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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  degree  at  this  the  thesis  in  University of  partial  fulfilment  of  of  department  this or  publication of  thesis for by  his  or  her  representatives.  of  Kng-in^-ring  The University of British Columbia Vancouver, Canada  Date  DE-6 (2/88)  It  this thesis for financial gain shall not  civil  A p r i l 28, 1998  for  an advanced  Library shall make it  agree that permission for extensive  scholarly purposes may be  permission.  Department  requirements  British Columbia, I agree that the  freely available for reference and study. I further copying  the  is  granted  by the  understood  head of  that  be allowed without  copying  my or  my written  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 NH -N which was believed to 3  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 contamination. In 3  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 concentrations well below the water quality guidelines 3  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 O P s . 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 w a s 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 LIST OF TABLES LIST OF FIGURES ACKNOWLEDGMENTS 1.0  2.0  3.0  ii viii xi xii  INTRODUCTION  1  1.1 Study Objectives  2  1.2 The Nicomekl River and Watershed 1.2.1 River Modifications 1.2.2 River Flow 1.2.3 Dissolved Oxygen Concentrations 1.2.4 Possible Sources of Nicomekl River Contaminants 1.2.5 Nicomekl River Fish Resources  4 4 6 10 14 15  1.3 Study Area  19  AGRICULTURAL LAND USE IN THE STUDY AREA  21  2.1 Predominant Crops Grown  21  2.2 Drainage Ditch Water Management  21  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 3.3.1 Acute Lethality to Fish and Invertebrates of Pesticides Likely Used on the Nicomekl Farmlands 3.3.2 Sublethal Effects of Pesticides to Fish 3.3.3 Sublethal Effects of Pesticides to Invertebrates 3.3.4 Environmental Factors Affecting Pesticide Toxicity 3.3.5 The Chemistry, Use, and Persistence of Diazinon 3.3.6 The Chemistry, Use, and Persistence of Chlorpyrifos 3.3.7 The Chemistry, Use, and Persistence of Prometryn  34 34 40 43 47 49 51 52  3.4 Ammonia Contamination 3.4.1 Farm Animals 3.4.2 Fertilizer  53 53 56  V  Page 3.5 Literature Review of Ammonia Toxicity 3.5.1 Terrestrial Sources and Speciation of Ammonia in the Aquatic Environment 3.5.2 Fate of Ammonia in the Aquatic Environment 3.5.3 Toxicity to Fish 3.5.4 Toxicity to Invertebrates  58 58 58 60 65  3.6 Pesticide, Ammonia, and Metal Water Quality Guidelines for the  4.0  5.0  Protection of Aquatic Life  68  3.7 The Relevance of the C. dubia Test  72  REVIEW OF PREVIOUS CHEMICAL AND BIOLOGICAL TESTING IN THE STUDY AREA  73  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 5.3.1 C. dubia Culture 5.3.2 Initial Full Strength (100%) Ditch and River Water Sample Chronic Testing for Inhibition of Survival and Reproduction . . . 5.3.3 Dilution Series Chronic Testing of Lethally Toxic Samples . . . . 5.3.4 Lethally Toxic Samples Biological Toxicity Identification Evaluation (TIE) 5.3.5 Toxicity Tests on Diazinon, Chlorpyrifos, and Prometryn 5.3.6 Reference Toxicant Testing.and Culture Health 5.3.7 Test Endpoints and Statistical Analyses 5.4 Chemical Analyses of Ditch and River Water Samples  88 88  5.4.1 5.4.2 5.4.3 5.4.4 5.4.5  Organics Metals Total Ammonia Dissolved Total/Inorganic/Organic Carbon Dissolved oxygen, pH, Conductivity, Hardness, and Colour . .  89 92 92 94 96 97 98 99 102 103 103 104  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 5.7.1 Test Method  106 107  5.7.2 Test Endpoints 5.8 Chemical Analyses of Sediment Samples  6.0  108 108  5.8.1 Percentage Organic Matter  108  5.8.2 Metals  108  RESULTS AND DISCUSSION  109  6.1 Water Samples' Ceriodaphnia dubia Chronic Toxicity Test Results . . . . 109 6.1.1 Lethally Toxic Samples 6.1.2 Sublethally Toxic Samples 6.1.3 C. dubia Culture Health and Reference Toxicant Evaluation . . 6.2 Lethally Toxic Samples' Biological Toxicity Identification Evaluation . . . 6.3 Ditch and River Water Samples' Chemical Analyses 6.3.1 Organics 6.3.2 Metals 6.3.3 Total Ammonia 6.3.4 Dissolved Total/Organic/Inorganic Carbon, pH, Conductivity, Hardness, and Colour  114 115 117 119 123 123 131 134  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  136  171  VII  Page APPENDIX 1  Study Site Drainage Ditches' Municipal Pump Station Records  APPENDIX 2  Statistical Comparison of C. dubia Chronic Survival  188  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 C. dubia Lethally Toxic Samples' Identified Diazinon and Chlorpyrifos Select-lon-Scan Chromatogram Peaks and Mass Spectra Diazinon and Chlorpyrifos Standards' Select-lon-Scan  APPENDIX 7 APPENDIX 8  206 208  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  APPENDIX 12  Samples'Chemistry South Cloverdale Ditch .& Rapid Mixing of South Cloverdale Ditch Water Into the Nicomekl River (Photos)  APPENDIX 13  Sediment Bioassays' Initial and Final Overlying Water Chemistry  215 220 221  VIII  LIST OF TABLES Table 1.  Page  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 Maximum 4-Day Average Total Ammonia.Concentration for the  70  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  Table 10b.  ix LIST OF TABLES Cont. Table 16.  Page 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) Results of 96-h C. dubia Piperonyl Butoxide Toxicity Identification on Lethally Toxic Samples (Stored at 4°C in glass for 5 months)  120  Table 20.  Table 21.  121  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 Sediment Samples Appearance, Odour, and Visible Indigenous Invertebrates, Prior to Toxicity Testing  138 152  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  Table 26.  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  Figure 2.  Rainfall Near the Study Area, May to November, 1997  Figure 3.  3 rcl Daily Nicomekl River Flow (m Is) at 203 St., 3 km Upstream of the Study Area From May to October, 1997  5 7-8  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 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.  Figure 15.  141  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  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 ammoniacontaminated 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 2 3 2 St. and 5 2 nd  nd  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 /s(EC, 1998). 3  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).  Precipitation Record  Precipitation Record  Cloverdale East, B C  Cloverdale East, B C  X* • Sampling  40  j | || - j, ii Ii  j  llso  I  c o 120  fio  ! ! | j i I | ! ' ! :1  • 3  ! i ! j I '•I j t r -—4-  • -M • 1  -1  i •  i;  f  i  S  7  I  i  .  1  !  1 •  n  May, 1997  I•  n  i  Ii1 n i l3n  5  7  9  June, 1997 Precipitation Record Cloverdale East, B C  i JL.i  ;i i  X Sampling =  An  I  E E,30  1  |  1  B O •520  I IS 20  .  fio  |  i  a.  |  (1 5  7  t  X  2i 27 July, 1997  11 13 IS .17 IS 231 23  9  j j|  tm  29 31  a. n  1  n I 3  5  7  9  t  Precipitation Record Cloverdale East, B C  !  I ! j \• • I iji | ,-| •i j| i s 1  3  5  |  1  7  I  9  —  , j  1  ••t~"  ; ' I  X ~ Sampling  ;  'l | J j i  I j  t  1 i  :  I10  23 2 5 27 29 31  Cloverdale East, B C 1  1120  e 21  13 15 17 19  —,  _  Precipitation Record 40  c O  11  j|  August, 1997  X = Sampling  |I30  I  j  V III I10  ....  II I  I  !  T" j  3  1 o_ n nn 11 13 15 17 19 21 23 2S 27 2S  Cloverdale East, B C  !  1  I  T  I  Precipitation Record  | ,  E E.30 c o  i i  i  X = Sampling  40  l  I  i  i  is 20 o.  I  ] |  E" E,30 c o  fio  11 13 15 17 1£ 21 23 25 2 7 29 31  9  X = Sampling  40  I i I  M —1 - Ul | |  •h I n 1  M  1IIn  1  ~  •—j—  n 11 13 15 17 19 21 23  September, 1997  1  n f— n  It  25 27 29  fllL 111  l J n 1  3  t  5  7  9  I  11 13 15 17 19 21 2 3 2 5 27 29 31  October, 1997  t  8  Figure 2 Cont.  Precipitation Record Cloverdale East, B C f = Sampling An  i  ....  •  E E^n e o  | _  I i  f 10 -rr 0.  •U 11 ml  1  L_  i  i  £  i  I !  ir " i  El  3 5  i  1 7  •j f -  I  i  i  ,  |  I  i !  I I !  II1 •  i !  iI  j  (  i  i J  i  i |  1  dJ I  J_ i  i  - 1 1 Li | Ji. J • 1.1 n- .t  ul  i  |  uk  It  9 11 13 15 17 19 21 23 25 27 29  November, 199 't  Figure 3.  Daily Nicomekl River Flow (m /s) at 203 St., 3 km Upstream of the Study Area From May to October, 1997 (EC, 1998). 3  rd  Nicomekl River Flow  Nicomekl River Flow  May, 1997  June, 1997  Y = Sampling  y = Sampling 16 14  i  j I  -3T12 <|10 •  il 1  3  4 2 0  5 7 9 11 13 15 17 19 21 23 25 27 29 31 May  t  r"  :  <  \1  1  ...  r'  \  \  3  45  Nicomekl River Flow  7  9 11 13 15 17 19 21 23 25 27 29 June f  Nicomekl River Flow  July, 1997  August, 1997  y' = Sampling  J = Sampling 16 14 1»12 |10  il 1  3 5  7 9 11 13 15 17 19 21 23 25 27 29 31 t  4 2 0  U-i.  1  3  July  Nicomekl River Flow  5  7 9 11 13 15 17 19 21 23 25 27 29 31 f August f  Nicomekl River Flow  September, 1997  October (Incomplete), 1997  X = Sampling  J = Sampling  16 14 44-4  i s 2 0  1  3  5  7  9 11 13 15 17 19 21 23 25 27 29 September  f  1  3 f  5 7  9 11 13 15 17 19 21 23 25 27 29 31 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).  LEGENDE  WATER WELLS  DEPTH IN FEET  DEPTH IN METRES  50-75  15.3-22.9  PUITS D'EAU  o  75-125  12.9-38.1  O  125-175  3*.1-53.3  175-225  53.3-68,6  129.5-1*4.8  A-  Roads:  Routes:  hard surface, all weather  ami highway  pavle, toute salson...  mora than 2 linw  jj'chamriaa Upartm hard surface, all weather  pave*, toute saison  loose or stabilized surface, afl weather  gravier agglom«re\ toute saison  loose surface, dry weather  degravier.piriodeseche  cart track  de terra...  trail or portage  senUer ou portage  piui da 2 vain SgBS»-ffMlj molrn de 2 VOMM  2voJti  2itn-»ormof8 tathanaianu 2rainou pi in mdrnt da 2 volas  ildini  stop  Railway, normal gauge, singta track  Chemm de ter, voie unique{ecartement normal)..-^m-mJLam»  Horizontal control point, with elevation  Point geodesique, avec cote  Bench mark, with elevation  Repire de nh/ellement avec cote  Spot elevation, precise  Point cot*; precis  Mine or open cut  Mine ou fosse a del ouvert  Building  Bltiment  School  £cole  Cemetery  Cimetfire  •  Church  454 & BM 157-*  f i£j  W £glise  1  Post Office  Bureau de poste  p.  Navigation Light  Feu de navigation  *  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  • "  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  itaOoji  ^a  450  Rock; bare or awash  Lake intermittent indefinite  iff  EQUIDISTANCE OES G O U R D E S I O PI EDS EQUIDISTANCE DES C O U R S E S 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  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  # Samples  (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 (O2 reduction) and to a much 2  lesser extent H S. If low bottom water and sediment DO concentrations exist in the river's 2  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 m /day and maximum 430 m /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. 3  3  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 a n d Murray C r e e k s from 0.5 to 10.6 km a n d 1 to 8.6 km u p s t r e a m of their c o n f l u e n c e with the N i c o m e k l , respectively ( S w a i n & H o l m s , 1988b). M a n y of the ditches in the region follow old temporary or permanent stream b e d s . F i s h c a n m o v e into or out of the 5 large municipally controlled d r a i n a g e ditches only w h e n their flood b o x e s are o p e n (low tide or l o w river f l o w s c o i n c i d i n g high ditch water-levels).  In addition to S w a i n a n d H o l m s '  (1988b) identification of Erickson Ditch a s a salmonid spawning region, E V S (1993) stated that B. Clark, of the B . C . Ministry of the Environment, reported to t h e m that cutthroat trout have b e e n observed spawning in Ericson Ditch. S i n c e E r i c s o n Ditch is a k n o w n s a l m o n i d s p a w n i n g corridor, its p u m p e d d i s c h a r g e is by w a y of s c r e w p u m p s , to offer further protection to s a l m o n i d fry/smolts m o v i n g d o w n s t r e a m into the N i c o m e k l River. E r i c s o n ditch offers suitable s a l m o n habitat d u e to its consistent flow a n d proliferation into s e v e r a l branches in its headlands. T h e other 4 municipally controlled d r a i n a g e ditches are likely not s u i t a b l e s a l m o n s p a w n i n g habitats d u e to their shorter lengths, lower a n d irregular flows, a n d fine g r a i n e d s e d i m e n t s ( E V S , 1993). W h i l e s a l m o n i d s may not be s p a w n i n g in t h e s e ditches, other l e s s s e n s i t i v e bottom s p e c i e s w h i c h d o not d e m a n d a s rigorous f l o w s a n d high D.O. concentrations, a n d c a n s p a w n in silty s e d i m e n t s or o n ditch m a c r o p h y t e s , may b e s p a w n i n g here.  E V S (1993) o b s e r v e d the p r e s e n c e of carp,  Cyprinus carpio in the Nicomekl River. Bourque a n d Hebert (1982) f o u n d the three s p i n e d s t i c k l e b a c k (Gasterosteus (Richardsonius  aculeatus),  prickly s c u l p i n {Cottus asper), r e d s i d e s h i n e r  balteatus), lamprey {Lampetra richardsoni), b r o w n b u l l h e a d (Ictalurus  nebulosus), peamouth chub (Mylocheilus caurinum), crayfish, a n d frogs in the S e r p e n t i n e system.  T h e s e s p e c i e s are likely f o u n d in the N i c o m e k l R i v e r a s w e l l , c o n s i d e r i n g its  proximity a n d similar g e o g r a p h y a n d w a t e r chemistry.  T h i s author o b s e r v e d a catfish  (appeared to be the brown bullhead, Ictalurus nebulosus) caught in the N i c o m e k l R i v e r in  18 the study site, and s t i c k l e b a c k (not three-spine) in both the 1 6 8  t h  St. N E Ditch a n d O l d  L o g g i n g Ditch.  nd T h e Nicomekl River h a s a salmon hatchery located at 2 3 2  nd St. a n d 5 2  A v e . In the fall,  the hatchery s p a w n s its returned coho, red Chinook, white Chinook, pink, a n d c h u m s a l m o n (Rhidine, 1997). In April of e a c h year the Nicomekl Hatchery releases all its fish at roughly 6 months of age. In April of 1997 the hatchery released approximately 50,000 r e d Chinook, 2 9 , 0 0 0 white Chinook, 5 0 , 0 0 0 c h u m , 3 0 , 0 0 0 pink, a n d 2 9 , 0 0 0 c o h o (Rhidine, 1997). T h e hatchery only rears its fish from fall to spring. Hatchery r e l e a s e d pink a n d c h u m a r e at the fingerling s t a g e a n d s h o u l d migrate to the o c e a n immediately following r e l e a s e (Iwama, 1991).  Hatchery released  Chinook a r e c l o s e to the smolt s t a g e (Rhidine, 1997), a n d  should migrate to the o c e a n immediately o r during t h e s u m m e r months. T h e hatchery's c o h o are a l w a y s released at the fingerling stage. T h e 1997 c o h o w e r e r e l e a s e d at 2 - 2 . 5 g r a m s w e i g h t (Rhidine, 1997). T h e s e fish s h o u l d s p e n d b e t w e e n 1 a n d 2 y e a r s in the N i c o m e k l R i v e r and/or its tributaries prior to smoltification a n d migration to t h e o c e a n (Iwama, 1991). Typical c o h o returns are 600-800 fish (Rhidine, 1997). T h e river's natural cutthroat a n d s t e e l h e a d trout populations s h o u l d h a v e river rearing.times of 5 a n d 1 - 2 y e a r s , respectively (Iwama, 1991).  T h e r e exists the potential that juvenile s a l m o n i d s s e e k i n g refuge a n d rearing in the lower r e a c h e s of the N i c o m e k l R i v e r or its d r a i n a g e ditches, or smolting fish migrating through the lower reaches of the river to the o c e a n , c o u l d b e e x p o s e d to c h e m i c a l l y contaminated agricultural runoff. W h i l e Ericson Ditch is the only k n o w n s a l m o n i d s p a w n i n g ditch, s o m e of the other large drainage ditches in the river's lower r e a c h e s may s e r v e a s refuge a r e a s  19  for rearing juvenile s a l m o n a s w e l l . A s a l r e a d y m e n t i o n e d , s t i c k l e b a c k w e r e o b s e r v e d in the O l d L o g g i n g Ditch a n d the 1 6 8  t h  St. N E ditch.  Furthermore, fish f o r a g i n g in  contaminated ditches or river z o n e s may be reduced d u e to the agricultural r u n o f f s impact o n s e n s i t i v e invertebrate o r g a n i s m s .  T h e r e a l s o exists the possibility that in the fall  s p a w n i n g adult fish c o u l d c o m e in contact with toxic river w a t e r within or d o w n s t r e a m of the study site.  S i n c e the depletion of natural c o h o a n d Chinook s t o c k s is currently of paramount c o n c e r n to the Department of F i s h e r i e s a n d O c e a n s ( D F O ) , further e m p h a s i s is b e i n g p l a c e d o n b o t h the protection a n d hatchery production of this fish.  