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Effects of cooling water discharge from a thermoelectric power plant on the nutrient and phytoplankton… Henry, Michael Francis 2005

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EFFECTS OF COOLING WATER DISCHARGE FROM A THERMOELECTRIC POWER PLANT ON THE NUTRIENT AND PHYTOPLANKTON DYNAMICS IN PORT MOODY ARM, BRITISH COLUMBIA, CANADA by M I C H A E L FRANCIS H E N R Y B.Sc. University of Western Ontario, London 1994 A THESIS SUBMITTED IN PARTIAL F U L F I L L M E N T OF THE REQUIREMENT FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE F A C U L T Y OF G R A D U A T E STUDIES (Oceanography) THE UNIVERSITY OF BRITISH C O L U M B I A December 2005 © Michael Francis Henry Abstract This thesis examines the influence of thermoelectric generation, particularly its nutrient loading effect, on the phytoplankton dynamics in Port Moody Arm (PMA), British Columbia, Canada, a shallow tidal inlet near the city of Vancouver. Spatiotemporal trends of phytoplankton biomass and composition were investigated over a 2V4 year period. These were related to 'natural' environmental factors within P M A and the influence of Burrard Generating Station (BGS), a 950 M W natural gas-fired electric utility that is permitted to withdraw 1.7xl0 6 m 3 d"1 from P M A and discharge the nutrient-rich cooling water at thermally elevated temperatures of 27°C. This study determined that P M A supports some of the highest phytoplankton standing stocks in BC coastal waters; the maximum biomass concentration (1,200 mg chl a m"2) was recorded during a bloom of the potentially ichthyotoxic raphidophyte, Heterosigma akashiwo. This was primarily due to its partial confinement and low light attenuance levels, which led to a highly stratified system where the seaward-flowing surface layer was contained entirely within the photic layer. Average chlorophyll a concentrations in P M A (95 mg chl a m~2 integrated over 10 m) were >3-fold higher than the contiguous waters of the Strait of Georgia and other adjacent inlets. The intake of cooling water through the BGS condenser system had profound effects on the entrained phytoplankton. Overall, -50% of the entrained phytoplankton biomass was destroyed during condenser passage, which was mainly due to in-plant cooling water chlorination. Dinoflagellates suffered the greatest mortality (55%), as compared to nanoflagellates (48%), and diatoms (34%). Since this cooling water was drawn from depth where biomass was low, and tidal flushing rates in P M A are high, this daily phytoplankton biomass loss can be considered inconsequential, amounting to <1% of the total P M A phytoplankton biomass. The greatest effect of BGS on the phytoplankton dynamics of P M A was related to the discharge of elevated nutrients into the surface layer of P M A during the summer when these waters were N-limited. Seasonal phytoplankton trends in P M A were characteristic of other BC inlets as winter populations were dominated by low levels of nanoflagellates and spring and fall diatom blooms occurred each year. However, the summer community composition differed as flagellate species dominated in 1999 and diatoms in 2000. During 1999, the spring freshet was near historical levels, creating a long-lasting strongly stratified system. As well, BGS operation was <50% of 2000 levels. Consequently, nutrient inputs into the P M A surface layer during 1999 were insufficient to support diatom growth during the summer months. In contrast, BGS operated near peak capacity throughout the summer of 2000 and stratification levels in P M A were moderate, resulting in a -3-fold increase in nutrient inputs relative to 1999. This allowed diatoms, specifically Skeletonema costatum, to dominate the summer phytoplankton assemblage. BGS operation directly contributed nutrients to the P M A surface layer through its cooling water discharge. In addition, BGS operation is the primary contributor to the estuarine circulation within PMA. Consequently, BGS was likely the largest nutrient source to the summer surface waters of P M A , and was therefore the primary cause of this summer phytoplankton species shift within P M A . If these added nutrients were entirely incorporated into photosynthetic biomass, this input would replace by an order of magnitude that which was destroyed due to cooling water intake. However, it is.unlikely this would lead to eutrophication within P M A because of the strong tidal effects within the estuary and the lack of a sill at the inlet mouth. This is the first study to directly link the importance of nutrient loading from thermoelectric power generation to the phytoplankton dynamics of aquatic system. Table of Contents Abstract ii Table of Contents iv List of Tables vii List of Figures ix List of Appendices xiii Acknowledgements xv Statement of Co-authorship xviii Chapter 1 - General Introduction and Thesis Outline 1 1.1 Electrical power generation - General Overview 1 1.2 Importance of phytoplankton in aquatic systems 5 1.3 Thermoelectric power plant operation 10 1.4 Effects of once-through cooling systems on phytoplankton 12 1.4.1 Intake and plume entrainment 12 1.4.2 Nutrient discharges 15 1.5 Burrard Generating Station .....17 1.5.1 Operation 17 1.5.2 Other studies related to BGS operation 20 1.6 Port Moody Arm 20 1.6.1 Overview , 20 1.6.2 Physical description 21 1.6.3 Biological description 24 1.7 Thesis Objectives 27 Chapter 2 - Phytoplankton Dynamics of Port Moody Arm, BC 29 2.1 Introduction 29 2.2 Materials and Methods 31 2.2.1 Study Sites and Sampling Frequency 31 2.2.2 Meteorological and hydrometric data 31 2.2.3 Physical oceanographic measurements 32 2.2.4 Water Chemistry 32 2.2.5 Phytoplankton biomass (chlorophyll a) 33 2.2.6 Phytoplankton collection, enumeration, and identification 33 2.3 Results 35 2.3.1 Climate in the Port Moody Region 35 2.3.2 Hydrodynamic features 38 2.3.3 Nutrients 45 2.3.4 Light 49 2.3.5 Phytoplankton biomass (chl a) and community composition 53 2.3.6 Influence of hydrography and fortnightly tidal cycles 58 iv 2.4 Discussion 63 2.4.1 Overview -63 2.4.2 Onset of the spring bloom 64 2.4.3 Role of the spring freshet 67 2.4.4 Tidal Influences 70 2.4.5 Harmful Algal Bloom (HAB) species 75 2.4.5.1 Heterosigma akashiwo 75 2.4.5.2 Pseudo-nitzschia spp 77 2.4.5.3 Dinophysis and Alexandrium spp 78 2.4.6 Role of pollution in P M A 79 Chapter 3 - Effects of Entrainment Through the Cooling Waters of Burrard Generating Station on the Phytoplankton of Port Moody Arm, BC 82 3.1 Introduction 82 3.2 Materials and Methods 84 3.2.1 Intake entrainment 84 3.2.1.1 Phytoplankton biomass (chl a) analysis 85 3.2.1.2 Enumeration and identification of phytoplankton community 85 3.2.2 Plume entrainment 86 3.3 Results 90 3.3.1 BGS operation 90 3.3.2 Effect of intake entrainment 92 3.3.2.1 Phytoplankton biomass 92 3.3.2.2 Phytoplankton community composition 98 3.3.2.3 Effects of BGS discharge rate and temperature on phytoplankton mortality ....98 3.3.3 Effect of plume entrainment on temperature, salinity and phytoplankton biomass in the P M A receiving waters 100 3.4 Discussion 109 3.4.1 Effects of intake entrainment on phytoplankton stocks and species composition.. 109 3.4.2 Effects of plume entrainment on phytoplankton stocks 113 Chapter 4 - Effects of Nutrient Inputs from Once-through Power Plant Generation and 'Natural' Entrainment on the Summer Phytoplankton Community of Port Moody Arm 118 4.1 Introduction 118 4.2 Materials and Methods 121 4.2.1 Phytoplankton composition and environmental gradients 121 4.2.2 Data Analysis 121 4.2.3 Nutrients in BGS cooling water discharge 123 4.2.4 Vertical Entrainment Model 124 4.2.5 Estimation of vertical nutrient fluxes 129 4.3 Results 131 4.3.1 Relationship between environmental factors and phytoplankton community composition during P M A summers 131 4.3.2 Nutrient concentrations in BGS cooling water discharge 140 4.3.3 Nutrient load discharged into P M A during BGS operation 144 4.3.4 Comparison of nutrient inputs due to 'natural' entrainment and BGS operation... 147 4.3.4.1 Pycnocline depths 147 4.3.4.2 Average temperatures for surface and bottom water layers 149 4.3.4.3 Estimates of site-specific surface layer outflow and estuarine entrainment.... 151 4.3.5 Relationship between vertical nutrient entrainment and direct BGS enrichment... 155 4.4 Discussion 159 4.4.1 Relationship between the summer phytoplankton community composition and environmental factors 159 4.4.2 Sources of nutrients for summer surface waters: BGS operation and estuarine circulation 160 4.4.3 Influence of vertical mixing in P M A during 2000 162 4.4.4 Influence of BGS on the P M A ecosystem during 2000 164 4.4.5 Atmospheric and riverine inputs. 169 4.4.6 Ammonium (NH 4) and the Role of Grazing 171 4.4.7 Ecological implications of BGS operation 172 Chapter 5 - Summary and conclusions 178 5.1 Phytoplankton Dynamics of Port Moody Arm 178 5.2 Burrard Generating Station 180 5.2.1 Intake and Plume Entrainment 180 5.2.2 BGS Nutrient Discharges and 'Natural' Entrainment in P M A 182 5.3 Future Studies ; 185 References 188 Appendices 213 vi List of Tables Table 1.1 Total electrical production during 2004 accumulated from 30 countries comprising the Organization for Economic Co-operation and Development (OECD). Table information includes grouped regional production, the ten greatest OECD electrical producing countries as well as three important non-OECD countries. Values are rounded to nearest 10 TWh except geothermal/other sources which were not altered when<10TWh 2 Table 1.2. Descriptive characteristics of Port Moody Arm. From Waldichuk (1965) 22 Table 2.1. Seasonal average chl a concentrations (ug 1"') at one sampled site in Burrard Inlet (SO) and five sampled sites in Port Moody Arm (S1-S5) from July 1998 to September 2000. Values are integrated levels (mg chl a irf2) averaged over depth (10 m at Site 0 through Site 4 and 5 m at Site 5) 59 Table 3.1. Effects of BGS intake entrainment on average chlorophyll a and phaeophytin a concentrations and the predicted particulate and dissolved carbon (P/DOC) discharged into the P M A receiving waters. Standard deviations are in italics 95 Table 3.2. Average cell densities of phytoplankton (±SD; italics) in the intake and discharge waters of BGS, their associated % decline of cell densities, and the significance level from the paired t-test. Bold denotes significance at p<0.05 99 Table 3.3. Pearson product-moment correlation matrix of % mortality (% declines in chl a) during BGS intake entrainment and selected BGS cooling water characteristics during the summer seasons of 1999-2000 (n=42) 101 Table 4.1. Summary of Detrended Correspondence Analysis (DCA) involving the summer phytoplankton of P M A , 1998-2000 ....133 Table 4.2. Summary of variance explained by environmental variables within Canonical Correspondence Analysis (CCA) and interset correlations between six forward-selected environmental variables and the first two canonical axes 134 Table 4.3. Pearsons correlation matrix of eight selected environmental variables used in C C A . * indicates significance at p<0.05 135 Table 4.4. Results of Canonical Correspondence Analysis (CCA) on common summer P M A phytoplankton and selected environmental variables 137 Table 4.5. Phytoplankton taxon abbreviations used in C C A biplot in Fig. 4.3 139 Table 4.6. BGS discharge rates and the nutrient concentrations and ratios within the cooling water discharge during 1999 and 2000. Student's t-tests were used to determine annual summer differences. Significance at p<0.05. ns- not significant. Summer is June 21to v i i September 21. Mean BGS discharge rates were calculated from daily discharge rates. 