, as indicated by records kept of neonate production, for 4 brood animals from each 14 day (maximum-age) set of 30 to 50 brood animals. An exception was the neonates used for the full strength ditchwater and rd riverwater tests commencing October 3 . As previously mentioned, the neonates used for the October 3 r d tests came fungus contaminated culture water, and unfortunately brood in poor health. The 4 adults from this population, monitored for reproduction, exhibited limited neonate production, and ended up dying prematurely. Based on the test criteria 118 set by EC (1992) this alone, irrespective of the < 15 mean neonates produced in the upstream control, invalidates the October 3 r d tests. Poor culture health was an intermittent th problem for short durations in September and October. However The September 17 , October 3 RETESTS, and October 16 t h tests, were all supplied with neonates from brood sets which appeared healthy, with adequate daily neonate production. In addition to the monitoring of brood health, reference toxicant tests were performed once a month throughout the track the culture's health. The IC50 (mg/L NaCI) endpoints ± 2 S.D. for the individual reference toxicants are shown in Appendix 3. EC (1992) states that providing an individual reference toxicant IC50 falls within the mean IC50 ± 2 S.D. from previous testing, then the culture can be assumed to be in good health at the time of this test. The reference toxicants performed were all within or marginally outside the mean IC50 ± 2 S.D. of 1371 ± 183 mg/L NaCI. Consequently, the sensitivity of the test organisms was consistent throughout the study. The toxicity laboratory which provided the C. dubia, B.C. Research, itself produced a similar mean IC50 ± 2 S.D. of 1476 ± 304 mg/L NaCI for its own reference toxicants, between April and November of 1997. The sixth th reference toxicant performed, which began on October 29 failed due to 30% control mortality on day 5 of the test, which may be either coincidental or due to fungal contamination. This reference toxicant's IC50 results are those up to day 4 of the test. 119 6.2 Lethally Toxic Samples' Biological Toxicity Identification Evaluation T h e C 1 8 treatment el iminated toxicity from the 1 7 6 t h St. and South C loverda le Ditch s a m p l e s . Tab le 18 shows the one-tai led t-test p-value results compar ing survival and reproduct ion of the C 1 8 treated toxic samp les with that in the culture/control water. A l l three chronic tests were conducted at the same time and used neonates from the same broods. Tab le 18. Resu l ts of C 1 8 C. dubia B io logical Toxicity Identification. June 26 t h Site 8,176 th St. Ditch Control Untreated C18 Treated Survival (%) 100 0 89 P-value (1-tailed) v. Cont. 0.084 Mean # Neonates +/- S.D 18.9+/-6.4 0 15.0+/-7.3 P-value (1-tailed) v. Cont. 0.110 July 16 Site 5, South Cloverdale Ditch Control Untreated C18 Treated Survival (%) 90 0 100 P-value (1-tailed) v. Cont. 0.084 Mean # Neonates +/- S.D. 15.4 +/-4.6 0 12.1 +/-6.3 P-value (1-tailed) v. Cont. 0.100 One- ta i led p-values are for culture/control versus C 1 8 treated samples . P-va lue > 0.05 indicates that there is no statistically signif icant dif ference. 120 Since both lethally toxic samples were statistically speaking completely detoxified using the C18 treatment, this suggested that the contaminant(s) responsible for the toxicity were organic and slightly non-polar in nature. Lethality testing was also performed on the June 26 t h 176th St. ditch (Site 8) water sample th and July 16 South Cloverdale Ditch (Site 5) water sample, treated with 200 ppb of piperonyl butoxide (PBO), to determine whether the toxicant(s) were metabolically active organophosphorous insecticide. Table 19 shows the results of the first PBO toxicity identification. The control water used for comparison was culture/control water from the White Rock pond. Table 19. Results of 96-h C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at -10 °C for Two Months in Plastic Bottles).. Sample % Survival +/-S.D. Comparison P-Value Result (Survival) Control 85 +/- 30 % Cont. v. MeOH Cont. 0.196 Cont. = MeOH Cont. MeOH Control 100 +1-0% Site 8 85 +/- 19% Site 8 v. Cont. 0.5 Site 8 = Cont. Site 8 PBO 85 +/- 19% Site 8 PBO v. Cont. 0.0002 * Site 8 PBO > Cont. Site 8 PBO v. MeOH Cont. 0.108 Site 8 PBO = MeOH Cont. Site 8 v. Site 8 PBO 1.000 Site 8 = Site 8 PBO Site 5 15+/-30% Site 5 v. Cont. 0.008 * Site 5 < Cont. Site 5 PBO 95 +/- 10% Site 5 PBO v. Cont 0.281 Site 5 PBO = Cont. Site 5 PBO v. MeOH Cont. 0.196 Site 5 PBO = MeOH Cont. Site 5 v. Site 5 PBO 0.004 * Site 5 < Site 5 PBO * P-value < 0.05 indicates that there is a statistically significant difference. 121 The 96-hour testing on the 2-month old samples which had been frozen in plastic showed that piperonyl butoxide statistically detoxified the Site 5 sample. This suggested the presence of metabolically active organophosphate insecticides as the toxicant(s) in the South Cloverdale Ditch. Significant mortality was not observed in the Site 8 sample, so th conclusions over the nature of the 176 St. Ditch's toxicant(s) could not be made using the PBO non-PBO comparison. It is believed that the Site 8 sample's toxicant(s) either had a higher affinity to bind to plastic or were less stable than the toxicant(s) in Site 5 sample. A second PBO toxicity identification was performed using the same samples at a later date. These tests were conducted on portions of the lethally toxic samples which had been stored in the dark at 4 °C in glass bottles for 5 months. The results of the second PBO tests are shown in Table 20. Table 20. Results of 96-h C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at 4°C in glass for 5 months). Sample % Survival +/- S.D. Comparison P-Value Result (Survival) Control 100 +/-0% Cont. v. MeOH Cont. 0.5 Cont. = MeOH Cont. MeOH Control 100 +1-0% Site 8 35 +/-19 % Site 8 v. Cont. 0.003* Site 8 < Cont. Site 8 PBO 100 +1-0% Site 8 PBO v. Cont. 0.5 Site 8 PBO = Cont. Site 8 PBO v. MeOH Cont. 0.5 Site 8 PBO = MeOH Cont. Site 8 v. Site 8 PBO 0.003 * Site 8 < Site 8 PBO Site 5 100 +1-0% Site 5 v. Cont. 0.5 Site 5 = Cont. Site 5 PBO 100 +1-0% Site 5 PBO v. Cont. 0.5 Site 5 PBO = Cont. Site 5 PBO v. MeOH Cont. 0.5 Site 5 PBO = MeOH Cont. Site 5 v. Site 5 PBO 0.5 Site 5 = Site 5 PBO * P-value < 0.05 indicates that there is a statistically significant difference. 122 For the 96-hour testing of the 5-month old samples which had been refrigerated in glass there was significant mortality in the Site 8 sample, and the piperonyl butoxide treatment significantly increased survival. This established the presence of metabolically active organophosphate insecticides as the toxicant(s) in the 176 t h St. Ditch. However, in this second round of testing significant toxicity was this time not observed in the Site 5 sample. Continuation of testing to 7 days eventually produced significant mortality in the Site 5 sample, which was again not present in the Site 5 sample treated with PBO. These 7-day test results are shown in Table 21. Table 21. Results of 7-day C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at 4°C in glass for 5 months). Sample % Survival +/- S.D. Comparison P Value Result (Survival) Control 5+/- 10% Cont. v. MeOH Cont. 0.006 * Cont. = MeOH Cont. MeOH Control 5+/- 10% Site 8 5+/- 10% Site 8 v. Cont. 0.5 Site 8 = Cont. Site 8 PBO 100 +/-0% Site 8 PBO v. Cont. 0.0002 * Site 8 PBO > Cont. Site 8 PBO v. MeOH Cont. 0.0002 * Site 8 PBO > MeOH Cont. Site 8 v. Site 8 PBO 0.0002 * Site 8 < Site 8 PBO Site 5 15+/-19% Site 5 v. Cont. 0.199 Site 5 = Cont. Site 5 PBO 85 +/- 30 % Site 5 PBO v. Cont. 0.004 * Site 5 PBO > Cont. Site 5 PBO v. MeOH Cont. 0.004 * Site 5 PBO > MeOH Cont. Site 5 v. Site 5 PBO 0.006 * Site 5 < Site 5 PBO * P-value < 0.05 indicates that there is a statistically significant difference. 123 Control organisms did not survive to 7 days, likely due to lack of food; however, PBO treated Site 8 and 5 C. dubia survival did not appear to be affected due to the lack of food. Natural food was likely plentiful in the ditch water. Reported metabolically activated OPs include malathion, diazinon, chlorpyrifos (Bailey et al., 1997), methyl parathion, dimethoate, and azinphos methyl (Matsumura, 1975), in decreasing order of the amount of each used for commercial agriculture in the Lower Mainland in 1991. Consequently, one of these, or a combination of these were likely responsible for the Site 8 and 5 toxicity. Malathion was reported to have a soil half-life of only 0.8 days (Rao and Davidson, 1980). Diazinon was reported to have a soil half-life of 32 to 48 days (Rao and Davidson, 1980). Chlorpyrifos persists in soils for 60 to 120 days (BCPC, 1991). Although malathion is perhaps used in high quantities in the study site, given diazinon and chlorpyrifos' much longer soil half-lives, it is more likely that these latter two compounds were responsible for the observed toxicity. 6.3 Ditch and River Water Samples' Chemical Analyses 6.3.1 Organics It was anticipated that toxic samples would be collected frequently and that the techniques used for the GC/MS analyses could be modified and perfected throughout the study, such that the concentration of pesticides in later toxic samples could be more accurately ascertained However, since only two lethally toxic samples were found early in the study much of the method development had to be performed on these two samples after being subjected to storage. Several different extraction and GC/MS techniques were performed in an attempt to support the presence of OP insecticides in the two lethally toxic samples, 124 as indicated by the piperonyl butoxide results. Unfortunately the accuracy of the results for the better techniques performed later in the study was weakened due to the storage time and storage techniques from the day of original sampling. The presence and approximate quantification of a few pesticides was determined in the samples exhibiting lethality. None of the solid phase extraction full-ion-scan gas chromatography/mass spectroscopy (GC/MS) analyses revealed the presence of any organophosphate. insecticides in either sublethally or lethally toxic samples. Numerous large GC peaks were produced, but none were identified as OPs using their mass spectra and the computer data base. At the time th of testing this was surprising since PBO testing clearly showed that toxicity in the 176 St. and South Cloverdale Ditch samples was likely due to OP insecticide contamination. Lack of detection is attributed to several possible causes. There may have been failure of the SPE tubes to extract the insecticides from solution, degradation of the insecticides adsorbed to the columns while in storage, or only partial release of the insecticides upon elution of the tubes. However, Lacorte et al. (1995) used a similar C 1 8 SPE technique to that above, in filtering groundwater samples spiked at 10 //g/L with 19 different OP pesticides. They showed that complete recovery of 16 of the 19 pesticides was observed for SPE columns kept at -20°C for 8 months, with the exception of mevinphos, dichlorvos, and phosmet for which degradation did occur, and recovery was only 28 to 58%. They also found complete recovery of the same 16 OP pesticides when SPE columns were stored at only 4°C for up to 3 months, with 22 to 79% recovery of mevinphos, dichlorvos, and phosmet. Due to the proven high recoveries of the SPE 125 technique used, the reason for failed detection was more likely that the large peaks from the samples' natural organics, and the GC's high background level, masked the detection of the low concentrations of OP insecticide(s) capable of producing toxicity. The SPE GC/MS full-ion-scans of the two lethally toxic samples did reveal the presence of the triazine herbicide, prometryn, in the Site 8 sample with a GC retention time of approximately 33.73 min. (Appendix 4). Based on the standard curve for prometryn in acetonitrile (Appendix 5), it was determined that the: concentration of prometryn was approximately 1.15 mg/L in the concentrated extract, or 2 /ug/L in the original sample. Direct liquid-liquid extractions were also performed on the two lethally toxic samples to determine if this technique would produce better results than the solid phase extractions. Liquid-liquid extractions were performed on the same portions of the two lethally toxic samples which had been frozen for 2 months in plastic that were used for the PBO analysis. The liquid-liquid extractions analysed by GC/MS full-ion-scan also did not reveal any OP insecticide(s) in the lethally toxic samples. The problem of excessive background noise with the full-ion-scan still existed, and the storage in plastic had reduced the concentrations of OP insecticide(s) in.these.samples, as evidenced in the first PBO tests. th The liquid-liquid extractions also revealed the presence of prometryn in the 176 St. (Site 8) sample. Based on the standard curve for prometryn in methylene chloride (Appendix 5), it was determined that the extract contained 1.175 mg/L of prometryn or 3 //g/L in the original sample. Consequently, it appears that the liquid-liquid extraction and solid phase extraction gave similar recoveries of prometryn. Prometryn is much more water soluble 126 and has a longer half-life than OP insecticides, and was likely less prone to adsorption to the plastic storage containers or degradation, allowing for a somewhat reasonable comparison between the SPE and LL extraction techniques. Another possible explanation for failed detection of the OP insecticide(s) likely present in the lethally toxic samples, was that their concentration(s) were below the detection limit of the full-ion-scan techniques used. The concentration factor could have been increased by filtering larger volumes of sample through the SPE column. This could not be achieved due to the plugging of SPE columns with suspended solids when filtering volumes of ditch and river samples much greater than 250 mL. Could the techniques used detect lethal concentrations of pesticides? Consider the case of diazinon and chlorpyrifos. Diazinon exhibited a 48-h LC50 of 0.4//g/L and chlorpyrifos exhibited a 48-h LC50 of 0.5 //g/L, in lethality testing of these two compounds (presented later). If one assumes that diazinon was the primary toxicant in the two toxic ditch samples, could there have been detection of this compound by the instrument if the samples concentration had been only 0.4 //g/L, using the