C o n s e q u e n t l y , w h e t h e r the  agricultural d r a i n a g e d i t c h e s in the N i c o m e k l R i v e r are d i s c h a r g i n g c o n t a m i n a n t s in c o n c e n t r a t i o n s affecting s a l m o n i d s or their f o o d s o u r c e o r g a n i s m s w a s o n e of the objectives promoting this study.  1.3  Study Area nd  T h e c h o s e n a r e a for this field investigation is located b e t w e e n 152 nd b e t w e e n 32  th a n d 184  St., a n d  th and 48  A v e . (Figure 6). T h i s region w a s c h o s e n d u e to its e x t e n s i v e  agricultural u s e , a b u n d a n c e of d r a i n a g e ditches, lack of p e r m e a b l e soil a n d high runoff, e a s e of site accessibility, location of the known salmonid spawning a r e a at the h e a d w a t e r s of Erickson Ditch, a n d p r e v i o u s w a t e r a n d s e d i m e n t s a m p l i n g by C o a s t l i n e (1989), W a n (1989), E V S (1993), a n d W a n et. al (1994, 1995) w h i c h r e v e a l e d the p r e s e n c e of p e s t i c i d e s in this region.  20  21  2.0  AGRICULTURAL LAND USE IN THE STUDY AREA  2.1  Predominant Crops Grown  T h e study a r e a is predominantly u s e d for commercial vegetable production, a n d to a l e s s e r extent, c o m m e r c i a l berry production, flower n u r s e r i e s , a n d h o b b y v e g e t a b l e farms. T h e v e g e t a b l e s o b s e r v e d to b e c o m m e r c i a l l y g r o w n during the study p e r i o d i n c l u d e d c o m , potatoes, lettuce, o n i o n s , carrots, beets, p u m p k i n s a n d z u c c h i n i .  Berry c r o p s w e r e  exclusively blueberries. F i g u r e 7 s h o w s the locations of the p r e d o m i n a n t v e g e t a b l e a n d blueberry c r o p s g r o w n in the study a r e a n e a r the d r a i n a g e d i t c h e s in J u l y of 1997.  2.2  Drainage Ditch Water Management nd  T h e r e are 5 large runoff/irrigation d i t c h e s in the study a r e a , b e t w e e n 1 5 2  th St. a n d 1 8 4  St., w h i c h e a c h h a v e their d i s c h a r g e a n d irrigation controlled by municipal p u m p i n g stations located at the junction between the ditches a n d the Nicomekl River. T h e s e 5 m a i n ditches are specifically " S o u t h C l o v e r d a l e Ditch", " E r i c s o n Ditch", " B u r r o w s Ditch", " O l d L o g g i n g Ditch", a n d " H a l l s Prairie Ditch" (Figure 7).  If a d i t c h e s w a t e r - l e v e l s b e c o m e  h i g h e r than the river's water-level, s u c h a s after significant rainfall a n d runoff e v e n t s , and/or at p e r i o d s of low tide w h e n the river's tidal g a t e s a r e o p e n e d a n d the river w a t e r level drops, then d i s c h a r g e from the d i t c h e s c a n o c c u r s by hydraulic h e a d through flood b o x e s at the p u m p i n g stations.  N o m u n i c i p a l r e c o r d s a r e kept for ditch d i s c h a r g e by  gravity through the flood b o x e s . If ditch w a t e r - l e v e l s are high e n o u g h that d i s c h a r g e is n e c e s s a r y to prevent farm l a n d from b e i n g f l o o d e d , but the river w a t e r - l e v e l is a l s o h i g h d u e to runoff upstream of the study site a n d / o r the c l o s u r e of the river's tidal g a t e s at high tide, t h e n gravity d i s c h a r g e is not p o s s i b l e a n d t h e s e 5 d i t c h e s a r e m e c h a n i c a l l y d i s c h a r g e d by p u m p s .  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 m /s, 3.98 m /s, 1.1 m /s, 2.26 m /s, and 3  3  3  3  3  1.9 m Is, respectively (Lalond, 1998). This is assuming all of the 2-3 pumps at each station are in operation. Single pump discharge flows are 1.2 m3/s, 2.0 m3/s, 0.55 m3Is, 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 waterlevel 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 structuralwood 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 T a b l e 3.  Most Probable Pesticides U s e d on the N i c o m e k l F a r m l a n d s during the 1997 G r o w i n g S e a s o n ( D e s c e n d i n g O r d e r of E s t i m a t e d A m o u n t of E a c h U s e d ) .  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 preemergence 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 preemergence  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 preemergence  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 postemergence 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 postemergence  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 preemergence  peas  31 Thiram  Fungicide  Hexazinone  Herbicide  Diclofop methyl  Herbicide  Pendimethalin  Herbicide  Seed treatment  corn, beans, peas  Sprayed when necessary  blueberries  Propionic acid (fatty-acid synthesis inhibitor)  Sprayed when necessary  beets, beans, peas  Dinitroanaline (cell division inhibitor)  Sprayed preemergence or postemergence  com  Dimethyldithiocarbamate  E V S (1993) conducted interviews with farmers in the study a r e a during the 1 9 8 9 growing s e a s o n . T h e farmers indicated that they u s e d glyphosate (applied prior to S p r i n g plowing), c h l o r p r o p h a m ( u s e d a s a top-killer) a n d chlorpyrifos ( E V S , 1993).  C h l o r p r o p h a m is  currently only r e c o m m e n d e d for u s e a s spot treatment against d o d d e r in alfalfa f o r a g e c r o p s ( B C M A F F , 1996a), h a d z e r o s a l e s in B . C . in 1995, a n d w a s unlikely u s e d in the study site in 1997. T h e author o b s e r v e d p e s t i c i d e s p r a y i n g o n J u l y 1 6 , 1997, of a white t h  mist (tractor-tank-boom setup) on potato crops located adjacent to B u r r o w s ditch, Site 11, th o n the field south of 4 0  A v e . O v e r - s p r a y into B u r r o w s ditch w a s o b s e r v e d at this time.  S a m p l i n g of B u r r o w s Ditch w a s performed o n this date. observed on July 29  th  Similar spraying w a s also  , 1997, o n the potato c r o p s to the north of 4 0  b e t w e e n the O l d L o g g i n g Ditch a n d 176  th  th  Ave., between  St. N o s a m p l e s w e r e c o l l e c t e d o n J u l y 2 9  th  .  F o r the study a r e a , the most a b u n d a n t l y u s e d i n s e c t i c i d e s are o r g a n o p h o s p h a t e a n d carbamate compounds.  T h e most a b u n d a n t l y u s e d h e r b i c i d e s are more v a r i a b l e in  c h e m i c a l c o m p o s i t i o n , but are primarily nitrogenous c o m p o u n d s s u c h a s a m i d e s a n d triazines.  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)  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 %  % Change  Class  Carb. Insect.  33 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 %  Methomyl  167  401  + 140%  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  178, 382  213,678  + 20 %  Pesticides  Dimethoate  TOTAL  Herbicide Carb. Insect.  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  Reference  (Mg/L) Glyphosate 12g/L(25 °C)  Caotan 3.3 mg/L (25 °C)  Atrazine 30 mg/L (20 °C)  Malathion 145 mg/L (room temp.)  Mancozeb 6-20 mg/L  Rainbow trout 96-h LC50 = 86,000 (BCPC, 1991) Rainbow trout (0.8 g) 96-h LC50 = 130,000 * (Johnson & Finley, 1980) Rainbow trout (0.8 g) 96-h LC50 = 130,000 * (Mayer & Ellersieck, 1986) Rainbow trout (1.0 g) 96-h LC50 = 8,300 (R) (Johnson & Finley, 1980) Rainbow trout 96-h LC50 = 50,000 (Folmar, 1976) Rainbow trout 96-h LC50 = 54,800 (Hildebrand etal., 1982) Channel catfish (2.2 g) 96-h LC50 - 130,000 * (Johnson & Finley, 1980) Daphnia 48-h LC50 > 780,000 (BCPC, 1991) (Johnson & Finley, 1980) Daphnia magna 48-h EC50 = 3,000 (R) (Mayer & Ellersieck, 1986) Daphnia magna 48-h LC50 = 2,950 (Mayer & Ellersieck, 1986) Midge, Chironomus plumosus 48-h LC50 = 55,000 (R) Roundup formulation, 3 -42 times more toxic than technical grade  (Johnson & Finley, 1980)  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)  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)  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 *  (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) & Finley, 1980) Amphipod, Gammarus asciatus 96-h LC50 = 0.76 (Johnson * Amphipod, Gammarus lacustris 48-h LC50 = 1.8 (Cope, 1966) (Cope, 1966) Stonefly, Simocephalus sp. 48-h LC50 = 3.0 (Cope, 1966) Mayfly, Baetis sp. 48-h LC50 = 6.0 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  (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)  Ceriodaphnia dubia 48-h LC50 Ceriodaphnia dubia 24-h LC50 Ceriodaphnia dubia 48-h LC50 Ceriodaphnia dubia 96-h LC50  (Amatoetal., 1992) (Bailey et al., 1997) (Bailey et al., 1997) (Bailey et al., 1997) (Bailey et al., 1997)  = 0.35 = 0.37 - 0.75 = 0.26 - 0.58 = 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 (Katznelson & Mumley,'97) & Finley, 1980) Amphipod, Gammarus fasciatus 96-h LC50 = 0.2 (Johnson * Amphipod, G. pseudolimnaeus 48-h LC50 = 4 (Morgan, 1976) Amphipod, Gammarus lacustris 96-h LC50 = 170 (Morgan, 1976) (Morgan, 1976) Amphipod, Hyallela azteca 48-h LC50 = 22 (FWPCA, 1968) Stonefly, Pteronarcys sp. 48-h LC50 = 60 (Morgan, 1976) Midge, Chironomus tentans 48-h LC50 = 0.1 (Morgan, 1976) Isopod, Asellus communis 96-h LC50 = 21 (Morgan, 1976) Mollusc, Helisoma trivolvis 7-day LC50 = 528 (Morgan, 1976) Crayfish, Orconectes propinquus 48-h LC50 = 537  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 *  (BCPC, 1991) (Holcombe, et al., 1982) (Maceketal., 1969) (Mayer & Ellersieck, 1986) (FWPCA, 1968) (Maceketal., 1969) (Johnson & Finley, 1980) (Johnson & Finley, 1980)  Daphnia pulex 48-h EC50 = 1.78  (Wood, 1997)  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  (Bailey etal., 1997) (Bailey et al., 1997) (Bailey et al., 1997)  (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,  (Bailey et al., 1997)  & Finley, 1980) Gammarus lacustris 96-h LC50 = 0.11 (Johnson *  Amphipod, Gammarus lacusths 24-h LC50 = 0.76(Sanders, 1969) 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  Benomvl 4 mg/L (25 °C)  Napropamide  (Moore et al., 1998) (Phipps et al., 1995) (Sanders & Cope, 1968) (Moore et al., 1998) (Ankley et al., 1994a)  Rainbow trout (1.2 g) 96-h LC50 = 170 * (Johnson & Finley, 1980) Rainbow trout 96-h LC50 = 170 (BCPC, 1991) Rainbow trout (1.0 g) 96-h LC50 = 310 (Johnson & Finley, 1980) (Johnson & Finley, 1980) Channel catfish (1.2 g) 96-h LC50 = 28 * (Mayer & Ellersieck, 1986) Daphnia magna 48-h LC50 = 2,800 (Mayer & Ellersieck, 1986) Amphipod, G. Pseudolimnaeus 96-h LC50 = 750 Midge, Chironomus plumosus 48-h EC50 = 7,000(Mayer & Ellersieck, 1986) Rainbow trout 96-h LC50 = 16,600  (BCPC, 1991)  Rainbow trout 96-h LC50 = 51,000 Carp 96-h LC50 = 68,000  (BCPC, 1991) (Chin & Sudderuddin,1979)  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, (BCPC, 1991) (Johnson & Finley, (Johnson & Finley, (Johnson & Finley,  1980)  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, (BCPC, 1991) (Johnson & Finley, (Johnson & Finley, (FWPCA, 1968) (Johnson & Finley, (Sanders, 1969)  1980)  73 mg/L (20 °C)  Methamidophos > 200 g/L (20 °C)  Dichlobenil 18 mg/L (20 °C)  Paraquat very soluble  1980) 1980) 1980)  1980) 1980) 1980)  38  Parathion (Methvl) 11 mg/L (20 °C)  Rainbow trout (1.1 g) 96-h LC50 = 3,700 * (Johnson & Finley, 1980) Rainbow trout 96-h LC50 = 2,750 (Macek & McAllister, 1970) Rainbow trout 96-h LC50 = 2,700 (BCPC, 1991) Rainbow trout 96-h LC50 = 2,800 (Palawski et al., 1983) Cutthroat trout (0.2 g) 96-h LC50 = 1,850 * (Johnson & Finley, 1980) Coho (1.0 g) 96-h LC50 = 5,300 (Macek & McAllister, 1970) Channel Catfish (1.4 g) 96-h LC50 = 5,240 (Johnson & Finley, 1980) Channel Catfish 96-h LC50 = 5,710 (Macek & McAllister, 1970) Carp (0.6 g) 96-h LC50 = 7,130 (Macek & McAllister, 1970) Daphnia magna 48-h LC50 = 0.14 * (Johnson & Finley, 1980) (Mayer & Ellersieck, 1986) Daphnia pulex 48-h EC50 = 0.60 Amphipod, Gammarus fasciatus 96-h LC50 = 3.8 (Johnson & Finley, 1980)  Linuron  Rainbow trout 96-h LC50 = 16,000  81 mg/L (24 °C)  Daphnia magna 48-h EC50 = 270 (Mayer & Ellersieck, 1986) Midge, Chironomus plumosus 48-h EC50 = 2,900(Mayer & Ellersieck, 1986)  Dicamba  Rainbow trout (0.8 g) 96-h LC50 = 28,000 * (Johnson & Finley, 1980) Rainbow trout 96-h LC50 = 135,000 (BCPC, 1991) (Bohmont, 1967) Rainbow trout 48-h LC50 = 35,000 (Bond etal., 1965) Coho 24-h LC50 = 151,000 (Johnson & Finley, 1980) Daphnia magna 48-h LC50 > 100,000 * (Sanders and Cope, 1966) Daphnia pulex 48-h LC50 = 11,000 Amphipod, G. fasciatus 96-h LC50 > 100,000 * (Johnson & Finley, 1980) (FWPCA, 1968) Amphipod, Gammarus lacustris 48-h LC50 = 5,800 (Johnson & Finley, 1980) Isopod, Asellus 96-h LC50 > 100,000 *  6.5 g/L (25 °C)  (BCPC, 1991)  MCPA Amine  Rainbow trout 96h-LC50 = 117,000  825 mg/L (room temp.)  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  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)  Salmon 48-h LC50 = 3,250,000 Daphnia magna E C = 23,000  (Bohmont, 1967) (Crosby & Tucker, 1966)  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)  620 mg/L (25 °C)  Amitrole 280 g/L (25 °C)  Carbofuran 320 mg/L (25 °C)  39 Dimethoate 25 g/L (21 °C)  Diquat 700 g/L (20 °C)  Carbaryl 40 mg/L (30 °C)  FluazifoD-butvl  Rainbow trout (1.5 g) 96-h LC50 = 6,200 * (Johnson & Finley, 1980) Rainbow trout (1.5 g) 96-h LC50 = 8,600 (Mayer & Ellersieck, 1986) Rainbow trout 24-h LC50 = 19,000 (Alabaster, 1969) Amphipod, Gammarus lacustris 96-h LC50 = 200 (Johnson * & Finley, 1980) Amphipod, Gammarus lacustris 24-h LC50 = 900 (Sanders, 1969) Daphnia magna 48-h LC50 = 3,320 (Song, et al., 1997) Daphnia magna 48-h LC50 = 2,500 (FWPCA, 1968) Stonefly, Pteronarcys sp. 48-h LC50 = 140 (Cope, 1965b) 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  (BCPC, 1991) (Folmar, 1976) (FWPCA, 1968) (Bohmont, 1967) (BCPC, 1991) (Crosby & Tucker, 1966) Daphnia magna EC = 7,100 Amphipod, G. fasciatus 96-h LC50 > 100,000 * (Johnson & Finley, 1980) Rainbow trout (1.5 g) 96-h LC50 = 4,340 (Macek & McAllister, 1970) Rainbow trout 96-h LC50 = 1,300 (BCPC, 1991) Rainbow trout 96-h LC50 = 1,100 (Mayer & Ellersieck, 1986) Cutthroat trout (0.5 g) 96-h LC50 = 7,100 * (Johnson & Finley, 1980) Coho (1.0 g) 96-h LC50 = 764 (Macek & McAllister, 1970) (Macek & McAllister, 1970) Channel catfish (1.5 g) 96-h LC50 = 15,800 (Mayer & Ellersieck, 1986) Channel catfish (1.5 g) 96-h LC50 = 7,790 (Macek & McAllister, 1970) Carp (0.6 g) 96-h LC50 = 5,280 * (Chin & Sudderuddin, 79) Carp 96-h LC50 = 1,700 (Johnson & Finley, 1980) Daphnia pulex 48-h EC50 = 6.4 * Amphipod, Gammarus fasciatus 96-h LC50 = 26 *(Johnson & Finley, 1980) Amphipod, Gammarus lacustris 96-h LC50 = 22 *(Johnson & Finley, 1980) (Sanders & Cope, 1968) Stonefly, Pteronarcella badia 24-h LC50 = 5.0 (Johnson & Finley, 1980) Isopod, Asellus 96-h LC50 = 280 * Rainbow trout 96-h LC50 = 1,370 Mirror Carp 96-h LC50 =1,310  (BCPC, 1991) (BCPC, 1991)  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)  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)  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  Rainbow trout 96-h LC50 > 180,000  33 g/L (25 °C)  Dapfcn/a 48-h LC50 = 151  (Mayer & Ellersieck, 1986) (BCPC, 1991)  1 mg/L  Methomyl 58 g/L (25 °C)  Prometryn 33 mg/L (20 °C)  Thiram 30 mg/L (room temp.)  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 r e a s d r a i n a g e d i t c h e s or the N i c o m e k l R i v e r itself a r e more f e a s i b l e than lethal effects, g i v e n that m u c h lower c o n c e n t r a t i o n s of p e s t i c i d e s are required.  W e i s s (1961) reported that fish e x p o s e d to 100 / / g / L O P  insecticides s h o w e d inhibition of brain a c e t y c h o l i n e s t e r a s e , a n d that regeneration of this e n z y m e to normal l e v e l s m a y take up to 3 0 d a y s .  W i t h r e s p e c t to d i a z i n o n , W e i s s  demonstrated that largemouth b a s s a n d f a t h e a d m i n n o w s e x p o s e d to 5 0 0 / / g / L d i a z i n o n that d i d not d y e h a d brain a c e t y c h o l i n e s t e r a s e activity r e d u c e d by 8 2 % . W e i s & W e i s (1975) found that regeneration of removed caudal fins of the killifish  Fundulus heteroclitus  w a s retarded after two w e e k s of exposure to only 1 0 / / g / L o f the O P i n s e c t i c i d e s malathion a n d parathion, a n d the carbamate insecticide carbaryl. D o d s o n & M a y f i e l d (1979) s h o w e d that r a i n b o w trout e x p o s e d to 5 0 0 / / g / L of the herbicide diquat for 2 4 hours s h o w e d a significant d e c r e a s e in s w i m m i n g s p e e d s , a n d modifications in rheotaxis, resulting in a n i n c r e a s e d i n c i d e n c e of d o w n s t r e a m drift. Hatfield a n d A n d e r s o n (1972) found that after 2 4 h o u r s of e x p o s u r e to 1,000 / / g / L of the O P insecticide fenitrothion, Atlantic s a l m o n  (Salmo salar) parr w e r e more vulnerable than u n e x p o s e d fish to predation by large brook trout  (Salvelinus fontinalis). S y m o n s (1973) reported that Atlantic s a l m o n e x p o s e d to 100  / / g / L a n d 1 0 0 0 / / g / L fenitrothion for 15 hours s h o w e d a 2 0 % a n d 5 0 % d e c r e a s e in the n u m b e r of fish holding territories 6 d a y s following the treatment, respectively. S y m o n s (1973) a l s o s h o w e d that y o u n g Atlantic s a l m o n regurgitated f o o d that w a s c o n t a m i n a t e d w i t h fenitrothion.  