143 Table 4.7. Nutrient loading into the P M A surface waters from BGS cooling water discharge during the 1999-2000 sampling regime. Student's t-tests were used to determine annual summer differences. Significance is at p<0.05. Summer for 1999 and 2000 is June 21 to September 21. BGS discharge means are calculated from daily rates (June 21-Sept 21) while nutrient loads are calculated from nutrient concentrations and BGS discharge rates for sampled dates 146 Table 4.8. Summary of summer outflow surface volumes and associated vertical entrainment rates during 1999 and 2000 and contribution of BGS and natural entrainment to the total P M A surface inputs. Note: Total input is the sum of the BGS discharge volume and the entrainment volume between sites 1 and 3 154 viii List of Figures Fig. 1.1. Comparison between total electrical production (solid line) and human population growth (circles) in Canada, 1950-2004 3 Fig. 1.2. Sources of electrical production from 1977-2000 in: A) Canada, and B) the province of British Columbia. * - Fossil-fuelled 4 Fig. 1.3. Time series of total water withdrawals in Canada from various anthropogenic sources, 1972-1996 6 Fig. 1.4. Location map of Burrard Inlet (top) and Port Moody Arm (middle) and associated sampling sites (S0-S5). Bottom diagram is bathymetry of sampling transect. A l l depths correspond to the lowest astronomical tide 7 Fig. 1.5. Schematic of the operation of a natural gas-fired thermoelectric generating station 11 Fig. 1.6. Electrical production at Burrard Generating Station from 1960-2000 .....18 Fig. 2.1. Daily fluctuations of five physical parameters in the Port Moody Arm region during 1998-2000: A) temperature; B) irradiance; C) wind speed; D) precipitation; and E) river discharge (from Seymour River, Environment Canada 2005). Circles denote monthly averages (precipitation is a monthly total) read from the right-hand axes. Vertical dotted lines separate different years. A l l data provided by the Greater Vancouver Regional District (GVRD) except where otherwise stated 36 Fig. 2.2. Vertical profiles of temperature (°C) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000 39 Fig. 2.3. Vertical profiles of salinities for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000 40 Fig. 2.4. Vertical profiles of density (ot) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 41 Fig. 2.5. Vertical stratification indices (SI) at Site 3 in Port Moody Arm from July 17, 1998 to September 29, 2000. Based on difference between the selected physical parameter (i.e. temperature, salinity, or density) from the surface to 10 m normalized to depth (10 m). Hatched line represents SI=0.4 42 Fig. 2.6. Spring and summer daily discharge rates for the Seymour River, BC, 1998-2000, including the 10-year average. Data from Environment Canada (2005). 44 i x Fig. 2.7. Vertical profiles of NO3 concentrations (uM) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000 46 Fig. 2.8. Vertical profiles of Si(OH)4 concentrations (u,M) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000 47 Fig. 2.9. Spatial and temporal variability of: A) light extinction coefficients (k; m 1 ) , and C) euphotic zone (1% light level) and pycnocline depth at Site 3 in P M A 50 Fig. 2.10. Boxplots of mean extinction coefficients (k) at the five sites sampled in P M A (Sl -S5) and one site (SO) in BI, 1998-2000. Solid line in box is the median and dashed line is the average. Boxes indicate the 25 and 75 t h percentiles, respectively, error bars represent the 10 and 90 t h percentiles, and dark circles the 5 t h and 95 t h percentiles. Numbers below individual boxes are mean values (n=43). Different letters associated with each mean value imply significant differences at p<0.05 (Kruskal-Wallis One-Way A N O V A on Ranks with Dunn's pair-wise multiple comparison test) 51 Fig. 2.11. Least squares linear regressions relating light extinction coefficients (k) to chl a concentrations through the single sampling site in Burrard Inlet (A; SO) and the five sampling sites in Port Moody Arm (B-F; S1-S5), July 17, 1998 to July 8, 2000. Significance of regression coefficient (R) at p<0.05 52 Fig. 2.12. Vertical profiles of chlorophyll a ([xg l"1) for the five sampled sites in P M A (Sl -S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000 54 Fig. 2.13. Seasonal phytoplankton community composition at Site 3 in P M A from July 17, 1998 to September 29, 2000 measured as: A) cell density and B) carbon-based biomass. Actual cell densities and carbon biomass in Henry and Harrison (2002) 55 Fig. 2.14. 1998-2000 time series (A-C) of: Top) river discharge (Seymour River; see text for explanation) and wind speeds; Middle) density stratification index (Ao t) at S3 (surface to 10 m normalized to depth); and Bottom) tidal range and chl a (at S3) in the P M A region. Scale on y-axis is the same for all graphs except chl a. Dashed line in middle graphs represents stratification index of 0.4 60 Fig. 3.1. Sampling grid design for the BGS discharge plume. Refer to Fig. 1.4 for grid location. Arrows at site 4 represent direction of discharge jet 87 Fig. 3.2. Daily values of: A) electrical production; B) discharge rates; and C) temperature characteristics (triangles - discharge temperature, circles - intake temperature, squares - temperature change during transit) during the operation of BGS from July 1998 to September 2000 91 Fig. 3.3. Concentrations of: A) chl a (\ig l"1) and B) phaeophytin a (\ig l"1) in the intake and discharge waters of BGS from June 14 to October 28, 1999, and C) the associated % declines in chl a during transit (i.e. discharge chl a concentrations divided by intake chl a concentrations x 100) 93 x Fig. 3.4. Concentrations of: A) chl a (u.g l"1) and B) phaeophytin a (u,g l"1) in the intake and discharge waters of BGS from March 24 to September 29, 2000, and C) the associated % declines in chl a during transit, (i.e. discharge chl a concentrations divided by intake chl a concentrations x 100) 94 Fig. 3.5. Relationship between the integrated chl a levels (mg m" ; grey bars) at Site 3 and BGS discharge rates (open circles) from July 17, 1998 to September 29, 2000. Individual discharge rates are calculated from 5 day means±l SD. Inset is a linear regression between the date-specific integrated chl a concentrations and 5 day mean discharge rates in primary graph. Regression was non-significant (p=0.78) 97 Fig. 3.6. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (u.g l"1) through the BGS discharge plume during a low discharge event, August 19, 1999. Discharge rate: 1.4xl0 4 m 3 d"1. Discharge temperature: 15.1°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3 102 Fig. 3.7. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (u.g 1"') through the BGS discharge plume during a medium discharge event, August 3, 2000. Discharge rate: 5.5xl0 5 m 3 d"1. Discharge temperature: 23.6°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3 103 Fig. 3.8. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (u,g l"1) through the BGS discharge plume during a high discharge event, September 8, 2000. Discharge rate: 1.5xl0 6 m 3 d"'. Discharge temperature: 26.2°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3 104 Fig. 3.9. Surface contours within the BGS discharge plume of: A) average temperature increase (°C) above minimum sampling grid temperature, B) average site-specific fraction of integrated chl a grid maximum, and C) average site-specific fraction of integrated phaeophytin a grid maximum. 'Discharge' at the top of the page indicates location of BGS discharge platform and the direction of the discharge. Refer to Fig. 3.1 for surface sites and transect locations within the sampling grid. Contours are averages of 12 sampling dates covering a variety of discharge regimes 108 Fig. 4.1. Two-layer estuarine flow within Port Moody Arm and the influence of the operation of Burrard Generating Station (BGS). Q = volume; T = temperature; out = outflowing; in = inflowing; d = discharge; int = intake. P 0 = initial pycnocline depth; Pf = final pycnocline depth 125 Fig. 4.2. Summer phytoplankton community composition (carbon-based biomass) at Site 3 in Port Moody Arm, 1998-2000. Values in parentheses on x-axis are average chl a concentrations for that particular summer (1998, n=6; 1999, n=12; 2000, n=21) 132 xi Fig. 4.3. Canonical Correspondence Analysis (CCA) biplot showing relationships between phytoplankton taxa (points) and environmental factors (arrows) during the summers of 1998-2000 in Port Moody Arm. Groups of algae are diatoms (dark circles), dinoflagellates (grey diamonds), Heterosigma akashiwo (dark diamond), unidentified nanoflagellates (dark square), and euglenoids (white square). Phytoplankton taxon abbreviations are listed in Table 4.5. BGS=Burrard Generating Station discharge rate, Strat=stratification. 138 Fig. 4.4. Concentrations of: A) N 0 3 , B) P 0 4 , and C) Si(OH) 4 in the cooling water discharge of Burrard Generating Station from June-October, 1999 and March-September, 2000. ....142 Fig. 4.5. Loading rates of: A) N 0 3 , B) P 0 4 , and C) Si(OH) 4 discharged from Burrard Generating Station into the surface waters of Port Moody Arm from June-October, 1999 and March-September, 2000 145 Fig. 4.6. Vertical profiles of NO3 concentrations (fiM) along the nearshore discharge plume (see Fig. 3.1 for transect sites) during: A) low (1.4xl0 4 m 3 d"1; August 19, 1999), B) medium (5.5xl0 5 m 3 d"1; August 3, 2000), and high BGS discharge events (1.5xl0 6 m 3 d"1; September 8, 2000). The 'Discharge' at top of diagram refers to the location and direction of discharge. Site 4 is the immediate discharge area and Site 1 receives no discharge and acts as a control. Sites 7, 10, and 13 are directly downstream of the discharge outfall. Each profile was obtained during periods of NO3 limitation. The distance between Site 1 and Site 13 is 500 m 148 Fig. 4.7 Physical features of P M A during the summer of 1999 and 2000 that were used in the entrainment model (Section 4.2.4); pycnocline depth (A, B), average surface layer temperature (C, D), average bottom layer temperature (E, F), and temperature difference (AT) between surface and bottom layer (G, H). Symbol key in Fig. 4.7F applies to each graph 150 Fig 4.8. Volume of seaward flowing surface water at each respective site within Port Moody Arm during the summer of: A) 1999, and B) 2000; vertical entrainment volumes between sites during the summer of C) 1999, and D) 2000; and the vertical flux of NO3, P 0 4 , and Si(OH) 4 into the surface waters of P M A during the summers of E) 1999, and F) 2000. Surface entrainment rates are the difference between outflowing volumes at respective sites. Vertical nutrient fluxes were calculated as the product of the entrainment rates and the respective bottom water nutrient concentration 153 Fig. 4.9. A) The relative contribution of BGS nutrient inputs compared to the total nutrient flux (BGS + 'natural' entrainment) for 1999 and 2000, and B) the total N 0 3 load contributed to Port Moody Arm from BGS and entrainment (PMA) sources during the summer of 1999 and 2000, respectively 157 Xll List of Appendices Appendix A . l . Tidal heights in Port Moody Arm, BC from January 1 to December 31, 1998 214 Appendix A.2. Tidal heights in Port Moody Arm, BC from January 1 to December 31, 1999 215 Appendix A.3. Tidal heights in Port Moody Arm, BC from January 1 to December 31, 2000 216 Appendix B . l . Ammonium (NH4) concentrations collected from one site in Burrard Inlet (SO) and five sites in Port Moody Arm (S1-S5) during March 1999 217 Appendix C. 1. Four-day nutrient addition experiments showing the effect of added NO3 (open circles), PO4 (dark circles), and no nutrient addition (control; dark triangle) on the phytoplankton growth (fluorescence) from samples collected from Site 3 in Port Moody Arm during 1999. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uEm" 2 s"1 .218 Appendix C.2. Four-day nutrient addition experiments showing the effect of added NO3 (open circles), PO4 (dark circles), and no nutrient addition (control; dark triangle) on the phytoplankton growth (fluorescence) from samples collected from Site 3 in Port Moody Arm during 2000. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uEiri s"1 219 Appendix C.3. Four-day nutrient addition experiments showing the effect of added NO3, PO4, and no nutrient addition (control) on the phytoplankton growth from samples collected from Site 3 in Port Moody Arm during 1999 and 2000. Positive sign (+) represents positive growth and's' represents stationary growth. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uE m"2 s"1 220 Appendix D . l . Chl a concentrations (u.g l"1) in the BGS discharge effluent collected 2-10 minutes after intake sampling. Samples collected on June 14, 1999. A l l discharge chl a concentrations were significantly lower than in the intake water (One-way A N O V A , df=17, F= 7.328, P=0.02; Tukey's Pairwise Multiple Comparison Test) and were statistically similar to each other. Error bars represent standard deviaton from mean and n=3 221 Appendix E . l . Vertical profiles of density (at) for three sites in P M A (SI-S3) from June 18 to July 15, 1999 222 Appendix E.2. Vertical profiles of density (ot) for three sites in P M A (SI-S3) from July 21 to August 11, 1999 223 Appendix E.3. Vertical profiles of density (c t) for three sites in P M A (SI-S3) from August 19 to September 24, 1999 224 xiii Appendix E.4. Vertical profiles of density (ot) for three sites in P M A (SI-S3) from June 6 to June 29, 2000 225 Appendix E.5. Vertical profiles of density (c t) for three sites in P M A (S1-S3) from July 6 to August 3, 2000 226 Appendix E.6. Vertical profiles of density (at) for three sites in P M A (SI-S3) from August 9 to September 9, 2000 227 Appendix E.7. Vertical profiles of density (ot) for three sites in P M A (SI-S3) from September 19 and September 29, 2000 228 Appendix F . l . Site-specific pycnocline depths (Z p y c ; mean±SD), average surface (T su rf) and bottom water (Tb0t) temperatures, and the temperature difference between the two layers (AT) in Port Moody Arm from June 18 to September 24, 1999 and June 6 to September 29, 2000. Summer is June 21 to September 21. Temperature data were used in 2D entrainment model (Eq. 4.6) and the determination of pycnocline depth and average surface and bottom layer temperatures are described in Section 4.2.4 229 xiv Acknowledgements I am first indebted to my supervisor, Dr. Paul J. Harrison for his support, encouragement, and patience through this scientific journey. Paul has been instrumental guiding my scientific interests all the while allowing me to explore the natural world beyond 200 u,m. The members of my committee all deserve thanks for keeping their doors open. Dr. John Dower for statistical advice and a common sense approach, and Dr. Max Taylor for Heterosigma questions, world-class phytoplankton identification and use of his lab. Early on, Dr. A l Lewis provided necessary interest and advice. It would have been impossible, due to the size and scope of this project, to accomplish this thesis without the aid of others. I'd like to thank A l Brotherston of BC Hydro for financial and personal support. The M E L P crew, specifically Liz Freyman, Dave Robertson, Deanna Lee, Brent Moore, Greg Kanya, and Ed Quilty, who donned their raincoats and filtered Port Moody Arm dry. David Fissel, Jianhua Jiang, and A l Taylor at A S L Environmental took the time to answer numerous questions, and Jeff Greenbank, Sherri Rendek, Melanie Clarke, and other BGS technicians collected samples at BGS. I cannot thank Behzad Imanian and Ming Guo enough for their help in the lab and field. Behzad, thanks for not killing me as you threatened. Nobody should have to work that hard. Hugh Maclean for always coming through in the clutch. The members of the Harrison lab provided a consistent source of encouragement and scholarship during my years at UBC. Early on Allen Milligan somehow got me into the -field, thanks to Diana Varela for things scholastic and otherwise (Luna was fat when I got her), and Anthony Fielding for providing a good steak. Thanks to my 'officemate' Nelson Sherry for patiently answering the questions early in my postgraduate career when all I xv knew was that the ocean was 'pretty salty'. During the formative years, Joe Needoba, Michael Lipsen, Tony Wagey, Shannon Harris (argh the nutrient analyzer! Thanks!), Rana El-Sabaawi, Tawnya Peterson, Heather Toews, and Adrian Marchetti, all made Rm 1469 a place of mystery and intrigue. Somehow the nocturnal hours in Rm 1469 just wouldn't have been the same without the presence of Robert Strzepek. No sweat about the blame. It was my pleasure. Graham Peers, who knows the importance of a good ski, great coffee, and too much Portuguese food, has made writing in Montreal what it should be. Thanks to those who have shared personal experiences along the way: The Bai, Jon and Dar, Dave and Sue, and the Mood (Duane). Mike Bentley for the birds and the stories. David Timothy for understanding the Zen of Pedro. The spirits of Monk, Miles, Trane, Mingus, and Zorn for carrying me through the midnight hours. Ken Morgan has brought me into the world of macrofauna, and while it has extended my stay at UBC, the cruises have given me moments David Attenborough would envy. Many thanks to Ken and Dr. Bi l l Sydeman at Point Reyes Bird Observatory for keeping food on my table and upping my lifelist. And GO PETE! Special thanks goes to Big C, Carol Adly, for all things good, for being an Edgar, and loving dollar Expos games. That faithful 4,800 km journey is the best one I ' l l ever make. I could have never completed this without your endless support and encouragement. I owe you. To Dr. Charlie Trick: I know where my aquatic roots originated from. Finally, my deepest thanks to my parents, Frank and Lin Henry, and my brothers John and Mark who have shared in everything I've done, in presence or in spirit. I can't thank you enough Dad for the hours of Mutual of Omaha's Wild Kingdom, National Geographic specials, or pulling us out to Rondeau Park and Point Pelee, whether it was xvi raining, -15°C, or sunshine. It set the stage for this entire journey. This thesis is dedicated to those at 54 Heather. xvn Statement of Co-authorship Chapter 2 - Phytoplankton Dynamics of Port Moody Arm, BC Co-author contributions: Paul J. Harrison - provided funding to M . Henry and helped in initial design I was responsible for 90% of the data collection and lab analysis and all manuscript preparation. Chapter 3 - Effects of Entrainment Through the Cooling Waters of Burrard Generating Station on the Phytoplankton of Port Moody Arm, BC Co-author contributions: Paul J. Harrison - provided funding to M . Henry I was responsible for the experimental design, 80% of the data collection, 90% of lab analysis, and all manuscript preparation. Chapter 4 - Effects of Nutrient Inputs from Once-through Power Plant Generation and 'Natural' Entrainment on the Summer Phytoplankton Community of Port Moody Arm Co-author contributions: Trish Armundrud - Aided in the development of entrainment model (Section 4.2.4) Paul J. Harrison - provided funding to M . Henry I was responsible 90% of analyses and 100% of data collection and manuscript preparation. xvm Chapter 1 - General Introduction and Thesis Outline 1.1 Electrical power generation - General Overview Thermoelectric power generation is the dominant source of electrical production contributing >80% of the worldwide electrical output (Table 1.1). Because of the inability to fully convert thermal energy into electricity, thermoelectric generating stations require enormous volumes of cooling water (ca. 2.6xl0 6 m 3 d"1 for 1,000 M W power plant) to remove excess heat from their condenser system (GESAMP 1984; Langford 1990). This makes thermoelectric generation the largest industrial user of water, and due to the temperature elevation of the cooling water, the largest anthropogenic source of thermal energy to aquatic habitats (Nalewajko and Dunstall 1994). As a result, 'once-through' power plants are located near marine, estuarine, and freshwater systems that can fulfill this cooling water requirement as well as accept its thermal effluent. Since substantial volumes of cooling water from this process are drawn from and ejected into aquatic habitats, thermoelectric generation can have important environmental impacts as aquatic organisms are subjected to the influence of thermally-elevated effluents and entrainment into the cooling water. As in most developed countries, increased industrialization and urbanization in Canada after World War II has. led to the rate of electrical production and consumption outstripping the rate of population growth (Fig. 1.1). During the last 50 years, the total electrical generation in Canada has increased by an order of magnitude, driven by increased production from all major electrical generating sources (Fig. 1.2). While hydroelectric generation remains the predominant method of electrical production in Canada (Table 1.1; Fig. 1.2), increased numbers of fossil fuel and nuclear power plants were also constructed 1 Table 1.1 Total electrical production during 2004 accumulated from 30 countries comprising the Organization for Economic Co-operation and Development (OECD). Table information includes grouped regional production, the ten greatest OECD electrical producing countries as well as three important non-OECD countries. Values are rounded to nearest 10 TWh except geothermal/other sources which were not altered when <10 TWh. Energy Production (TWh) Per Fossil OECD Country Total Capita3 Fuels Nuclear Hydro Geothermal/Other OECD Total 9510 5840 2250 1290 130 North America 4670 3070 880 620 100 Europe 3210 1730 940 520 20 Pacific 1630 1050 420 150 10 USA 3960 14.1 2810 790 260 100 Japan 1053 8.3 660 290 100 3 Canada 570 17.8 150 90 330 France 550 9.2 60 430 60 Germany 530 6.4 350 160 20 U K 370 6.2 290 70 10 South Korea 340 7.1 210 130 10 Italy 285 4.9 230 50 5 Spain 240 •6.0 140 60 40 Australia 190 9.5 170 20 OECD Country China b 1572 1.2 1240 20 310 2 Russia b 850 5.9 530 130 180 3 India b 554 0.5 460 20 70 4 - based on World Almanac 2002 populations b - 2002 production totals. Electrical production values from www.cslforum.org. (Carbon Sequestration Leadership Forum) 2 Year Fig. 1.1. Comparison between total electrical production (solid line) and human population growth (circles) in Canada, 1950-2004. 