full-ion-scan, and only 360X or 500X concentration by the liquid-liquid extraction and SPE, respectively? For the higher 500X concentration, if 0.4 //g/L of diazinon was present in either of the toxic samples, then after SPE, SPE elution, and solvent concentration, 200 //g/L of diazinon should have injected into the GC MS. Is the instrument capable of measuring 200 //g/L of diazinon? Standards of 80 //g/L, 160 //g/L and 800 //g/L of diazinon in methylene chloride analyzed by the GC/MS in the full-ion-scan mode produced peak areas of 1369, 14,874, and 127 222,874, respectively. Similarly, standards of chlorpyrifos at 13.2 //g/L, 26.4 //g/L, 66 //g/L, 132 //g/L, and 1,320 //g/L were analyzed by the GC/MS in full-ion-scan testing. The three lower concentrations of chlorpyrifos were entirely not detectable in these standard samples, and the concentrations of 132 //g/L and 1,320 //g/L chlorpyrifos produced peak areas of only 13,831 and 1,038,597. There appears to be a non linear relationship between concentration and peak area when comparing standards with near detection limit concentrations to higher concentrations. In theory following a 500X concentration 0.16 //g/L (80 //g/L standard) diazinon, and :0.26 //g/L (132 //g/L standard) chlorpyrifos can produce detectable peak areas with the full-ion-scan. However, these peak areas are very low. To put them into perspective, there were several impurities in the methylene chloride standards which themselves produced peaks with areas over 300,000. The 500X concentrated liquid-liquid extraction full-ion-scans of the Sites' 8 and 5 lethally toxic samples had several distinguished peaks, most of which had peak areas much greater than the 222,874 area peak produced by even the 800 //g/L diazinon standard. For ditchwater and riverwater samples, the background baseline of the chromatogram using the full-ion-scan was at an abundance great enough that a peak from a diazinon concentration of 200 //g/L (corresponding to 0.4 //g/L in the original -sample) injected into the instrument was likely not observable. For the ditchwater, and riverwater samples, it was hard to identify even distinguishable small peaks of areas as high as 1,000,000 by looking at the peak's mass/charge spectra, again due to the excessive amount of background spectra picked up by the full-ion-scan. The SPE full-ion-scan technique used was unable to detect 0.8 //g/L diazinon and 0.132 //g/L chlorpyrifos in upstream Nicomekl Riverwater using the SPE full-ion-scan technique. 128 In a further attempt to determine what pesticides were present in the 176 t h St. and South Cloverdale Ditch lethally toxic samples, their stored liquid-liquid extractions were concentrated to 3600X the original samples. It was expected that the concentration to 3600X would allow for identification of the OP pesticides suspected in these samples, although quantification for the original sample would be influence by storage of the original samples prior to the liquid-liquid extraction, and storage of the extractions prior to further concentration. However, again using the full-ion-scan, the presence of pesticides could still not be distinguished amongst the multitude of large peaks and the high baseline, since the impurities had also been further concentrated. Consequently, the GC/MS was programed to perform a select-ion-scan analysis of the 3600X concentrated toxic samples looking specifically for diazinon and chlorpyrifos by scanning for ions with the mass to charge ratio of 137,179, 197, 304, and 314 characteristic of these compounds. Diazinon's mass spectra has an abundance of the 137, 179, and 304 mass to charge ratios as seen in the full ion scan of the 800 fxg/L diazinon standard and the computer's database mass spectra for this compound. Chlorpyrifos' mass spectra has an abundance of the 197 and 314 mass to charge ratios as seen in the full ion scan of the 1,320 //g/L chlorpyrifos standard and the computer's database mass spectra for this compound. The select-ion-scan of the lethally toxic samples had much less background noise'than the full-ion-scan and small peaks of lower peak area were more distinguishable (Appendix 6). Diazinon and chlorpyrifos standards as low as 16 //g/L and 26.4 //g/L, respectively, were detectable using the select-ion-scan analysis. The 3600X concentrated toxic samples revealed the presence of diazinon and chlorpyrifos in the Site 8 sample and diazinon in the Site 5 sample. The pesticides were identified in 129 the samples based on two observations. Peaks were observed for the samples with similar retention times as found for the diazinon (31.354 min.) and chlorpyrifos (34.989 min.) standards. The Site 8 sample produced distinguishable peaks at 31.348 and 34.992 min. (Appendix 7). The Site 5 sample produced a distinguishable peak at 31.348 min. (Appendix 7). These peak locations closely matched those of the diazinon and chlorpyrifos standards using the select-ion-scan (Appendix 8). In addition to identification by peak location, the mass to charge spectra for the 5 select ions at each of these peaks (Appendix 7) when compared to that of the standards (Appendix 8) was further indicative that the peaks were in fact diazinon and chlorpyrifos. Quantification of the diazinon and chlorpyrifos in the samples was determined. The diazinon standards of 16 //g/L, 32 //g/L, 80 //g/L, and 800 //g/L in methylene chloride, using the select ion scan analysis produced peak areas of 1931, 3305, 16121, and 563,923, respectively. Based on a curve drawn as linear segments between the above standards (Appendix 9), the diazinon concentration in the Site 8 sample was at least 0.023 //g/L and the diazinon concentration in the Site 5 sample was at least 0.033 //g/L. The chlorpyrifos standards of 26.4 //g/L, 132 //g/L in methylene chloride, usingselect ion scan analysis produced peak areas of 2902 and 1815, respectively. The chlorpyrifos standard of 1,320 //g/L using the full ion scan analysis produced a peak area of 1,038,597. Based on a standard curve drawn as linear segments between the above standards (Appendix 10), the chlorpyrifos concentration in the Site 8 sample was at least 0.067 //g/L. For the chlorpyrifos standard curve, the 1,320 //g/L standard's full-ion-scan had to be used, since the computer accidentally overwrote the 1,320 //g/L standard's select-ion-scan. Due to the fact that the liquid-liquid extractions were performed on samples which had been frozen 130 for 2 months in plastic containers, and the additional fact that the liquid-liquid extractions themselves were stored in the freezer for 3 months before they were concentrated to 3600X and analysed with the select ion scan, it is believed that the concentrations of diazinon and chlorpyrifos in the original samples were likely higher at the time of initial toxicity testing. The PBO toxicity tests used the same solutions as the liquid-liquid extractions and showed 96-h survival of 85% and 15% in the Site 8 and Site 5 samples, compared to the 0% 96-h survival observed in both of the original fresh samples. Furthermore, standards curves used were not highly accurate since they were developed using a limited number of standards, and assumed linearity between the standards, for simplicity. Since the select-ion-scan was required to identify and quantify pesticides at the low concentrations found, the SPE 500X concentration technique was re-evaluated on standards of low diazinon and chlorpyrifos concentrations in upstream Nicomekl River water. Based on the standard curve for diazinon (Appendix 9) the select-ion-scan analysis of the 500X concentration SPE technique for prepared solutions of 0.8 //g/L and 0.4 //g/L diazinon in Site 13 upstream control water quantified diazinon at only 0.4 //g/L and 0.1 //g/L, respectively. In other words, diazinon was under-quantified using the SPE technique. Likewise, based on the standard curve for chlorpyrifos (Appendix 10) the select-ion-scan of the 500X SPE technique for 0.066 //g/L chlorpyrifos in upstream control water quantified Chlorpyrifos at 0.16 //g/L. In other words, chlorpyrifos was over-quantified using the SPE technique. Therefore, while the SPE technique is capable of detecting pesticides at low concentrations with only a 500X concentration when using the select-ion-scan, this minimal concentration of the sample, and the use of low concentration standards 131 does not produce accurate quantification, as shown above. Consequently, the author recommends that for low concentration pesticide analyses, at least 1 L should be filtered through the SPE columns. If suspended solids prevent such high filtration volumes due to SPE tube clogging, then Q 1 8 filtration disks should be used over tubules. Immunoassay techniques are another faster and cheaper approach, if only identification and not quantification is required. 6.3.2 Metals th The results of the ICP/MS metal scans on the toxic 176 St: and South Cloverdale Ditch samples, their control samples, and non-toxic samples from the same sites, are shown in Table 22. The results of the ICP metal scans performed on the two controls biologically tested alongside each toxic sample, and ditch samples collected from Sites 8 and 5 on the sampling dates prior to or following those which produced toxicity (for which no toxicity was observed) are also shown in Table 22 for comparison. The hardness of these water samples was also included such that these metal concentrations could be compared to the water quality guidelines presented in Table 11. 132 Table 22. ICP Metal Scan Results for Toxic, Non-Toxic, and Control Samples. Metal Site 8 June 26 Site 8 July 16 Site 13 June 26 Cult/C. June 26 Site 5 July 16 Site 5 Aug. 7 Site 13 July 16 Cult/C. July 16 Ag (jug/L) < 0.05 0.15 < 0.05 < 0.05 0.05 0.05 0.05 < 0.05 Al (mg/L) 0.211 0.55 1 0.35 1 0.32 1 0.78 1 0.21 1 0.17 1 0.14 1 A s (ug/L) 4 2 1 2 3 1 1 1 Ba (ug/L) 6.95 20.8 71.8 11.20 18.05 11.6 13.3 12.8 Be (ug/L) < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 <0.5 < 0.5 Bi Cug/L) < 0.05 < 0.05 < 0.05 < 0.05 < 0.05 < 0.05 < 0.05 < 0.05 C a ^ g / L ) 21.9 52.9 14.2 16.7 26.4 24.1 16.95 10.65 Cd (Mg/L) < 0.1 0.5 < 0.1 < 0.1 0.3 < 0.1 < 0.1 < 0.1 Co (^g/L) 0.92 18.3 0.54 0.30 8,84 1.34 0.22 0.18 Cr ^ g / L ) 2.5 1 3.5 1 3.0 2.5 2.5 2.0 2.5 1 2.0 Cu (ug/L) 9.4 1 14.7 1 4.4 1 9.9 1 11.1 1 5.5 1 5.6 1 9.2 1 Fe (mg/L) 0.84 1 0.87 1 0.72 1 0.67 1 2.43 1 0.62 1 0.48 1 0.34 1 Hg (Mg/L) < 1 < 1 < 1 < 1 < 1 < 1 < 1 < 1 K (mg/L) 5.00 4.90 1.75 1.80 4.15 3.40 1.40 1.10 Mg(mg/L) 9.54 22.20 5.31 8.10 12.15 8.60 5.38 3.03 Mn (Mg/L) 41.0 352 23.1 88.1 204.0 75.8 27.6 60.0 Mo (ug/L) 2.0 1.3 0.6 2.1 1.2 1.7 0.5 1.0 Na (mg/L) 34.6 34.5 7.40 18.05 27.6 27.2 7.25 5.10 Ni (Mg/L) 13.0 63.4 2.0 1.6 31.4 7.8 1.2 1.0 P (mg/L) 0.1 OA < 0.1 < 0.1 0.1 < 0.1 < 0.1 < 0.1 Pb (ug/L) <2 <2 6 1 6 1 2 4 1 <2 2 Sb (Mg/L) 0.25 0.35 0.15 0.55 0.25 0.25 0.10 0.40 Se (ug/L) 3 1 3 1 < 1 < 1 1 1 < 1 < 1 Sn (ug/L) < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 < 0.5 Sr (Mg/L) 129.0 282.0 91.3 62.8 163.5 130.0 88.6 63.1 Ti (ug/L) 6 10 13 12 10 7 7 5 Tl (ug/L) <0.05 < 0.05 < 0.05 < 0.05 <0.05 < 0.05 < 0.05 < 0.05 u (Mg/L) 0.20 0.20 < 0.05 < 0.05 0.15 0.25 < 0.05 < 0.05 v (ug/L) 1 1 2 1 1 1 1 1 Zn (Mg/L) 11.5 80.5 1 6.0 20.0 52.5 1 17.0 10.5 12.0 Hardness 108 265 63 83 138 103 66 80 Note: June 26th culture/control was hardened to 80 mg/L as CaC03, July 16th culture/control water was unhardened. 1 Water Quality Guidelines (Table 11) exceeded. Hardness 60-120 mg/L as CaC0 3 . 133 The Canadian Water Quality guidelines for aluminum, chromium, copper, and iron were exceeded in almost all of the ditch, river, and control water samples tested. Zinc concentrations exceeded these guidelines in the non-toxic Site 8 sample and toxic Site 5 sample. However, could these metals have contributed to the observed lethal toxicity? The following are daphnid toxicity endpoints for the metal in question: Aluminum [Ceriodaphnia dubia 48-h LC50 = 300-500 /ugll (Shephard, 1983)]; chromium [Daphnia magna 96-h LC50 = 15.3 //g/L (Call et al., 1981); copper [Ceriodaphnia dubia 7-day LC50 = 49 - 67 //g/L (Anonymous, 1997)]; iron [Daphnia magna 21 -day LC50 = 5.9 mg/L (Biesinger and Christensen, 1972)]; and zinc [Ceriodaphnia dubia 7-day LC50 = 230 - 250 //g/L (Anonymous, 1997)]. Al and Fe may have stressed the C. dubia in the Site 5, South Cloverdale Ditch sample, since Al exceeded, and Fe were close to the above daphnid lethality endpoints. Stress may have increased their susceptibility to the OP insecticides present. If one observes the metal concentrations for all of the metals, and all of the samples in Table 22, the only metals found in higher concentrations in a lethally toxic sample than any of the non-toxic samples were Al and Fe, and this occurred only in the th th July 16 Site 5, South Cloverdale Ditch water. The small ditch parallel to 40 Ave., which connects to the South Cloverdale ditch was visibly contaminate with iron, in the form of ferric hydroxide. This author's findings are similar to those of EVS (1993), which reported that Nicomekl River and ditchwater samples always exceeded the Al water quality guidelines, and periodically exceeded the Cd, Cu, Fe, Pb, Mn, Hg, and Zn water quality guidelines. Swain and Holms (1988b) reported that aluminum concentrations are naturally high in the Nicomekl system. The concentrations of Cr, Cu, Fe, and Se may also naturally exceed 134 CCME (1986) and BCME (1989) water quality criteria in the Nicomekl system, since both upstream (Site 13) and White Rock pond concentrations of these metals exceeded these water quality criteria. The more drastic elevated concentrations of these metals in the runoff ditches, above those in the upstream river or White Rock pond samples, may in part be due to leaching from the farm soils following precipitation, and may in part be due to road runoff. Zinc concentrations were only observed to violate water quality guidelines in ditch samples, and not in the upstream or White Rock pond control samples. Again, the source of this Zn enrichment may be due to soil leaching or road runoff. 