Bull a n d M c l n e r n e y (1974) found that c o h o  (Oncorhynchus kisutch)  e x p o s e d to fenitrothion for 2 h o u r s at c o n c e n t r a t i o n s of 100 to 2 3 0 / i g / L exhibited d e c r e a s e d feeding a n d those e x p o s e d to 4 8 0 to 750 / / g / L exhibited d e c r e a s e d locomotion a n d inability to maintain position 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-been demonstrated at concentrations much lower r  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. T h e number of insects colonizing plate s a m p l e r s w a s r e d u c e d by 7 5 % , mayfly colonization w a s severely reduced, a n d caddisfly colonization w a s completely eliminated.  L i k e w i s e , the 2.39 / / g / L concentration of chlorpyrifos killed 5 5 % of the 2 7 5  bluegills a n d 4 6 % of the 2 7 5 largemouth b a s s w h i c h h a d b e e n introduced to the p o n d for the test. Muirhead-Thomson (1978) f o u n d the a m p h i p o d  Bammarus pulex to be the most  prone to c h a n n e l drift at sublethal concentrations of chlorpyrifos.  P u s e y et al. (1994) reported that a n artificial stream d o s e d for 6 hours with chlorpyrifos at 0.1 //g/L h a d no significant effect o n stream macroinvertebrate communities: h o w e v e r , 6 h o u r d o s i n g at 5.0 / / g / L did r e d u c e the a b u n d a n c e of s e v e r a l s p e c i e s . F o r s u b s e q u e n t testing, W a r d et al. (1995) reported that a n artificial stream e x p o s e d to chlorpyrifos for 21 d a y s at the s a m e concentrations exhibited a d e c r e a s e d n u m b e r of taxa a n d total i n v e r t e b r a t e a b u n d a n c e for both the high (5.0 //g/L) a n d low d o s e s (0.1 //g/L).  The  individual a b u n d a n c e s of 9 of the 36 n o n - c h i r o n o m i d taxa a n d 13 of the chironomid taxa c o u n t e d w e r e significantly r e d u c e d by the study's insecticide application ( W a r d et a l . , 1995).  E a t o n et a l . (1985) reported r e d u c e d taxon r i c h n e s s in a n artificial stream for  repeated acute exposures to 3.1 //g/L chlorpyrifos. a n d a 100 d a y e x p o s u r e to 0.22 //g/L.  W i t h r e s p e c t to O P pesticides,  C. dubia d o e s not a p p e a r to demonstrate significant  inhibition of reproduction to diazinon in the 7-day chronic test, other than that d u e to death of the o r g a n i s m s .  K a t z n e l s o n a n d M u m l e y (1997) o b s e r v e d that C.  dubia e x p o s e d to  urban runoff s a m p l e s from residential a n d mixed land u s e catchments w h e r e d i a z i n o n w a s s h o w n to be a major c a u s e of toxicity p r o d u c e d n u m e r o u s offspring before they d i e d , in s a m p l e s with d i a z i n o n concentrations low e n o u g h to a l l o w time for reproduction, t h o s e  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 s a m e organism often produce different toxicity endpoints d u e to variations in the c h e m i c a l formulation, a g e of organism, test apparatus, etc. A brief description of the environmental (and test) parameters w h i c h affect pesticide toxicity follows. M a n y p e s t i c i d e s a r e a v a i l a b l e for purchase by farmers in different forms. D i a z i n o n is a v a i l a b l e a s a wettable p o w d e r (in w h i c h diazinon is 5 0 % of the powder by weight) or a s a n emulsifiable concentrate (500 g/L diazinon).  In g e n e r a l , emulsifiable c o n c e n t r a t e s a r e m o r e toxic t h a n t e c h n i c a l g r a d e  material, w h e r e a s dust a n d powder formulations are l e s s toxic (Murty, 1986), likely d u e to their reduced solubility. Commercial pesticide products are often more toxic than t e c h n i c a l grade material a s the active ingredient is often c o m b i n e d with s y n e r g i s t s to i n c r e a s e the p r o d u c t s effectiveness.  F o r e x a m p l e , J o h n s o n a n d F i n l e y (1980) reported that the  h e r b i c i d e R o u n d u p w a s 3 - 4 2 times m o r e toxic t h a n the t e c h n i c a l g r a d e of its active ingredient, g l y p h o s a t e . M o s t toxicologists u s e technical g r a d e material for their testing; h o w e v e r , e v e n t e c h n i c a l g r a d e material will vary in its percent purity. T h e a g e of the chemical (or stock solutions) u s e d by the farmer or toxicologist m a y affect its toxicity d u e to h y d r o l y s i s , biological d e g r a d a t i o n , a d s o r p t i o n to storage v e s s e l s ,  a n d the p o s s i b l e  formation of more toxic degradation byproducts. T h e toxicity of a g e d solutions of t e c h n i c a l grade d i a z i n o n in w a t e r h a s actually b e e n s h o w n to i n c r e a s e d u e to the formation of the m o r e toxic b r e a k d o w n c o m p o u n d d i a z o x o n (Murty, 1986). Toxicity tests c o n d u c t e d in p l a s t i c v e s s e l s m a y p r o d u c e higher L C 5 0 ' s than t h o s e performed in g l a s s v e s s e l s d e p e n d i n g o n the propensity of the p e s t i c i d e s to bind to plastics. W i t h the e x c e p t i o n of malathion, the activity of most O P insecticide c o m p o u n d s is not influenced by p H , a n d p H o n l y a f f e c t s toxicity with r e s p e c t to its control o n the rate of hydrolysis (Murty, 1986; H e n d e r s o n & P i c k e r i n g , 1957).  M o s t p e s t i c i d e s a r e m o r e toxic to f i s h at higher  t e m p e r a t u r e s , e g chlorpyrifos ( M a c e k et a l . , 1969) a n d d i a z i n o n (Murty, 1986).  Water  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  (^12^21^2^3^^)'  i s  a  0-2-isopropyl-6-methylpyrimidin-4-yl c  '  e a r  phosphorothioate)  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 Protection a s a wettable p o w d e r ( W P ) ( 5 0 % d i a z i n o n ) or a n emulsifiable concentrate ( E C ) (500 g/L diazinon) ( B C M A F F , 1996a). It h a s a l s o b e e n s o l d u n d e r the b r a n d n a m e s of N e o c i d a l , N u c i d o l , a n d K n o x Out ( B C P C , 1991), a n d D a z z e l , D i a g r a n , D i a n o n , Diaterrfos, Diazajet, D i a z i d e , Didazital, D i a z o l , D i z i n o n , D y z o l , G a r d e n t o x , K a y a z o l , N i p s a n , Sarolex, a n d Spectracide (Smith, 1987). T h e B C M A F F r e c o m m e n d s the p o w d e r e d d i a z i n o n b e u s e d a s a s e e d treatment for corn, m i x e d with a fungicide ( 7 5 % captan or 7 5 % thiram), a n d that a solution of d i a z i n o n m a d e from wettable p o w d e r ( W P ) or emulsifiable concentrate ( E C ) be u s e d a s a pre-planting e m e r s i o n treatment for potato pieces or sprayed with g r o u n d equipment o n potato a n d corn c r o p s w h e n insects a p p e a r ( B C M A F F , 1996a).  D i a z i n o n is u s e d to protect corn from root m a g g o t s a n d protect  potatoes from aphids, C o l o r a d o potato beetles, f l e a b e e t l e s , leafhoppers, a n d leafminers ( B C M A F F , 1996a). Diazinon's recommended B . C . u s a g e rate is 1.11 k g / h a a n d 1.11 L/ha (445 m L / a c r e ) , W P a n d E C , respectively ( B C M A F F , 1996a).  D i a z i n o n is reported to h a v e a soil half-life of 32 to 4 8 d a y s ( R a o a n d D a v i d s o n , 1980). In a q u e o u s solutions, diazinon degrades by hydrolysis. At 2 0 ° C it h a s a hydrolysis half-life of 11.8 hours (pH 3.1), 185 d a y s (pH 7.4) a n d 6.0 d a y s (pH 10.4) ( B C P C , 1991). M o r g a n (1976) reported a hydrolysis half-life of 4 3 d a y s at 1 6 ° C a n d p H 7.6. h y d r o l y s i s , d i a z i n o n c a n be biologically d e g r a d e d .  In addition to  K a t z n e l s o n a n d M u m l e y (1997)  reported a half-life of 7 to 4 0 d a y s in s u r f a c e w a t e r s a m p l e s . S e t h u n a t h a n a n d P a t h e k (1972) found that for p r e p a r e d solutions of d i a z i n o n in w a t e r from rice fields w h i c h h a d previously b e e n e x p o s e d to r e p e a t e d d i a z i n o n d o s i n g , bacterial d e g r a d a t i o n of d i a z i n o n by Arthrobacter a n d Flavabactehum  s p . w a s rapid (23 m g / L to 0 m g / L in 5 d a y s ) ( 2 3 °C),  w h e r e a s incubation of d i a z i n o n in w a t e r from untreated rice fields s h o w e d little bacterial  51 degradation of d i a z i n o n (23 mg/L to 2 0 mg/L in 12 days). Further, they f o u n d that w a t e r f r o m fields previously treated with chlorpyrifos or carbofuran did not a c c e l e r a t e the degradation of d i a z i n o n . H e n c e , the d e v e l o p m e n t of bacteria specifically a b l e to rapidly d e g r a d e d i a z i n o n may o c c u r in a q u a t i c environments w h i c h h a v e r e c e i v e d n u m e r o u s previous diazinon applications. S e t h u n a t h a n a n d P a t h a k (1972) a l s o f o u n d that d i a z i n o n persisted longer in solutions w h i c h h a d b e e n incubated with soil than t h o s e without soil, stating that adsorption to soil particles d e c r e a s e d biological degradation/hydrolysis. D i a z i n o n h a s b e e n reported to a c c u m u l a t e in c r e e k s e d i m e n t s a n d 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-trichloro-2-pyridyl phosphorothioate) ( C g H ^ C ^ N O s P S ) is a colourless crystal a n d a l s o a c t s o n a n i m a l s by w a y of a c e t y c h o l i n e s t e r a s e inhibition. It h a s a v a p o u r p r e s s u r e of 2.5 m P a ( 2 5 ° C ) a n d a solubility in w a t e r of 2 mg/L ( 2 5 ° C ) ( B C P C , 1991). Chlorpyrifos is currently sold in B . C . for agriculture u n d e r the brand n a m e s of L o r s b a n 4 E , P y r i n e x 4 8 0 E C by D o w E l a n c o a n d United A g r i P r o d u c t s , s o l e l y a s a n e m u l s i f i a b l e concentrate ( E C ) (480 g/L, 1 0 L jugs) ( B C M A F F , 1996a). It h a s a l s o b e e n sold under the brand n a m e s of Dursban, Spannit, a n d S i l f r i f o s . ( B C P C , 1991) a n d B r o d a n , D o w c o 179, E r a d e x , a n d Killmaster (Smith, 1987).  T h e B C M A F F r e c o m m e n d s that  chlorpyrifos be s p r a y e d with ground equipment w h e n insect d a m a g e a p p e a r s o n corn, p o t a t o e s , a n d s u g a r beets for the control of c u t w o r m s a n d C o l o r a d o potato b e e t l e s ( B C M A F F , 1996a). Chlorpyrifos' r e c o m m e n d e d B . C . d o s a g e rate is 1.0 to 2.4 L/ha (405 to 9 7 0 m L / a c r e ) for v e g e t a b l e c r o p s ( 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 oH gN S) is 1  1  5  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  NA  NA  1632  NA  333  325  0  1088  78  5147  81  637  591  0  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  Richmond Delta Maple Ridge  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 lands south of 32  nd  th th 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 extend into this area. Further, this region between 32  th 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 St., and the headlands of Anderson Creek at 216 St. and 24 Ave. as locations nd  tn  th  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 (NH ), is a cellular metabolic waste, produced by both carnivorous 3  and herbivorous animals during the amino acid deamination step of protein digestion. Ammonia is excreted by animals as urea [NH -(C=0)-NH ] within their urine. Once urea 2  2  enters aquatic environments (and likely moist soils), it undergoes hydrolysis to form unionized ammonia (NH ).  Feces contains proteins, as the animals' unassimilated  3  consumed protein and the animals' own dead blood cell and intestinal tissue. These proteins are converted to NH through protein deamination described above. Runoff from 3  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 (NH ) is converted to ammonium (NH ) +  3  4  in the following reaction: NH (g) + nH 0(l) 3  NH (nH 0)(aq.)  2  3  2  NH  + 4  + OH" + (n-1)H 0(l) 2  Water pH, and to a lesser extent temperature, are the predominant water quality parameters which control this reaction and the relative concentrations of N H and N H . +  3  4  At higher pH's and temperatures more total ammonia is in the form of unionized ammonia (NH ). Very roughly, for a one unit pH increase and a 1 °C temperature increase, the 3  unionized ammonia (NH ) concentrations increase 10 fold and 1.1 fold, respectively. 3  59  Under aerobic conditions, unionized ammonia may be converted to nitrite ( N 0 ) by 2  Nitrosomonas sp. bacteria, and the nitrite may then be simultaneously converted to nitrate ( N 0 ) by Nitrobacter sp. bacteria. Nitrite is very toxic to fish [96-h LC50 rainbow trout 3  Onchorhynchus mykiss = 0.19 - 0.39 mg/L N 0 (CCME, 1986)]; however, the nitrite/nitrate 2  conversions generally happen simultaneously, such that N H is essentially converted 3  directly to the relatively non-toxic N 0 " [96-h LC50 rainbow trout Onchorhynchus mykiss 3  = 6.0 g/L (CCME, 1986)] without elevating aquatic N 0 " concentrations. 2  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 (NH ). Thurston & Russo (1981a) found that N H 3  + 4  may be toxic to salmonids  as well as NH ; however, they concluded that N H is 300 to 400 times more toxic than 3  NH  3  + 4  Laboratory measurements of ammonia often measure total ammonia (NH and NH ) +  3  4  using the colorimetric "Phenate Method" and typically report the concentration as NH -N 3  (total ammonia as N). This nomenclature should not be confused with the concentration of unionized ammonia (NH ) or unionized ammonia as N (NH -N). Ammonia toxicity test 3  3  endpoints are commonly reported as NH -N or N H (mg/L). To aid in the comparison of 3  3  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 e v i e w e d will b e in terms of the c o n c e n t r a t i o n of u n i o n i z e d a m m o n i a a s nitrogen m g / L N H - N . T h e n e c e s s a r y c o n v e r s i o n s h a v e a l r e a d y b e e n m a d e from the original literature 3  w h e r e required.  3.5.3  Toxicity to Fish  Lethal Toxicity Arthur et a l . (1987) r a n k e d the sensitivity of 5 different f r e s h w a t e r fish to u n i o n i z e d a m m o n i a , from most to least sensitive, with their respective m e a n 9 6 - h L C 5 0 ' s a s r a i n b o w trout {Onchorynchus  mykiss)  (0.44 m g / L N H _ N ) > w a l l e y e {Stizostedion 3  m g / L N H - N ) > c h a n n e l catfish (Ictalurus punctatus) 3  (Catastomus  commersoni)  vitreum) (0.54  (0.71 m g / L N H - N ) > white s u c k e r 3  (1.26 m g / L N H - N ) > f a t h e a d m i n n o w (Pimephales 3  promelas)  (1.78 mg/L N H - N ) , all tests at p H 7.7 to 8.3). 3  S a l m o n i d s are extremely s e n s i t i v e to u n i o n i z e d a m m o n i a . T h u r s t o n a n d R u s s o (1983) found 96-h L C 5 0 ' s for rainbow trout ranging from 0.17 to 0.37 mg/L N H - N ( @ 10 - 1 3 °C 3  a n d p H 7.7 - 7.9).  O b s e r v a b l e a c u t e r e s p o n s e s for s a l m o n i d s to N H  3  are typically  hyperexcitability, hyperventilation, violent erratic m o v e m e n t s , c o n v u l s i o n a n d c o m a , l e a d i n g to death. T h e s e effects a p p e a r to b e . c a u s e d by n e u r o l o g i c a l disorder.  Acute  lethality will u s u a l l y o c c u r within 9 6 - h of testing, if it is to o c c u r at all, a s s h o w n by Thurston a n d R u s s o (1983) w h i c h s h o w e d little difference b e t w e e n 3 5 - d a y L C 5 0 v a l u e s and 96-h L C 5 0 values.  