3 600 ^ 500 O © 400 I 1991-2000 H 1981-1990 3 1977-1980 I c u 3 O U a S I I 300 H 200 A 100 80 fj 6 0 B I 1991-2000 3 1981-1990 3 1977-1980 I I I u S o u a "« 40 20 H ni i (0) Combustion Internal Thermal* Nuclear Hydro Turbine Combustion Total Electrical Source Fig. 1.2. Sources of electrical production from 1977-2000 in: A) Canada, and B) the province of British Columbia. * - Fossil-fuelled 4 during this period to meet this energy demand. Consequently, thermoelectric generation now draws more water (-64%; Kienholz et al. 2000) from Canadian water-ways than all other anthropogenic uses combined (Fig. 1.3). This brings into focus the potential ecological influence that this industry may impart upon affected aquatic systems. In British Columbia, thermoelectric power generation is a minor component of the electrical production contributing -6% of the total electrical profile over the last 25 years (Fig. 1.2B). However, >90% of this production occurs at a single plant located on the north shore of Port Moody Arm (PMA), a small tidal inlet and appendage of the Burrard Inlet system near the city of Vancouver (Fig. 1.4). This thermal power plant, Burrard Generating Station (BGS), is permitted to discharge 1.7xl0 6 m 3 d"1 of thermally-elevated effluent into P M A , which is four times the volume discharged by the other 32 industrial permit-holders combined (BIEAP 1997a). Because of this massive infusion of thermal effluent, a study of the physical, chemical, and biological parameters within P M A is an essential step in understanding the potential impact that this discharge may have on the inlet. 1.2 Importance of phytoplankton in aquatic systems Phytoplankton are among the most ecologically important organisms on the planet, accounting for -40% of the global primary production (Falkowski 1994). These organisms are thought to mediate atmospheric CO2 concentrations over the open ocean through the 'biological carbon pump'- the conversion of inorganic carbon into organic biomass in the photic zone and its subsequent export to the deep ocean, resulting in the net transfer of carbon from the atmosphere to the ocean (see Chisholm 2000). Marine phytoplankton also generate the sulphur-precursors (dimethylsulphoniopropionate, DMSP; dimefhylsulphide, DMS) 5 50000 40000 E e ^ 30000 E •o | 20000 S-V 08 10000 I I M i l Mining Municipal Agri- Manu- Thermal Total and private culture facturing power residential Source Fig. 1.3. Time series of total water withdrawals in Canada from various anthropogenic sources, 1972-1996. 6 I I I I I I I 1 I I 1 I I I I • I . I 0 1000 2000 3000 4000 5000 6000 7000 8000 9000 Distance (m) Fig. 1.4 Location map of Burrard Inlet (top) and Port Moody Arm (middle) and associated sampling sites (S0-S5). Bottom diagram is bathymetry of the sampling transect and follows the long axis of middle map. A l l depths correspond to the lowest astronomical tide. 7 responsible for the formation of cloud condensation nuclei over remote regions of ocean, ultimately leading to an increase in the earth's albedo (Charlson et al. 1987). Consequently, phytoplankton play a large role in regulating global climate through their participation in these biogeochemical cycles. In addition, phytoplankton utilize inorganic nutrients and light energy to manufacture the necessary organic materials that fuel the vast majority of heterotrophic production in aquatic habitats. Indeed, zooplankton biomass and fish yields have been significantly correlated with autotrophic production and biomass (Iverson 1990; Ware 2000; Ware and Thomson 2005). Within the marine environment, the coastal zone contributes disproportionately to global primary production as its surface area comprises -10% of the oceanic habitat, yet it accounts for -20-27% of its annual primary production (Longhurst et al. 1995; Pauly and Christensen 1995). This production supports major fisheries as nearly 20% of the global catch occurs in these regions (Pauly and Christensen 1995). Therefore, factors contributing to alterations in phytoplankton production and/or community composition in this region can have far-reaching ecological consequences. The greatest threat to the coastal habitat arises from the dense human settlements that exist in these areas. In the U.S., over half the population lives in coastal cities, which comprise only 17% of the continental area (National Research Council 2000). In Canada, >75% of the population resides near major freshwater and marine systems (Fuller et al. 1995). As a result, coastal ecosystems have been impacted by a variety of anthropogenic practices. This includes excessive nutrient inputs from both river-borne (Peierls et al. 1991; Caraco and Cole 1999; Turner et al. 2003a, b) and atmospheric sources (Paerl et al. 2002a), as well as the introduction of exotic and invasive species (Alpine and Cloern 1992), harvesting of wild stocks (Pauly and Palomares 2005), proliferation of aquaculture (Naylor et al. 2000), altered land-use practices (Meeuwig 1999), 8 deforestation (Hopkinson and Vallino 1995), river diversion and dam construction (Nilsson et al. 2005; Scudder 2005), and the discharge of industrial effluents (Langford 1990). These impacts have led to the eutrophication of some coastal areas (Nixon 1995; Elmgren 2001; Cloern 2005 and references therein), reductions in primary productivity in others (Cloern 1996), and community shifts that have included the increased presence of nuisance and potentially toxic phytoplankton species (Smayda 1990; Hallegraeff 1993; Smayda 1997; Anderson et al. 2002). Phytoplankton have high growth rates, on the order of hours to days, and diverse species-specific tolerances to environmental stimuli. As a consequence, phytoplankton are ideal subjects when attempting to elucidate the short term variability within an aquatic system. This is particularly important when the variability originates from an anthropogenic source. Because of the vast quantities of thermal effluent discharged into the coastal zone, the operation of thermoelectric generating stations may pose a serious threat to coastal water bodies (GESAMP 1984). Accordingly, an understanding of the effects of thermoelectric generation on the phytoplankton dynamics within an affected environment is critical due to their vital role in aquatic ecosystems. This thesis presents findings related to the operation of Burrard Generating Station (BGS) and its effect on the phytoplankton standing stocks of Port Moody Arm (PMA), British Columbia. In the remainder of this chapter, I will present an overview of thermoelectric generating processes and their effects on phytoplankton as well as summaries of the operation of BGS and the physical and ecological setting of P M A . An outline of the subsequent chapters is presented in Section 1.7. 9 1.3 Thermoelectric power plant operation During the process of thermoelectric generation (also known as steam generation), fresh water is superheated through the combustion of fossil fuels (natural gas, coal, and oil) or through nuclear fission within a nuclear reactor. The resulting high-pressure steam jet drives a turbogenerator, which converts the thermal energy into mechanical and ultimately electrical energy (Fig. 1.5). However, due to the second law of thermodynamics1, a substantial portion of thermal energy is not converted to electricity, and the excess heat must be removed from the system. In general, thermoelectric power plants have thermal efficiencies of only 34-39%, requiring the removal of >60% of the thermal energy as waste heat (GESAMP 1984; Langford 1990). This removal is accomplished largely through two methods: 1) closed or recirculating cooling water systems, and 2) once-through or direct cooling systems. In closed systems, cooling channels, ponds, or large towers are utilized to contain the thermal effluent and heat is released into the atmosphere through evaporation. The cooled water is subsequently recirculated through the power plant and the cycle is repeated continuously. In once-through cooling systems, cold subsurface water is drawn from an adjacent water body, passed through a condenser system (where heat is exchanged), and the heated effluent is released directly into the nearby receiving water body. Each system has its own benefits and drawbacks. With regard to closed systems, >95% of the cooling water can be recirculated (Langford 1990), therefore preventing any potential deleterious effects related to the thermal effluent. However, the construction and maintenance of the canals, ponds, and towers required to cool the intake waters is a multi-million dollar expense and often makes these facilities cost-prohibitive. The construction of 1 The second law of thermodynamics states that all forms of energy can be completely converted into heat energy, although the reverse process is not possible. 10 Steam and NOx/SOx out Stacks Lake Catalytic Converter Freshwater in • Natural Gas in Electricity out Boiler Steam Turbines Generator W Steam Condenser Warm water out Cold water in Fig. 1.5. Schematic of the operation of a natural gas-fired thermoelectric generating station. 11 'once-through' systems negates this expense; however, temperature increases of 5-20°C during passage through the condenser system lead to effluent temperatures approaching 30°C, and historically 40-50°C (Gibbons and Sharitz 1974; Reetz 1982). This thermal discharge has contributed to significant environmental impacts including fish kills (Young and Gibson 1973; Kennish 1992 and references therein; Hadderringh and Jager 2002), destruction of Cypress forest (Gibbons and Sharitz 1974), coral reefs (Jokiel and Coles 1974), and sea grass communities (Roessler 1971). Consequently, electrical utilities must cope with the biological effects caused by the thermal discharges as well as the environmental impact studies and negative public relations that accompany this type of operation. Because of the increased environmental and cultural sensitivity to the thermal pollution created by 'once-through' systems, closed-cycle cooling systems have now become more common (50%) than direct cooling systems in the U.S.A (44%; Veil 2000). 