6.3.3 Total Ammonia Ammonia analyses were performed on entire sets of samples for four sampling dates, those which exhibited lethal toxicity, and some of those dates producing sublethal effects. These total ammonia measurements are shown in Table 23. The samples demonstrating lethal toxicity have been flagged (L) and the samples demonstrating sublethal toxicity where reproduction was significantly less than the Site 13 (river upstream control) have been marked (S). With respect to unionized ammonia, Nimmo et al. (1989) reported that C. dubia neonates had a 48-h LC50 of 1.06 mg/L NH3-N. Nimmo et al. also reported that C. dubia exhibited reduced reproduction for a 7-day exposure to 0.68 - 0.88 mg/L NH3-N (pH 8.0 @ 25 °C). The lethally toxic samples had total ammonia concentrations of only 0.25 mg/L (pH 7.0) and 0.15 mg/L (pH 7.1) as nitrogen. At an assumed temperature of 15 °C, unionized ammonia concentrations would have been only 0.27% and 0.34% total ammonia (0.0007 mg/L and 0.0005 mg/L NH3-N, respectively), both well below that causing lethal or sublethal effects to C. dubia. 135 Table 23. Total Ammonia (mg/L NH3-N) of All Water Samples, on Dates Exhibiting Toxicity in Select Samples. Sample June 26 July 16 Oct. 3 Nov. 21 Site 1 0.022 0.074 0.086 0.068 Site 2 0.017 0.162 0.117 0.051 Site 3 0.023 0.182 0.085 0.057 Site 4 0.009 0.106 (S) 0.113 0.056 Site 5 0.129 (S) 0.152 (L) 0.238 0.206 Site 6 0.069 0.048 Site 7 0.015 0.084 0.068 0.052 Site 8 0.248 (L) 0.126 0.273 0.115 (S) Site 10 0.007 0.051 0.172 0.050 Site 11 0.020 (S) 0.095 0.156 0.066 Site 12 10.820 (S) 0.118 0.280 0.048 Site 13 0.062 0.062 0.052 0.048 Culture Control 0.270 0.020 0.058 0.034 S = Sublethally toxic sample, Inhibited reproduction relative to upstream control. L = Lethally toxic sample th The June 26 , Site 12, Old Logging FJitch sample had a high •ammonia concentration, at 10.82 mg/L total ammonia as nitrogen. At this ditches' measured pH of 8.5, and an assumed temperature of 15°C, this equates to an unionized ammonia concentration of 7.97% total ammonia (0.86 mg/L NH3-N) (Thurston et al., 1979), a value acutely lethal to salmonids. This concentration is within the range for which Nimmo et al. (1989) showed chronic sublethal effect (0.68 - 0.88 mg/L NH3-N). Consequently, the sublethal C. dubia toxicity for this Old Logging Ditch sample was most likely due to ammonia contamination. 136 6.3.4 Dissolved Total /Inorganic/Organic Carbon, pH, Conductivity, Hardness, and Colour These ditch and water sample chemical measurements are shown in Appendix 11. The th Site 10 nearby-field-puddle collected October 16 had a pH of only 5.6, beyond the range of acceptable test pHs suggested by EC (1992). This is believed to be the cause of its observed sublethal toxicity. All other chemical measurements suggested conditions suitable for the suitable for the survival and reproduction of C. dubia. 6.4 Sensitivity of C. dubia to Detected Pest ic ides There existed the possibility that the C. dubia chronic bioassays may have not been sensitive enough to detect the presence of pesticides at concentrations found to cause lethality by other researchers, due to the fact the chronic testing necessitated the addition of food to test solutions (EC, 1992) and used plastic medicine cups. Bailey et al.'s (1997) C. dubia testing of diazinon and chlorpyrifos which produced the 96-h L C 5 0 endpoints of 0.32 - 0.35 //g/L and 0.055 //g/L, respectively, was performed without the addition of food to test solutions (U.S. EPA, 1993a) in glass micro-cuvettes. Since pesticides can adsorb to suspended solids and plastics, the food and medicine cups used could have scavenged pesticides from the ditchwater and riverwater test solutions, thus limiting their bioavailability and masking their presence. Acute (US EPA, 1993a) and chronic (EC, 1992) tests were performed on the GC/MS identified pesticides diazinon, chlorpyrifos, and prometryn, using the same plastic medicine cups used to test the field samples. The chlorpyrifos 48-h acute test was also fed to recreate the conditions of the river and ditch water testing. The lethality endpoints from these tests are shown in Table 24. 137 Table 24. C. dubia Lethality to Detected Pesticides. Pesticide 24-h LC50 (+/- 95% C.l.) 48-h LC50 (+/- 95% C.l.) 96-h LC50 (+/- 95% C.l.) 6-day LC50 (+/- 95% C.l.) Diazinon All results from 6-day test. > 0.8 Mg /L (highest cone.) Chronic Test Daily changeover Daily feeding 0.64 M g / L 2 (0.4, 0.8) Chronic Test Daily changeover Daily feeding 0.57 M g / L 1 (0.4, 0.8) Chronic Test Daily changeover Daily feeding 0.28 M g / L 1 (0.2, 0.4) Chronic Test Daily changeover Daily feeding Chlorpvrifos Results from 48-h test and 6-day test. 0.77 //g/L 3 (0.57, 1.03) 48-h Acute Test No changeover Fed at 0-h 0.50 M g / L 1 (0.33, 0.66) 48-h Acute Test No changeover Fed at 0-h > 0.132 Mg /L (highest cone.) Chronic Test Daily changeover Daily feeding > 0.132 Mg /L (highest cone.) Chronic Test Daily changeover Daily feeding Prometryn All results from 96-h test. 7.59 mg/L 3 (5.78, 9.92) Changeover at 48-h, fed at-2-h and 2-h prior to changeover 6.22 mg/L1 (2.5, 10.0) Changeover at 48 h,fed at-2-h and 2-h prior to changeover 4.58 mg/L1 (2.5, 10.0) Changeover at 48 h,fed at-2-h and 2-h prior to changeover Note: The chronic chlorpyrifos test did not produce lethality even at the highest concentration of 0.132 //g/L. In lieu of repeating the chronic test using higher concentrations an acute (48-h) test was performed using higher concentrations. Computer program used Binomiahriethod to.calculate the LC50. Computer program used Linear interpolation method to calculate the LC50. Computer program used Probit method to calculate the LC50. The sublethal inhibition of reproduction test results (NOEC, LOEC, IC25) for the diazinon and chlorpyrifos testing are shown in Table 25. 138 Table 25. Diazinon and Chlorpyrifos C. dubia Chronic Test (6-day) Reproduction Inhibition Test Endpoints. Insecticide NOEC LOEC IC25 (+/- S.D.) Diazinon 0.2 //g/L 0.1 //g/L 0.26 +/- 0.002 //g/L Chlorpyrifos 0.132 //g/L > 0.132 //g/L > 0.132 //g/L For chlorpyrifos, the highest concentrations tested, 0.132 //g/L, showed no effect on reproduction. Diazinon actually showed a statistically significant stimulation of reproduction at 0.1 //g/L (LOEC). The mean neonate production for 0.2, 0.1, 0.05, 0.025 //g/L and the upstream Site 13 control was 32.5, 35.2, 22.5, 22.2, and 22.2 neonates, respectively. The 0.2 //g/L diazinon concentration did not have statistically greater reproduction than the control using a = 0.05; however, its mean neonate production appears to have been enhanced as well. Hormesis, may have occurred. Further testing should be done to determine if diazinon produces hormesis. Likewise, the 6-day test on chlorpyrifos should be repeated at higher concentrations to better evaluate whether this pesticide causes reproduction inhibition. The author's chronic test 96-h LC50 test endpoint for diazinon (0.57 //g/L) was similar to that of Bailey et al.'s (1997) acute test (0.32 - 0.35 //g/L). Likewise, the test endpoints for prometryn were similar to that reported for Daphnia in U.S. EPA (1996). However, for chlorpyrifos, the author's chronic test failed to produce lethality after 6 days at a concentration (0.132 //g/L) more than double Bailey et al.'s acute test's 96-h LC50 (0.055 //g/L). The author's 48-h LC50 for chlorpyrifos (0.5 //g/L) from acute testing which included feeding was roughly seven-fold higher than Bailey et al.'s (1997) 48-h LC50 (0.058 - 0.079 //g/L) obtained using glass test vessels. Of the three pesticides tested, chlorpyrifos was 139 the least soluble and likely the most prone to adsorption to test feed and plastic test vessels. Katznelson and Mumley (1997) reported that chlorpyrifos has a greater tendency to adsorb to solid surfaces than diazinon. This likely accounts for the author's test's lower sensitivity to this insecticide. Wood (1997) also reported that the sensitivity of Daphnia pulex to chlorpyrifos, with testing conducted in plastic medicine cups, was lower than that reported in the literature for testing in glass. 6.5 Relationship Between Rainfall and Toxic Samples A record of precipitation throughout the study period was obtained from Environment Canada for a nearby gauging station in Cloverdale. Daily precipitation was plotted for each month of the 1997 growing season in Figure 2. There was no toxicity in samples collected on June 6 t h , following the significant rainfall of late May (as high as 25 mm/day). The two lethally toxic samples and four of the 11 sublethally toxic samples were collected within 1 week of the significant rainfall events of June (> 15 mm/day) and July (> 25 mm/day). Four of the 11 sublethally toxic samples were collected within one day of the th September 16 rainfall (35 mm). It is believed that the observed ammonia and pesticide contamination in June and July was a result of the runoff fromthe significant rainfall events which had occurred in the week prior to sampling. It is unlikely that the pesticide contamination was due to over-spray caused by insecticide re-application following the rainfall events, since the fields were puddled and muddy for some time following the rainfalls, and tractor-tank-boom type re-application this soon after the rainfalls would have been difficult, if not impossible. The soil and aquatic degradation times for chlorpyrifos and diazinon are long enough to support the hypothesis that the contamination was due to the crop/soil wash-off of inecticides applied prior to the rainfall events. 140 6.6 Dilution Calculations for Discharge from Contaminated Ditches The daily cumulative hours of pumping for all five of the major ditches in the study site, throughout the study period, are shown in Appendix 1. Discharge is known to have occurred from the Old Logging Ditch and South Cloverdale Ditch following the rainfall events which likely produced their observed ammonia and pesticide contamination, respectively. Unfortunately, there is no municipal record of whether or not the OP insecticide contaminated 176 t h St. ditch was discharged, since its flows are regulated privately. However, had this ditch been discharged while contaminated, its low discharge flows (observed on other dates at approximately 0.01 - 0.04 m /s) should have been sufficiently diluted to prevent adverse biological effects on the river's most sensitive organisms. Figures 12a and 12b show the number of hours the Old Logging Ditch and South Cloverdale Ditch were pumped into the Nicomekl River, prior to the sampling date on which each exhibited ammonia and OP insecticide contamination, respectively. Figures 12a and 12b provide evidence that the Nicomekl River may have received flows of toxic ditch water from the onset of rainfall to, and beyond, the dates of sample collection. The discharges shown in Figures 12a and 12b are only discharge via pumping. Gravity driven flows through these ditches' flood boxes may have also occurred. For the periods in 3 question, the Old Logging Ditch was pumped at 0.75 m/s, and the South Cloverdale Ditch 3 was pumped at 2.4 m Is. 141 0 E? r CT) Q c o C O) o o +3 —I a a o (sjnoLj) B u j d i u n d e B j e i p s i Q Q bl) Q 1. GO s E CO co CD co C M o CD CO CO IO CD CD CD tazmzazm a i azzag gga C N ^ajn^c -j -s o CN LO O LO O ( t u t u ) uoj;e;jdpajd 142 ( s j n o i | ) B u i d m r i d a B j e i p s i Q 0) G) CO o> £ C D O ^ CO >^ s 1 5 Q C CD 0 > _ o O o o ° CD > Q GO "HH 6 cd GO 1 i j . JZT ! 1 1 H m .... §i i i Q L | i Ql • x " P c ] i — 1 L o J S i k L 1 c 1 0 1 +-» _JSL • MB - .gnu,' QL • o p Or ml mtmmm C • i E B . , a i — • E n 1 1 : E i { I CO o CO CD CM CO CM CM co CM m CM -a-CM CO CM CM CM CM O CM Oi CO \~-CD m co CM .** CO o> CO CO CD m co CM o o CO o CN 13 (mm) uojieijdpajd =3 LL 143 Dilutions of the Old Logging Ditch discharge and South Cloverdale Ditch discharge were calculated to assess whether the Nicomekl's most sensitive aquatic life could have been adversely effected, had the ditches discharged waters contaminated to the degree observed in the samples used for the toxicity tests. Two dilution scenarios were used. The first was a worst case, minimum dilution, scenario, which assumed that the discharges were solely into the flow of the Nicomekl River recorded at 203 St. (Figure 3), upstream of the study site's 5 large drainage ditches. The second was a best case, maximum rd dilution, scenario, and assumed that the discharges were into the 203 St. river flow, plus possible additional river flow contributed by the upstream study site ditches. There are no rd other significant drainage ditches/water sources between the study site and the 203 St. monitoring station. The maximum attainable flow contribution to the river by the upstream ditches was determined from Appendix 1's municipal pumping records on a daily basis (i.e. based on which ditches discharged on each day, and at what rate for each discharge). These worst case and best case Nicomekl River flows at the Old Logging Ditch and the South Cloverdale Ditch, during their respective June and July discharges are shown in Figures 13a and 13b. For simplicity, and to hypothesize non-local downstream effects, all dilution calculations assumed complete mixing of ditch discharge with Nicomekl River water. At least with respect to the South Cloverdale Ditch, this assumption appears to reflect the actual river conditions within a short distance of the discharge, under low river flow conditions. Appendix 12 pictorially shows the rapid mixing of the South Cloverdale Ditch discharge with the Nicomekl River. 144 Figure 13a. Minimum and Estimated Maximum Nicomekl River Flows at the Old Logging Ditch, Between Rainfall and Observed Ammonia Contamination. 