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 NH -N (the highest concentration tested), but 3  til  that at the end of the alevins' yolk absorption (50 day of development) the 24-h TLm was only 0.059 mg/L NH -N, the same value they reported for adult trout. Rice and Stokes 3  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 upon the 3  transition to the fry stage is likely due to physiological changes following yolk absorption. The sensitivity of rainbow trout fry and adults to N H is similar; however, there is some 3  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 N H decreased 3  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 NH -N for fish weights 3  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 occurred at roughly 1-2 g fish weight (96-h LC50 approximately 0.5 - 0.7 3  62  mg/L NH _ ) (Thurston and Russo, 1983). This weight is the approximate size of the 3  N  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 with 96-h LC50's of approximately 0.2 - 0.3 mg/L NH -N 3  3  (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 NH -N for 10 days (@ 15 °C, pH 7.3) exhibited a decreased 3  resistance to fungal infection. Smith and Piper (1975) exposed rainbow trout to 0.013 mg/L NH -N for 9.5 months. There was no mortality in test fish at 9.5 months; however, 3  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 month, fish exposed to N H incurred bacterial gill disease th  3  infection, resulting in severe mortalities among the population. Thurston et al. (1984) found that adult rainbow trout exposed to > 0.06 mg/L NH -N for 4 months (@ 9 °C, pH 3  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 is available to 3  cause toxicity. However, these parameters play an additional role than just controlling the NH <-*• N H 3  + 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 NH -N (pH 7.0), 0.54 mg/L NH -N (pH 7.3), 0.71 mg/L 3  3  NH -N (pH 7.6) and 0.88 mg/L NH -N (pH 7.9), @ 15 °C. 3  3  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 NH -N (4 °C), 0.50 mg/L NH -N (10 °C), 3  3  and 0.86 mg/L NH -N (19 °C), @ pH 7.7 - 8.3. 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 exposure and uptake. The second is that increased ventilation 3  flushes the gill chamber of excreted C 0 , which raises gill chamber pH and causes a shift 2  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 NH -N (D.O. = 3.6 mg/L), 3  0.48 mg/L NH -N (D.O. = 6.6 mg/L), and 0.62 mg/L NH -N (DO. = 8.6 mg/L), (@ 12 - 13 3  3  °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 N H (Haywood, 1983). 3  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 increased the tolerance of these fish to lethal 3  levels of NH , but this resistance only lasted 3 days (perhaps by activating the above 3  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 N H in a gradient tank. 3  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 N H concentrations is less 3  clear. Green sunfish were not repelled by sublethal N H concentrations and stickleback 3  were attracted to sublethal NH concentrations. Hepner (1959) showed that carp avoided 3  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 NH -N) to 31 invertebrate species. 3  Organism  96-h LC50 (mg/L NH -N)  Reference  3  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 N H of the cladocerans. In addition to the above 3  values, Arthur et al. (1987) reported a 48-h LC50 for Simocephalus vetulus of 1.41 mg/L NH -N. The 48-h LC50 for Daphnia magna was reported as 2.94 mg/L NH -N (pH 8.3 3  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 NH -N (pH 8.1 @ 20.4°C). Nimmo et al. (1989) 3  reported a Ceriodaphnia dubia 48-h LC50 value of 1.06 mg/L NH -N (pH 7.8 @ 25°C). 3  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 NH -N (pH 8 @ 25°C). 3  Nimmo et al. (1989) pointed out that there is excellent agreement between the range of "acute" toxicity N H to fishes (their fish toxicity endpoints were for johnny darters, larval 3  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 N H toxicity to some fish. Although, the author points out that the 3  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  Glyphosate Captan Atrazine Metolachlor Simazine Dicamba MCPA Amine Bromoxynil Carbofuran Dimethoate Diclofop methyl  Use  (MAC) Mg/L  Herbicide Fungicide Herbicide Herbicide Herbicide Herbicide Herbicide Herbicide Insecticide-Carbamate Insecticide-OP Herbicide  65 2.8 2 8 10 10 2.6 5 1.75 6.2 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 Diazinon (OP) Chlorpyrifos (OP)  Acute 0.08  0.04  1  0.083  0.1 0.065  2  3  0.013  2  RMC  1  0.041  2  Malathion (OP) Parathion (OP)  Chronic  2  Carbaryl (Carbamate)  0.02  Azinphos-methyl (OP)  0.01  4  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  4  (US EPA, 1986) (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 (NH -N), for comparison with the ammonia 3  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 6.5 6.75 7.00 7.25 7.50 7.75 8.00 8.25 8.50 8.75 9.00  Table 10b.  10°C  15°C  20°C  25°C  25.48 23.02 20.55 16.40 12.7 13.3 5.83 3.37 1.89 1.13 0.68  24.66 22.19 19.73 16.19 12.25 8.63 5.67 3.29 1.89 1.13 0.71  23.84 22.19 18.91 15.78 12.00 8.47 5.59 3.21 1.89 1.17 0.75  . 16.44 15.28 13.48 11.01 8.38 5.92 3.95 2.30 1.41 0.88 0.59  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 6.5 6.75 7.00 7.25 7.50 7.75 8.00 8.25 8.50 8.75 9.00  10°C  15°C  20°C  25°C  1.81 1.81 1.81 1.81 1.81 1.73 1.13 0.64 0.37 0.21 0.13  1.81 1.81 1.81 1.81 1.81 1.81 1.09 0.62 0.36 0.22 0.13  1.22 1.22 1.22 1.22 1.23 1.15 0.76 0.44 0.26 0.16 0.11  0.85 0.85 0.85 0.85 0.86 0.81 0.54 0.32 0.19 0.12 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  CCME (1986)  Al  0.1 mg/L, pH > 6.5, C a  As  50 //g/L  CCME (1986)  Ba  1,000  BCME (1989)  Be  1,100 /ug/L for hardness > 75  CCME (1986)  Ca  < 8,000  BCME (1989)  Cd  0.2 /ug/L for hardness 0 - 60  M  g/L 2 +  > 4.0 mg/L  /ug/L  fxg/L  0.8 /ug/L for hardness 60-120 1.3 /ug/L for hardness 120-180  CCME (1986)  CCME (1986)  1.8 /ug/L for hardness > 180 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  CCME (1986)  3 //g/L for hardness 120-180 4 /ug/L for hardness > 180 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  CCME (1986)  65 /ug/L for hardness 60-120 110 /ug/L for hardness 120-180 150 /ug/L for hardness > 180 Pb  1 /ug/L for hardness 0 - 60 2 /ug/L for hardness 60 -120  CCME (1986)  4 /ug/L for hardness 120-180 7 /ug/L for hardness > 180 Se  1  Zn  30  /ug/L /ug/L  CCME (1986) 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 St NE ditch th  water contained dinoseb at 5 //g/L, and the 176 St. inner SE ditch sample exhibited trace th  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 St.), 36 //g/kg (inner SE ditch at 176 St.), 37 //g/kg (Nicomekl th  th  River, 50m above 176 St.), and 29 //g/kg (NE ditch at 184 St.) (Coastline, 1989). th  th  Endosulfan was not detected in any of the river sediments, but was found at 428 //g/kg in sediments from the inner 176 St. ditch and at trace concentrations (< 10 //g/kg) in the th  sediments from the SE 176 St., NW 176 St., and inner SE 176 St. ditches (Coastline, th  th  th  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 St. over these 6 months, 4 contained endosulfan; th  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 St. Nicomekl River ditch th  (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 St. to downstream of the Old Logging Ditch). Their colonization studies th  were conducted from Dec. 14 26 to July 4  th  to Jan 29  (1990), and Nov. 2 to 27  th  (1989), Feb. 2 3 to April 12 (1990), April rd  th  (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 anusup 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 ) (EC, 1996). Nevertheless, Athur et 2  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 NH -N) in the Nicomekl River.  Location  3  Dates  3  Total NH -N Range  Total NH -N Mean  <0.01 -0.13  < 0.01  3  3  Nicomekl at 99A Dam (Downstream)  1976- 1979  Nicomekl at 99A Dam (Downstream)  1988- 1992  28 Primarily Fall 2  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  0.005-0.177  0.049  Nicomekl at 184th St.  12/1989 to 11/1990  6  3  Burrows Ditch  12/1989 to 11/1990  6  3  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  # Samples  (BCME, 1997) (EVS, 1993)  20  1  28 Primarily Fall 2  0.03-0.13 0.08 - 0.27  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 1  Description of Location th  4  Ave. bridge th Nicomekl River, north side of river beneath 168 St. bridge th Un-named drainage ditch, NE 168 St. Nicomekl River, 5 m downstream of South Cloverdale Ditch discharge point  5  South Cloverdale Ditch, a t  6  Nicomekl River, 40 m upstream of South Cloverdale Ditch discharge point  7  9  Nicomekl River, south side of river 50 m upstream of the 176 St. bridge, downstream of Site 8 ditch discharge point. th Un-named drainage ditch, SE 176 St; inner ditch (furthest from road) Nicomekl River, 50 m upstream of Site 8 ditch discharge point  10  Ericson Ditch, at 40  11  Burrows Ditch, at 40 Ave.  12  Old Logging Ditch, at 40 Ave.  13  Nicomekl River, at 64 Ave. east o f  2 3  8  W.R.P.  Nicomekl River, 50 m upstream of the 40  4 0 t h  Ave.  th  Ave. And 180  St.  th  th  th  2 1 6 t h  St. (upstream control site)  th White Rock pond, Southmere Park, 16 site)  Ave. and Oxford St. (culture/control  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 St. ditch at one time, on May 6. A sample of this th  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 St. ditch's flap gates was observed on the August 7 and September 17 sampling th  th  th  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 (1520) 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 ) from a small pond located in 3  Southmere Village Park, White Rock (Figure 1), at the north side of 16 Ave. just east of th  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 fullstrength (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  Dilution Series Chronic Testing of Lethally Toxic Samples  5.3.3  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.  Lethally Toxic Samples Biological Toxicity Identification Evaluation.  5.3.4  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 set of samples; the th  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. IC o's were calculated despite the EC (1992) recommendation for 5  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 2 0 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 nonconforming 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  C h e m i c a l A n a l y s e s o f Ditch and River Water S a m p l e s  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 liquidliquid 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 acidpreserved 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-NH G (AWWA, 1995). Testing was conducted using the Civil 3  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 using an infrared detector, 2  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 (de 2  March, 1981) and a 30-day LC50 of < 0.3 mg/L 0 (Nebeker et al., 1992). West et al. 2  (1993) found that H. azteca was more sensitive than C. tentans and L. variegatus in 10day 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.  Discharge Site 8  7d  100  24.2 +/- 6.7  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  Reproduction Comparison With Controls > Culture Cont.  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 6d  100  24.6 +/- 3.5  = Upstream, = Culture Cont.  *  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 10 Site 11  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  Ill  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 Site 1  *  Site 2 Site 3  *  Site 4 Site 5  *  Site 6 Site 7  *  Site 8  Test Duration  Survival %  Mean # Neonates +/- S.D.  Reproduction Comparison With Controls  6d  100  16.3 +/-4.5  < Upstream, = Culture Cont.  6d  100  19.1 +/-5.5  = Upstream, > Culture Cont.  6d  100  16.3 +/- 6.8  < Upstream, = Culture Cont.  6d  100  21.0 +/-3.7  = Upstream, > Culture Cont.  6d  100  17.2 +/-4.0  < Upstream, > Culture Cont.  6d  100  20.0 +/- 4.7  = Upstream, > Culture Cont.  6d  90  16.6 +A8.7  < Upstream, = Culture Cont.  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  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.  Oct. 3 Samples  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  113  Table 15 Cont. 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  © E X  Lethally toxic sample. Sublethally toxic sample. Effect observed; partial paralysis by test end of 3 of 6 surviving organisms. 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 , 1997, and Site 5, South Cloverdale Ditch at 4 0 Ave., collected th  th  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  Site 8, June 26 176th St. Ditch  32%  NA  42.7 % (+/- 3.5)  > 100%  Site 5, July 16 South Cloverd. Ditch  32%  NA  19.9% (+/-12.8)  92.9 %  96-h LC50 6-day LC50  74.8 % (56.0,100.0)  39.9 % (32.0,56.0)  52.2 % (32.0,100.0)  36.5 % (28.6, 46.3)  1  2  1  1  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  2  Computer program used Binomial method to calculate the LC50.  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 ' upstream control. Since these controls came close to rd  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 riverwater sample from Site 4 had a 7-day survival of 60%. This survival th  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  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  Site 12, Old Logging Ditch  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 tests came fungus contaminated culture water, and unfortunately brood r d  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 tests. Poor culture health was an intermittent r d  th  problem for short durations in September and October. However The September 17 , October 3 RETESTS, and October 16 tests, were all supplied with neonates from brood th  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 eliminated toxicity from the 1 7 6 samples.  t h  St. a n d S o u t h C l o v e r d a l e Ditch  T a b l e 18 s h o w s the one-tailed t-test p-value results c o m p a r i n g survival a n d  r e p r o d u c t i o n of the  C  1  8  treated toxic s a m p l e s with that in the culture/control water.  All  three c h r o n i c tests w e r e c o n d u c t e d at the s a m e time a n d u s e d n e o n a t e s from the s a m e broods.  T a b l e 18.  R e s u l t s of  C  1  8  C.  dubia B i o l o g i c a l Toxicity Identification.  June 26 Site 8,176 St. Ditch th  th  Control Survival (%)  Untreated  100  0  89  P-value (1-tailed) v. Cont. Mean # Neonates +/- S . D  0.084 18.9+/-6.4  0  15.0+/-7.3  P-value (1-tailed) v. Cont.  July 16  C18 Treated  0.110  Site 5, South Cloverdale Ditch Control Survival (%)  Untreated  90  0  P-value (1-tailed) v. Cont. Mean # Neonates +/- S.D.  C18 Treated 100 0.084  15.4 +/-4.6  0  P-value (1-tailed) v. Cont.  O n e - t a i l e d p-values are for culture/control v e r s u s  12.1 +/-6.3 0.100  C  1  8  treated s a m p l e s .  P - v a l u e > 0.05 indicates that there is no statistically significant difference.  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 26 176 St. ditch (Site 8) water sample th  th  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.  Control  85 +/- 30 %  MeOH Control  100 +1-0%  Site 8 Site 8 PBO  Comparison  P-Value  Result (Survival)  Cont. v. MeOH Cont.  0.196  Cont. = MeOH Cont.  85 +/- 19%  Site 8 v. Cont.  0.5  Site 8 = Cont.  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.  Control MeOH Control Site 8 Site 8 PBO  100 +/-0% 100 +1-0%  Cont. v. MeOH Cont.  0.5  Cont. = MeOH Cont.  35 +/-19 % 100 +1-0%  Site 8 v. Cont. Site 8 PBO v. Cont.  0.003* 0.5  Site 8 < Cont. Site 8 PBO = Cont.  100 +1-0% 100 +1-0%  Site 8 PBO v. MeOH Cont. 0.5 Site 8 v. Site 8 PBO 0.003 * Site 5 v. Cont. 0.5 Site 5 PBO v. Cont. 0.5  Site 8 PBO = MeOH Cont. Site 8 < Site 8 PBO Site 5 = Cont. Site 5 PBO = Cont.  Site 5 PBO v. MeOH Cont. 0.5 Site 5 v. Site 5 PBO 0.5  Site 5 PBO = MeOH Cont. Site 5 = Site 5 PBO  Site 5 Site 5 PBO  Comparison  P-Value  Result (Survival)  * 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 St. Ditch. However, in this th  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.  Control MeOH Control Site 8  5+/- 10% 5+/- 10%  Cont. v. MeOH Cont.  0.006 *  Cont. = MeOH Cont.  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 Site 5 v. Cont.  0.0002 * 0.199  Site 8 < Site 8 PBO Site 5 = Cont.  0.004 *  Site 5 PBO > Cont.  0.004 *  Site 5 PBO > MeOH Cont. Site 5 < Site 5 PBO  Site 5 Site 5 PBO  Comparison  15+/-19% 85 +/- 30 % Site 5 PBO v. Cont.  Site 5 PBO v. MeOH Cont. Site 5 v. Site 5 PBO  P Value  0.006 *  Result (Survival)  * 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 St. and South th  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-ionscan 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-ionscan, 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. Metal  ICP Metal Scan Results for Toxic, Non-Toxic, and Control Samples. 