1.4 Effects of once-through cooling systems on phytoplankton 1.4.1 Intake and plume entrainment The biotic zone of influence of once-through power plants can extend from the beginning of the intake system to the extent of the discharge plume and include the smallest prokaryotes to the largest nektonic species. Typically, the strongest impacts are related to the impingement of aquatic organisms onto intake screens, transit through the condenser pipes, and subsequent entrainment into the thermal plume. While impingement onto intake screens is not a concern for phytoplankton (as they easily pass through these screens), the massive fish kills associated with impingement, plant entrainment, and thermal exposure during the 1950s through the early 1970s (e.g. Kennish 1997) set the stage for much of the ensuing research involving aquatic organisms and power plant operation. 12 Because of their ability to pass through intake screens, phytoplankton are entrained within the cooling waters of once-through power plants and are thus subjected to the conditions therein. This intake entrainment subjects the algae to sudden and potentially lethal increases in temperature, excessive shearing stresses and pressures, in addition to the exposure to biocidal chemicals. Over the last 30 years, abundant research has focused on the effects of intake entrainment on phytoplankton and a majority of this research has sought to elucidate the influence of temperature and chlorination, or the synergistic effects of both, on the entrained algal communities. Temperature trends indicate that phytoplankton communities suffer the greatest losses in biomass and primary productivity when cooling waters are heated above 25°C. This typically exceeds the cellular thermal tolerances of most temperate phytoplankton (Eppley 1972) leading to enzymatic damage and cell lysis (Briand 1975; Flemer and Sherk 1977; Bienfang and Johnson 1980; Jordan et al. 1983). In contrast, productivity is often stimulated in cooler intake waters when discharge temperatures are <25°C such as during the winter or spring season (Morgan and Stross 1969; Miller et al. 1976; Keskitalo 1987). In tropical waters where phytoplankton species naturally survive near their thermal limits, even small increases in cooling water temperatures can cause dramatic losses (Saravanane et al. 1998). Despite the thermal evidence, the greatest losses to phytoplankton are undoubtedly due to the biocidal treatment of the cooling waters (typically in the form of chlorination), which is used to reduce the formation of surface biofilms and mollusc accrual on the condenser pipes. Substantial reductions in phytoplankton biomass and productivity have been reported during periods of chlorination (Hamilton et al. 1970; Brook and Baker 1972; Fox and Moyer 1975; Eppley et al. 1976; Sanders 1984) with decreases of >80% being common (Carpenter et al. 1972; Ahamed et al. 1993). Some recovery of the phytoplankton 13 appears to occur after chlorination, as Goldman and Quinby (1979) found natural populations recovered successfully after chlorine exposure ceased, although at much lower growth rates. This may be species specific, however, as Hirayama and Hirano (1970) showed the green alga Chlamydomonas sp. recovered after exposure to chlorination while the marine diatom Skeletonema costatum did not. The effect of intake entrainment on phytoplankton within once-through cooling water systems also extends to the community level. Sellner et al. (1984) found that the productivity of the naked nanoflagellate Cryptomonas acuta declined significantly after plant passage, whereas productivity was unaltered in the diatoms Cyclotella caspia and Thalassionema nitzschioides as well as the armoured dinoflagellate Prorocentrum mariae-lebouria (now Prorocentrum minimum). Jordan et al. (1983) also found that cryptophyte mortality increased significantly relative to other algal groups during cooling pipe transit at the Surrey Power Plant in Virginia. In a Finland nuclear power plant, Keskitalo (1987) reported that production of the dominant spring diatom Chaetoceros wighamii was stimulated during plant passage, while summer production of Cryptomonas sp. and Pyramimonas sp. was severely reduced. Briand (1975) speculated that these species-specific responses to cooling water passage may actually reinforce the dominance of certain species. Since he found that dinoflagellates had greater survival rates following entrainment than did diatoms, Briand hypothesized that a 'red tide' of the dinoflagellate Gonyaulax polyedra may have formed along the coast of California in response to its selective tolerance during passage through the cooling pipes. The final noteworthy feature of cooling water processes is plume entrainment, which results from the turbulent mixing of the discharge jet with the surrounding receiving waters 14 and the subsequent far-field diffusion of the plume as the turbulence subsides (GESAMP 1984). Therefore, any plankton located in the ambient receiving area that are entrained within the discharge waters are subjected to the conditions present within the discharge plume. The effects of this type of entrainment on planktonic organisms has been described as 'extremely localized' and 'of minimal ecological importance', as the discharge plume is quickly mixed to near ambient conditions beyond the immediate discharge area (GESAMP 1984). This occurs because many nektonic organisms can actively avoid potentially harmful environments, while planktonic organisms are only at risk within the immediate 'boil' where temperatures and remnant biocides may still exist at extreme levels. This exposure is usually brief, thus impacts on plankton can be considered transient. Studies have shown that sessile benthic communities such as coral reefs (Jokiel and Coles 1974), sea grasses (Roessler 1971) and periphyton (Bamber and Spencer 1984; Bamber 1995) are at the greatest risk within the discharge plume. 1.4.2 Nutrient discharges The greatest threat to the coastal zone is the excessive addition of nutrients from anthropogenic sources (Nixon 1995; Nixon 1998; National Research Council 2000; Howarth et al. 2002). The manifestation of these allochthonous inputs can result in increased eutrophication or autotrophic community shifts, leading to alterations in food web dynamics, increased occurrences of nuisance and potentially toxic algal species, and an overall degradation of the aquatic environment (National Research Council 2000). In shallow coastal environments, the thermal effluent discharged by power plants is ejected into the surface layer as a buoyant jet (Sobey et al. 1988; Jiang et al. 2002; 2003). This feature may have important implications for the phytoplankton ecology of an inlet, since the 15 intake waters are drawn from the deep water habitat where temperatures are reduced, thus increasing the heat absorbing capacity of the cooling water. Because this deep water is nutrient-rich, and is released as a buoyant jet within the photic layer, the nutrient-laden discharge is available for phytoplankton acquisition. In effect, the operation of the once-through cooling process can mimic an artificial upwelling or a nutrient-rich riverine system and can have similar effects within these systems. Despite more than 30 years of intensive research into biological and power plant interactions, no detailed studies involving the direct nutrient loading influence of power plant operation have ever been published, although its potential effect as well as circumstantial evidence have been reported (GESAMP 1984; Langford 1990; Jahn et al. 1998; Socal et al. 1999). Under suitable light levels, phytoplankton production in temperate estuarine and coastal areas is generally related to the supply of NO3 (Ryther and Dunstan 1971; Howarth 1988; Vitousek and Howarth 1991). The influence of NO3 inputs to the northern temperate coastal zone depends on its loading rate, which is a function of the flow rate, the N O 3 concentration of the nutrient source, and the volume and residence time of the receiving waters (Pickney et al. 2001). Consequently, when power plants operate at near maximal capacities over long durations, particularly in estuaries with small surface areas and depths, the potential for nutrient loading is considerable. These effects may be particularly striking during summer, when temperate waters are characteristically limited by NO3, leading to flagellate-dominated communities (Margalef 1978). Since BGS cycles vast quantities of subsurface water into the small surface area and shallow surface layer of P M A , and because this system is typically NOvlimited during the summer season (Stockner and Cliff 1979), the process of vertical nutrient cycling may be the most important consequence of thermoelectric generation in the inlet. Thus, P M A is a model system within which to study 16 the contrasting effects of phytoplankton removal during cooling water transit and the potential for enhanced phytoplankton growth from the nutrient-rich thermal discharges. 1.5 Burrard Generating Station 1.5.1 Operation Burrard Generating Station is a natural gas-fired thermal electric plant located on the north shore of Port Moody Arm (Burrard Inlet), British Columbia, midway along the longitudinal axis of the inlet (Fig. 1.4). Commencing production in 1962, peak levels of operation occurred from 1965-75 and again from 1989 to the present (Fig. 1.6). Overall, this facility is capable of producing 950 M W through the operation of six generating units, each of which contains a steam boiler, turbine, and generator. Water for steam production is 3 1 drawn from nearby Buntzen Lake (limit: 8 000 m d" ), and the excess heat from its condenser system is removed via a once-through seawater system cooled with intake water from P M A . Based on Permit PE-07178 (BCMELP 1995), the maximum allowable discharge limit is 1.7xl0 6 m 3 d~', to be discharged at a temperature no higher than 27°C. This represents -14% of the daily tidal prism in P M A (~ 4% of the total inlet volume), and the discharge temperature is 4-10°C higher than ambient summer levels, approaching 20°C during winter (Jiang and Fissel 2004). At maximal capacity, BGS is capable of producing about 12% of British Columbia's electrical budget (BC Hydrol996), and is used largely as supplemental energy when hydroelectric power is reduced due to low runoff. BGS is also used occasionally for energy trade, producing additional electricity for export (largely to the U.S). As hydroelectric production in British Columbia is being fully utilized due to increased demand, it is predicted that BGS will be required to operate at greater capacity until new generating facilities are added to the provincial grid (BC Hydro 2005). 