16 14 m 12 | 10 o 8 6 4 2 0 Nicomekl River Flow at Old Logging Ditch 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 June Min. Possible Flow Max. Possible Flow Figure 13b. Minimum and Estimated Maximum Nicomekl River Flows at South Cloverdale Ditch, Between Rainfall and Observed OP Contamination. Nicomekl River Flow at South Cloverdale Ditch Min. Possible Flow Max. Possible Flow 145 The maximum dilution scenario assumed that the daily discharges of all the upstream ditches, and the discharge of the ditch being diluted, occurred at the same time. However, this was very unlikely, since the upstream ditches, and ditches in question, often only discharged sporadically for a few cumulative hours per day. Consequently, while the actual river conditions likely fell somewhere between the best and worst case scenario, the worst case scenario was the most probable. The full-strength Old Logging Ditch waters had a high enough unionized ammonia concentration (0.86 mg/L NH3-N) to be acutely lethal to salmonids (Section 3.5.3). In the immediate vicinity of the Old Logging ditch, ammonia concentrations may have exceeded the Canadian acute water quality guidelines for ammonia during discharge ( 2 - 3 hours/day). However, downstream effect is of greater concern. These waters should have mixed fairly rapidly with the Nicomekl River given the discharge flow of 0.75 m3/s. The estimated fully-mixed total ammonia concentrations of the Nicomekl River during the discharge of the Old Logging Ditch following the late June rainfall event are shown under the minimum and maximum dilution scenarios in Figure 14. The pH of the ditch (8.75) was also diluted accordingly with the typical pH of the river (7.25), to arrive at minimum and maximum dilution scenario river pH's of approximately 7.3 and 7.8, respectively. For simplicity, the change in pH with dilution was assumed to be linear. Under each possible river pH scenario the U.S. EPA (1985a) acute (1-hour) Maximum Ammonia Concentration (MAC) for the Protection of Salmonids and Other Cold Water Fish Species (Table 10a) was plotted. The chronic total ammonia MAC (Table 10b) was also plotted. Its value does not vary with the two possible river pHs. The MACs plotted, were based on an assumed river temperature of 15 °C. 146 CD O CD CD ro _c o to CO CO f 3 I E < o co E I— .9 J £ z 2 2 Q. +3 • J C CO "~ LU 7 CN CD CO O O C O T J - C M O C O C O ^ C M O h/Buj) N-CHN o co co co X oo X Q. I— ro c < O « 5 o ro o x ro co ro co X ro -*—' o o cz o O C D -«—< =5 O < CO X Q_ ro O < "3 o < o < o 'cz o 1— JZ O CO c o LL > CD E o o CD CO h~ CZ CD O CD -*—' T— ro . J= CD II o f= O -z S co ro X c 2 ! i o) E E < ro ro c o O H E _ i E o) < E ro co ^ ro ro r-Q o S ro Q I I ro o E —' ^8 LU O E ro CD t-4—< co S jS x: -fc o to .-~ Q. T3 =3 to CD .c o E o CD TJ •I E o 2 to J -o -I E 2 Z 5 C 0 II ro E o c " ° TJ ® CD C0:9 5 c >< E CD -*—' CD CO to o CL E = 5 E co CO o CL ro .5= Q . .5 E x o o CO CO CO c c c E E E 3 U ZJ CO CO CO co to to < < < aTS"cr CD ZJ O) LL 147 Figure 14 provides evidence that under the probable minimum dilution scenario using lower river flows, the downstream fully-mixed Old Logging Ditch waters' total ammonia concentration should not have exceeded the maximum acceptable concentration (MAC) for acute exposure to cold water fish species and other sensitive life, but likely would have exceeded the MAC for chronic exposure. However, since the discharge from this ditch was only for 2-3 hours per day the Nicomekl River would have had slugs of fully mixed ammonia concentrations moving downsteam. Possible exposure time to slugs of ammonia contaminated water is unknown, but reflects both the length of the slug (length of continuous discharge), and the velocity of the river. However, given that daily cumulative discharge was only 2 -3 hours per day, the number of slugs or length of slugs was unlikely that significant. Although fully mixed ammonia concentrations may have exceeded chronic exposure water quality guidelines, it is unlikely that relatively stationary river organisms would have been exposed to a fully-mixed ammonia contaminated slugs for periods which would have produced adverse biological effect. On the south side of the Nicomekl In the immediate vicinity of the Old Logging Ditch, ammonia concentrations may have exceeded acute water quality guidelines before slugs of discharge were completely mixed vertically or horizontally in river space. However, based on the findings of Jones (1948) and Summerfelt and Lewis (1967) fish likely would have avoided this local region of high ammonia concentrations. Table 23 reported the total ammonia concentration roughly 3 km downstream of the Old Logging Ditch at Site 1, 40 t h Ave. on June 26 t h as only 0.022 mg/L NH3-N. The Old Logging Ditch was reported to have discharged for 2-3 hours on the previous days. The 148 Site 1 collection point was approximately 3 km downstream of the Old Logging Ditch. The 3 Nicomekl's flow was roughly 1 m /s during late June. Using an assumed river width of 25 m and assumed depth of 3 m, the river would have been flowing at approximately 0.01 th m/s, and would have taken approximately 3.5 days to reach this point. The low 40 Ave ammonia measurement likely represents a segment of river water which passed by the Old rd Logging Ditch on June 23 while the ditch was not discharging. Similar dilution calculations were performed for the discharge of the South Cloverdale Ditch for the days prior to its detected OP insecticide contamination. OP concentrations in the South Cloverdale Ditch should not have been high enough to adversely affect fish living in this ditch. Diazinon was chemically identified in this ditch water at only 0.02 //g/L, however concentrations are believed to have been higher, and other OP insecticides may have also been present. Consequently, without accurate ditch OP concentration(s), determination of downstream river insecticide concentrations based on the 0.02 //g/L of diazinon detected would have been futile. Instead, the estimated fully-mixed peak percentage of South Cloverdale Ditch water in the Nicomekl River during pumping, was calculated for the days between the rainfall event and the sampling event, again using the minimum and maximum dilution scenarios. The estimated peak daily percentages of ditch th water in the river (Figure 15) were compared with the June 26 South Cloverdale Ditch water sample's LC50s, and the literature review on toxicity of low concentrations of OP insecticides to invertebrates, to determine the possible biological effects of these South Cloverdale Ditch discharges. 149 CD CD 3 —D C CD > a: CD E o o CD o o m CO " D i_ CD > O O _c -*—< Z J O C O o co c o TO c CD O c o O >^  CD Q CD CD Q_ " D £ CD E —• co LU E CD CD i_ -*—' CO s 1 C J C O .ti C L T J =5 CO CD E o C D T3 £ E o 2 o £ M— D co E 5 CD E o c = 2 © CD CJ>:9 ] Q X E Q. E o o co CO o Q. E Z J E x CD E CO CO O C L E rs E c I C O C O O ) c c c E 13 E E CO CO CO CO CO CO < < < CD Xi O LO CD Z J 150 The dilution calculations suggest that following complete mixing (believed to have occurred in a short distance downstream) the Nicomekl River likely consisted of 10 % to th th 80 % South Cloverdale Ditch water, between July 5 and July 16 for those hours of discharge identified in Figure 12b. The maximum number of hours the South Cloverdale Ditch was pumped per day, following the July rainfall, was 13 hours on July 8 t h . However, 3 th th the high river flows (2.5 to 15 m Is) from July 5 to 11 significantly diluted the South Cloverdale Ditch water pumped into the Nicomekl over these dates. Good river flow during this period likely would have restricted lengths of downstream exposure to fully mixed slugs. The worst case minimum dilution scenario for July 8 t h estimates that there should have been no more than 20% ditch water in the Nicomekl River due to this discharge. The 48-h, 96-h, and 6-day LC50's of the toxic South Cloverdale Ditch sample were 92.9 %, 52.2 %, and 36.5 %. Therefore, downstream ditchwater concentrations in the river should not have been high enough, nor long enough, to kill river organisms as sensitive as C. dubia following the most intensive period of South Cloverdale Ditch pumping. Of course, this assumes that earlier discharges has the same concentrations th of OP insecticide(s) that were in the July 16 sample. th th The situation is more questionable for the later discharge flows of July 12 to 16 . Although the pumping on these dates were extremely short in cumulative duration (1-2 hours/day), the estimated percentages of ditch water in the Nicomekl for these dates were likely much higher (perhaps as high as 80%), since river flow at these time was very low 3 (< 1 m /s). Further lower river flows would have increased exposure times to full-mixed slugs. However, again its is unlikely that the length of exposure to these concentrations would have been long enough to kill river invertebrates as sensitive as C. dubia. 151 There is the possibility that all the mixed South Cloverdale Ditch water may not have kept moving down the system, increasing exposure durations. Pockets of South Cloverdale Ditch water may have remained for longer periods in slow flowing portions of the river. Lethality of sensitive invertebrates could have occurred in more stagnant river zones with exposure to the mixture of South Cloverdale Ditch water and river water for greater than 24 hours. On July 16 t h the Site 4 sample, collected immediately downstream of the South Cloverdale Ditch discharge point appears to have been contaminated with OP insecticides (sublethal toxicity in the 7-day C. dubia test; 40% mortality, and 30% of surviving test organisms exhibited paralysis symptoms characteristic o f pesticide contamination). Discharge o f the South Cloverdale Ditch was not occurring at the time the Site 4 sample was collected. The samples collected 2 km upstream (Site 7) and 4 km downstream (Site 1) from Site 4 on this date produced no adverse effects to C. dubia. The river at Site 4 may have had remnants of an earlier discharge. However, since effect was not observed in the Site 4 sample until the 6 t h day of its bioassay, the effect of remnant discharge to organisms as sensitive as C. dubia should not have occurred at this site, since it is unlikely that this pocket of the river would have remained stagnant for the next 6 days. Since this sample was collected so close (5 m) to the South Cloverdale Ditch discharge point, the observed river contamination at this time may more likely have been caused due to leakage from the pump station's flood boxes at the time of sampling, and not the previous discharge events. Was there likely a sublethal impact on benthic invertebrates as a result of the South Cloverdale Ditch discharge? This is a difficult question to answer. Morgan (1976), Arthur 152 et al. (1983). Ward et al. (1995) and Eaton et al. (1985) all provided evidence of benthic invertebrate drift following acute exposure to OP insecticides > 3 ^ g/L. It is unlikely that the duration and OP concentration of the pulsed discharges from the South Cloverdale Ditch produced drift effects. OP concentrations and exposures were not liekly high enough. Howver, even if drift of the more sensitive benthic organisms did occur [i.e., amphipods, mayflies, caddisflies, damselflies, Arthur et al. (1983)], recolonization should have been fairly rapid (Morgan, 1976) and there should not have been a significant overall impact on fishes' benthic food supply. 6.7 Hyallela azteca Sediment Toxicity Tests and Sediment Chemistry Large (5 -10 cm) freshwater mussels (believed to be Anodonta kennerlyi) were observed in the Nicomekl river at the extreme upstream Site 13 at the time of sediment collection, suggesting this site was relatively uncontaminated. Table 26 provides a brief description of each of the collected sediment samples prior to initiating their toxicity tests. Table 26. Sediment Samples Appearance, Odour, and Visible Indigenous Invertebrates, Prior to Toxicity Testing. Site Sediment Description 1 Silty, No odour, Large Chironomids 2 Sandy with Rocks, No odour, Twigs and Grass, No Visible Invertebrates 5 Clay, No Odour, No Visible Invertebrates 7 Sandy with Clay Clumps, No Odour, Twigs and Grass, No Visible Invertebrates 8 Muddy with Clay Clumps, Slight Oily and H 2S Odour, No Visible Invertebrates 10 Sandy with Rocks, Very Slight H 2S Odour, Twigs and Grass, No Visible Invertebrates 11 Silty and Clayey, Slight H 2S Odour, Twigs and Grass, No Visible Invertebrates 12 Silty and Clayey, Oily Odour, Twigs, No Visible Invertebrates 13 Sandy with Rocks, No Odour, Leaves and Woody Debris, No Visible Invertebrates 153 Shortly into the sediment toxicity tests, all of the ditch and river samples (except Site 5) were teaming with Oligochaetes, believed to be Tubifex sp. Chironomids, archiannelid worms, small snails, isopods (Physellus sp.), and ostracods were observed in the sediments (except Site 5) during testing, but in much lower abundance than the oligochaetes, and in no distinguishable pattern between samples. The percent survival and mean dry weight of the 10 Hyallela in each of the 4 replicates of sediment tested from each site are shown in Table 27. Table 27. Sediment Bioassays % Survival and Growth (Test-End Mean Hyallela Dry Weight). Sample A % Survival A Mean Dry Weight M9 B % Survival B Mean Dry Weight M9 C % Survival C Mean Dry Weight D % Survival D Mean Dry Weight ^g Site 1 100 285 90 293 60 285 80 195 Site 2 50 178 90 201 90 174 0 Site 5 100 250 100 294 90 316 100 264 Site 7 100 365 100 279 90 220 100 253 Site 8 100 248 90 329 90 321 90 206 Site 10 90 218 90 216 100 252 90 201 Site 11 100 178 100 162 90 227 70 306 Site 12 100 129 90 127 80 134 50 157 Site 13 100 241 70 257 80 225 50 138 Silica Sand 0 0 0 0 154 There was complete lethality in the silica sand control, the cause of which is unknown but may have been due to inadequate prior washing. The upstream control sediment did not meet EC's (1996) valid control requirement of 80% mean survival. Nevertheless, since its mean survival was 75%, and it met the growth requirement of a mean Hyallela weight of 100 //g by the end of the test, Table 28 presents the t-test results comparing the Hyallela survival and growth (test-end mean dry weights) in the ditch and river sediment samples from the study area with the sediment from the upstream river control location, Site 13. Table 28 includes the results of the analyses for sediment organic matter (%). Table 28. T-test Results for Hyallela azteca 14 Day Survival and Growth in Ditch and River Sediment Samples Versus the Upstream Sediment Sample, Plus Sediments' Organic Matter (%). Sediment Sample Location Mean % Survival +/-S.D. Mean Weight fog) +/-S.D. Survival, Versus Site 13 (P=, 1 tail) Weight, Versus Site 13 (P=, 1 tail) Organic Matter (%) Site 1 83 +/- 17 265 +/- 46 0.299 0.106 7.1 Site 2 58 +/- 43 184 +/- 15 0.251 0.165 7.9 Site 5 98 +/- 5 281 +/- 30 0.063 0.042 * 3.0 Site 7 98 +/- 5 279 +/- 62 0.063 0.084 8.0 Site 8 93 +/- 5 276 +/- 59 0.100 0.089 18.5 Site 10 93 +/- 5 222 +/- 22 0.100 0.416 4.1 Site 11 90 +/- 14 218+/-65 0.143 0.473 15.7 Site 12 80 +/- 22 137 +/- 14 0.375 0.032 * 6.9 Site 13 75 +/- 21 215+/-53 * P-value < 0.05 indicates that there is a statistically significant difference. 