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  A g (jug/L)  < 0.05  0.15  < 0.05  < 0.05  0.05  0.05  0.05  < 0.05  A l (mg/L)  0.21  0.55  0.35  0.32  A s (ug/L)  4  2  1  2  3  1  1  1  B a (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  Ca ^g/L)  21.9  52.9  14.2  16.7  26.4  24.1  16.95  10.65  C d (Mg/L)  < 0.1  0.5  < 0.1  < 0.1  0.3  < 0.1  < 0.1  < 0.1  C o (^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  3.0  2.5  2.5  2.0  2.5  1  2.0  C u (ug/L)  9.4  1  9.9  11.1  1  5.5  5.6  1  Fe (mg/L)  0.84  2.43  1  0.62  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  P b (ug/L)  <2  <2  6  <2  2  S b (Mg/L)  0.25  0.35  S e (ug/L)  1  3  1  1  1  1  1  14.7  1  4.4  0.87  1  0.72  3  1  1  1  1  0.67  6  1  1  1  0.78  1  2  0.21  4  1  1  1  1  0.17  0.48  1  0.14  9.2 1  1  1  0.34  1  0.15  0.55  0.25  0.25  0.10  0.40  1  < 1  < 1  1  1  < 1  < 1  S n (ug/L)  < 0.5  < 0.5  < 0.5  < 0.5  < 0.5  < 0.5  < 0.5  < 0.5  S r (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  6.0  20.0  17.0  10.5  12.0  Hardness  108  63  83  103  66  80  80.5  1  265  52.5  1  138  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 NH -N. Nimmo et al. also reported that C. dubia exhibited 3  reduced reproduction for a 7-day exposure to 0.68 - 0.88 mg/L NH -N (pH 8.0 @ 25 °C). 3  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 NH -N, respectively), both well below that causing lethal or 3  135  sublethal effects to C. dubia.  Table 23.  Total Ammonia (mg/L NH -N) of All Water Samples, on Dates Exhibiting Toxicity in Select Samples. 3  Sample  June 26  July 16  Oct. 3  Nov. 21  Site 1 Site 2 Site 3 Site 4 Site 5 Site 6 Site 7 Site 8 Site 10 Site 11 Site 12 Site 13 Culture Control  0.022  0.074  0.086  0.068  0.017  0.162  0.117  0.051  0.023  0.182  0.085  0.057  0.009  0.106 (S)  0.113  0.056  0.129 (S)  0.152  (L)  0.238  0.206  0.069  0.048  0.015  0.084  0.068  0.052  0.248 (L)  0.126  0.273  0.115 (S)  0.007  0.051  0.172  0.050  0.020 (S)  0.095  0.156  0.066  10.820 (S)  0.118  0.280  0.048  0.062  0.062  0.052  0.048  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 NH -N) (Thurston et al., 1979), a value acutely lethal to 3  salmonids. This concentration is within the range for which Nimmo et al. (1989) showed chronic sublethal effect (0.68 - 0.88 mg/L NH -N). Consequently, the sublethal C. dubia 3  toxicity for this Old Logging Ditch sample was most likely due to ammonia contamination.  136  6.3.4  D i s s o l v e d Total/Inorganic/Organic C a r b o n , 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 P e s t i c i d e s  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.  Pesticide  Diazinon  All results from 6-day test.  C. dubia Lethality to Detected Pesticides.  24-h LC50 (+/- 95% C.l.) > 0.8 M g / L (highest cone.) Chronic Test Daily changeover Daily feeding  48-h LC50 (+/- 95% C.l.)  96-h LC50 (+/- 95% C.l.)  0.28 M g / L (0.2, 0.4)  Chronic Test Daily changeover Daily feeding  Chronic Test Daily changeover Daily feeding  Chronic Test Daily changeover Daily feeding  0.50 M g / L (0.33, 0.66)  > 0.132 M g / L (highest cone.)  > 0.132 M g / L (highest cone.) Chronic Test Daily changeover Daily feeding  (0.4,  Mg/L  2  0.8)  1  0.77 //g/L (0.57, 1.03)  Results from 48-h test and 6-day test.  48-h Acute Test No changeover Fed at 0-h  48-h Acute Test No changeover Fed at 0-h  Chronic Test Daily changeover Daily feeding  Prometryn  7.59 mg/L (5.78, 9.92)  6.22 mg/L (2.5, 10.0)  4.58 mg/L (2.5, 10.0)  All results from 96-h test.  Changeover at 48-h, fed at-2-h and 2-h prior to changeover  Changeover at 48 h,fed at-2-h and 2-h prior to changeover  3  (+/- 95% C.l.)  0.57 M g / L (0.4, 0.8)  0.64  Chlorpvrifos  3  6-day LC50  1  1  1  1  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  Diazinon Chlorpyrifos  NOEC  LOEC  IC25 (+/- S.D.)  0.2 //g/L  0.1 //g/L  0.26 +/- 0.002 //g/L  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 , following the significant rainfall of late May (as high as 25 mm/day). th  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 St. ditch was discharged, since its flows are regulated th  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  (sjnoLj)  B u j d i u n d  e B j e i p s i Q  0 E? r  Q  CT)  c co  o  C  o  CD  Q  Q O)  bl)  CD CD  1.  CD co  GO  s  o  CM  E  +3 —I  o CD  CO  a a  CO  o  CO IO  ^ a jgga n ^ cC N tazmzazm a i azzag -j  o  LO  CN (tutu)  O  LO  uoj;e;jdpajd  -s O  142  (sjnoi|)  B u i d m r i d  a B j e i p s i Q  i  1  JZT 1H 1  i L  Q  £  CO O CO  ^ >^  o  CD  0 >  •  | i —  x  1  "HH cd GO  .  CO  \~-  MB  QL o  CD  Or  m  ml  mtmmm  •  C  E B  ,  co CM  a  i  _o  .** CO o> CO  O  i —  o  CM -aCM CO CM CM CM  Oi  1  p  6  m  O CM  1  - .gnu,' •  CM  CM  1  +-» _JSL •  i  oJ  0  GO  CM  co  c  Q  CO CD CM CO CM  i  Si kL  CD  CD >  C  c]  L  o>  s 1 5 Q  Ql  "P  o  !  ....  m §i  0) G)  CO  j.  •  CO  13  °  E  CD  n  1  1  E  :  m  co  i  CM  {  I  o  o  CO  o CN  (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.  Nicomekl River Flow at Old Logging Ditch 16 14 m 12 | 10 8 o 6 4 2 0  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  c o  LL  o  CO CO  t-  c  f3  IE  <  J£  E o o  x: -fc  CD  T3  S jS  o to .-~ Q.  CO h ~  <  «5  x ro  -«—< =5 O  CZ CD O CD  CD TJ  •I E  -*—' T—  o 2  ro .  CD  II  to J -  J= CD  o -I E2  o f=  O  -  z  II  S co ro X c  o  7  .9  z  CN  co X CD  ro -*—'  2 2  CO  Q . +3 •J C CO "~ LU  o o cz o  O  O O C O T J - C M O C O C O ^ C M O  h/Buj) N-CHN  Z5  o) E ro c  CO  X Q_  ro O <  "3 o <  E < ro o  O  o <  o 'cz o  1—  JZ  O  " ° TJ ® CD  H  E _i E o) < E ro co  C0  ro o E c  2!i  co ro  =3  E to CD o .c o  ro  <  co E  — I  oo  O  o  co  CD  ro X o Q.  — I  CD 4—<  co co  co X  E ro  >  C0:9  c  CO  E  CL  ><  5  to co CO  o o  E -*—' CD E Q..5 E ro x o o  CL  CD  ^ ro ro r-  Q o roSQ  ro o E —'  ^8 CD ZJ O)  LL  .5=  CO CO CO  c c c E E E  II  LU  =5  O  3  CO  U  ZJ  CO CO  co to to  <<< aTS"cr  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 Ave. on June 26 as only 0.022 mg/L NH -N. The Old th  th  3  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:  E CD -*i_ —'  CD  CD  E o o  CO  s  CD  CO  T J  =5  .ti  E o CD  o o  1  CJ  C L  CO CD  T3  £ E o 2  m CO  "D  o £  i_ CD  M—  >  O  D c o E 5  O _c -*—<  CD  o E c  ZJ  O C O  o  =  co  2  c o  TO  c CD O  c o  O >^ CD  Q CD CD  Q_ "D  £ CD  E  —•  co  LU  LO CD ZJ  ©  CD  CJ>:9 ] Q co CO CO X o CO  E Q. E E rs Q. E E E x c o CD o E c c c E E E O  C L  ZJ  C O  C O  I  O )  13  CO CO CO CO CO CO  < << CD Xi O  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 . However, 3 th th the high river flows (2.5 to 15 m Is) from July 5 to 11 significantly diluted the South th  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 estimates that there th  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 The situation is more questionable for the later discharge flows of July 12  th 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 the Site 4 sample, collected immediately downstream of the South Cloverdale th  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 day of its bioassay, the effect of remnant discharge to organisms as sensitive as th  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 S Odour, No Visible Invertebrates  10  Sandy with Rocks, Very Slight H S Odour, Twigs and Grass, No Visible Invertebrates  11  Silty and Clayey, Slight H S 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  2  2  2  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.  Sample  Sediment Bioassays % Survival and Growth (Test-End Mean Hyallela Dry Weight).  A % Survival  A Mean Dry Weight  B % Survival  M9  Site 1 Site 2 Site 5 Site 7 Site 8 Site 10 Site 11 Site 12 Site 13 Silica Sand  100 50 100 100 100 90 100 100 100 0  285 178 250 365 248 218 178 129 241  B Mean Dry Weight  C % Survival  C Mean Dry Weight  D % Survival  M9  90 90 100 100 90 90 100 90 70 0  293 201 294 279 329 216 162 127 257  D Mean Dry Weight ^g  60 90 90 90 90 100 90 80 80 0  285 174 316 220 321 252 227 134 225  80 0 100 100 90 90 70 50 50 0  195 264 253 206 201 306 157 138  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  Site 1 Site 2 Site 5 Site 7 Site 8 Site 10 Site 11 Site 12 Site 13  Mean % Survival +/S.D.  83 +/- 17 58 +/- 43 98 +/- 5 98 +/- 5 93 +/- 5 93 +/- 5 90 +/- 14 80 +/- 22 75 +/- 21  Mean Survival, Weight, Weight Versus Site Versus Site 13 13 fog) +/(P=, 1 tail) (P=, 1 tail) S.D.  265 +/- 46 184 +/- 15 281 +/- 30 279 +/- 62 276 +/- 59 222 +/- 22 218+/-65 137 +/- 14 215+/-53  0.299 0.251 0.063 0.063 0.100 0.100 0.143 0.375  Organic Matter (%)  0.106 0.165 0.042 * 0.084 0.089 0.416 0.473 0.032 *  * P-value < 0.05 indicates that there is a statistically significant difference.  7.1 7.9 3.0 8.0 18.5 4.1 15.7 6.9  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.)  Cd  0.24 +/- 0.01  0 +/- 0.23  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 %  Author's (mg/kg +/- S.D.)  Error -4 %  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.  Metal  Sediment Sample's Hypothetical True Dry Weight Metal Content (mg/kg).  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  502.5  206.4  339.7  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.  Metal  Fraser River Sediment Background Trace Metal Concentrations, Canadian Sediment Quality Guidelines Trace Metal TEL, and Other Studies Nicomekl Sediment Metal Concentrations. Cdn. Crit. (TEL)  Natural Background Fraser River 1  Nic. 184th St. (1990)  1  (mg/kg)  2  (mg/kg)  (mg/kg)  Nic. Below Old Logging Ditch (1990) (mg/kg)  Old Logging Ditch (1990)  Burrows Ditch (1990)  2  2  (mg/kg)  2 (mg/kg)  Cd  0.6  < 0.25 - 0.50  NM  NM  NM  NM  Cr  37  20-45  NM  NM  NM  NM  Cu  36  18-40  20-43  25-38  45-58  42-53  Pb  35  3-11  30-60  15-58  18-70  19-87  Ni  18  7-54  NM  NM  NM  NM  Zn  123  45-95  40 - 300  80 - 300  2 0 0 - 1100  1 9 0 - 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. M o u t h ( 1 9 8 9 - 1 9 9 3 ) approx. 5 s a m p l e d a t e s (mg/kg) 3  N i c , 6 locations (1979)  Cd  0.28 - 0.38  < 1.0  Cr  23.2-61.8  21 - 360  Cu  6.25-29.10  8-22  Pb  1.71 - 16.5  16-71  Ni  15.0-55.7  16-28  Zn  31.7-105  41 - 5 7  Mo  0  < 1 - 2.0  Co  14.6-18.3  14-22  Ca  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  4 (mg/kg)  All metal measurements are total mg/kg dry weight. 1  (Smith et al., 1995); (EVS, 1993); (BCME, 1997); (Swain and Holms, 1988b)  NM = Not Measured  2  3  4  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 , proved to be lethally toxic with 48-h, 96-h and 6-day LC50's of > 100%, 74.8%, and th  39.9% for the 176 St. sample and 92.9%, 52.2%, and 36.5% for the South Cloverdale th  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 St. ditch sample at approximately 0.02 //g/L, 0.07, and 3 //g/L, th  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 overspray 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 NH -N. This equated to an unionized 3  ammonia concentration of approximately 0.86 mg/L NH -N, based on the ditchwater's 3  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 NH -N) and above that reported 3  to cause lethality to salmonids [rainbow trout, 1-2 g fish 96-h LC50 approximately 0.5 - 0.7 mg/L N H . , 10 - 20 g fish 96-h LC50 approximately 0.2 - 0.3 mg/L N H (Thurston and 3  N  3  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  th  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 (48h) 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 th  St. and  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).  Invertebrate species abundance and diversity could be determined in sediment samples collected both upstream and downstream of one ditch (author suggests South Cloverdale Ditch) both prior to and subsequent to an observed period of OP insecticide contaminated ditch water. Although it is difficult to predict if, when and where such an event will occur, so prior benthos sampling downstream of a ditch discharge would be difficult.  More research is necessary on the toxicity of combinations of insecticides, herbicides, humic substances, suspended sediments, and possibly metals, as drainage ditches may contain multiple compounds.  171  REFERENCES Alabaster, J.S. 1969. Survival offish in 164 herbicides, insecticides, fungicides, wetting agents, and miscellaneous substances. Int. Pest Control 11 (2): 29-35. As Cited In Pimentel (1971). Aller, R.C., and J.Y. Yingst. 1983. Biogeochemistry of tube-dwellings: a study of the sedentary polychaete Amphitrite ornata (Leidy). J. Mar. Res. 36: 201-254. As Cited In Lee (1991). Allison, D.T., and R.O. Hermanutz. 1977. Toxicity of Diazinon to Brook Trout and Fathead Minnows. Prepared for the Environmental Research Laboratory/Duluth, Office of Research and Development, U.S. Environmental Protection Agency, Duluth, Minnesota. EPA-600/3-77-060. May, 1977. Amato, J.R., D.I. Mount, E.J. Durhan, M.T. Lukasewycz, G.T. Ankley, and E.D. Robert. 1992. An example of the identification of diazinon as a primary toxicant in an effluent. Environ. Toxicol. Chem. 11: 209-216. Ankley, G.T., D.J. Call, J.S. Cox, M.D. Kahl, R.A. Hoke, and PA. Kosian. 1994a. Organic carbon partitioning as a basis for predicting the toxicity of chlorpyrifos in sediments. Environ. Toxicol. Chem. 13(4): 621-626. Ankley, G.T., D.A. Benoit, J.C. Balogh, T.B. Reynoldson, K.E. Day, and R.A. Hoyke. 1994b. Evaluation of potential confounding factors in sediment toxicity tests with three freshwater benthic invertebrates. Environ. Toxicol. Chem. In press (1994). As Cited In EC (1996). Ankley, G.T., J.R. Dierkes, D.A. Jensen, and G.S. Peterson. 1991. Piperonyl butoxide as a tool in aquatic toxicological research with organophosphate insecticides. Ecotoxicol. Environ. Saf. 21: 266-274. Anonymous, 1997. Anonymous C. dubia chronic test results, using water samples spiked with Cu and Zn. Author had personal involvement with the test. Arthur, J.W., C.W. West, K.N. Allen, and Steven F. Hedtke. 1987. Seasonal toxicity of ammonia to five fish and nine invertebrate species. Bull. Environ. Contam. Toxicol. 38: 324-331. Arthur, J.W., J.A. Zischke, K.N. Allen, and R.O. Hermanutz. 1983. Effects of diazinon on macroinvertebrates and insect emergence in outdoor experimental channels. Aquatic Toxicology 4: 283-301. th AWWA, 1995. Standard Methods for the Examination of Water and Wastewater. 19 Edition. Prepared and published by the American Public Health Association, American Water Works Association, and Water Environment Federation. Edited by Mary Ann H. Franson. American Public Health Office, Washington, D.C.  172  Bailey, H.C., J.L. Miller, M.J. Miller, L.C. Wiborg, L. Deanovic, and T. Shed. 1997. Joint acute toxicity of diazinon and chlorpyrifos to Ceriodaphnia dubia. Environ. Toxicol. Chem. 16(11): 2304-2308. Bailey, H.C., C. Digiorgio, K. Kroll, J.L. Miller, D.E. Hinton, and G. Starrett. 1996. Development of procedures for identifying pesticide toxicity in ambient waters: carbofuran, diazinon, chlorpyrifos. Environ. Toxicol. Chem. 15 (6): 837-845. Bathe, R., K. Sachsse, L. Ullmann, W.D. Hoermann, F. Zak, and R. Hess. 1975. Evaluation offish toxicity in the laboratory. Proc. Eur. Soc. Toxicol. 16: 113-124. As Cited In CCME (1989). BCMAFF, 1996a. Field Crop Guide to Weed, Disease, Insect, Bird and Rodent Control for Commercial Growers. 1996 Edition. British Columbia Ministry of Agriculture, Fisheries and Food. Province of British Columbia, Victoria, B.C., Canada. BCMAFF, 1996b. Berry Production Guide for Commercial Growers. 1996/97 Edition. British Columbia Ministry of Agriculture, Fisheries and Food. Province of British Columbia, Victoria, B.C., Canada. BCME, 1997. Environmental Monitoring System Detailed Results Report (Database Printout). B.C. Ministry of Environment. Surrey, B.C. EMS ID 0300060, 0300062 (water samples), E207869 (sediment samples). December, 1997. BCME, 1989. Approved and Working Criteria for Water Quality. B.C. Ministry of Environment, Water Management Branch, Resource Quality Section, Victoria, B.C. Prepared by L.W. Pommen. As Cited In EVS (1993). BCME. 1982. Map of Burrows Ditch Drainage Area, by the B.C. Ministry of Environment, Water Management Branch. Prepared by K.W. Wilson, Engineer, to accompany report on District of Surrey Burrows Ditch Project Key Plan Showing Watershed Boundaries. File No. 0242512-110B, DWG. No. 5339-1. BCMELP, 1997. Survey of Pesticide Use in-British Columbia. 