17 5000 4000 3000 H 2000 A IOOO H 0 •J3-C3. n i l r n n I I . . . . i i i | l l i l l l l—l— 11 1 111 T 1 1960 1970 1980 1990 2000 Year Fig. 1.6. Electrical production at Burrard Generating Station from 1960-2000. Cooling water for BGS is drawn from P M A about 30 m offshore at a depth of ~11 m. The intake water is drawn through a series of rotating screens (mesh size of 4 mm), which are designed to restrict larger debris and organisms from entering the cooling system. Therefore, all mesoplankton and smaller size class plankton can be expected to pass through the mesh. Intake waters travel through and exit the plant in roughly 2-8 min, faster during maximal operation and slower during minimal generating periods (Duval 1998). When ambient bottom water conditions in P M A reach 10°C, the intake water is chlorinated to reduce the accrual of mussel larvae and other biofilms on the inner surfaces of the cooling pipes. Before discharge, the residual chlorine within the heated cooling water is removed from the effluent using a sulphur dioxide dechlorination system. Chlorine discharge concentrations, monitored daily, are lower than those of municipal drinking water (i.e. <0.02mgl"'). After heating, the BGS cooling water is expelled through two sets of discharge pipes. Each contains two individual pipes measuring 2.45 m in diameter with one set discharging effluent along the shoreline in an eastward direction and the other towards the main channel of P M A . The cooling water from the shoreward pipes is discharged at 3 m above the seabed while seaward discharge occurs at 2.5 m below mean low tide level with initial flow moving in a southeasterly direction. One of the most important characteristics of the thermal effluent is that it emerges as a buoyant jet due to the heating of the intake waters (Jiang et al. 2002; 2003). The discharge jet surfaces -20-30 m from the outfall at core velocities approaching 40 cm s"1 during peak operation. As this discharge progresses up-inlet, velocities fall below 10 cm s"1 before turning westward, flowing seaward out of P M A into Burrard Inlet. Without BGS operation, residual seaward surface flow in P M A is -1 cm s"1, but reaches 5 cm s"1 during moderate 19 plant operation and nearly 10 cm s"1 during maximal operation (Jiang et al. 2003). To compensate for the increased seaward flow, as well as the intake water being drawn from depth, landward bottom water flow also increases during BGS operation. Thus, BGS operation is the primary contributor to the positive estuarine circulation in P M A . A full account of the influence of BGS on the circulation of P M A is presented by Jiang et al. (2002; 2003) and Seaconsult (1995). 1.5.2 Other studies related to BGS operation In response to requirements outlined under Permit PE-07178 (BCMELP 1995), BC Hydro prepared a six-component environmental assessment plan encompassing the biological and physical aspects of Port Moody Arm. The environmental assessment plan included a series of laboratory and field studies intended to define the potential effects of BGS on the biota of Port Moody Arm and largely involved the influence of thermal discharges on salmon behaviour and physiology (Birtwell et al. 2001a, 2001b; Greenbank et al. 2001a, b; Korstrom et al. 1998). This assessment plan also included a comprehensive study of the thermal budget of P M A and the contribution of BGS to this budget (Taylor et al. 2001). Additional important reports involving the role of BGS in P M A are found in Hodgins and Webb (1991), Seaconsult (1995), and Duval (1998). 1.6 Port Moody Arm 1.6.1 Overview P M A (49°17'N, 122°53'W; Figure 1.4) is the easternmost extent of the Burrard Inlet system and is bordered by the municipalities of Port Moody and Burnaby. Despite being situated in an area of increasing population growth and industrial activity, P M A has only 20 recently received attention with regards to biological research (Belan 2003; Je et al. 2003; Levings et al. 2003; Stehr et al. 2003), despite having some of the highest concentrations of heavy metals (Cd, Cu, Pb) and polycyclic aromatic hydrocarbons recorded within BI sediments (BIEAP 1998). Before these recent publications, there have only been two peer-reviewed publications involving P M A , which included single studies of physical (Waldichuk 1965) and biological processes (i.e., phytoplankton and zooplankton ecology; Stockner and Cliff 1979). Gilmartin (1964) published the first regional measurements of primary production in the adjacent Indian Arm, and Buchanan (1966) provided a comprehensive inventory of the phytoplankton composition in the same inlet. Further plankton studies in Indian Arm include chlorophyll a measurements (Weigand and Pond 1979) and a study linking tidal activity to zooplankton transport (Lewis and Thomas 1986). Much of the information available on P M A is in the form of 'grey literature' such as environmental reports from local government and consulting agencies. 1.6.2 Physical description P M A is a small positive estuary with a length of 6.5 km and average width of 0.9 km (Table 1.2). A maximum depth of 24 m at the mouth of P M A quickly shoals over the eastern third of the embayment, giving an average depth of only 8.8 m over the length of the Arm at low tide. The inlet is semi-confined, bordered to the south by Burnaby Mountain, and to the north by the Coast Mountain Range and is surrounded by largely glacial and alluvial deposits dominated by Vashon tills, Quadra sands, and Coquitlam drift (Lian et al. 2001). Winds are predominantly east/west because of this semi-confinement. P M A is located in the Coastal Western Hemlock biogeoclimatic zone (CWH), which is characterized by mild temperatures and abundant rainfall during the winter season followed 21 Table 1.2. Descriptive characteristics of Port Moody Arm. From Waldichuk (1965). Descriptive Parameter Tide Measurement Length 6.5 km Width, average 0.9 km Depth, average 8.8 m Area Low 3.9 x 106 m 2 High 5.7 x 10 6 m 2 Volume Low 3 . 4 x l 0 7 m 3 High 4.9 x 107 m 3 Tidal prism Low 5.1 x 1 0 6 m 3 High 14.5 x 10 6 m 3 High* 12.2 x 10 6 m 3 *Jiang et al. (2003) by cool dry summers (Pojar et al. 1991). Consequently, freshwater inputs into P M A are intermittent and flow mainly from two creeks: Mossom Creek on the north shore and Noons Creek at the head (although there are several other small creeks that flow into the Arm). Riverine input is greatest during peak snowmelt in May and June resulting in the spring freshet, but is greatly reduced in mid-summer, leading to occasional negative estuarine flow as surface evaporation exceeds freshwater input (Waldichuk 1965). Due to the low riverine influence, tidal activity is the dominant process affecting the physical, chemical and biological properties within PMA. Tides in P M A are mixed semi-diurnal in nature and are generally mesotidal (2-4 m; Monbet 1992) with an average tidal range of 3.3 m. During winter and summer, tidal ranges can reach macrotidal status (>4 m) as spring tides approach 5 m (Appendix A). The tidal prism in P M A has been estimated at 12.2xl0 6 m 3 (Jiang et al. 2002), which is roughly one-third of the total Arm volume. Overall, seaward surface flows are tidally driven and can reach 30 cm s"1 during maximum ebb tides, while landward bottom water flow is slower at 10 cm s"1. Because of the short length and strong tidal activity in P M A , tidal exchange is large such that an estimated 50-83% of the water is exchanged daily, with the inlet completely flushing within about one week (Waldichuk 1965). During very low spring tides, roughly 28% of the surface area of the inlet is exposed as mudflats, which may contribute up to 50% of the summer thermal budget of P M A (Taylor and Fissel 1999). Sediments throughout P M A are largely fine silts and clays and have a distinct H 2 S odour in the upper horizon, suggesting anoxia. Because of frequent shipping activity, P M A has a history of periodic dredging activity in the eastern portion of the inlet with the dredged materials often being deposited at the mouth of the Arm (T. Pedersen, University of Victoria, pers. comm.). Most of the oceanographic properties of P M A are derived from Burrard Inlet and Indian Arm, resulting 23 from the intensive tidal mixing that occurs at the Second Narrows (Isachsen and Pond 2000) and freshwater flow from Indian Arm. 1.6.3 Biological description Stockner and Cliff (1979) provide the only report of phytoplankton dynamics within Port Moody Arm, which was included as part of a larger spatial study of the entire Burrard Inlet system. The primary finding of the two-year time series was the presence of high autotrophic biomass and primary production in P M A as standing stocks were about twice that measured in the seaward section of Burrard Inlet, and primary production was -25-100% higher. Based on the trophic classification of Nixon (1995) and the primary production estimate reported by Stockner and Cliff (1979), P M A is considered a hypertrophic inlet (>500 g C m"2 y"1) and, together with Saanich Inlet (Timothy and Soon 2001), has the highest annual primary production levels reported for British Columbia (Harrison et al. 1983) and possibly the west coast of North America. Stockner and Cliff (1979) speculated that the high primary production in P M A resulted from a stable mixed-layer coupled with an elevated nutrient supply from sewage inputs, and a low flushing rate. While the first point is undoubtedly true (due to the semi-confinement of PMA), the later two points probably contribute little to the high production of the inlet. In the case of flushing rates, Waldichuk (1965) indicated that tidal flushing in P M A is strong, and is an effective mechanism for removing suspended contaminants. Considering a tidal prism that constitutes about one-third of the total volume of the inlet, P M A should be highly flushed. As for sewage inputs, all municipal waste bound for Burrard Inlet was diverted into the Fraser River at Annacis Island in 1974, and any contribution from combined sewer overflows (CSOs) would occur during the winter rainy season (i.e. when 24 phytoplankton would be effectively limited by light availability as opposed to nutrient levels). Since P M A is a largely pristine watershed with no agricultural activity, and only receives intermittent riverine inputs, it seems unlikely that these factors could contribute substantially to the inlet production when compared to other areas of Burrard Inlet. Instead, high primary production is likely due to the stable mixed layer, a favourable light regime caused by low levels of suspended particulates in the water column, as well as the increased benthic/pelagic coupling from its shallow depth and strong tidal mixing, which support nutrient replenishment into the surface waters. While P M A had one of the highest light extinction coefficients in Burrard Inlet (&=0.47) during Stockner and Cliff's study, this value is still relatively low by estuarine standards (see Cloern 1987). The high attenuance is more likely related to the high standing stocks in P M A rather than a high sediment load, which is usually observed downstream in the outer section of Burrard Inlet due to the influence of the sediment transport from the Fraser River. The seasonal progression of phytoplankton stocks in P M A is similar to that for the Strait of Georgia and its surrounding waters (Harrison et al. 1983 and references therein; Haigh et al. 1992), as well as other temperate, coastal habitats (Margalef 1958; Guillard and Kilham 1977). Phytoplankton biomass and production levels are at seasonal lows during winter due to light limitation, but the spring diatom bloom commences by mid-March or early