155 Based on the t-test results, Hyallela azteca survival was not significantly reduced in any of the sediments from the study site than in the upstream control sediment. Growth appears to have been enhanced in the Site 5 sample and inhibited in the Site 12 sample. The growth enhancement of the Site 5 sample was likely due to the fact that this sample was clay, and had no indigenous life present during testing. These two factors gave the Site 5 test sediment's overlying waters the highest hardness, and lowest total ammonia concentration of all the sediment samples during testing (Appendix 13), which may have improved the growth Hyallela. Further, since no indigenous life was present in the Site 5 sample, this sample's Hyallela did not have to compete with other organisms for food. The growth inhibition of the Site 12 sediment may be due to PAH contamination as its oily odour suggests. There does not appear to have been a strong relationship between survival/growth and the sediments' organic matter (%), as predicted in the literature review. Survival in the Site 2 sample may have been lower than that of the upstream site, although the large standard deviation for this site's survival prevented this from showing statistically. It is unfortunate that all the organisms in the silica sand control died, as Nicomekl system sediments could not be compared to non-system sediments. However, considering that all of the sediment samples from the Nicomekl system produced Hyallela azteca weights in excess of the valid control requirement of 100 /j.g, the sediments appear to be relatively uncontaminated. Table 29 shows the total metal content (mg/kg) of the sediment samples. 156 Table 29. Sediment Sample's Measured Dry Weight Metal Content (mg/kg). Metal Site 1 Site 2 Site 5 Site 7 Site 8 Site 10 Site 11 Site 12 Site 13 Cd 0.11 0.03 0 0.07 0.34 0 0.03 0.01 0 Cr 16.68 17.63 16.17 16.16 18.81 10.81 16.40 16.28 9.24 Cu 10.15 17.91 20.27 14.93 23.62 7.63 17.01 17.70 3.45 Pb 5.72 18.72 6.67 14.15 16.31 3.28 9.76 7.55 2.88 Ni 20.34 29.69 32.28 27.14 40.71 13.73 26.40 27.25 11.16 Zn 46.53 74.53 51.86 83.41 83.95 36.75 53.16 53.23 32.18 Mo 0 0 0 0 0 0 0 0 0 Co 6.58 12.05 8.97 10.11 11.61 4.40 7.36 8.99 3.67 Ca 10.61 7.08 27.95 3.54 10.26 15.92 56.96 14.86 1.42 Mn 234.8 529.1 238.4 367.2 226.9 194.8 255.4 377.8 155.2 Mg 4823 6707 8268 5230 5809 3812 6795 6601 2999 Al 12019 14130 13624 13376 14243 8178 12707 13432 7704 Fe 18399 20203 19441 16882 19255 10960 16443 20669 7949 The author's measured total metal concentrations for the NRC supplied MESS-2 reference sediment, along with the NRC's reported metal concentrations for thissediment, are shown in Table 30. 157 Table 30. MESS-2 Reference Sediment's Dry Weight Metal Content (mg/kg), Author's Results Compared to NRC's. Metal NRC's (mg/kg +/- 95% C.l.) Author's (mg/kg +/- S.D.) Error Cd 0.24 +/- 0.01 0 +/- 0.23 -4 % Cr 106 +/-8.0 14.34 +/-0.15 Cu 39.3 +/- 2.0 20.02 +/-0.18 - 49 % Pb 21.9+/- 1.2 14.95 +/- 0.26 - 32 % Ni 49.3+/- 1.8 33.79 +/- 0.32 - 31 % Zn 172 +/- 16 101.96 +/-0.02 - 41 % Mo 2.85 +/-0.12 0 +/- 0.66 Co 13.8+/- 1.4 8.61 +/-0.12 - 38 % Mn 365 +/- 21 242.84 +/- 1.91 - 33 % In general, the author's measured dry weight sediment total metal concentrations were lower than those reported by the NRC (NRC, Date unknown) for the MESS-2 sample. The author's technique was unable to detect molybdenum in the MESS-2 sample, explaining the zero molybdenum values for the Nicomekl sediments. BCME (1997) and Swain and Holms (1988b) also reported no or limited detection of Mo in Nicomekl sediments, respectively. The author's technique grossly under-quantified chromium for the MESS-2 sample. Consequently, the chromium concentrations in the Nicomekl sediments may have been much higher than that shown. BCME (1997) and Swain and Holms (1988b) reported higher concentrations of Cr in their testing of Nicomekl River sediments. The discrepancies between the authors results and NRC's results for the MESS-2 sediment are likely a result of differences in the sediment digestion/metal extraction procedures used. The hypothetical true metal concentrations of the Nicomekl sediments, correcting for the degrees of under-quantization for each metal listed in Table 30, are shown in Table 31. 158 Table 31. Sediment Sample's Hypothetical True Dry Weight Metal Content (mg/kg). Metal Site 1 Site 2 Site 5 Site 7 Site 8 Site 10 Site 11 Site 12 Site 13 Cd 0.11 0.03 0 0.07 0.35 0 0.03 0.01 0 Cu 15.2 26.7 30.2 22.2 35.2 11.4 25.3 26.4 5.1 Pb 7.6 24.7 8.8 18.7 21.5 4.3 12.9 10.0 3.8 Ni 26.6 38.9 42.3 35.6 53.3 18.0 34.6 35.7 14.6 Zn 65.6 105.1 73.1 117.6 118.4 51.8 75.0 75.1 45.4 Co 9.1 16.6 12.4 13.9 16.0 6.1 .10.2 12.4 5.1 Mn 312.3 703.7 317.1 488.4 301.8 259.1 339.7 502.5 206.4 For comparison with the Nicomekl sediment metal concentrations, the natural/background sediment metal concentrations for the Fraser River Estuary are shown in Table 32. Table 32 also shows the Canadian Sediment Quality Guidelines for trace metal threshold effect level (TEL) (Smith et al., 1995), and sediment metal concentrations of the Nicomekl system reported by EVS (1993), BCME (1997), and Swain and Holms (1988b). 159 Table 32. Fraser River Sediment Background Trace Metal Concentrations, Canadian Sediment Quality Guidelines Trace Metal TEL, and Other Studies Nicomekl Sediment Metal Concentrations. Metal Cdn. Crit. (TEL) 1 (mg/kg) Natural Background Fraser River 1 (mg/kg) Nic. 184th St. (1990) 2 (mg/kg) Nic. Below Old Logging Ditch (1990) 2 (mg/kg) Old Logging Ditch (1990) 2 (mg/kg) Burrows Ditch (1990) 2 (mg/kg) Cd 0.6 < 0.25 - 0.50 NM NM NM NM Cr 37 2 0 - 4 5 NM NM NM NM Cu 36 1 8 - 4 0 2 0 - 4 3 2 5 - 3 8 4 5 - 5 8 4 2 - 5 3 Pb 35 3 - 1 1 3 0 - 6 0 1 5 - 5 8 1 8 - 7 0 1 9 - 8 7 Ni 18 7 - 5 4 NM NM NM NM Zn 123 4 5 - 9 5 40 - 300 80 - 300 2 0 0 - 1100 190 - 1040 Mn NM NM 450 - 650 525 - 625 570 - 690 530 - 570 Fe NM NM 230 - 475 350 - 425 340 - 380 275 - 370 Metal Nic. Mouth ( 1 9 8 9 - 1 9 9 3 ) approx. 5 sample dates (mg/kg) 3 N i c , 6 locations (1979) 4 (mg/kg) C d 0.28 - 0.38 < 1.0 Cr 23 .2 -61 .8 21 - 360 Cu 6 .25-29 .10 8 - 2 2 Pb 1.71 - 16.5 1 6 - 7 1 Ni 15 .0 -55 .7 1 6 - 2 8 Zn 3 1 . 7 - 1 0 5 41 - 5 7 Mo 0 < 1 - 2.0 C o 14 .6 -18 .3 1 4 - 2 2 C a No Total Done 246 - 424 Mn 172 - 408 209 - 438 Mg 5300 - 7500 4640 - 7220 Al 16900-21300 7 5 8 0 - 10900 Fe 10200-33400 14000-26600 All metal measurements are total mg/kg dry weight. 1 (Smith et al., 1995); 2 (EVS, 1993); 3 (BCME, 1997); 4 (Swain and Holms, 1988b) NM = Not Measured 160 This author's metal results are similar to those reported by these authors. For each sediment tested, all of the measured metals were below the Canadian Sediment Quality Guidelines TEL, except for Ni. Ni exceeded the TEL criteria for all the ditch and river sediments except Site 13 (upstream river) and Site 10 (Burrows Ditch). Nicomekl sediments had metal concentrations close to the natural background level of deep, historic sediments collected from the Fraser River (FREMP, 1996) for Cd, Cr, Cu, Zn, and Ni. Lead concentrations in 3 of the Nicomekl sediments did not exceed TEL guidelines; however, they were higher than the natural background concentrations for the Fraser River. Lead concentrations appear to be lower than those reported in the earlier studies. th Of all the sediments tested, the Site 8, 176 St. ditch sediments appears the most contaminated with respect to trace metals. The Site 13, upstream river sediment had much lower metal concentrations than any of the sediments from within the study site. 6.8 Summary This study determined whether the ditches draining the agricultural lands alongside the Nicomekl River, Surrey, B.C. are discharging waters toxic to the sensitive freshwater test organism Ceriodaphnia dubia. Environment Canada's (1992) chronic (7 ± 1 days) test methods for measuring inhibition of C. dubia reproduction and survival were performed on a total of 50 water samples collected from 6 runoff ditches and 35 receiving water samples from 5 river locations, within an area of extensive vegetable and blueberry farm land use. Sampling was performed at regular three week intervals, from May to November of the 1997 growing season. Ditch and river sediments were also tested once in the fall for toxicity using Environment Canada's (1996) 14-day chronic test for inhibition of the amphipod Hyallela azteca's survival and growth. 161 Of the 85 ditch and riverwater samples tested only two ditch samples, one collected from th th a 176 St. ditch on June 26 , and one collected from the South Cloverdale Ditch on July 16 t h, proved to be lethally toxic with 48-h, 96-h and 6-day LC50's of > 100%, 74.8%, and 39.9% for the 176 t h St. sample and 92.9%, 52.2%, and 36.5% for the South Cloverdale Ditch sample, respectively. A proven (Bailey et al., 1997; Ankley et al., 1991) biological toxicity identification evaluation using piperonyl butoxide, showed that the toxicity of these samples was likely caused by metabolically active organophosphorous (OP) insecticides. Malathion, diazinon, chlorpyrifos, methyl parathion, dimethoate and/or azinphos methyl are metabolically active OP insecticides used in the Lower Mainland (in decreasing order of the abundance used in 1991). One of these, or a combination of these (Bailey et al., 1997) was likely present in the two lethally toxic samples. Gas chromatography/mass spectroscopy revealed the presence of diazinon, chlorpyrifos, and prometryn in the 176 th St. ditch sample at approximately 0.02 //g/L, 0.07, and 3 //g/L, respectively, and the presence of diazinon in the South Cloverdale Ditch sample at approximately 0.03 //g/L. There is evidence that the chemical testing under quantified the amount of these two OP insecticides present in the original samples used for the toxicity tests. Under quantification due to degradation or adsorption to storage vessels was likely, since the select-ion-scan work necessary to detect these compounds was performed following full-ion-scan testing, after several months of frozen storage in plastic bottles. The PBO analyses performed on the same frozen samples used for the select-ion-scan work showed reduced toxicity over the original fresh samples, supporting this belief. Other OP insecticides which were not screened for using select-ion-scan GC/MS may also have also been present. 162 Ammonia concentrations in the lethally toxic samples were well below C. dubia effect levels and water quality guidelines. Metal concentrations in the lethally toxic samples were below Canadian and Provincial water quality criteria, with the exception of Al, Cr, Cu, and Zn; however, the concentrations of Cr, Cu, and Zn were lower in the toxic samples than that measured in non-toxic ditch waters. The lethally toxic South Cloverdale Ditch sample did have high Al and Fe concentrations, approaching daphnids chronic effect levels, at 0.785 mg/L and 2.43 mg/L, respectively. These metals may have stressed the C. dubia test organisms, making them more sensitive to the OP insecticide contamination. The crops grown in the vicinity of the insecticide contaminated ditches were largely corn and potatoes, which had recommended OP insecticide applications both as a pre-planting seed treatment and spray when insects appear. The observed ditch insecticide contamination coincided with the two large rainfall events of June and July. Sampling following May, August, September, October, and November rainfall episodes showed no lethal toxicity. It appears as though the incidents of OP contamination were not from insecticides used for corn seed treatment or potato-piece treatment, since toxicity was not observed following the late May rainfall event but arose well into the growing season, once crops were established. Toxicity was likely from the wash-off from crops and land run-off of previous aerially applied insecticides. It is believed that toxicity was not a result of over-spray during insecticide re-application following the rainfall events, since the grounds were generally puddled and muddy for some time following the significant June and July rainfall events, and it is unlikely that farmers would have driven their tractors into the fields given these conditions. 163 The suggested water quality criteria for the maximum acute/chronic concentrations of diazinon and chlorpyrifos (0.08/0.04) in aquatic environments were likely exceeded in the lethally toxic ditches. However, the ditches' low concentrations of insecticide contamination should not have had an adverse biological effect on any fish rearing or seeking refuge in these two drainage ditches. Concentrations of less than 1 //g/L of OP insecticide are capable of producing the observed C. dubia LC50s. Only the South Cloverdale Ditch was shown to have discharged OP contaminated waters into the Nicomekl River. Fully-mixed dilution calculations of the South Cloverdale Ditch discharge for the period of time between the rainfall event, which is believed to have initiated the toxic conditions, and the date of sample collection, showed that the Nicomekl River would have consisted of 20% to 80% South Cloverdale Ditch downstream of the discharge flow. Localized percentages of ditch water in the immediate zone of discharge may have been higher. OP insecticide(s) concentrations in the river likely exceeded the acute water quality criteria for diazinon. It is difficult to surmise what effect the South Cloverdale ditch discharge would have had on the invertebrate community in the Nicomekl River downstream of the South Cloverdale Ditch. It is unlikely that there would have been mortality of invertebrates as sensitive as C. dubia within the river. River invertebrates exposure time to South Cloverdale Ditch water was not likely long enough to produce the toxic effects observed in the bioassays. The maximum cumulative daily discharge of South Cloverdale Ditch water during the period in question was only for 13 hours. OP insecticide concentrations and exposure durations in the river were not likely high enough to cause drift of the more sensitive benthic invertebrates. 