1995. Prepared for Ministry of Environment, Lands and Parks Pollution Prevention and Remediation Branch and Environment Canada Fraser River Action Plan. Prepared by Norecol Dames & Moore. Technical Report DOE FRAP #1997-16. BCMELP, 1993. Survey of Pesticide Use in British Columbia: 1991. Prepared for Ministry of Environment, Lands and Parks Pesticide Management Branch and Environment Canada Conservation and Protection. Prepared by Norecol Environmental Consultants Ltd.. Pesticide Management Program, Publication #93-3. BCPC, 1991. The Pesticide Manual. A World Compendium. 9 Edition. Ed. Worthing, C.R., and R.J. Hance. The British Crop Protection Council, Farnham, Surrey, Great Britain. th  173  Becker, D.S., G.R. Bilyard, and T.C. Ginn. 1990. Comparisons between sediment bioassays and alterations of benthic macroinvertebrate assemblages at a marine superfund site: Commencement Bay, Washinton. Environ. Toxicol. Chem. 9: 669685. As Cited In EC (1996). Biesinger, K.E., and G.M. Christensen. 1972. Effects of various metals on survival, growth, reproduction and metabolism Daphnia magna. J. Fish. Res. Board Can. 29: 1690-1700. As Cited In CCME (1986). Bohmont, B.L. 1967. Toxicity of herbicides to livestock, fish, honeybees, and wildlife. 20th Western Weed Control Conf., Proc. 21: 25-27. As Cited In Pimentel (1971). Bond, C.E., J.D. Fortune, and F. Young. 1965. Results of preliminary bio-assays with kurosal-SL and dicamba. Progr. Fish-Cult. 27: 49-51. As Cited In Pimentel (1971). Bourque, S., and Hebert, G. 1982. A Preliminary Assessment of Water Quality and Biota in the Serpentine River and Nicomekl Rivers and Mahood Creek, 1974-1975. B.C. Ministry of Environment and Environmental Protection Service. As Cited In Robinson (1988). Bull, C.J., and J.E. Mclnerney. 1974. Behavior of juvenile coho salmon (Onchorhynchus kisutch) exposed to Sumithion (fenitrothion), an organophosphate insecticide. J . Fish. Res. Board. Can. 31: 1867-1872. Burdick, G.E., E.J. Harris, H.J. Dean, T.M. Walker, J. Skea, and D. Colby. 1964. The accumulation of DDT in lake trout and the effects on reproduction. Trans. Am. Fish. Soc. 93:127-136. As Cited In Rice and Stokes (1975). Burton, G.A. Jr., and C.G. Ingersoll. 1994. Evaluating the Toxicity of Sediments, in: The ARCS Assessment Guidance Document, Report EPA/905-B94/002, U.S. Environmental Protection Agency, Chicago, III. As Cited In EC (1996). Call, D.J., L.T. Brooke, N. Ahmad, and D.D. Vaishnav. 1981. Aquatic Pollution Hazard Assessments and Development of a Hazard Prediction Technology by Quantitative Structure Activity Relationships. Second Quarterly Report to U.S. EPA Center for Lake Superior Environmental Studies, University of Wisconsin-Superior, Superior, Wisconsin. As Cited in CCME (1986). Canfield, T.J., N.E. Kemble, W.G. Brumbaugh, F.W. Dwyer, C.G. Ingersoll, and J.F. Fairchild. 1994. Use of benthic invertebrates community structure and sediment quality triad to evaluate metal-contaminated sediment in the Upper Clark Fork River, MT. Environ. Toxicol. Chem. 13: 1999-2012. Carballo, M., and M.J. Munoz. 1991. Effect of sublethal concentrations of four chemicals on susceptibility of juvenile rainbow trout (Onchorhynchus mykiss) to saprolegniosis. Appl. Environ. Microbiol. 57: 1813-1816.  174  CCME, 1995. Canadian Water Quality Guidelines: Updates (March, 1995). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers o f the Environment. CCME, 1994. Canadian Water Quality Guidelines: Updates (March, 1994). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers o f the Environment. CCME, 1993a. Canadian Water Quality Guidelines: Updates (October, 1993). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers o f the Environment. CCME, 1993b. Canadian Water Quality Guidelines: Updates (March, 1993). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers o f the Environment. CCME, 1991. Canadian Water Quality Guidelines: Updates (April, 1991). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers o f the Environment. CCME, 1989. Canadian Water Quality Guidelines: Updates (September, 1989). Prepared by the Task Force on Water Quality Guidelines o f the Canadian Council o f Ministers of the Environment. CCME, 1986. Canadian Water Quality Guidelines. 3.0 Freshwater Aquatic Life. Prepared by the Task Force on Water Quality Guidelines of the Canadian Council of Ministers of the Environment. Chin, Y. N., and K.I. Sudderuddin. 1979. Effect of methamidophos on the growth rate and esterase activity of the common carp Cyprinus carpio L. Environ. Pollut. 18: 213221. Clemens, H.P., and K.E. Sneed. 1959. Lethal doses of several commercial chemicals for fingerling channel catfish. U.S. Fish Widl. Serv., Spec. Sci. Rep.: Fisheries 316. 10p. As Cited In Pimentel (1971). Coastline. 1989. Preliminary Evaluation of Selected Pesticides in the Nicomekl River Watershed, Lower Fraser Valley, B.C., and Proposed Approaches for Assessing Toxicity to Aquatic Biota. Prepared for Environment Canada by Coastline Environmental Services Ltd. June, 1989. Cope, O.B. 1966. Contamination of the freshwater ecosystem by pesticides. Supplement on pesticides in the environment and their effects on wildlife. 3: 33-44. As Cited in Pimentel (1971). Cope, O.B. 1965a. Sport fisheries investigations. Pesticide wildlife studies, U.S. Fish Wildl. Serv. Circ. 226: 51-63. As Cited In Pimentel (1971).  175  Cope, O.B. 1965b. Sport fishery investigation. The effect of pesticides on fish and wildlife. U.S. Fish Wildl. Serv. Circ. 226: 51-64. As Cited In Pimentel (1971). Crosby, D.G., and R.K. Tucker. 1966. Toxicity of aquatic herbicides to Daphnia magna. Science 154: 289-291. As Cited In Pimentel (1971). Davies, C.S., and R.J. Richardson. 1980. Organophosphorous Compounds. Pages 527544 in P.S. Spencer and H.H. Schaumberg, eds. Experimental and Clinical Neurotoxicology. William and Wilkins, Baltimore, Md. 929pp. As Cited In Smith (1987). Davis, D., D. Serdar, and A. Johnson. 1997. Assessment of Cranberry Bog Drainage Pesticide Contamination. Results from Chemical Analyses of Surface Water, Tissue, and Sediment Samples Collected in 1996. By the Environmental Investigations and Laboratory Services Program, Washington State Department of Ecology, Olympia, Washington. Publication No. 97-329. July, 1997. de March, B.G.E. 1981. Hyallela azteca (Saussure), in: Manual for the Culture of Selected Freshwater Invertebrates, Canad. Spec. Publ. Fish. Aquat. Sci. No. 54, Department of Fisheries and Oceans, Ottawa, Ont. As Cited In EC (1996). Dodson, J.J., and C.l. Mayfield. 1979. Modification of the reotrophic response of rainbow trout (Salmo gairdneri) by sublethal doses of the aquatic herbicides diquat and simazine. Environ. Pollut. 18: 147-157. Donigian, A.S. Jr., D.W. Meir, and P.P. Jowise. 1986. Stream transport and agricultural runoff of pesticides for exposure assessment: a methodology. Part A--text and appendices A through F. Prepared for U.S. EPA. EPA/600/3-86/011 a. March 1986. Eaton, J., J. Arthur, R. Hermanutz, R. Kiefer, L. Mueller, R. Anderson, R. Erikson, B. Nordling, J. Rogers, and H. Pritchard. 1985. Biological effects of continuous and intermittent dosing of outdoor experimental streams with chlorpyrifos. Aquatic Toxicology and Hazard Assessment 8th Symposium. (R.C. Banner and D.J. Hansen, Eds.) pp. 85-118: American Society for Testing and Materials, Philadelphia. As Cited in Ward et al. (1995). Environment Canada (EC). 1998. Nicomekl River at 203 St., Langley, Station No. 08MH155,. (Unpublished Data). Hydrometric and Environmental Data, Environmental Services and Applications, Environment Canada. January 30, 1998. Environment Canada (EC). 1997. Precipitation Record, Cloverdale East, B.C. AES National Headquarters ID: 1101708, Regional ID: 1640J (Unverified Data). Climate Data Services, Applications and Services, Environment Canada. November 17, 1997.  176  Environment Canada (EC). 1996. Test for growth and survival in sediment using the freshwater amphipod Hyallela azteca. Preview to Final Manuscript. Methodology Development and Application Section, Technology Development Directorate, Environment Canada, Ottawa, Canada, December 1996. Environment Canada (EC). 1994. Guidance Document on Collection and Preparation of Sediments for Physiochemical Characterization and Biological Testing. Report EPS/RM/29, 132p, December, 1994, Prepared by Gladys Stephenson for Technology Development Directorate, Environment Canada, Ottawa, ON. Environment Canada (EC). 1992. Biological Test Method: Test of Reproduction and Survival Using the Cladoceran Ceriodaphnia dubia. Report EPS/1/RM/21, 72 p., February, 1992, Environment Canada, Conservation and Protection, Ottawa, ON. EVS, 1993. The Effects of Dinoseb and Endosulfan from Agricultural Drainage on the Biota in the Nicomekl River Watershed: Volume 1. Prepared for Environmental Conservation Integrated Programs Branch, Environment Canada. Prepared by EVS Consultants, April, 1993. Fitzmayer, K.M., J.G. Geiger, and M.J. van den Avyle. 1982. Effects of chronic exposure to simazine on the cladoceran, Daphnia pulex. Arch. Environ. Contam. Toxicol. 11 (5): 603-609. As Cited In McLeay (1988). Folmar, L.C., H.O. Sanders, and A.M. Julin. 1979. Toxicity of the herbicide glyphosate and several of its formulations to fish and aquatic invertebrates. Arch. Environ. Contam. Toxicol. 8 (3): 269-278. As Cited In McLeay (1988). Folmar, LC. 1976. Overt avoidance reaction of rainbow trout fry to nine herbicides. Bull. Environ. Contam. Toxicol. 15 (5): 509-514. Frear, D.E., and J. Boyd. 1967. Use of Daphnia magna for the microbioassay of pesticides. I. Development of standardized techniques for rearing Daphnia and preparations of dosage-mortality curves for pesticides. J. Econ. Entomol. 60: 12281236. As Cited In CCME (1991). Fredeen, F.J.H., A.P. Arnason, and B. Berck. 1953. Adsorption of DDT on suspended solids in river water and its role in black-fly control. Nature 171: 700. As Cited In Murty (1986). FREMP. 1996. The Fraser River Estuary. Environmental Quality Report, by the Fraser River Estuary Management Program, Burnaby, B.C. October, 1996. FWPCA. 1968. Water Quality Criteria. Report of the National Tech. Adm. Comm. To Seer, of the Interior Fed. Water Pollution Contr. Adm. U.S.D.I. 234p. As Cited In Pimentel (1971). George, J.P., and H.G. Hingorani. 1982. Herbicide toxicity to fish-food organisms. Environ. Pollut. Ser. A. 28: 183-188.  177  Gersich, F.M. And D.L. Hopkins. 1986. Site-specific and chronic toxicity of ammonia to Daphnia magna straus. Env. Toxicol, and Chem. 5: 443-437. Goodnight, C.J. 1973. The use of aquatic macroinvertebrates as indicators of stream pollution. Trans. Amer. Micros. Soc. 92 (1): 1-13. Goodnight, C.J., and L.S. Whitley. 1961. Oligochaetes as indicators of pollution. Proc. 15th Indust. Waste Conf, Purdue Univ. Eng. Ext. Sen, 106 (45): 139-142. As Cited In Goodnight (1973). Gulley, D., A. Boelter, and H. Bergman. 1989. Toxstat 3.2, Colour Version (Computer Program). To be Used in Conjunction with U.S. EPA (1994a). Fish Physiology and Toxicology Laboratory, Department of Zoology and Physiology, University of Wyoming, Laramie, WY. GVRD, 1988. Report of the Combined Sewer Overflow and Urban Runoff Committee. Liquid Waste Management Plan. Burnaby: Greater Vancouver Regional District. 285p. As Cited In Hagen (1990). Hagen, M.E. 1990. Agricultural Runoff Contamination in the Fraser River Estuary. Waste Management Activity Program Discussion Paper. Fraser River Estuary Management Program. December, 1990. Hartman, W.A., and D.B. Martin. 1985. Effects of four agricultural pesticides on Daphnia pulex, Lemna minor, and Potamogeton pectinatus. Bull. Environ. Contam. Toxicol.  35: 646-651. As Cited In CCME (1989).  Halstead, E.C. 1986. Figure 9, Well Location Map and Hydrogeological Fence Diagram, Township 7, Surrey and Langley District Municipalities, British Columbia. In Groundwater Supply - Fraser Lowland, British Columbia. Inland Waters Directorate - Scientific Series No. 145. Prepared for National Hydrology Research Institute, Environment Canada - NHRI Paper No. 26. Halstead, E.C. 1978. Nicomekl-Serpentine Basin Study, British Columbia. Inland Waters Directorate - Scientific Series No. 94. Prepared for Fisheries and Oceans Canada. Hansen, S.R., and Associates. 1994. Identification and Control of Toxicity in Storm Water Discharges to Urban Creeks. Final Report prepared for Alemeda County Urban Runoff Clean Water Program, Hayward CA. August 1994. As Cited In Katznelson and Mumley (1997). Hatfield, C.T., and J.M. Anderson. 1972. Effects of two insecticides on the vulnerability of Atlantic salmon (Salmo salar) parr to brook trout (Salvelinus fontinalis) predation. J. Fish. Res. Bd. Canada 29: 27-29. Haywood, G.P. 1983. Ammonia Toxicity in Teleost Fishes: A Review. Canadian Technical Report of Fisheries and Aquatic Sciences No. 1177. June, 1983.  178  Henderson, C , and Q.H. Pickering. 1957. Toxicity of organic phosphorous insecticides to fish. Trans. Am. Fish. Soc. 87: 39-51. Hepner, B. 1959. Use of aqueous ammonia in fertilizing fish ponds. Bamidgeh 11:71 -80. As cited in Haywood (1983). Hildebrand, L.D., D.S. Sullivan, andT.P. Sullivan. 1982. Experimental studies of rainbow trout populations exposed to field applications of Roundup® herbicide. Arch. Environm. Contam. Toxicol. 11: 93-98. Holcombe, G.W., G.L. Phipps, and D.K. Tanner. 1982. The acute toxicity of kelthane, dursban, disulfoton, pydrin, and permethrin to fathead minnows Pimephales promelas and rainbow trout Salmo gairdneri. Environ. Pollut. Ser. A. 29: 167-178. Huggins, A.K., G. Skutsch, and E. Baldwin. 1969. Ornithine-urea cycle enzymes in teleostean fish. Comp. Biochem. Physiol. 28(2): 587-602. As cited In Haywood (1983). Hughes, D.N., M.G. Boyer, M.H. Papst, C D . Fowle, G.A.V. Rees, and P. Baulu. 1980. Persistence of three organophosphorous insecticides in artificial ponds and some biological implications. Arch. Environ. Contam. Toxicol. 9: 269-279. Humburg, N.E., S.R. Colby, E.R. Hill, L.M. Kitchen, R.G. Lym, W.J. McAvoy, and R. Prasad. 1989. Herbicide Handbook of the Weed Science Society of America. Sixth Edition. Champaign, Illinois: Weed Science Society of America, 1989. Ingersoll, C.G. and M.K. Nelson. 1990. Testing sediment toxicity with Hyallela azteca (Amphipoda) and Chironomus ripahus (Diptera), p. 93-109 in: Aquatic Toxicology and Risk Assessment, W.G. Landis and W.H. van der Schalle (eds). 13th Volume, ASTM STP 1096, American Society for Testing and Materials, Philadelphia, Pa. As Cited In EC (1996). Iwama, G.K. 1991. Intensive Fish Production. Animal Science 480, Intensive Fish Production, Course Manual. University of British Columbia, Canada. Johnson, W.W., and M.T. Finley. 1980. Handbook of Acute Toxicity of Chemicals to Fish and Aquatic Invertebrates. Summaries of Toxicity Tests Conducted at Columbia National Fisheries Research Laboratory, 1965-78. United States Department of the Interior Fish and Wildlife Service, Resource Publication 137, Washington, D.C., 1980. Jones, J.R.E. 1948. A further study of the reactions of fish to toxic solutions. J. Exp. Biol. 25: 22-34. As cited In Haywood (1983).  179  Katznelson, R., and T. Mumley. 1997. Diazinon in Surface Waters in the San Francisco Bay area: Occurrence and Potential Impact. Prepared for the California State Water Resources Control Board by Woodward Clyde Consultants and the California Regional Water Quality Control Board San Francisco Bay Region. June, 1997. Kendall, A.W., Green, W.J. and David Pascoe. 1986. Studies on the acute toxicity of pollutants to freshwater macroinvertebrates. Arch. Hydrobiol. 106 (1): 61-70. Kersting, K., and H. van der Honing. 1981. Effect of the herbicide Dichlobenil on the feeding and filtering rate of Daphnia magna. Verhandlungen Int. Verein. Theor. Angew. Limnol. 21 (2): 1135-1140. As Cited In McLeay (1988). Lacorte, S., N. Ehresmann, and D. Barcelo. 1995. Stability of organophosphorous pesticides on disposable solid-phase extraction precolumns. Environ. Sci. Technol. 29: 2834-2841. Lalond, V. 1998. Personal Communication with V. Lalond, Planning Division, Department of Engineering, City of Surrey. January, 1998. Lee, H. 1991. A clam's eye view of the bioavailability of sediment-associated pollutants. Chapter 5. Organic Substances and Sediments in Water. Volume 3, Biological. Lewis Publishers, Chelsea, Michigan, pp. 73-93. Levi, G., G. Morisi, A. Caletti, R. atanzaro. 1974. Free amino acids in fish brain: normal levels and changes upon exposure to high ammonia concentrations in viro and upon incubation of brain shoes. Comp. Biochem. Physiol. 49A: 623-636. As Cited in Haywood (1983). Lintott, D.R. 1992. Master's Thesis. University of Saskatchewan, Saskatoon. As Cited In CCME (1993a) Lloyd, R. and L.D. Orr. 1969. The diuretic response by rainbow trout to sub-lethal concentrations of ammonia. Water. Res. 3: 335-344. As Cited In Haywood (1983). Lorz, H.W., S.W. Glenn, R.H. Williams, C M . Kunkel, and L.A. Norris. 