April as incident radiation and surface stratification increases. Spring blooms are dominated by Skeletonema costatum and Thalassiosira spp., which can last into mid-May or early June. A second diatom bloom occurs during the early fall, dominated largely by S. costatum, and typically has higher biomass levels than the spring peak. Thus, most of the annual primary production in P M A is controlled by diatom growth. Between the two diatom-25 dominated peaks, NO3 can become limiting, which leads to phytoplankton communities dominated by dinoflagellates between July and mid-August. This trend was also noted by Buchanan (1966) in Indian Arm. There was no mention of harmful algal species in the study of Stockner and Cliff (1979), despite evidence that H A B species do inhabit the Burrard Inlet complex (Taylor and Haigh 1993). Annual blooms of the ichthyotoxic raphidophyte Heterosigma akashiwo have been reported just downstream of P M A in English Bay and Jericho Beach (Taylor and Haigh 1993) as well as in Deep Cove in Indian Arm (F.J.R. Taylor, University of British Columbia, pers. comm.). The presence of the paralytic shellfish poison (PSP)-producing dinoflagellate, Alexandrium tamarense, in Burrard Inlet has also been mentioned (Taylor and Haigh 1993). Stockner and Cliff (1979) did reveal that the summer assemblage in P M A contained a high proportion of Nitzschia spp. (22% of phytoplankton community), which could have been a member of the potentially neurotoxic Pseudo-nitzschia complex. Nonetheless, it is possible that the presence, or even blooms, of important H A B species could have been missed during Stockner and Cliff's monthly sampling regime. The peak zooplankton biomass in Burrard Inlet was also found in P M A . The highest zooplankton biomass followed the spring bloom but not the fall peak, which Stockner and Cliff (1979) claimed was due to the presence of large autumn populations of jellyfish and hydroid medusae that maintained the herbivorous zooplankton at low densities. As a result, it appears likely that zooplankton grazing (likely copepods) plays an important role controlling the spring diatom bloom, while this grazing pressure is removed during the fall season leading to the larger fall diatom blooms. Although P M A generally has a healthy and robust plankton community, this is not the case for the benthic community (Belan 2003; Je et al. 2003; Stehr et al. 2003). As stated 26 previously, the P M A sediments have the highest concentration of anthropogenically-derived contaminants in Burrard Inlet (BIEAP 1998). The primary constituents of this benthic pollution include unusually high levels of polycyclic aromatic hydrocarbons (PAH), polychlorinated biphenols (PCB), dichlorodiphenyltrichloroethane (DDT) and associated compounds, as well as the trace metals As, Cd, Cu, Hg, Pb, and Zn (BIEAP 1997b). The result is P M A has a largely depauperate benthic community comprised mainly of pollution-tolerant polychaetes (i.e. Tharyx multifilis; Belan 2003) with virtually no molluscs, despite the presence of a vast subtidal and intertidal area and large molluscan populations throughout the remainder of Burrard Inlet (Je et al. 2003). Given that bivalve grazing can impart heavy grazing losses to phytoplankton populations in shallow estuaries (Cloern 1982; Alpine and Cloern 1992) and other water bodies (Madenjian 1995; Nicholls et al. 1999), their virtual absence in P M A may play an important role in the phytoplankton standing stocks within the inlet. 1.7 Thesis Objectives The operation of once-through thermoelectric power stations in coastal locations has led to serious ecological impacts through the discharge of thermal effluents and the entrainment of aquatic organisms. The purpose of this thesis was to investigate the impacts that thermoelectric generation may have on the nutrient and phytoplankton dynamics of Port Moody Arm, BC. 27 Specifically, the objectives were to investigate: 1) the spatiotemporal variability of the phytoplankton standing stocks and community composition in P M A , and to determine the environmental factors that contribute to this variation. 2) the effects of intake entrainment on the phytoplankton communities and to evaluate the primary causes of mortality during transit through the condenser pipes. 3) the contribution of the BGS nutrient discharges relative to that 'naturally' entrained within P M A and to determine the resulting influence of this nutrient loading within the summer phytoplankton community. 4) whether this nutrient source is capable of supplying sufficient nutrients for phytoplankton to offset the phytoplankton mortality resulting from intake entrainment. This thesis is organized as follows: Chapter 2 presents the phytoplankton and physicochemical dynamics in P M A over a two-year period. In Chapter 3, the effects of intake and plume entrainment on the phytoplankton are discussed. In Chapter 4, a simple two-dimensional nutrient entrainment model is compared with the nutrient loading capability of BGS to determine the primary interannual source of nutrients into the summer surface waters of P M A . Chapter 5 provides the summary and conclusions of the study. 28 Chapter 2 - Phytoplankton Dynamics of Port Moody Arm, BC 2.1 Introduction Burrard Inlet (BI) is the major marine inlet intersecting the city of greater Vancouver, a metropolis of over two million people. As a result, this long (-30 km) and narrow (1-5 km) estuary is bordered by urban settlement and industrial activity along its perimeter and supports the bulk of the shipping operations in the region. Stockner and Cliff (1979) were the first to address any anthropogenic or physico-chemical influence on the phytoplankton ecology of BI. During their two-year study, they found that Port Moody Arm (PMA) was the most productive section of the inlet despite the strong anthropogenic presence in the estuary. This includes petroleum and chemical industries (Petro-Canada, Imperial, General Chemical Canada, and Esso), ship-loading and berthing operations, and the occurrence of marinas and a fueling station within the inlet. Based on municipal discharge limits into P M A , it is possible that Burrard Generating Station poses the greatest threat to the inlet due to its discharge of vast quantities of thermally-elevated effluent. Since phytoplankton production is the main source of organic material supporting heterotrophic production, and because alterations in either production rates or community composition can impart cascading effects through the marine food web (Turner et al. 1998; Chavez et al. 2002), it is necessary to examine this trophic level when discerning potential anthropogenic impacts. Given that there is abundant evidence suggesting that increased human influence on the coastal zone may result in greater instances of harmful algal species (Smayda 1990; Hallegraeff 1993; Anderson et al. 2002), gaining an understanding of the mechanisms that drive phytoplankton dynamics is critical to identifying and reacting to potential anthropogenic impacts. 29 In the absence of toxic elements, phytoplankton growth in coastal areas is constrained largely by light and nutrient availability (Cloern 1999). In shallow tidal inlets the interplay between these features is spatially and temporally variable and often controlled by riverine inputs and fortnightly tidal cycles (Trigueros and Orive 2000). The hydrodynamic influence of these sources is capable of supplying nutrients to the surface waters, controlling light levels through sediment inputs and its influence on water column stability, as well as determining the residence time within the inlet through tidal flushing and positive estuarine flow. The goal of this chapter was to characterize the temporal and spatial phytoplankton dynamics in Port Moody Arm; specifically, how the physico-chemical environment is influenced by freshwater inputs and tidal cycles, and how this in turn influenced the phytoplankton biomass and community composition within P M A . This chapter provides the baseline data required to determine whether the operation of Burrard Generating Station has had any effect on the phytoplankton dynamics in P M A between 1998 and 2000. 30 2.2 Materials and Methods 2.2.1 Study Sites and Sampling Frequency Five sites spanning Port Moody Arm (PMA) and one site (Site 0; SO) west of the confluence of Burrard Inlet (BI) and Indian Arm (IA) (Fig. 1.4) were selected for this study. The five sites within P M A are diverse in character, extending from the mouth of the inlet (Sites 1 and 2; SI and S2) to the shoaling (Site 4; S4) and intertidal (Site 5; S5) areas. Site 3 (S3) is located adjacent to Burrard Generating Station in the main channel. The study commenced in mid-July 1998 and was completed in September 2000. Sampling generally occurred weekly from early March to the end of August, with fortnightly sampling in September and October and monthly sampling in winter. In all, 72 cruises were undertaken, covering a variety of tidal regimes and sampled at different times during their cycles. A l l cruises were aboard watercraft supplied by British Columbia Ministry of Environment, Lands and Parks (BCMELP). A suite of biological, chemical, and physical parameters was measured at each station with sampling always occurring between 10:00 and 15:00 PST, beginning at SO and ending at S5. 2.2.2 Meteorological and hydrometric data Hourly measurements of temperature, irradiance, and precipitation were collected at Rocky Point near the head of Port Moody Arm and provided by the Greater Vancouver Regional District (GVRD). Hourly discharge data for Seymour River, BC were obtained from Environment Canada (2005) and hourly tidal data for Port Moody Arm, BC were (C) obtained using WXTide32 , version 4.0. Tidal data for 1998-2000 is presented in Appendix A . 31 2.2.3 Physical oceanographic measurements Two instruments were used for gathering in situ hydrographic data. A YSI model 6920 multi-parameter sonde with a YSI 610-D portable display capable of measuring temperature, salinity, dissolved oxygen and pH was used from July 17, 1998 until April 8, 1999. The sonde was lowered on a weighted line with hydrographic readings taken at discrete 1 m intervals until the sediment layer was reached. Thereafter, continuous vertical profiles for temperature, salinity, density and depth were performed using an InterOcean® S4 CTD. A l l instruments were tested frequently to ensure appropriate calibration. 