164 Inhibition of C. dubia reproduction relative to the upstream river control water was observed in 10 of the 83 non-lethally toxic ditchwater and riverwater samples. One of the th sublethally toxic ditch samples, collected from The Old Logging Ditch on June 26 , had a total ammonia concentration of 10.8 mg/L NH3-N. This equated to an unionized ammonia concentration of approximately 0.86 mg/L NH3-N, based on the ditchwater's measured pH (8.5) and estimated temperature (15 °C). This unionized ammonia concentration was within the range known to cause chronic reproductive effects to C. dubia (C. dubia 7-day chronic value = 0.68 - 0.88 mg/L NH3-N) and above that reported to cause lethality to salmonids [rainbow trout, 1-2 g fish 96-h LC50 approximately 0.5 - 0.7 mg/L N H 3 . N , 10 - 20 g fish 96-h LC50 approximately 0.2 - 0.3 mg/L NH 3 (Thurston and Russo, 1983)]. Consequently, if any salmonids were taking refuge in the Old Logging th Ditch around June 26 they likely would have been killed. Discharge of the Old Logging Ditch was shown to have occurred around the date of this ditches ammonia toxicity. Assuming that this ditch contamination was produced as a result of the previous rainfall, high ammonia concentrations may have been discharged into the Nicomekl River for 1 to 2 hours per day in the week prior to sampling. Dilution calculations of these discharges estimated that the fully-mixed ammonia concentrations in the Nicomekl River should not have exceeded the U.S. EPA's acute (1-hour) maximum acceptable concentration (MAC) of total ammonia for the protection of salmonids and other cold water fish species. Chronic (4-day) MAC'S may have been exceeded; however it is unlikely that fish exposure to the slugs of fully-mixed waters could have been 4-days, given that the river maintained a low flow during this period. All other ditchwater and riverwater sample ammonia measurements performed during the study were well below both water quality guidelines. 165 One sublethally toxic sample was collected on July 16 t h at Site 4 in the Nicomekl River, immediately downstream of the discharge point for the South Cloverdale Ditch (lethally toxic on this date, but not discharging at the time of collection). The inhibited reproduction in this sample was largely due to its 40% test organism mortality and surviving organism paralysis, for which insecticide(s) from the South Cloverdale Ditch are the suspected contaminant. Prepared solutions of diazinon in upstream river control water produced a chronic test (6 day) 48-h LC50, 96-h LC50, and 6-day LC50 of 0.64, 0.57, and 0.28 //g/L, respectively. Prepared solutions of chlorpyrifos in upstream river control water yielded an acute test (48-h) 24-h LC50 and 48-h LC50 of 0.77 and 0.50 //g/L, respectively. Prepared solutions of prometryn in upstream river control water yielded an acute test (96-h) 48-h LC50 and 96-h LC50 of 6.22 mg/L and 4.58 mg/L, respectively. Testing in plastic medicine cups with feeding did not appear to substantially mask the toxicity of diazinon or prometryn to that found in the literature, but did reduce the toxicity of chlorpyrifos approximately 5-fold to that reported by Bailey et. al (1997) for testing in glass vessels without the addition of test food. This study found that the C. dubia 7-day reproduction inhibition endpoint is not a good indicator of sublethal concentrations of organophosphate insecticides. Neither testing using dilutions of the two toxic ditchwater samples, or lab made solutions of diazinon and chlorpyrifos in upstream control water, showed noticeably reduced reproduction, even at organophosphate concentrations where organisms eventually died. IC25 endpoints for reproduction appear to be largely, if not entirely, due to lethality. Hence, the sublethal endpoint in the chronic test is likely not useful in determining whether receiving water samples are contaminated with diazinon and chlorpyrifos (possibly other OP insecticides) 166 at concentrations above water quality guidelines but below that causing lethality to C. dubia. The chronic diazinon test actually showed reproduction stimulation at sublethal concentrations. This could have been due to coincidence, hormesis, or the elimination of adverse bacteria or fungus from used river water by sublethal diazinon concentrations. Stimulation has not been reported in the literature by other researchers. More testing is required to see if this phenomenon repeats itself. The author recommends that if C. dubia testing is being used only to identify the presence or absence of OP insecticides that the neonate production of the chronic test should be forgone. The chronic test does appear to be more useful than an acute test as mortality was observed to take greater than 96 hours to manifest given low OP concentrations. Sediment toxicity testing using Hyallela azteca failed to show a significant difference in survival between 5 runoff ditch and 3 river bottom sediments versus sediment from the upstream control site. Growth was significantly reduced in the Old Logging Ditch sediment sample, with test-end mean +/- S.D. Hyallela weights of 137 ± 14 //g in the ditch sediment versus 215 ± 53 fj.g in the upstream control sediment. All of the sediments had metal concentrations which met Canadian Sediment Quality threshold effect level (TEL) guidelines except for Ni. Ni exceeded the TEL in all of the sediments tested; however, the Ni concentrations were within the naturally occurring levels reported for prehistoric Fraser River Sediments. The Old Logging Ditch exhibited oil contamination both visibly and by odour. The sublethal toxicity at this site may be due to PAH contamination of this ditch's sediments. The sediments of the Nicomekl River and its ditches do not appear to have accumulated pesticides over the 1997 growing season to concentrations toxic to sensitive invertebrate species such as Hyallela azteca. 167 GENERAL CONCLUSIONS The majority of Nicomekl River and drainage ditch samples, collected every three weeks for six months during the 1997 growing season, were non-toxic (lethally or sublethally) to the sensitive freshwater test organism Ceriodaphnia dubia. Sediments collected in October, 1997 from the Nicomekl River and its drainage ditches were generally non-toxic (lethally or sublethally) to the sensitive amphipod Hyallela azteca. nd The commercial vegetable and blueberry farms on the lands between 152 St. and th 184 St. in Surrey do not appear to be creating a significant toxicity problem in the drainage ditches of this region, or in the Nicomekl River itself. Organophosphorous insecticide contamination of the drainage ditches to levels lethal to C. dubia was observed on two occasions (2 of 50 ditchwater samples). The discharge of one of these contaminated ditches was shown to have occurred in the days between the previous rainfall event and the observed toxicity. Ammonia contamination of one of the study site ditches was observed only on one occasion, at ammonia concentrations chronically sublethally toxic to C. dubia, and likely acutely lethally toxic to salmonids. Municipally pumped discharge flows from the drainage ditches occur during the summer. There can be minimal dilution of these flows in the river, given the high flow rates of pumped ditch water, relative to the low summer baseline river flows. 168 There should have been no significant effects on the rivers most sensitive biota due to the discharge of OP contaminated or ammonia contaminated ditch water during the study period. It is concluded that for the study site investigated, there is no need to treat drainage ditch waters, or to restrict their discharges to periods of high river flow. While this study surmised that the 1997 agricultural activities did not adversely impact the Nicomekl's aquatic biota, future growing seasons could result in greater pesticide use. This, combined with more frequent or more severe rainfall events, could produce effects of consequence. Diazinon and chlorpyrifos (possibly other OP insecticides as well) do not appear do significantly inhibit C. dubia reproduction in the chronic test, at concentrations below that producing mortality. However, the chronic C. dubia test, and toxicity identification evaluation using piperonyl butoxide, allow for identification of the presence of metabolically active OPs at concentrations non detectable by conventional gas chromatography analyses following solid phase extraction, using the low filtration volumes these tubes are limited to (due to the presence of suspended solids in ditchwater and river water samples). There is the need to develop and implement Canadian water quality criteria for specific OP insecticides (eg diazinon, chlorpyrifos, malathion), given that these compounds are highly toxic and are finding their way into receiving waters. 169 SUGGESTED FUTURE STUDIES In situ C. dubia and fish bioassays should be positioned in the Nicomekl River and its drainage ditches during the growing season to further identify the sporadic events of pesticide and ammonia contamination. In situ collection of pesticides from the river, downstream of drainage ditches, onto columns filled with CIS-bonded resins should be performed over 1 week periods during the growing season to integrate the pesticide contamination over time and identify contamination events which may escape a conventional sampling schedule. While it was possible to calculate the dilution of pumped discharges into the Nicomekl River; the drainage ditches are also known to discharge via gravity feed when ditch water-levels are above that of river water-levels. More information is required on the discharge via the flood boxes (IE - which ditches allow flow in this manner in the summer, and what are the flow rates and consequent flow dilutions). The degree to which all the drainage ditches are used, or could be used, as fish habitats should be determined. The Nicomekl's benthic Invertebrate abundance and species diversity upstream, within, and downstream of the study area should be measured before, during, and after a growing season to determine if the study site's 1990 downstream declining EP/Chironomidae ratio (EVS, 1993), which was most severe in summer months, is still observable. 170 An extensive sampling of river bottom-water D.O. and temperature should be performed on the Nicomekl system before, during, and following the growing season to determine if hypoxic/anoxic conditions exist (majority of sampling to date has been surface sampling), and if so surmise the probable effects of these conditions on the system's fish and invertebrate biota (Is low D.O. possibly responsible for EVS (1993)'s reported declining EP/Chironomidae ratio). 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Englewood Cliffs, New Jersey: Prentice Hall. 188 Appendix 1 Study Site Drainage Ditches' Municipal Pump Station Records Drainage Ditch Pump Station's Daily Hours of Operation South Cloverdale Ditch (Site 5^  Mav - Aug.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible solely due to pump hours being recorded at a slightly later time the following day. Discharge always equaled 2.4 m3/s at this site. Pump. = discharge, no irrigation at this site. May Pump June Pump July Pump Aug. Pump 1 1.9 1 4.5 1 0 1 0 2 1.7 2 1.9 2 0 2 0 3 1.7 3 3.7 3 0 3 0 4 2.4 4 1 4 0.8 4 0 5 8.8 5 1 5 0.8 5 0 6 12.2 6 0 6 0.8 6 0 7 1.1 7 0 7 2.8 7 0.6 8 1.1 8 0 8 13.3 8 0.6 9 1.1 9 0 9 6.9 9 0.6 10 1.1 10 0 10 3.7 10 0.6 11 1.1 11 0 11 1.5 11 0.8 12 0 12 0.4 12 1.5 12 0.7 13 0.4 13 0.4 13 1.5 13 0 14 0.6 14 0.4 14 0.8 14 0 15 0.6 15 0.4 15 1 15 0 16 0.7 16 1.7 16 0 16 0 17 0.7 17 0.7 17 0 17 0 18 0.7 18 0.1 18 0 18 0 19 0.7 19 0 19 0 19 0 20 0.5 20 0.2 20 0 20 0 21 0.2 21 0.1 21 0 21 0.3 22 0 22 0.1 22 0 22 0.1 23 0 23 0 23 0 23 0 24 0 24 0.1 24 0 24 0 25 0 25 0 25 0.3 25 0 26 0 26 0.4 26 0.3 26 3 27 0 27 0.5 27 0.3 27 0.8 28 0.1 28 0.5 28 0 28 0 29 1.9 29 0.5 29 0 29 0 30 6.4 30 0 30 0 30 0 31 6.4 31 0 31 0 189 Appendix 1 Cont. Drainage Ditch Pump Station's Daily Hours of Operation South Cloverdale Ditch (Site 5) Sept. - Nov.. 1997 Times shown are the hours of pumping between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible solely due to pump hours being recorded at a slightly later time the following day. Discharge always equaled 2.4 m3/s at this site. Pump. = discharge, no irrigation at this site. Sept. Pump Oct. Pump Nov. Pump 1 0 1 3 1 2.3 2 0 2 3 2 2.3 3 0.3 3 4.1 3 5.5 4 0.4 4 8.2 4 1.8 5 0.3 5 4.1 5 2.6 6 0.3 6 1.5 6 4.2 7 0.3 7 0.8 7 1.8 8 0.3 8 3.1 8 1.8 9 0.1 9 1.6 9 1.8 10 0.1 10 1 10 1.1 11 0 11 1.1 11 1.1 12 0 12 1.1 12 0.6 13 0 13 1.7 13 0.4 14 0 14 2 14 0 15 0.2 15 1.7 15 0 16 2.7 16 1.5 16 0 17 3.6 17 0.2 17 0.8 18 0.8 18 0.2 18 0.8 19 0.2 19 .0.2 19 3.9 20 0.2 20 0.2 20 4 21 0.1 21 0 21 3.4 22 0 22 0.1 22 3.4 23 0 23 0.7 23 3.4 24 0.1 24 0.4 24 6.3 25 1 25 0.6 25 5 26 0.9 26 3.8 26 7.4 27 0.8 27 1.7 27 7.4 28 4.4 28 4.9 29 1.1 29 10 30 2 30 8.2 31 2.3 190 A p p e n d i x 1 C o n t . E r i c s o n D i t c h ' s ( S i t e 1 0 ) M u l t i p l e P u m p s ' C u m u l a t i v e D a i l y H o u r s o f O p e r a t i o n M a v - A u g . . 1 9 9 7 T i m e s s h o w n a r e t h e h o u r s b e t w e e n t h e m o r n i n g o f t h e d a y s h o w n a n d t h e m o r n i n g o f t h e f o l l o w i n g d a y . G r e a t e r t h a n 2 4 p u m p - h o u r s o n a s i n g l e d a y i s p o s s i b l e s i n c e t h e r e a r e 2 p u m p s a t t h i s s t a t i o n . I n a d d i t i o n , g r e a t e r t h a n 2 4 h o u r s o f p u m p i n g o n a s i n g e d a y i s p o s s i b l e s o l e l y d u e t o p u m p h o u r s b e i n g r e c o r d e d a t a s l i g h t l y l a t e r t i m e t h e f o l l o w i n g d a y . A v e r a g e d i s c h a r g e = 2 . 0 m 3 / s p e r p u m p . P u m p . = d i s c h a r g e , I r r i g . = i n t a k e . May Pump Irrig June Pump Irrig July Pump Irrig Aug. Pump Irrig 1 6.9 0 1 45.1 0 1 0 0 1 0 11.2 2 7.5 0 2 13.3 0 2 0 0 2 0 11.2 3 7.5 0 3 12.2 0 3 0.2 0 3 0 11.2 4 7.5 0 4 5.8 0 4 1.5 0 4 0 11.2 5 39.7 0 5 3.1 0 5 1.5 0 5 0 12.7 6 19.4 0 6 0.9 0 6 1.5 0 6 0 8.2 7 8.3 0 7 0.9 0 7 9.7 0 7 0 7.3 8 4.9 0 8 0.9 0 8 48.4 0 8 0 7.7 9 1 0 9 1.8 0 9 45.9 0 9 0 7.7 10 1 0 10 1.7 0 10 47.9 0 10 0 7.7 11 1 0 11 1.6 0 11 5.3 0 11 0 5.9 12 1.2 0 12 1.9 0 12 5.3 0 12 0 9.8 13 2 0 13 0.5 0 13 5.3 0 13 0 9.3 14 0 0 14 0.5 0 14 5.6 0 14 0 8.5 15 0 0 15 0.5 0 15 3.2 0 15 0 10.5 16 0.6 0 16 3.9 0 16 1.9 0 16 0 10.5 17 0.6 0 17 0.9 4.9 17 0.5 0 17 0 10.5 18 0.6 0 18 0 0 18 0.9 0 18 0 9.9 19 0.6 0 19 0 0 19 0.9 0 19 0 11.5 20 0 0 20 1.6 0 20 0.9 0 20 0 6.1 21 0 0 21 1.6 0 21 0 0 21 0 5.7 22 0 0 22 1.6 0 22 0 0 22 0 6.4 23 0 0 23 0 0 23 0 0 23 0 6.4 24 0 0 24 0 0 24 0 0 24 0 6.4 25 0 0 25 0 0 25 0 0.7 25 0 0.4 26 0 0 26 0.8 0 26 0 0.7 26 0.5 1.3 27 0 0 27 0.8 0 27 0 0.7 27 0.5 1.3 28 0 0 28 0.8 0 28 0 13.1 28 0 0 29 0 0 29 0 0 29 0 11.4 29 0 0 30 45.1 0 30 0 0 30 0 10.4 30 0 0 31 45.1 0 31 0 9.1 31 0 0 191 Appendix 1 Cont. Ericson Ditch's (Site 10) Multiple Pumps' Cumulative Daily Hours of Operation Sept. - Nov.. 1997 Times shown are the total daily pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 2 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 2.0 m3/s per pump. Pump. = discharge, Irrig. = intake. Sept. Pump Irrig Oct. Pump Irrig Nov. Pump Irrig 1 0 0 1 7.5 0 1 5.1 0 2 0 0 2 4.7 0 2 5.1 0 3 0 0 3 9.3 0 3 12 0 4 0 11.7 4 9.3 0 4 5.8 0 5 0 8.2 5 9.3 0 5 6.1 0 6 0 8.2 6 3.8 0 6 7.5 0 7 0 8.2 7 1.8 0 7 4.8 0 8 0 8.9 8 6.5 0 8 4.8 0 9 0 9.2 9 4.4 0 9 4.8 0 10 0 10.2 10 3 0 10 3.4 0 11 0 10 11 3 0 11 3.4 0 12 0 11.5 12 3 0 12 2.7 0 13 0 11.5 13 3 0 13 1.5 0 14 0 11.5 14 4.1 0 14 0.4 0 15 0 3.4 15 6.6 0 15 0.4 0 16 0.8 0 16 3 0 16 0.4 0 17 1.4 0 17 1.3 0 17 2.8 0 18 0 0 18 1.3 0 18 4.4 0 19 1.7 0 19 1.3 0 19 9.8 0 20 1.7 0 20 0 0 20 9.7 0 21 1.7 0 21 0 0 21 7.5 0 22 0 0 22 1.8 0 22 7.5 0 23 2.2 0 23 2.3 0 23 7.5 0 24 1.5 0 24 3.2 0 24 10 0 25 2.3 0 25 3.2 0 25 9.8 0 26 4.8 0 26 3.2 0 26 5.7 0 27 4.8 0 27 3.7 0 27 24.2 0 28 4.8 0 28 9.6 0 29 3.3 0 29 27.4 0 30 3.3 0 30 13 0 31 5.1 0 192 Appendix 1 Cont. Burrows Ditch's (Site 11) Multiple Pumps' Cumulative Daily Hours of Operation May - Aug.