1979. Effects of selected herbicides on smolting of coho salmon. U.S. Environ. Protection Agency. Rep. EPA/600/3-79/071. Corvallis, Oreg. As Cited In McLeay (1988). Macek, K.J., D.F. Walsh, J.W. Hogan, and D.D. Holz. 1972. Toxicity of the insecticide Dursban to fish and aquatic invertebrates in ponds. Trans. Amer. Fish. Soc. 3: 420427. Macek, K.J., and W.A. McAllister. 1970. Insecticide susceptibility of some common fish family representatives. Trans. Amer. Fish. Soc. 99 (1): 20-27.  180  M a c e k , K . J . , C . H u t c h i n s o n , a n d O . B . C o p e . 1969. T h e effects of temperature o n the susceptibility of blue-gills a n d rainbow trout to s e l e c t e d pesticides. Bull. Environ. C o n t a m . T o x i c o l . 3: 1 7 4 - 1 8 3 . A s C i t e d In Pimentel (1971). Marchini, S., L. Passerini, D. C e s a r e o , a n d M.L. Tosato. 1988. Herbicidal triazines: a c u t e toxicity on Daphnia, fish, and plants a n d a n a l y s i s of its relationships with structural factors. E c o t o x i c o l . Environ. Saf. 16: 148-157. A s C i t e d In C C M E (1991). M a t t h i e s s e n , P., D. S h e a h a n , R. H a r r i s o n , M. Kirby, R. Rycroft, A . Turnbull, C . V o l k n e r , a n d R. Williams. 1995. U s e of a Gammarus pulex bioassay to m e a s u r e the effects of transient carbofuran runoff from farmland. Ecotoxicol. Environ. Saf. 3 0 : 111-119. Matsumura. F. 1975. Toxicology of Insecticides. P l e n u m , N e w York. A s C i t e d in A n k l e y et al. (1991). Mayer, F.L. Jr., a n d M . R . Ellersieck. 1986. M a n u a l of A c u t e Toxicity: Interpretation a n d D a t a B a s e for 4 1 0 C h e m i c a l s a n d 6 6 S p e c i e s of F r e s h w a t e r A n i m a l s . United States Department of the Interior F i s h a n d Wildlife S e r v i c e . R e s o u r c e P u b l i c a t i o n 160. W a s h i n g t o n , D . C . , 1986. M c L e a y , D . J . 1988. D e v e l o p m e n t of a B i o a s s a y Protocol for E v a l u a t i n g the T o x i c R i s k to R e g i o n a l F i s h e r i e s R e s o u r c e s P r o p o s e d by F o r e s t - U s e H e r b i c i d e s . C a n a d i a n Forestry Service and the British C o l u m b i a Ministry of F o r e s t s , O c t o b e r 1988, F R D A Report 0 3 9 , I S S N 0 8 3 5 0 7 5 2 . M C P A T a s k F o r c e . 1987. Static A c u t e Toxicity of T e c h n i c a l M C P A Dimethylamine Salt (2-methyl-4-chlorophenoxyacetic acid, dimethylamine salt) to R a i n b o w Trout, Salmo gairdneri. Industry T a s k F o r c e o n M C P A R e s e a r c h D a t a . R e s e a r c h T r i a n g l e Park, North C a r o l i n a . A s C i t e d In C C M E (1995). M e n c o n i , M., a n d C . C o x . 1994. H a z a r d a s s e s s m e n t of the insecticide d i a z i n o n to a q u a t i c organisms in the S a c r a m e n t o - S a n Joaquin River system. Administrative Report 9 4 2, C a l i f o r n i a Department of F i s h ; a n d G a m e , R a n c h o C o r d o v a , C A . A s C i t e d In D a v i s et al. (1997). Moore, M.T., D.B. Huggett, W . B . G i l l e s p i e Jr., J . H . R o d g e r s Jr., a n d C M . C o o p e r . 1998. Comparative toxicity of chlordane, chlorpyrifos, a n d aldicarb to four a q u a t i c testing o r g a n i s m s . A r c h . Environ. C o n t a m . T o x i c o l . 34: 152-157. M o r g a n , H . G . 1976. S u b l e t h a l Effects of D i a z i n o n o n S t r e a m Invertebrates. Dissertation. Univ. G u e l p h , Ontario, C a n a d a , 157 pp.  Ph.D.  Mount, D.I., a n d L. A n d e r s o n - C a r n a h a n . 1988. Methods for A q u a t i c Toxicity Identification E v a l u a t i o n s : P h a s e I Toxicity C h a r a c t e r i z a t i o n P r o c e d u r e s . E P A - 6 0 0 / 3 - 8 8 - 0 3 4 . Environmental R e s e a r c h Laboratory, Duluth, M N . A s C i t e d in A n k l e y et al. (1991).  181  Muirhead-Thomson, R.C. 1978. Lethal and behavioral impact of chlorpyrifos methyl and temephos on select stream macroinvertebrates: experimental studies on downstream drift. Arch. Environm. Contam. Toxicol. 7: 139-147. Mulkey, L.A., and A.S. Donigan, Jr. 1984. Modeling alachlor behavior in three agricultural river basins. U.S. EPA, Environmental Research Laboratory, Athens, GA. As Cited In Donigian (1986). Murphy, S.D. 1975. Pesticides. Pages 408-453 in L.J. Casarett and J. Doull, eds. Toxicology, The Basic Science of Poisons. Macmillan Publishing Co., New York. 768 pp. As Cited In Smith (1987). Murty, A.S. 1986. The Toxicity of Pesticides to Fish. Vol. 2. Boca Raton, Florida: CRC Press, 1986. NAS. 1973. Water Quality Criteria, 1972. A Report of the Committee on Water Quality Criteria, Environmental Studies Board, National Academy of Sciences, National Academy of Engineering, Washington, D.C. As Cited In Davis et al. (1997). Nebeker, A.V., S.T. Onjukka, D.G. Stevens, G.A. Chapman, and S.E. Dominguez. 1992. Effects of low dissolved oxygen on survival, growth and reproduction of Daphnia, Hyallela and Gammarus. Environ. Toxicol. Chem. 11:373-379. As Cited In EC (1996). Nimmo, D.R., D. Link, L.P. Parrish, G L . Rodriguez, and William Wuerthele. 1989. Comparison of on-site and laboratory toxicity tests: derivation of site-specific criteria for un-ionized ammonia in a Colorado transitional stream. Env. Toxicol. Chem., 8: 1177-1189. NRC, Date unknown. National Research Council MESS-2 Certified Reference Material. Marine Sediment Reference Materials for Trace Metals and Other Constituents. Institute for Environmental Research and Technology, National Research Council of Canada, Ottawa, Ontario, Canada. Ort, M.P., J.F. Fairchild, and S.E. Finger. 1994. Acute and chronic effects of four commercial herbicide formulations on Ceriodaphnia dubia. Arch. Environ. Contam. Toxicol. 27: 103-106. Palawski, D., D.R. Buckler, and F.L. Mayer. 1983. Survival and condition of rainbow trout (Salmo gairdneri) after acute exposures to methyl parathion, triphenyl phosphate, and DEF. Bull. Environ. Contam. Toxicol. 30: 614-620. Pennak, M. 1978. Fresh-Water Invertebrates of the United States, 2nd ed. John Wiley & Sons, New York, NY. As Cited In EC (1992). Phipps, G.L., V.R. Mattson, and G.T. Ankley. 1995. Relative sensitivity of three freshwater benthic macroinvertebrates to ten contaminants. Arch. Environ. Contam. Toxicol. 28: 281-286. As Cited In Moore et al. (1998).  182  Pimentel, D. 1971. Ecological Effects of Pesticides on Non-Target Species. Executive Office of the President. Office of Science and Technology. June 1971. U.S. Gov. Printing Office, Stock No. 4106-0029. Presing, M. 1981. On the effects of dikonirt (sodium salt of 2,4-dichlorophenoxy-acetic acid) on the mortality and reproduction of Daphnia magna. Hydrobiologia 83 (3): 511-516. As Cited In McLeay (1988). Pusey, B.J., A.H. Arthington, and J. McLean. 1994. The effects of a pulsed application of chlorpyrifos on macroinvertebrate communities in an outdoor artificial stream system. Ecotoxicol. Environ. Saf. 27: 221-250. As Cited in Ward et al. (1995). Rao, P.S.C., and J.M. Davidson. 1980. Estimation of pesticide retention and transformation parameters required in nonpoint source pollution models. In Environmental Impact of Nonpoint Source Pollution. Eds. M. R. Overcash and J.M. Davidson. Ann Arbor, Michigan. Read, L.J. 1971. The presence of high ornithine-urea cycle enzyme activity in the teleost Opsanus tau. Comp. Biochem. Physiol. 39(2B): 409-413. As Cited In Haywood (1983). Rice, S.D. and Robert M. Stokes. 1975. Acute toxicity of ammonia to several developmental stages of rainbow trout, Salmo gairdneri. Fish. Bull. 73 (1): 207-211. Rhidine, Miles. 1997. Personal Communication with Miles Rhidine, Manager of Nicomekl Fish Hatchery. Rhone-Poulenc. 1992. MCPA DMAS-Acute toxicity to daphnids (Daphnia magna) under flow-through conditions. SLI Report #92-4-4235. Rhone-Poulenc AG Company, Research Triangle Park, North Carolina. As Cited In CCME (1995). Robinson, ST. 1988. Impacts of Agricultural Drainage and an Assessment of Diffused Aeration in the Serpentine River, British Columbia. M.A.Sc. Thesis, Department of Civil Engineering, University of British Columbia, Canada. October, 1988. 133 pp.. Sanders, H.O. 1969. Toxicity of pesticides to the crustacean, Gammarus lacustris. Tech. Paper 25, Bur. Sprot Fish. Wildl., U.S.D.I. 18p. As Cited In Pimentel (1971). Sanders, H.O., and O.B. Cope. 1968. The relative toxicities of several pesticides to naiads of three species of stoneflies. Limnol. Oceanogr. 13:112-117. As Cited In Pimentel (1971). Sanders, H.O., and O.B. Cope. 1966. Toxicities of several pesticides to two species of cladocerans. Trans. Am. Fish. Soc. 95: 165-169. As Cited In CCME (1993a)(1986).  183  Schober, U. and W. Lampert. 1977. Effects of sublethal concentrations of the herbicide atrazine on growth and reproduction of Daphnia pulex. Bull. Environ. Contam. Toxicol. 17 (3): 269-277. As Cited in McLeay (1988). Sethunathan, N., and M.D. Pathak. 1972. Increased biological hydrolysis of diazinon after repeated application in rice paddies. J. Agr. Food Chem. 20 (3): 586-589. Shephard, B. 1983. The Effect of Reduced pH and Elevated Aluminum Concentrations on Three Species of Zooplankton: Ceriodaphnia reticulata, Daphnia magna and Daphnia pulex. U.S. Environmental Protection Agency, Duluth, Minnesota. 14pp. As Cited In CCME (1996). Skidmore, J.F. 1965. Resistance to zinc sulphate of the zebrafish (Brachydanio rerio Hamilton-Buchanan) at different phases of its life history. Ann. Appl. Biol. 56: 4753. As cited in Rice and Stokes (1975). Smith, C.E. and R.G. Piper. 1975. Effects of Metabolic Products on the Quality of Rainbow Trout. U.S. Dept. Interior Fish and Wildlife Service Leaflet No. 4: 10p. As cited in Haywood, (1983). Smith, G.J. 1987. Pesticide Use and Toxicology in Relation to Wildlife: Organophosphorous and Carbamate Compounds. United States Department of the Interior Fish and Wildlife Service, Resource Publication 170, Washington, D.C. 1987. Smith, S.L., D.D. Macdonald, K.A. Keenleyside, and C.L. Gaudet. 1995. Development and Implementation of Canadian Sediment Quality Guidelines. Ecovision World Monograph Series (in press). As Cited in FREMP (1996). Song, M. Y., J.D. Stark, and J.J. Brown. 1997. Comparative toxicity of four insecticides, including imidacloprid and tebufenozide, to four aquatic arthropods. Environ. Toxicol. Chem. 16: 2494-2500. Stephan, C.E. 1977. Methods for Calculating an L C , P. 65-84 in: Aquatic Toxicology and Hazard Evaluation, F.L. Mayer and J.L. Hamelink (eds.), American Society for Testing and Materials, ASTM STP 634, Philadelphiaa, PA. 50  Suedel, B.C., and J.H. Rodgers Jr. 1994. Responses of Hyallela azteca and Chironomus tentans to particle-size distribution and organic matter content of formulated and natural freshwater sediments. Environ. Toxicol. Chem. 13:1639-1648. As Cited In EC (1996). Summerfelt, R.C. and W.M. Lewis. 1967. Repulsion of green sunfish by certain chemicals. J. Wat. Pollut. Control. Fed. 39: 2030-2038. As Cited In Haywood (1983). Supelco, 1996. Supelco Chromatography Products. Sales Catalogue. Supelco Inc. pp. 359-367.  184  Swain, L.G., and D.G. Walton. 1994. Survey of Sediments and Tissues from Boundary Bay and Roberts Bank. B.C. Ministry of Environment, Lands, and Parks. March, 1994. Swain, L.G., and G.B. Holms. 1988a. Fraser-Delta Area Boundary Bay and its Tributaries: Water Quality Assessment and Objectives. Prepared for B.C. Ministry of Environment and Parks, Water Management Branch, Province of British Columbia, February, 1988. Swain, L.G., and G.B. Holms. 1988b. Water Quality Assessment and Objectives FraserDelta Area Boundary Bay and its Tributaries. Technical Appendix. Prepared for B.C. Ministry of Environment and Parks, Water Management Branch, Province of British Columbia, February, 1988. Swartz, R.C., F.A. Cole, J.O. Lamberson, S.P. Ferraro, D.W. Schults, W.A. Deben, H. Lee, and R.J. Ozretich. 1994. Sediment toxicity, contamination, and amphipod abundance at a DDT and dieldrin-contaminated site in San Francisco Bay. Environ. Toxicol. Chem. 13: 949-962. As Cited in EC (1996). Swartz, R.C., F.A. Cole, D.W. Schults, and W.A. Deben. 1986. Ecological changes in the southern California bight near a large sewage outfall: benthic conditions in 1980 and 1983. Marine Ecology 31: 1-13. As Cited in EC (1996). Swartz, R.C., D.W. Schults, G.R. Ditsworth, W.A. Deben, and F.A. Cole. 1985. Sediment Toxicity, Contamination, and Macrobenthic Communities Near a Large Sewage Outfall, p. 152-175 in: Validation and Predictability of Laboratory Methods for Assessing the Fate and Effects of Contaminants in Aquatic Ecosystems. T.P. Boyle, (ed). ASTM STP 865, American Society for Testing and Materials, Philadelphia, Pa. As Cited in EC (1996). Swartz, R.C., W.A. Deben, K.A. Sercu, and J.O. Lamberson. 1982. Sediment toxicity and the distribution of amphipods in Commencement Bay, Washington, USA. Mar. Poll. Bull. 13: 359-364. As Cited In EC (1996). Symons, P.E.K. 1973. Behavior of young Atlantic salmon (Sa/mo salar) exposed to or force-fed fenitrothion, an organophosphate insecticide. J. Fish. Res. Board Can. 30: 651-655. Szumski, D.S., D.A. Barton, H.D. Tutnam, and R.C. Polta. 1982. Evaluation of EPA Unionized Ammonia Toxicity Criteria. J. Water Pollut. Cont. Fed. 54: 281-291. Thurston, R.V., R.C. Russo, R.J. Luedtke, C.E. Smith, E.L. Meyn, C. Chakoumakos, K.C. Wang, and C.J.D. Brown. 1984. Chronic toxicity of ammonia to rainbow trout. Trans. Am. Fish. Soc. 113: 56-73. Thurston, R.V. and R.C. Russo. 1983. Acute toxicity of ammonia to rainbow trout. Trans. Am. Fish. Soc. 112: 696-704.  185  Thurston, R.V. and R.C. Russo. 1981a. Ammonia toxicity to fishes. Effect of pH on the toxicity of the un-ionized ammonia species. Env. Sci. and Technol., 15 (7): 837840. Thurston, R.V., G.R. Phillips, and R.C. Russo. 1981b. Increased toxicity of ammonia to rainbow trout (Salmo gairdneri) resulting from reduced concentrations of dissolved oxygen. Can. J. Fish. Aquat. Sci. 38: 983-988. Thurston, R.V., R.C. Russo, and K. Emerson. 1979. Aqueous Ammonia Equilibrium Tabulation of Percent Un-ionized Ammonia. Prepared for U.S. Environmental Protection Agency, Office of Research and Development, Environmental Research Laboratory, Duluth, Minnesota. EPA 600/3-79-091, August, 1979. Town, C A . 1986. Instream Aeration of the Serpentine River. M.A.Sc. Thesis, Department of Civil Engineering, University of British Columbia, Canada. June, 1986. 117 pp.. U.S. Department of Agriculture (USDA). 1984. Simazine. In Pesticide Background Statements. Vol. 1. Herbicides. Agricultural Handbook No. 633, U.S. Department of Agriculture Forest Service, Washington, D.C. pp. S1-S92. As Cited In CCME (1991). U.S. EPA. 1996. Reregistration Eligibility Decision (RED), Prometryn. Office of Prevention, Pesticides and Toxic Substances, United States Environmental Protection Agency. July, 1996. EPA-738-R-95-028. U.S. EPA. 1994a. Short Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms. Third Edition. Environmental Monitoring Systems Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, Cincinnati, Ohio. EPA/600/4-91/002. U.S. EPA. 1994b. Methods for measuring the Toxicity and Bioaccumulation of SedimentAssociated Contaminants with Freshwater Invertebrates, 133p., U.S. Environmental Protection Agency, Duluth, Minn. EPA 600/R-94/024. As Cited In EC (1996). U.S. EPA. 1993a. Methods for Measuring the Acute Toxicity of Effluents and Receiving Waters to Freshwater and Marine Organisms. Fourth Edition., Environmental Monitoring Systems Laboratory, Cincinnati Office of Research and Development, U.S. Environmental Protection Agency, Cincinnati, Ohio. EPA/600/4-90/027F. August, 1993. U.S. EPA. 1993b. ICPIN 2.0 (Computer Program). A Linear Interpolation Method for Sublethal Toxicity: The Inhibition Concentration (Icp) Approach. Environmental Research Laboratory, U.S. Environmental Protection Agency, Duluth, MN. June, 1993. US EPA, 1986. Quality Criteria for Water. U.S. Environmental Protection Agency, Washington, D.C EPA 440/9-76-023. As Cited In Davis et al. (1997).  186  US EPA, 1985a. Ambient Water Quality Criteria for Ammonia -1984. U.S. Environmental Protection Agency, Office of Research and Development, Environmental Research Laboratory, Duluth, Minnesota. EPA 440/5-85-001, January, 1985. U.S. EPA. 1983. Water Quality Criteria for the Protections of Aquatic Life and its Use: Ammonia. Final Draft. Office of Research and Development. Environmental Research Laboratory, Duluth, MN. As Cited in Gersich and Hopkins (1985). Vilkas, A.G. 1976. Acute Toxicity of CGA-24705 Technical to the Water Flea Daphnia magna. (Unpublished study received Nov. 23, 1976, under 100-587; prepared by Aquatic Environmental Sciences, Union Carbide Corp., for Ciba-Geigy Corp., Greensboro, N.C.; CDL: 226955-C). As Cited In CCME (1991). Wan, M.T., S. Szeto, and P. Price. 1995. Distribution of endosulfan residues in the drainage waterways of the Lower Fraser Valley of British Columbia. J. Environ. Sci. Health. 30 (3): 401-433. Wan, M.T., S. Szeto, and P. Price. 1994. Organophosphorous insecticide residues in farm ditches of the Lower Fraser Valley of British Columbia. J. Environ. Sci. Health. 29 (5): 917-949. Wan, M.T. 1989. Levels of selected pesticides in farm ditches leading to rivers in the Lower Mainland of British Columbia. J. Environ. Sci. Health, 24 (2): 183-203. Ward, S., A.H. Arthington, and B.J. Pusey. 1995. The effects of a chronic application of chlorpyrifos on the macroinvertebrate fauna in an outdoor artificial stream system: species responses. Ecotoxicol. Environ. Saf. 30: 2-23. Ware, G.W. 1978. The Pesticide Book. Chapter 6: Insecticides. W.H. Freeman and Company, San Francisco. WCC, 1996. Sediment Diazinon Special Study. Report prepared by Woodward Clyde Consultants for Alameda Countywide Clean Water Program, Hayward CA, December 1996. As Cited In Katznelson and Mumley (1997). Weis, J.S., and P. Weis. 1975. Retardation of fin regeneration in Fundulus by several insecticides. Trans. Amer. Fish. Soc. 104: 135-137. Weiss, C M . 1961. Physiological effect of organic phosphorous insecticides on several species offish. Trans. Amer. Fish. Soc. 90: 143-152. Wenger, D.P. 1973. The Effects of Endrin on the Developmental Stages of the Rainbow Trout, Salmo gairdneh. M.S. Thesis, Kent State Univ., Kent, Ohio. As cited in Rice and Stokes (1975).  187  West, C.W., V.R. Mattson, E.N. Leonard, G.L. Phipps, and G.T. Ankley. 1993. Comparison of the relative sensitivity of three benthic invertebrates to coppercontaminated sediments from Keweenaw Waterway. Hydrobiologia 262: 57-63. As Cited In EC (1996). Wood, B.J. 1997. Excerpts from-A Bioassessment of Agrochemicals Entering a Freshwater System: Laboratory and In Situ Bioassays Using Daphnia pulex. M.Sc. Thesis, Environmental Science and Regional Planning Program, Washington State University. Zar, J.H. 1984. Biostatistical Analysis. Second Edition. 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 m /s at this site. Pump. = discharge, no irrigation at this site. 3  May  Pump  June  Pump  July  Pump  Aug.  Pump  4.5  1  0  1  0  1  1.9  1  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 0  11  0  11  1.5  11  0.8  12  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  18  0  18  0.7  18  0.1  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 m /s at this site. Pump. = discharge, no irrigation at this site. 3  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  7.4  27  0.8  27  1.7  28  4.4  28  4.9  29  1.1  29  10  30  2  30  8.2  31  2.3  190  Appendix  1  Cont.  Ericson Ditch's (Site 10) Multiple Pumps' Cumulative Daily Hours of M a v - A u g . . 1 9 9 7 Times shown are the hours between the mornin the following day. Greater than 24 pump-hours are 2pumps at this station. In addition, greater day is possible solely due to pump hours being following day. Average d i s c h a r g e =2 . 0 m / s p e r intake. 3  g of the on asin than 24 recorded pump.  Operation  day shown and the morning of gle day is possible since there hours of pumping on a singe at aslightlylater time the P u m p . =d i s c h a r g e , I r r i g .=  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 m /s per pump. Pump. = discharge, Irrig. = intake. 3  Sept.  Pump  Irrig  Oct.  Pump  Irrig  1  0  0  1  7.5  2  0  0  2  4.7  Nov.  Pump  Irrig  0  1  5.1  0  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 m /s per pump. Pump. = discharge, Irrig. = intake. 3  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 m /s per pump. Pump, discharge, Irrig. = intake. 3  Sept.  Pump  Irrig  Oct.  Pump  Irrig  Nov.  Pump  Irrig  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0.8 9.1 2.7 0 0 0 0 0 0 0 0 3.9 3.9 9.4 13.5  0 0 0 0 0 0 0 0 25.3 4.7 24 4.8 4.8 4.8 9.4 0.4 0 0 0 0 0 0 0 0 0 0 0 0 0 0  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31  8.4 5.6 12.2 12.2 12.2 5.5 3.5 5.8 5.7 3.4 3.4 3.4 3.4 4.6 7.9 4.6 2.1 2.1 2.1 0.5 0.8 2.2 2 3.8 3.8 3.8 5.3 9.4 22.1 18.4 7.3  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27  7.3 7.3 4.5 7.5 7.4 9.6 5.8 5.8 5.8 0.2 0.2 11.1 2.4 0.9 0.9 0.9 4 6.8 6.8 9.7 8.2 8.2 8.2 13 11.3 9.2 20.9  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 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 m /s per pump. Pump. = discharge, Irrig. = intake. 3  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  2.6  0  27  0  0  27  1.6  0  27  0  10.6  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 m /s per pump. Pump, discharge, Irrig. = intake. 3  Sept.  Pump  Irrig  Oct.  Pump  Irrig  Nov.  Pump  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30  0 0 0 0 0.5 0.5 0.5 0 0 0 0 0 0 0 0 5.3 8.4 3.1 2.6 2.6 2.6 0 1.7 0.9 2.7 6.2 6.2 6.2 3.6 2.0  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31  9 5.1 13.6 13.6 13.6 5.4 3.2 6.4 5.5 3.4 3.4 3.4 3.4 5.4 10.5 4.7 2.6 2.6 2.6  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27  9.8 9.8 19.5 10 10.6 14.7 6.9 6.9 6.9 4.6 4.6 3.2 3.3 1.5 1.5 1.5 7 6.4 13.5 17.5 12.2 12.2 12.2 21.4 20.3 12 14.2  .1.1 1.6 3.4 3.2 5 5 5 6.9 11.6 26.1 22.5 9.8  Irrig  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 m /s per pump. Pump, discharge, no irrigation at this ditch this particular year. 3  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 m /s per pump. Pump, discharge, no irrigation at this ditch this particular year. 3  Sept.  Pump  Oct.  Pump  Nov.  Pump  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31  0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 5.4 1.3 0.4 0.4 0.4 0 0.3 0.1 0.9 1.4 1.4 1.4 1.2 0.7 0.7  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31  2.9 1.6 4.8 4.8 4.8 2.3 1.6 2.5 1.7 1.2 1.2 1.2 1.2 1.9 3.3 1.7 0.9 0.9 0.9 0.3 0.5 0.9 0.7 1.5 1.5 1.5 1.6 4.9 11.5 7.7 3.4  1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27  3.4 3.4 6.7 3.4 3.5 4.6 2.3 2.3 2.3 1.3 1.3 1 1.2 0.5 0.5 0.5 2 1.9 4.1 3.9 3.7 3.7 3.7 7.4 6.2 4 11.6  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  Discharge From Site 8  21.2  2.5  Culture Control  18.1  6.3  June 6 1997  Mean # Neonates Per Female  Standard Deviation  Versus Culture Control (p=, 1 tail)  Versus Culture Control (p=,2 tail)  0.0007 *  0.001 *  Versus Site 13 Control  Versus Site 13 Control  (P=, 1 tail)  (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  Versus Site 13 Control  (P=, 1 tail)  (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" *  2.3 X 10 *  1.5 X10" *  2.9 X10" *  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  July 16 1997  Mean # Neonates Per Female  Standard Deviation  7  7  6  0.278  0.553  Versus Site 13 Control  Versus Site 13 Control  (p=, 1 tail)  (P=, 2 tail)  Versus Culture Control. (P=, 1 tail)  6  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  Versus Site 13 Control  (p=, 1 tail)  (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  Aug. 29 1997  Mean # Neonates Per Female  Standard Deviation  0.214  0.428  Versus Site 13 Control  Versus Site 13 Control  (p=, 1 tail)  (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  Versus Site 13 Control  (p=, 1 tail)  (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"*  6.8 X 10 *  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" *  9.5 X 10 *  Site 10  22.6  2.9  0.308  0.616  2.4 X10" *  4.9 X10" *  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" *  7.4 X 10" *  Site 13  22.0  2.4  8.6 X 10" *  1.7 X 10" *  W.R.P. Cont.  13.6  2.2  Oct. 3 1997  Mean # Neonates Per Female  Standard Deviation  5  9  7  5  8  8.6X10 *  1.7 X10" *  Versus Site 13  Versus Site 13  (P=, 1 tail)  (P=, 2 fail)  8  s  9  7  5  7  7  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  Versus Site 13 Control  (P=, 1 tail)  (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  Oct. 16 1997  Mean # Neonates Per Female  Standard Deviation  0.370  0.740  Versus Site 13 Control  Versus Site 13 Control  (P=, 1 tail)  (P=, 2 tail)  Versus Culture Control (p= 1 tail)  0.962  0.395  0.791  0.0007 *  0.001 *  0.370  0.740  Site 5  18.8  5.3  0.481  Site 10Nearby Field Puddle  8.6  4.5  1.8 X 10" *  Site 13  18.9  4.0  Culture Control  18.1  6.3  Nov. 21 1997  Mean # Neonates Per Female  Standard Deviation  5  3.7 X10" * 5  0.370  0.740  Versus Site 13 Control  Versus Site 13 Control  (p=, 1 tail)  (p= 2 tail)  Versus Culture Control (P=, 1 tail)  Versus Culture Control (P=,2 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  Ceriodaphnia dubia Successive Reference Toxicant Results 2000 1800 1600 cn E 1400 1200 o LO 1000 g 800 g 600 o 400 T3 O 200 i_ Q_ 0 CD O ro  «» ...  -  .__  f : t"  «»  1  »  —  «  -  I  4»  <»  II  « .  |!  «•  '  o - ....  -  . ...  _  -  T  •  '  1  -  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 St. Ditch. June 26 Sample th  th  8000070000600005000040000 30000-  60.00  Site 5. South Cloverdale Ditch. July 16th Sample l&bundanoe 140000 -I  120000]  100000]  80000-^  60000-^  40000-^  TIC: 0716S5.D  205  Appendix 5 Prometryn GC/MS. Full-lon-Scan. Standards' Curves  Prometryn GC MS Standards' Curves Including Site 8, J u n e 26, 1997 10000000 9000000 8000000 7000000 6000000 5000000 4000000 3000000 2000000 1000000 0  CO 0  CO CD Q_ CO  o o  0  3  4  5  6  7  10  mg/L Prometryn in GC-MS Acet. Std.  M. Chi. Std. -figr Site 8 - LL  Site 8-SPE  Prometryn GC MS Standards' Curves Including Site 8, J u n e 2 6 , 1997 1000000 900000 800000 700000 $ 600000 tt. 5 0 0 0 0 0 CO 40000CT ^ 300000 O 200000 100000 0  m  $ <  III : . | r . -  3  —m  0  0.2  0.4  0.6  0.8  1.2  1.4  1.6  1.8  mg/L Prometryn in GC-MS Acet. Std.  • M. Chi. Std. -Egg-  Site 8 - LL  - 0.003 mgfL  —|— Site 8-SPE 0.002 mgl.  206  Appendix 6 C. dubia Lethally Toxic Samples' GC/MS Chromatograms. Select-Ion Scan. 3600X Concentration Site 8. 176 St. Ditch. June 26 Sample th  th  |Abundance.  TIC: 9.D  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 Mirny-*  '  I  '''•i••••i••  10.00  15.00  20.00  • ' l  ' • ' • l '.  25.00  30.00  '  '  i  li  35.00  '  i  ' I  40.00  i  ' ' ' l > '  45.00  ' ' I  50.00  i  '  i  i  l ' ' ' ' l  55.00  60.00  i  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 TIC: 9.D  3000 2500 2000 1500 1000 500 'I 'I I ' I ' ' I I I ' I ' ' I ' I ' I ' I ' I I I ' I 0 ' 3ll00 i | ' i ' i i ' i ' ' | ' 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 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  h m e r »  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 St. Ditch. June 2 6 Sample CHLORPYRIFOS th  th  Abundance  Average of 34.984 to 34.995 min.: 9.D 1*7  1800 1600 1400  314  1200-1 1000 137 800I7S 600400 200 304 '  130  140  150  160  170  180  '  '  | • ' '  190  |  '  200  I I I | ' II  210  I | I '  220  I ' | ' ' I I | I I I I | I I I I | | | ,  230  240  250  260  I | | ,  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 Sample DIAZINON th  Abundance  TIC: 8.D  2400  2200  2000  1800  1600  1400  1200  1000  fnme-*  Imfc-^  31122  0 'i | '  31 !24 31J26 31128  1 ' 1 | ' 1  ••1  1 1  31130131132 31134  i 11 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 | 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  ''1 ''''1 ''''1 ''''1 • '''1 ''''1  ' ' ' ' I  ''' I' 1  ' ' ' I ' '  1  ' I ' ' ' ' I ' ' ' ' I ' ' ' ' I  ''''I'  130 140 150 160 170 180 190 200 210 220 230 240 250 260 270 280 290 300 310  ' ' ' I '  1  ' ' I  320  ' ' 1  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 ' i 1 ' i 1 ' 1 1 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 1 1  rrime->  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  212  Appendix 8 Cont. Diazinon and Chlorpyrifos Standards' Select-lon-Scan Chromatogram Peaks and Mass Spectra CHLORPYRIFOS. 66 uall Standard  Average of 34.988 to 34.999 min.: 3.D I £7  abundance 130  314  120 110 100 90 80 70 60 50 40  137  30f  179  30 20 10 0  1  1  130  1  1  1•  140  11 1  1 ••1  150  11  160  1 1 11  170  1  11  180  1 •• 1• 1  190  1  200  1  1  210  1 , 11  1  220  1111  1  230  1 11 1  1  240  1111  1  250  1 1 1 1  1  260  1 1 1 1  1  270  1 1 1 1  1  280  1 1 1 1  1  290  1 1 1 1  1  300  1 1 1 1  1 ' 1 1  310  1 1  320  213  Appendix 9 Diazinon's Select-lon-Scan GC/MS Standards' Curve  Diazinon GC MS Standards' Curve 0.8 micrograms/L SPE, 500X Cone.  600000 CD 550000 2500000 <450000 -£400000 2350000 ooooooc oooooo oooooo oooooo  d SIAI-OFJ  cl.,".  0  .|  1 - j" -"• •• • j ••'' -•'...-; '-y | • " j - - •' " • : : I • - • • j '•• : . ' | J • : " ' ' • •.: j . '. : • . :  1  .>|  100 200 300 400 500 600 700 800 micrograms/L Diazinon in GC-MS Chlorpyrifos Stds.  -CO  -  900  0.8 micrograms/L SPE, 500X Cone.  = 0.4 micrograms/L  Diazinon GC MS Standards' Curve Including Sites 8 June 26 & 5 July 16  _50000 <D 45000 <40000 35000 g30000  Q-25000  CO 20000 ^15000 O10000 O 5000 0 40 60 80 100 micrograms/L Diazinon in GC-MS -g^- O  Chlorpyrifos Stds. s i t e  = 0.023 micrograms/L  8, 3600X Cone.  -FR- Site 5, 3600X Cone.  = 0.033 micrograms/L  0.4 micrograms/L SPE, 500X Cone. :  0.1 micrograms/L  120  140  214  Appendix 10 Chlorpyrifos' Select-lon-Scan GC MS Standards' Curve  Chlorpyrifos GC MS Standards' Curve Including Site 8, June 26, 1997  1100000 (0 1000000 £ 900000 800000 lo 700000 ® 600000 500000 CO 400000 A 300000 O 200000 100000 0 <  0  400  600  m i c r o g r a m s / L Chlorpyrifos Stds.  800  1000 1200  C h l o r p y r i f o si n  G C - M S  -gg$- Site 8, 3600X Cone. = 0.067 micrograms/L  Chlorpyrifos GC MS Standards' Curve 0.066 micrograms/L S P E , 5 0 0 X C o n e . 30000 CO CD 27000 24000 21000 CO CD 18000 D_ 15000 CO 12000 9000 o 6000 3000 O 0 60  80  100  120  micrograms/L Chlorpyrifos in GC-MS Chlorpyrifos Stds.  0.066 micrograms/L SPS = 0.16 micrograms/L based on standards  140  215  Appendix 11 C. dubia Chronic Test. Ditch. River, and Control Water Samples' Chemistry May 6 1997  Initial DO (mg/L), 1st Day of Aeration  Initial PH  ?, 1  Total Hardness (as mg/L CaC03)  (//S/cm)  6.1  75  8.4  ?  Initial DO (mg/L), 1st Day of Aeration  Site 1  Cond.  TDC  TDIC  TDOC  Color  (mg/L)  (mg/L)  (mg/L)  (mg/L Pt)  190  ?  ?  ?  ?  68  ?  ?  ?  ?  ?  Initial PH  Total Hardness (as mg/L CaC03)  Cond. OuS/cm)  TDC (mg/L)  TDIC (mg/L)  TDOC (mg/L)  Color (mg/L Pt)  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  Discharge From Site 8 Culture Control  June 6 1997  1  1  216  Appendix 11 Cont  Initial DO (mg/L), 1st Day of 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  Initial DO (mg/L), 1st Day of 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  June 26 1997  July 16 1997  1  1  217  Appendix 11 Cont  Aug. 7 1997  Initial D O (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  Site 1  9.1, 5  7.8  Site 2  8.9, 5  Site 3  Cond. (jUS/cm)  TDOC (mg/L)  TDIC (mg/L)  TDIC (mg/L)  Color (mg/L Pt)  93  275  27.2  17.9  9.3  75  7.7  90  265  24.3  17.6  6.7  40  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  Site5  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  Initial D O (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  Cond. (piS/cm)  TDOC (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  Aug. 29 1997  1  1  218  Appendix 11 Cont Sept. 17 1997  Initial DO (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  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  Initial DO (mg/L), 1st Day of 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  Oct. 3 1997  1  1  Cond.  219  Appendix 11 Cont. Initial D O (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  GuS/cm)  TDOC (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  ?  ?  ?  ?  ?  Initial D O (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  Cond. (jUS/cm)  TDOC (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  ?  ?  ?  ?  TDIC (mg/L)  TDIC (mg/L)  Color (mg/L Pt)  Oct. 3 1997 Retests  Oct. 16 1997  Culture Control  Nov. 21 1997 Site 1  1  1  Cond.  ?, 1  7.2  80  ?  Initial D O (mg/L), 1st Day of Aeration  Initial PH  Total Hardness (as mg/L CaC03)  Cond.  10.0, 1  7.3  67  204  28.8  9.3  19.5  100  1  TDOC (pcS/cm) (mg/L)  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 Final Initial Total Total Cond. NH3-N NH3-N //S/cm mg/L mg/L  Final Cond. //S/cm  Initial Final Total Total Hard. Hard. mg/L as mg/L as CaC03 CaC03  D.O. Range mg/L  Initial Final pH 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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