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 2 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.55 m3/s per pump. Pump. = discharge, Irrig. = intake. May Pump Irrig June Pump Irrig July Pump Irrig Aug. Pump Irrig 1 6.2 0 1 15.4 0 1 0.3 0 1 0 14.2 2 5.4 0 2 5.7 0 2 0.1 0 2 0 15.8 3 5.4 0 3 8.5 0 3 0 0 3 0 15.8 4 5.4 0 4 5.9 0 4 2.6 0 4 0 15.8 5 22.4 0 5 4.5 0 5 2.6 0 5 0 14.8 6 19 0 6 1.5 0 6 2.6 0 6 0 13.9 7 7.8 0 7 1.5 0 7 8.1 0 7 0 14.3 8 5.5 0 8 1.5 0 8 26.7 0 8 0 10.4 9 2.5 0 9 1 0 9 18.4 0 9 0 10.4 10 2.5 0 10 1.4 0 10 12.8 0 10 0 10.4 11 2.5 0 11 0.9 0 11 5.8 0 11 0 7.3 12 1.8 0 12 2.4 0 12 5.8 0 12 0 11.4 13 1.6 0 13 0.8 0 13 5.8 0 13 0 13.3 14 0 0 14 0.8 0 14 3 0 14 0 14.2 15 0 0 15 0.8 0 15 1.9 0 15 0 15.4 16 0 0 16 2.2 0 16 1.5 0 16 0 15.4 17 0 0 17 2.3 0 17 0 0 17 0 15.4 18 0 0 18 1.3 0 18 0 0 18 0 18 19 0 0 19 0.7 0 19 0 0 19 0 16.7 20 0 0 20 0.7 0 20 0 0 20 0 12.2 21 0 0 21 0.7 0 21 0 24.6 21 0 14.2 22 0 0 22 0.7 0 22 0 11.3 22 0 10.9 23 0 0 23 1 0 23 0 6.7 23 0 10.9 24 0 0 24 0.8 0 24 0 8.2 24 0 10.9 25 0 0 25 0.8 0 25 0 6.9 25 0 4.5 26 0 0 26 1 0 26 .0 6.9 26 5 7 27 0 0 27 1.2 0 27 0 6.9 27 0 0 28 0 0 28 1.2 0 28 1.9 15.1 28 0 0 29 0 0 29 1.2 0 29 2 14.6 29 0 0 30 0 0 30 0.3 0 30 0 22.9 30 0 0 31 15.4 0 31 0 14.2 31 0 0 193 Appendix 1 Cont. Burrows Ditch's (Site 11) Multiple Pumps' Cumulative Daily Hours of Operation Sept. - Nov.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 2 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.55 m3/s per pump. Pump, discharge, Irrig. = intake. Sept. Pump Irrig Oct. Pump Irrig Nov. Pump Irrig 1 0 0 1 8.4 0 1 7.3 0 2 0 0 2 5.6 0 2 7.3 0 3 0 0 3 12.2 0 3 4.5 0 4 0 0 4 12.2 0 4 7.5 0 5 0 0 5 12.2 0 5 7.4 0 6 0 0 6 5.5 0 6 9.6 0 7 0 0 7 3.5 0 7 5.8 0 8 0 0 8 5.8 0 8 5.8 0 9 0 25.3 9 5.7 0 9 5.8 0 10 0 4.7 10 3.4 0 10 0.2 0 11 0 24 11 3.4 0 11 0.2 0 12 0 4.8 12 3.4 0 12 11.1 0 13 0 4.8 13 3.4 0 13 2.4 0 14 0 4.8 14 4.6 0 14 0.9 0 15 0 9.4 15 7.9 0 15 0.9 0 16 0.8 0.4 16 4.6 0 16 0.9 0 17 9.1 0 17 2.1 0 17 4 0 18 2.7 0 18 2.1 0 18 6.8 0 19 0 0 19 2.1 0 19 6.8 0 20 0 0 20 0.5 0 20 9.7 0 21 0 0 21 0.8 0 21 8.2 0 22 0 0 22 2.2 0 22 8.2 0 23 0 0 23 2 0 23 8.2 0 24 0 0 24 3.8 0 24 13 0 25 0 0 25 3.8 0 25 11.3 0 26 0 0 26 3.8 0 26 9.2 0 27 3.9 0 27 5.3 0 27 20.9 0 28 3.9 0 28 9.4 0 29 9.4 0 29 22.1 0 30 13.5 0 30 18.4 0 31 7.3 0 194 Appendix 1 Cont. Old Logging Ditch's (Site 12) Multiple Pumps' Cumulative Daily Hours of Operation May - Aug.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 3 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.75 m3/s per pump. Pump. = discharge, Irrig. = intake. May Pump Irrig June Pump Irrig July Pump Irrig Aug. Pump Irrig 1 7.8 0 1 19.3 0 1 0.7 0 1 0 15.7 2 10 0 2 10.3 0 2 0.7 0 2 0 15.7 3 10 0 3 12.6 0 3 0 0 3 0 15.7 4 10 0 4 8 0 4 5.8 10.1 4 0 15.7 5 48.6 0 5 5.8 0 5 5.8 10.1 5 0 21.4 6 30 0 6 2.3 0 6 5.8 10.1 6 0 27.5 7 13.4 0 7 2.3 0 7 8.7 0 7 0 16 8 8.3 0 8 2.3 0 8 42.8 0 8 0 0 9 4.4 0 9 1.3 0 9 23.6 0 9 0 0 10 4.4 0 10 1.7 0 10 18.5 0 10 0 0 11 4.4 0 11 1.7 0 11 6 0 11 0 18.9 12 3.3 0 12 2.1 0 12 6 0 12 0 0 13 2.4 0 13 1.7 0 13 6 0 13 0 22.1 14 0 0 14 1.7 0 14 6 0 14 0 26.2 15 0 0 15 1.7 0 15 2.8 0 15 0 23.6 16 1.8 0 16 3.8 0 16 2.8 0 16 0 23.6 17 1.8 0 17 4.1 0 17 1.6 10.3 17 0 23.6 18 1.8 0 18 2.8 0 18 0 10.3 18 0 22.3 19 1.8 0 19 1.1 0 19 0 10.3 19 0 24 20 2.1 0 20 1.8 0 20 0 10.3 20 0 24.3 21 0.3 0 21 1.8 0 21 0 20 21 0 24.3 22 0 0 22 1.8 0 22 0 20 22 0 0.7 23 0 0 23 1.9 0 23 0 34.2 23 0 0.7 24 0 0 24 1.1 0 24 0 14.2 24 0 0.7 25 0 0 25 1.4 0 25 0 10.6 25 0 20.3 26 0 0 26 1.8 0 26 0 10.6 26 2.6 6.4 27 0 0 27 1.6 0 27 0 10.6 27 2.6 0 28 2.3 0 28 1.6 0 28 0 20.6 28 0 0 29 5 0 29 1.6 0 29 0 4.8 29 0 0 30 19.3 0 30 0.7 0 30 0 21.2 30 0 0 31 19.3 0 31 0 0 31 0 0 195 Appendix 1 Cont. Old Logging Ditch's (Site M) Multiple Pumps' Cumulative Daily Hours of Operation Sept. - Nov.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 3 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.75 m3/s per pump. Pump, discharge, Irrig. = intake. Sept. Pump Irrig Oct. Pump Irrig Nov. Pump Irrig 1 0 0 1 9 0 1 9.8 2 0 0 2 5.1 0 2 9.8 3 0 0 3 13.6 0 3 19.5 4 0 0 4 13.6 0 4 10 5 0.5 0 5 13.6 0 5 10.6 6 0.5 0 6 5.4 0 6 14.7 7 0.5 0 7 3.2 0 7 6.9 8 0 0 8 6.4 0 8 6.9 9 0 0 9 5.5 0 9 6.9 10 0 0 10 3.4 0 10 4.6 11 0 0 11 3.4 0 11 4.6 12 0 0 12 3.4 0 12 3.2 13 0 0 13 3.4 0 13 3.3 14 0 0 14 5.4 0 14 1.5 15 0 0 15 10.5 0 15 1.5 16 5.3 0 16 4.7 0 16 1.5 17 8.4 0 17 2.6 0 17 7 18 3.1 0 18 2.6 0 18 6.4 19 2.6 0 19 2.6 0 19 13.5 20 2.6 0 20 .1.1 0 20 17.5 21 2.6 0 21 1.6 0 21 12.2 22 0 0 22 3.4 0 22 12.2 23 1.7 0 23 3.2 0 23 12.2 24 0.9 0 24 5 0 24 21.4 25 2.7 0 25 5 0 25 20.3 26 6.2 0 26 5 0 26 12 27 6.2 0 27 6.9 0 27 14.2 28 6.2 0 28 11.6 0 29 3.6 0 29 26.1 0 30 2.0 0 30 22.5 0 31 9.8 0 196 Appendix 1 Cont. Halls Prairie Ditch's Multiple Pumps' Cumulative Daily Hours of Operation Mav-Aug.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 2 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.95 m3/s per pump. Pump, discharge, no irrigation at this ditch this particular year. May Pump June Pump July Pump Aug. Pump 1 1.9 1 4.8 1 0.3 1 0 2 2.1 2 2.8 2 0 2 0 3 2.1 3 4.3 3 0 3 0 4 2.1 4 2.5 4 0 4 0 5 19.1 5 2.1 5 0 5 0 6 6.7 6 0.9 6 0 6 0 7 3.4 7 0.9 7 7.1 7 0 8 3.4 8 0.9 8 17.8 8 0 9 1.7 9 0.4 9 6.9 9 0 10 1.7 10 0.6 10 5.5 10 0 11 1.7 11 0.3 11 2.3 11 0 12 2.5 12 0.9 12 2.3 12 0 13 2.4 13 0.3 13 2.3 13 0 14 2.3 14 0.3 14 1.4 14 0 15 1.7 15 0.3 15 1.4 15 0 16 1.4 16 1.1 16 0.9 16 0 17 1.4 17 0.9 17 1.5 17 0 18 1.4 18 0.7 18 0 18 0 19 1.4 19 0.3 19 0 19 0 20 1.4 20 0.5 20 0 20 0 21 1.1 21 0.5 21 0 21 0 22 0.7 22 0.5 22 0 22 0 23 0.2 23 1 23 0 23 0 24 0.2 24 0.3 24 0 24 0 25 0.2 25 0.4 25 0 25 0 26 0 26 0.6 26 0 26 0 27 0 27 0.4 27 0 27 0 28 0.7 28 0.4 28 0 28 0 29 1.4 29 0.4 29 0 29 0 30 4.8 30 0.3 30 0 30 0 31 4.8 31 0 31 0 197 Appendix 1 Cont. Halls Prairie Ditch's Multiple Pumps' Cumulative Daily Hours of Operation Mav-Aug.. 1997 Times shown are the total pump-hours between the morning of the day shown and the morning of the following day. Greater than 24 pump-hours on a single day is possible since there are 2 pumps at this station, and in addition, greater than 24 hours of pumping on a singe day is possible solely due to pump hours being recorded at a slightly later time the following day. Average discharge = 0.95 m3/s per pump. Pump, discharge, no irrigation at this ditch this particular year. Sept. Pump Oct. Pump Nov. Pump 1 0 1 2.9 1 3.4 2 0 2 1.6 2 3.4 3 0 3 4.8 3 6.7 4 0 4 4.8 4 3.4 5 0 5 4.8 5 3.5 6 0 6 2.3 6 4.6 7 0 7 1.6 7 2.3 8 0 8 2.5 8 2.3 9 0 9 1.7 9 2.3 10 0 10 1.2 10 1.3 11 0 11 1.2 11 1.3 12 0 12 1.2 12 1 13 0 13 1.2 13 1.2 14 0 14 1.9 14 0.5 15 0 15 3.3 15 0.5 16 0 16 1.7 16 0.5 17 5.4 17 0.9 17 2 18 1.3 18 0.9 18 1.9 19 0.4 19 0.9 19 4.1 20 0.4 20 0.3 20 3.9 21 0.4 21 0.5 21 3.7 22 0 22 0.9 22 3.7 23 0.3 23 0.7 23 3.7 24 0.1 24 1.5 24 7.4 25 0.9 25 1.5 25 6.2 26 1.4 26 1.5 26 4 27 1.4 27 1.6 27 11.6 28 1.4 28 4.9 29 1.2 29 11.5 30 0.7 30 7.7 31 0.7 31 3.4 198 Appendix 2 Statistical Comparison of C. dubia Chronic Survival and Reproduction Between Samples and Controls May 6 1997 Mean # Neonates Per Female Standard Deviation Versus Culture Control (p=, 1 tail) Versus Culture Control (p=,2 tail) Discharge From Site 8 21.2 2.5 0.0007 * 0.001 * Culture Control 18.1 6.3 June 6 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (P=, 1 tail) Versus Site 13 Control (p=, 2 tail) Versus Culture Control (P=, 1 tail) Versus Culture Control (P=,2 tail) Site 1 37.2 5.3 0.430 0.860 0.006 * 0.011 * Site 2 41.0 6.5 0.047 0.094 0.001 * 0.003 * Site 3 44.8 8.4 0.251 0.502 0.011 * 0.022 * Site 4 34.4 13.4 0.315 0.630 0.044 * 0.087 Site 5 33.6 6.7 0.125 0.251 0.027 * 0.054 Site 7 40.3 5.9 0.064 0.128 0.002 * 0.003 * Site 8 33.5 4.0 0.058 0.117 0.023 * 0.046 * Site 10 35.8 9.8 0.409 0.818 0.016* 0.032 * Site 11 41.1 9.0 0.090 0.180 0.002 * 0.004 * Site 12 40.6 7.5 0.082 0.164 0.002 * 0.003 * Site 13 36.6 4.4 0.007 * 0.014* Culture Control 26.2 11.1 0.007 * 0.014* * P-value < 0.05 indicates that there is a statistically significant difference. 199 Appendix 2 Cont June 26 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (P=, 1 tail) Versus Site 13 Control (p= 2 tail) Versus Culture Control (p=, 1 tail) Versus W.R.P. Control (P=,2 tail) Site 1 21.9 6.7 0.357 0.714 0.420 0.841 Site 2 18.3 8.2 0.078 0.156 0.190 0.379 Site 3 21.5 7.3 0.314 0.628 0.475 0.949 Site 4 21.5 8.6 0.158 0.317 0.477 0.954 Site 5 15.7 6.3 0.006 * 0.013* 0.033 * 0.067 Site 7 19.3 8.4 0.133 0.266 0.279 0.559 Site 8 0 0 1.2 X 10"7* 2.3 X 10 7 * 1.5 X10"6* 2.9 X10"6* Site 10 24.6 3.5 0.203 0.406 0.091 0.182 Site 11 16.4 7.5 0.019* 0.038 * 0.068 0.136 Site 12 12.5 5.0 0.0001 * 0.0003 * 0.002 * 0.004 * Site 13 22.9 5.2 0.277 0.553 Culture Control 21.3 6.5 0.278 0.553 July 16 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (p=, 1 tail) Versus Site 13 Control (P=, 2 tail) Versus Culture Control. (P=, 1 tail) Versus Culture Control (p=,2 tail) Site 1 23.1 11.0 0.132 0.265 0.053 0.105 Site 2 25.8 6.7 0.024 * 0.048 * 0.005 * 0.011 * Site 3 18.4 12.9 0.447 0.894 0.274 0.547 Site 4 10.9 6.0 0.042 * 0.084 0.115 0.229 Site 5 0 0 0.0002 * .0.0003 * 0.0003 * 0.0006 * Site 7 20.3 11.1 0.294 0.588 0.146 0.292 Site 8 13.5 7.1 0.146 0.293 0.317 0.633 Site 10 17.0 11.7 0.443 0.887 0.362 0.724 Site 11 19.2 7.2 0.352 0.704 0.155 0.311 Site 12 18.0 11.1 0.475 0.950 0.282 0.564 Site 13 19.7 8.2 0.292 0.584 Culture Control 15.3 9.3 0.292 0.584 * P-value < 0.05 indicates that there is a statistically significant difference. 200 Appendix 2 Cont Aug. 7 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (p=, 1 tail) Versus Site 13 Control (P=, 2 tail) Versus Culture Control (p=, 1 tail) Versus Culture Control (p=,2 tail) Site 1 13.0 8.2 0.452 0.905 0.209 0.417 Site 2 19.1 3.2 0.013* 0.027 * 0.075 0.145 Site 3 15.1 5.1 0.260 0.521 0.408 0.816 Site 4 15.7 4.7 0.189 0.375 0.5 1 Site 5 16.2 6.6 0.176 0.352 0.432 0.864 Site 7 10.9 6.0 0.190 0.380 0.047 * 0.096 Site 8 19.2 3.6 0.105 0.210 0.252 0.504 Site 10 12.7 5.0 0.394 0.789 0.125 0.250 Site 11 9.9 6.7 0.125 0.250 0.030 * 0.061 * Site 12 14.3 7.2 0.386 0.773 0.324 0.649 Site 13 14.9 4.7 0.214 0.428 Culture Control 15.7 6.2 0.214 0.428 Aug. 29 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (p=, 1 tail) Versus Site 13 Control (p=, 2 tail) Versus Culture Control (P=, 1 tail) Versus Culture Control (P=,2 tail) Site 1 19.7 4.5 0.256 0.511 0.194 0.397 Site 2 17.2 5.3 0.163 0.325 0.204 0.409 Site 3 19.8 5.3 0.484 0.968 0.036 * 0.073 Site 4 18.9 7.1 0.392 0.784 0.102 0.205 Site 5 16.5 6.7 0.133 0.265 0.311 0.622 Site 7 16.8 6.5 0.151 0.301 0.271 0.542 Site 8 18.8 2.4 0.327 0.654 0.042 * 0.085 Site 10 17.0 5.3 0.144 0.287 0.226 0.452 Site 11 22.0 3.4 0.146 0.291 0.003 * 0.005 * Site 12 13.2 7.0 0.011 * 0.022 * 0.156 0.312 Site 13 19.7 5.7 0.044 * 0.089 Culture Control 15.1 5.7 0.044 * 0.089 * P-value < 0.05 indicates that there is a statistically significant difference. 201 Appendix 2 Cont Sept. 17 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (p=, 1 tail) Versus Site 13 Control (p=, 2 tail) Versus Culture Control (P=, 1 tail) Versus Culture Control (P=,2 tail) Site 1 16.3 4.5 0,002 * 0.003 * 0.056 0.112 Site 2 19.1 5.5 0.075 0.149 0.006 * 0.012* Site 3 16.3 6.8 0.015* 0.029 * 0.129 0.258 Site 4 21.0 3.7 0.240 0.481 3.4 X 10"5* 6.8 X 10 s* Site5 17.2 4.0 0.003 * 0.006 * 0.014* 0.027 * Site 6 20.0 4.7 0.125 0.250 0.0009 * 0.002 * Site 7 16.6 8.7 0.044 * 0.087 0.158 0.315 Site 8 23.7 2.3 0.060 0.121 4.7 X10"9* 9.5 X 109* Site 10 22.6 2.9 0.308 0.616 2.4 X10"7* 4.9 X10"7* Site 11 24.3 6.4 0.172 0.343 0.0002 * 0.0004 Site 12 21.4 3.9 0.341 0.681 3.7 X10"5* 7.4 X 10"5* Site 13 22.0 2.4 8.6 X 10"8* 1.7 X 10"7* W.R.P. Cont. 13.6 2.2 8.6X10 8* 1.7 X10"7* Oct. 3 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 (P=, 1 tail) Versus Site 13 (P=, 2 fail) Versus W.R.P. Cont. (P=, 1 tail) Versus W.R.P. Cont. (P=,2 tail) Site 1 18.5 5.0 0.041 * 0.081 0.018* 0.036 * Site 2 14.5 8.9 0.364 0.728 0.007 * 0.014* Site 3 20.8 6.6 0.013* 0.027 * 0.150 0.301 Site 4 16.5 6.8 0.158 0.316 0.008 * 0.017* Site5 16.2 7.8 0.196 0.392 0.012* 0.023 * Site 6 11.3 8.5 0.301 0.602 0.0007 * 0.001 * Site 7 21.1 9.0 0.024 * 0.047 * 0.228 0.456 Site 8 16.1 4.4 0.154 0.308 0.001 * 0.002 * Site 10 16.7 6.1 0.134 0.267 0.007 * 0.014* Site 11 16.6 4.1 0.173 0.346 0.018* 0.037 * Site 12 17.5 8.8 0.127 0.254 0.039 * 0.078 Site 13 13.2 7.5 0.001 * 0.002 * Culture Control 23.6 5.1 0.001 * 0.002 * * P-value < 0.05 indicates that there is a statistically significant difference. 202 Appendix 2 Cont Oct. 3 1997 Retests Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (P=, 1 tail) Versus Site 13 Control (P=, 2 tail) Versus Culture Control (P=, 1 tail) Versus Culture Control (P=,2 tail) Site 2 21.2 2.5 0.070 0.140 0.088 0.176 Site 6 17.7 4.3 0.263 0.527 0.436 0.872 Site 13 18.9 4.0 0.370 0.740 Culture Control 18.1 6.3 0.370 0.740 Oct. 16 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (P=, 1 tail) Versus Site 13 Control (P=, 2 tail) Versus Culture Control (p= 1 tail) Versus Culture Control (P=,2 tail) Site 5 18.8 5.3 0.481 0.962 0.395 0.791 Site 10-Nearby Field Puddle 8.6 4.5 1.8 X 10"5* 3.7 X10"5* 0.0007 * 0.001 * Site 13 18.9 4.0 0.370 0.740 Culture Control 18.1 6.3 0.370 0.740 Nov. 21 1997 Mean # Neonates Per Female Standard Deviation Versus Site 13 Control (p=, 1 tail) Versus Site 13 Control (p= 2 tail) Versus Culture Control (P=, 1 tail) Versus Culture Control (P=,2 tail) Site 1 21.2 9.1 0.252 0.504 0.437 0.874 Site 2 25.1 9.6 0.310 0.620 0.100 0.200 Site 3 28.4 7.7 0.048 * 0.096 0.006 * 0.013* Site 4 23.4 6.5 0.500 1 0.127 0.254 Site5 20.8 3.1 0.076 0.153 1.734 0.947 Site 6 22.2 5.0 0.291 0.582 0.227 0.454 Site 7 20.2 7.1 0.124 0.249 0.423 0.846 Site 8 11.4 6.5 0.0001 * 0.0002 * 0.0007 * 0.001 * Site 10 22.4 7.5 0.361 0.722 0.264 0.527 Site 11 26.6 7.5 0.132 0.263 0.021 * 0.042 * Site 12 19.7 5.7 0.063 0.127 0.323 0.646 Site 13 23.4 4.5 0.078 0.155 Culture Control 20.7 3.6 0.078 0.155 * P-value < 0.05 indicates that there is a statistically significant difference. 203 Appendix 3 C. dubia Successive Reference Toxicant Results O ro cn E o LO g g o T3 O i_ Q_ CD Ceriodaphnia dubia Successive Reference Toxicant Results 2000 1800 1600 1400 1200 1000 800 600 400 200 0 1 « » — « - ... . _ _ f -: t" « » « • 1 I « |! » 4 » < » . ' I I o - .... . ... _ -- - • -T ' 1 3 4 5 6 Reference Toxicant # 8 Individual +/- 2 S.D. Shown Thick Solid Lines = Mean IC50 +/- Mean 2 S.D. Reference Toxicant # 1 May 6 - May 13 2 May 23 - May 30 3 June 19 - June 25 4 July 25 - August 1 5 September 4 - September 11 6 October 29 - November 3 7 November 7 -November 14 8 December 15 - December 22 May to November reference toxicants coincide with the study site sample tests. The December reference toxicant coincides with the pesticide toxicity tests. 204 Appendix 4 C. dubia Lethally Toxic Sample's GC/MS Chromatograms. Full-lon-Scan 500X Concentration Site 8. 176 t h St. Ditch. June 26 t h Sample 80000-70000-60000-50000-40000 30000-60.00 th Site 5. South Cloverdale Ditch. July 16 Sample l&bundanoe 140000 -I TIC: 0716S5.D 120000] 100000] 80000-^  60000-^  40000-^  205 Appendix 5 Prometryn GC/MS. Full-lon-Scan. Standards' Curves CO 0 CO CD Q_ CO o o 10000000 9000000 8000000 7000000 6000000 5000000 4000000 3000000 2000000 1000000 0 Prometryn GC MS Standards' Curves Including Site 8, June 26, 1997 0 3 4 5 6 7 mg/L Prometryn in GC-MS 10 Acet. Std. M. Chi. Std. -figr Site 8 - LL Site 8-SPE 1000000 $ 900000 < 800000 700000 $ 600000 tt. 500000 CO 40000CT ^ 300000 O 200000 0 100000 0 Prometryn GC MS Standards' Curves Including Site 8, June 26, 1997 m III : . | r . - 3 0.2 —m 0.4 0.6 0.8 1.2 1.4 1.6 1.8 Acet. Std. mg/L Prometryn in GC-MS • M. Chi. Std. -Egg- Site 8 - LL —|— Site 8-SPE - 0.003 mgfL 0.002 mgl. 206 Appendix 6 C. dubia Lethally Toxic Samples' GC/MS Chromatograms. Select-Ion Scan. 3600X Concentration Site 8. 176 t h St. Ditch. June 26 t h Sample |Abundance. 23000 22000 21000 20000 19000 18000-1 17000 16000 15000 14000 13000 12000 11000 10000 9000 8000 7000 6000 5000 4000 3000 2000 1000 0 TIC: 9.D ' I ' ' ' • i • • • • i • • • ' l ' • ' • l '. ' ' i l i ' i ' I i ' ' ' l > ' ' ' I i ' i i l ' ' ' ' l i 10.00 15.00 20.00 25.00 30.00 35.00 40.00 45.00 50.00 55.00 60.00 M i r n y - * 207 Appendix 6 Cont. Site 5. South Cloverdale Ditch. July 16 Sample 208 Appendix 7 C. dubia Lethally Toxic Samples' Identified Diazinon and Chlorpyrifos Select-lon-Scan Chromatogram Peaks and Mass Spectra Site 8. 176 St. Ditch. June 26 Sample DIAZINON 3000 TIC: 9.D 2500 2000 1500 1000 500 0 1 ' i | ' i ' i i ' i ' ' | ' 'I 'I I ' 1 1 1 I ' 1 1 ' I 1 1 1 1 I 1 1 1 1 I ' 1 1 1 I 1 ' 1 ' I ' 1 1 1 I 1 1 ' 1 I 1 1 1 ' I ' 1 1 1 I 1 1 1 1 I 1 1 1 1 I 1 1 1 ' I 1 1 h m e r » 3ll00 3l!05 3l!l0 3l!l5 31.20 31.25 31.30 31.35 31.40 31.45 31.50 31.55 31.60 31.65 31.70 31.75 31.80 31.85 0 1 I | I I I I | I I I I | I I. I I | I I I l'| I I I I | I I I'. | I I . I | I I I I | T I I I | I I I I | I I I I | I I I I | I I I I | 1 I I I | I I I I | I I I I | I I I I | I I I I | ton/** 130 140 150 160 170 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 209 Appendix 7 Cont. C. dubia Lethally Toxic Samples' Identified Diazinon and Chlorpyrifos Select-lon-Scan Chromatogram Peaks and Mass Spectra Site 8. 176 t h St. Ditch. June 26 t h Sample CHLORPYRIFOS Abundance 1800 1600 1400 1200-1 1000 800-600-400 200 Average of 34.984 to 34.995 min.: 9.D 1*7 137 I7S 314 304 ' ' ' | • ' ' | ' I I I | ' I I I | I ' I ' | ' ' I I | I I I I | I I I I | | | , I | | , | | | | | | | | , , , | , . . . . . 130 140 150 160 170 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 210 Appendix 7 Cont. C. dubia Lethally Toxic Samples' Identified Diazinon and Chlorpyrifos Select-lon-Scan Chromatogram Peaks and Mass Spectra Site 5. South Cloverdale Ditch. July 16 t h Sample DIAZINON Abundance 2400 2200 2000 1800 1600 1400 1200 TIC: 8.D 1000 i 1 1 i | i . • i | i i i i | i i i i | i • i i | i i i i | i 11 i i i i | i i i i | i i i i | i i i i | fnme-* 31122 31 !24 31J26 31128 31130131132 31134 31 !36 31 !38 31 AO 31142 3144 31 A6 31A8 31.50 31.52 31.54 31.56 31.58 31.60 31.62 31.64 0 'i | ' 1 ' 1 | ' 1 • • 1 1 1 ' ' 1 ' ' ' ' 1 ' ' ' ' 1 ' ' ' ' 1 • ' ' ' 1 ' ' ' ' 1 ' ' ' ' I ' ' ' 1 I ' ' ' ' I ' ' 1 ' I ' ' ' ' I ' ' ' ' I ' ' ' ' I ' ' ' ' I ' ' ' ' I ' 1 ' ' I ' 1 ' ' I Imfc-^  130 140 150 160 170 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 211 Appendix 8 Diazinon and Chlorpyrifos Standards' Select-lon-Scan Chromatoaram Peaks and Mass Spectra DIAZINON. 80 uo/L Standard 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ' 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ' 1 1 1 1 1 1 ' i 1 1 1 1 1 1 1 1 ' i 1 1 1 1 1 1 ' 1 1 1 1 1 1 1 1 1 1 1 1 1 rrime-> 30.85 30.90 30.95 31.00 31.05 31.10 31.15 31.20 31.25 31.30 31.35 31.40 31.45 31.50 31.55 31.60 31.65 31.70 31.75 31.80 212 Appendix 8 Cont. Diazinon and Chlorpyrifos Standards' Select-lon-Scan Chromatogram Peaks and Mass Spectra CHLORPYRIFOS. 66 uall Standard abundance 130 120 110 100 90 80 70 60 50 40 30 20 10 0 137 Average of 34.988 to 34.999 min.: 3.D I £7 179 314 30f 1 1 1 1 1 • 1 1 1 1 1 1 • • 1 1 1 1 1 1 1 1 1 1 • • 1 1 • 1 1 1 , 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 ' 1 1 1 130 140 150 160 170 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 213 Appendix 9 Diazinon's Select-lon-Scan GC/MS Standards' Curve 600000 CD 550000 2500000 <450000 -£400000 2350000 Diazinon GC MS Standards' Curve 0.8 micrograms/L SPE, 500X Cone. ooooooc oooooo oooooo oooooo d SIAI-OFJ c l . , " . . | . > | 1 - j" -"• •• • j ••'' -•'...-; '-y | • " j - - '• " • : : I • - • • j '•• 1 : . ' | J • : ": ' ' • •.: j . '. : • . 0 100 200 300 400 500 600 700 800 900 micrograms/L Diazinon in GC-MS Chlorpyrifos Stds. - C O - 0.8 micrograms/L SPE, 500X Cone. = 0.4 micrograms/L _50000 <D 45000 <40000 35000 g30000 Q-25000 CO 20000 ^15000 O10000 O 5000 0 Diazinon GC MS Standards' Curve Including Sites 8 June 26 & 5 July 16 - g ^ - Chlorpyrifos Stds. - O s i t e 8, 3600X Cone. 40 60 80 100 micrograms/L Diazinon in GC-MS -FR- Site 5, 3600X Cone. 120 140 = 0.033 micrograms/L 0.4 micrograms/L SPE, 500X Cone. = 0.023 micrograms/L : 0.1 micrograms/L 214 Appendix 10 Chlorpyrifos' Select-lon-Scan GC MS Standards' Curve 1100000 (0 1000000 £ 900000 < 800000 lo 700000 ® 600000 500000 400000 A 300000 O 200000 100000 0 CO Chlorpyrifos GC MS Standards' Curve Including Site 8, June 26, 1997 0 400 600 800 1000 1200 m i c r o g r a m s / L C h l o r p y r i f o s i n G C - M S Chlorpyrifos Stds. -gg$- Site 8, 3600X Cone. = 0.067 micrograms/L CO CD CO CD D_ CO o O 30000 27000 24000 21000 18000 15000 12000 9000 6000 3000 0 Chlorpyrifos GC MS Standards' Curve 0.066 micrograms/L SPE, 500X Cone. 60 80 100 120 micrograms/L Chlorpyrifos in GC-MS 140 Chlorpyrifos Stds. 0.066 micrograms/L SPS = 0.16 micrograms/L based on standards 215 Appendix 11 C. dubia Chronic Test. Ditch. River, and Control Water Samples' Chemistry May 6 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L CaC03) Cond. (//S/cm) TDC (mg/L) TDIC (mg/L) TDOC (mg/L) Color (mg/L Pt) Discharge From Site 8 ?, 1 6.1 75 190 ? ? ? ? Culture Control 8.4 ? 68 ? ? ? ? ? June 6 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L CaC03) Cond. OuS/cm) TDC (mg/L) TDIC (mg/L) TDOC (mg/L) Color (mg/L Pt) Site 1 8.1, 5 7.3 85 240 28.2 10.7 17.5 160 Site 2 8.5, 4 7.1 90 245 27.0 12.0 15.0 150 Site 3 8.6, 5 7.1 95 260 28.7 12.6 16.1 120 Site 4 8.8, 3 7.3 84 225 23.8 11.0 12.8 110 Site5 8.4, 4 7.1 173 465 28.1 13.6 14.5 140 Site 7 8.9, 3 7.1 80 215 25.7 12.6 13.1 120 Site 8 8.5, 5 6.7 240 620 33.4 8.8 24.6 180 Site 10 9.4, 1 7.3 85 255 31.0 12.2 18.8 160 Site 11 8.7, 5 6.8 105 260 48.1 13.5 34.6 300 Site 12 10.9, 1 7.3 90 210 30.4 13.5 16.9 120 Site 13 9.6, 1 7.5 70 144 18.6 10.8 7.8 60 Culture Control ?, 1 7.7 74-80 195 19.9 13.0 6.9 30 216 Appendix 11 Cont June 26 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L CaC03) Cond. CuS/cm) TDOC (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pf) Site 1 8.8, 5 7.0 78 225 23.2 12.9 10.3 80 Site 2 8.8, 4 7.1 76 228 19.8 13.7 6.1 70 Site 3 8.6, 4 7.0 92 265 21.0 13.6 7.4 80 Site 4 8.9, 4 7.1 80 234 17.2 14.3 2.9 70 Site5 9.8, 1 7.0 155 442 19.2 12.2 7.0 100 Site 7 8.8, 4 7.2 75 228 16.5 14.7 1.8 50 Site 8 8.3 7.0 108 360 24.2 14.6 9.6 100 Site 10 9.8, 1 7.2 83 256 17.2 13.4 3.8 80 Site 11 9.2, 3 7.1 95 260 35.0 18.4 16.6 180 Site 12 8.1, 4 8.5 83 270 22.3 0 22.3 80 Site 13 9.3, 3 7.5 63 140 11.6 0 11.6 50 Culture Control ?, 1 7.1 83 225 16.0 13.7 2.3 25 July 16 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L CaC03) Cond. (//S/cm) TDOC (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 8.6, 5 7.2 95 274 26.7 12.8 13.9 130 Site 2 8.2, 5 7.2 120 352 30.0 16.8 13.2 140 Site 3 8.2, 5 7.1 133 400 33.1 17.9 15.2 160 Site 4 8.8, 4 7.2 113 339 26.4 14.2 12.2 110 Site5 8.9 7.1 138 406 24.4 12.9 11.5 130 Site 7 8.0, 5 7.2 103 311 30.6 15.4 15.2 100 Site 8 10.0, 1 7.0 265 660 26.8 6.0 20.8 80 Site 10 9.7, 1 7.4 93 237 21.3 11.3 10.0 70 Site 11 9.0, 4 7.1 105 283 45.8 15.6 30.2 210 Site 12 8.8, 5 7.4 90 211 28.2 15.9 12.3 70 Site 13 9.7, 1 7.6 66 157 17.0 12.1 4.9 40 Culture Control ?, 1 7.6 80 214 19.8 13.7 6.1 40 217 Appendix 11 Cont Aug. 7 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L C a C 0 3 ) Cond. (jUS/cm) T D O C (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 9.1, 5 7.8 93 275 27.2 17.9 9.3 75 Site 2 8.9, 5 7.7 90 265 24.3 17.6 6.7 40 Site 3 8.5, 5 7.5 98 296 24.3 17.6 6.7 50 Site 4 8.8, 4 7.6 90 260 25.8 17.9 7.9 35 S i t e5 9.3, 5 7.8 103 330 25.9 17.0 8.9 60 Site 7 8.8, 5 7.4 88 257 23.4 17.4 6.5 55 Site 8 8.7, 5 7.3 90 255 23.8 16.0 7.8 50 Site 10 8.9, 5 7.3 103 278 20.9 11.5 9.4 60 Site 11 7.7, 6 7.3 93 262 27.0 18.4 8.6 50 Site 12 7.0, 6 7.2 90 205 29.4 21.3 8.1 35 Site 13 8.7, 6 7.4 68 160 18.8 12.5 6.3 45 Culture Control ?, 1 7.3 79 240 22.8 14.0 8.8 35 Aug. 29 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L C a C 0 3 ) Cond. (piS/cm) T D O C (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 6.8 7.1 63 195 18.9 10.2 8.7 55 Site 2 6.6, 6 7.0 78 280 25.0 13.4 11.6 75 Site 3 6.1, 6 7.0 70 242 22.1 11.7 10.4 65 Site 4 7.4, 5 7.1 98 340 28.4 15.2 13.2 90 Site 5 6.8, 6 7.1 130 432 33.3 15.6 17.7 120 Site 7 7.5, 5 7.1 113 384 29.3 15.4 13.9 90 Site 8 7.0 7.0 169 465 30.7 11.9 18.8 120 Site 10 8.9, 4 7.4 94 265 22.7 12.1 10.6 85 Site 11 5.6 6.8 120 385 61.2 18.4 42.8 270 Site 12 5.8 7.1 110 262 34.7 20.0 14.7 80 Site 13 9.1, 3 7.6 65 162 16.9 12.0 4.9 30 Culture Control ?, 1 7.5 79 215 20.9 14.0 6.9 20 218 Appendix 11 Cont Sept. 17 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L CaC03) Cond. OuS/cm) TDOC (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 8.7, 4 7.3 90 280 28.2 17.4 10.8 40 Site 2 8.9, 5 7.2 70 242 24.8 13.5 11.3 65 Site 3 8.6, 5 7.2 75 242 25.5 13.9 11.6 55 Site 4 9.3, 5 7.1 62 180 22.1 9.0 13.1 70 Site 5 8.9, 5 7.0 68 195 22.9 8.0 14.9 130 Site 6 8.7, 5 7.1 60 186 25.5 10.5 15.0 80 Site 7 8.4, 5 7.0 74 212 23.6 9.1 14.5 70 Site 8 7.7 6.8 123 340 27.8 10.2 17.6 90 Site 10 9.1 7.1 85 204 34.9 7.4 27.5 160 Site 11 6.4 6.9 85 274 51.2 17.8 33.4 200 Site 12 8.4, 4 7.0 85 206 35.8 13.1 22.7 110 Site 13 9.1, 1 7.2 50 113 27.2 6.1 21.1 130 Culture Control ?, 1 7.5 83 237 22.5 12.5 10.0 25 Oct. 3 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial pH Total Hardness (as mg/L CaC03) Cond. OuS/cm) TDOC (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 7.9, 5 6.9 100 305 27.4 13.0 14.4 70 Site 2 8.3, 5 6.9 89 275 29.3 12.0 17.3 90 Site 3 8.2, 5 6.9 87 280 27.7 12.1 15.6 80 Site 4 8.8, 5 6.8 104 334 29.9 12.3 17.7 90 Site5 8.2, 5 6.8 178 545 30.1 12.3 17.8 100 Site 6 8.7, 5 6.9 83 263 31.1 12.4 18.7 90 Site 7 8.7, 5 6.9 80 255 37.6 12.4 25.2 80 Site 8 7.2, 5 6.4 209 560 22.6 5.0 17.6 80 Site 10 8.0, 5 6.9 114 340 36.4 12.9 23.5 130 Site 11 6.7, 5 6.8 149 410 54.4 14.0 40.4 410 Site 12 7.3, 5 6.9 130 315 38.3 14.4 23.9 100 Site 13 9.2, 1 7.1 62 160 26.8 10.9 15.9 160 Culture Control ?, 1 7.0 81 220 20.9 12.0 8.9 30 219 Appendix 11 Cont. Oct. 3 1997 Retests Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L C a C 0 3 ) Cond. GuS/cm) T D O C (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 2 7.6, 6 7.1 89 275 29.3 12.0 17.3 90 Site 6 8.1, 5 7.2 83 263 31.1 12.4 18.7 90 Site 13 8.5, 3 7.4 62 160 26.8 10.9 15.9 160 Culture Control ?, 1 7.2 80 ? ? ? ? ? Oct. 16 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L C a C 0 3 ) Cond. (jUS/cm) T D O C (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 5 9.0, 2 6.4 215 570 ? ? ? ? Site 10 Actual Farm Puddle 7.4, 6 5.6 115 364 ? ? ? ? Site 13 8.5, 3 7.4 62 160 26.8 10.9 15.9 160 Culture Control ?, 1 7.2 80 ? ? ? ? ? Nov. 21 1997 Initial DO (mg/L), 1st Day of 1 Aeration Initial PH Total Hardness (as mg/L C a C 0 3 ) Cond. (pcS/cm) T D O C (mg/L) TDIC (mg/L) TDIC (mg/L) Color (mg/L Pt) Site 1 10.0, 1 7.3 67 204 28.8 9.3 19.5 100 Site 2 10.1, 1 7.2 68 210 30.1 8.6 21.5 100 Site 3 10.2, 1 7.2 68 190 27.6 9.2 18.4 100 Site 4 10.2, 1 7.3 63 186 29.7 9.4 20.3 100 Site 5 9.0, 2 6.8 181 550 37.6 11.3 26.3 100 Site 6 10.0, 1 7.2 58 190 26.3 9.3 17.0 100 Site 7 10.2, 1 7.3 64 204 28.1 9.6 18.5 120 Site 8 9.1, 2 6.3 130 375 27.2 6.8 20.4 120 Site 10 10.6, 1 7.3 93 292 34.5 11.3 23.2 120 Site 11 8.7, 2 6.7 119 310 51.5 11.4 39.9 210 Site 12 9.8, 1 7.1 109 255 32.9 10.9 22.0 90 Site 13 10.6, 1 7.5 50 140 21.7 9.8 11.9 60 Culture Control ?, 1 7.5 83 220 22.3 14.0 8.3 20 220 221 Appendix 13 Sediment Bioassays' Initial and Final Overlvina Water Chemistry Sample Initial Total NH3-N mg/L Final Total NH3-N mg/L Initial Cond. //S/cm Final Cond. //S/cm Initial Total Hard. mg/L as CaC03 Final Total Hard. mg/L as CaC03 D.O. Range mg/L Initial PH Final pH Temp. Range °C Site 1 0.34 0.00 189 330 67 115 5 . 6 - 7 . 5 7.3 7.3 24.2 - 26.0 Site 2 0.94 0.25 189 400 66 136 3.6-7.1 7.2 7.6 24.0 - 24.7 Site 5 0.07 0.09 302 705 97 260 4 . 7 - 7 . 0 7.3 7.3 24.0 - 25.0 Site 7 1.28 0.18 176 286 66 90 3 .6 -6 .9 7.4 7.5 24.0 - 25.0 Site 8 5.80 0.20 342 440 80 90 3 .3 -6 .6 7.4 7.2 24.0 - 24.9 Site 10 1.12 0.06 256 350 77 100 5.1 -6.1 7.3 7.6 23.9 - 24.9 Site 11 1.36 0.29 284 445 76 84 5 .9 -7 .2 7.6 7.7 23.7 - 24.2 Site 12 1.76 0.08 202 322 63 90 3.3-7.1 7.5 7.6 24.0 - 25.0 Site 13 0.39 0.09 120 162 50 50 4 .4 -7 .1 7.4 7.1 24.3 - 25.0 

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