2.2.4 Water Chemistry After each vertical CTD profile, water samples were collected for dissolved nutrients, specifically nitrate-nitrogen (NO3) , orthophosphate (PO4) and silicic acid (Si(OH)4). Water samples were collected at the surface, 2, 5 and 10 m depths using a Van Dorn water sampling device. Samples were drawn using acid-rinsed 60 ml plastic syringes, and purged through combusted (460°C for 4 h) 0.7 fim GF/F filters mounted on 25 mm acid-cleaned Millipore Swinnex® filter holders into acid-cleaned 30 ml Nalgene polyethylene bottles. The bottles were placed into an ice-filled cooler and returned to the lab where they were frozen at -20°C until further analysis. N O 3 , P 0 4 , and Si(OH) 4 were analyzed using standard techniques on a Technicon Auto Analyzer® II (Hager et al. 1968) within four months of collection. Ammonium (NH 4) samples were collected on three occasions during March 1999 and analyzed according to the methods of Parsons et al. (1984a). Results are presented in Appendix B. 32 2.2.5 Phytoplankton biomass (chlorophyll a) Phytoplankton biomass was determined by collecting duplicate chlorophyll a (chl a) samples from each depth described above. The precombusted GF/F filters from nutrient collection were quickly wrapped in aluminium foil and placed into an ice-filled cooler and returned to the lab and stored at -20°C for future analysis. Within 1 month of sample collection, chl a was extracted in 90% acetone for 24 h at -20°C after initially being sonicated in an ice-bath for 10 min. Chl a was analyzed by in vitro fluorometry using a 10-A U Turner Designs fluorometer (Parsons et al. 1984a). 2.2.6 Phytoplankton collection, enumeration, and identification Vertically integrated phytoplankton samples were collected from 0-10 m at S3 using a segmented integrated pipe sampler (SIPS) (Sutherland et al. 1992). This system involves a series of 3.3 m interlocking sections of PVC pipe which can be joined and dismantled to depths exceeding 30 m. Three sections reaching to 10 m were used for P M A sampling. Upon collection, the sections were dismantled aboard the boat and drained into a 15 L bucket that had been rinsed previously with site-specific seawater. The water was homogenized and 250 ml samples were collected in amber glass bottles and preserved with acidified Lugol's solution. Enumeration was conducted on settled samples using inverted microscopy. Briefly, 10 ml or 25 ml subsamples were allowed to settle for 24 h and were counted at 200x using a Zeiss IM inverted microscope. Single or multiple transects across the field of view were screened until at least 300 individual cells were counted. Up to 1,000 cells were counted when a single species dominated an assemblage. Cell numbers were converted to cellular carbon equivalents using the formulations of Taylor and Haigh (1996) based on calculations by Strathmann (1967) and Montagnes et al. (1994). Cupp (1943) and 33 Tomas (1997) were consulted for taxonomic identification of the phytoplankton and F.J.R. Taylor (University of British Columbia) provided additional assistance. Because the primary interest of the study was to identify trends in phytoplankton groups (i.e. diatoms, dinoflagellates, and other flagellates) in Port Moody Arm, phytoplankton taxonomy was performed to the genus level except in cases such as Skeletonema costatum and Heterosigma akashiwo where accurate identification to species level was clear. An inventory of the phytoplankton cell counts, densities, and carbon biomass is available in Henry and Harrison (2002). 34 2.3 Results 2.3.1 Climate in the Port Moody Region Atmospheric temperatures and average daily irradiances in the P M A region were highest during the summer months of July and August and declined into the winter season (Fig. 2.1 A, B). As a consequence of the 1997-98 El Nino, average monthly temperatures were always higher during 1998 than comparable months during 1999 and 2000, peaking in July and August 1998 (>19.7°C). No other monthly temperatures during the study exceeded 18.6°C (July 1999). Mean wind speeds between 1998 and 2000 averaged 5.1+3.3 m s"1 and also showed strong interannual and seasonal variability (Fig. 2.1C). Most notable in this variation were the above average sustained winds that occurred during the 4 month period between February and May 1999. These winds were -20-40% greater than the study mean with the maximum average daily wind speeds often reaching >10-15 m s"1. Freshwater inputs into P M A resulted from direct precipitation into the inlet as well as subsequent river discharge and run-off from the surrounding watershed. Figs. 2. ID and 2.IE show the precipitation and riverine discharge for the region. As mentioned in Section 1.6.2, riverine inputs into P M A originate largely from Mossum and Noons Creeks. Since no discharge data were available for these creeks, data from the Seymour River (see Section 2.2.2), located just downstream of SO, were used to illustrate seasonal trends. Although discharge rates between the Seymour River and the two creeks probably differed, the trends were comparable as the systems are situated only a few kilometres apart. 35 J F M A M J J A S O N D J F M A M J J A S O N D J F M A M J J A S O N D 1 9 9 8 1 9 9 9 2000 Fig. 2.1. Daily fluctuations of five physical parameters in the Port Moody Arm region during 1998-2000: A) temperature; B) irradiance; C) wind speed; D) precipitation; and E) river discharge (from Seymour River, Environment Canada 2005). Circles denote monthly averages (precipitation is a monthly total) read from the right-hand axes. Vertical dotted lines separate different years. A l l data provided by the Greater Vancouver Regional District (GVRD) except where otherwise stated. 36 Freshwater pulses into P M A showed a biphasic annual trend. Elevated inputs occurred consistently during late October through December, peaking in November, and were a direct result of the heavy rainfall and the ensuing river discharge that followed. Precipitation levels peaked over a four-month period during the winter of 1998/99 (Fig. 2.ID) with the total precipitation for each month between November 1998 and February 1999 exceeding that recorded during any other month through the observational period. A large portion of the winter precipitation accumulated as snow on the neighbouring coastal mountains, leading to a considerable release of snowmelt as temperatures increased in the spring. This annual freshet generally began during the month of April and produced the second seasonal pulse of freshwater discharge into P M A . More importantly, it accounted for the peak seasonal discharge rates into P M A (usually during June) with the effects persisting over a longer period than those caused by winter precipitation. Fig. 2.IE shows that the high precipitation experienced during the winter of 1998/99 led to elevated discharge rates in the following months with the above-average discharge lasting over a full month longer than either the summers of 1998 or 2000. This had a profound effect on the hydrodynamic regime within P M A . Overall, the climatic and hydrologic characteristics in P M A were temporally diverse during this study. Notable features include: 1) elevated temperatures and low riverine inputs during 1998; 2) substantial winter precipitation (and the resulting freshet) coupled with the strong winds observed in early 1999; and 3) moderate temperature, precipitation, and riverine inputs during the course of 2000 (relative to those in 1998 and 1999). 37 2.3.2 Hydrodynamic features The temperature, salinity, and density profiles presented in Figs. 2.2-2.4 demonstrate that the hydrodynamic environment in P M A exhibited strong spatial and temporal variability. The study began with the presence of strong vertical gradients of temperature (Fig. 2.2) and salinity (Fig. 2.3) with a pycnocline established near 6 m. Temperatures in both the surface and bottom waters during this time were the highest measured during the study, reaching a surface peak of 25.2°C at S5 during late July 1998. The highest temperatures were always observed in the intertidal region, decreasing progressively toward the mouth of P M A and into BI. Although temperature can contribute significantly to the buoyancy characteristics of estuaries, the P M A density profiles were clearly driven by salinity (least-squares linear regression; R =0.93, p<0.001) even during the summer months when solar insolation was greatest. Salinities between 21 in the surface and 25 in the bottom waters in mid-July 1998 provided strong stratification to the inlet at the beginning of the study. As summer progressed, the stratification index (SI=A at*h"'), defined here as the density (at) difference between the surface and 10 m, and normalized to depth (h), decreased rapidly (Fig. 2.5) as riverine flow dropped to seasonal lows (Fig. 2. IE), signifying the end of the influence of the spring freshet. Thereafter, the intrusion of Strait of Georgia deepwater into Burrard Inlet dominated the density profile into autumn as bottom water salinities >27 flowed into P M A at depth. Intense precipitation during November restored the vertical stratification, though surface waters were continually incorporated into the returning deep waters in the coming winter months via increased wind mixing and the turbulent mixing that occurred at the Second Narrows of BI (Isachsen and Pond 2000). 38 Fig. 2.2. Vertical profiles of temperature (°C) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 39 Fig. 2.3. Vertical profiles of salinities for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 40 Fig. 2.4. Vertical profiles of density (ot) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 41 J A S O N D J F M A M J J A S O N D J F M A M J J A S 1998 1999 2000 Date Fig. 2.5. Vertical stratification indices (SI) at Site 3 in Port Moody Arm from July 17, 1998 to September 29, 2000. Based on difference between the selected physical parameter (i.e. temperature, salinity, or density) from the surface to 10 m normalized to depth (10 m). Hatched line represents SI=0.4. 42 Throughout the spring and summer of 1999, the physical structure of P M A was influenced dramatically by the largest snowmelt in recent history (Fig. 2.6) due to four months of high winter rainfall. With the commencement of the spring freshet, salinities declined sharply beginning in late April, with the effects of the freshwater input lasting into September. Summer salinities for both the surface and bottom waters were consistently the lowest measured during the course of the study. From late June until late August, salinity levels were as low as 18 in the surface and 23 in the bottom waters, and dipped below 22 for several weeks during late July. At only one other time during the study did bottom water salinities fall below 24 at S3, during a single week in mid-July 2000. Notably, the magnitude of stratification during the summer of 1999 was similar to the peak stratification values during the summers of 1998 and 2000, despite the large freshet (Fig. 2.5). The duration of this stratification, however, was approximately 45 days longer than that observed during the summer of 2000. Spatially, waters in BI (SO) and at the mouth of P M A (SI) were colder and more brackish than in the central (S2 and S3) and eastern regions (S4 and S5; Fig. 2.3). Following the decline of the riverine inputs into P M A , waters became increasingly mixed during autumn and remained that way until the November rains re-established brackish (salinity=l 1) surface waters. . The hydrodynamic regime in P M A during 2000 showed comparable seasonal trends to the previous years with cold, moderately buoyant winter waters becoming increasingly stratified due to greater seasonal insolation and the initiation of the freshet in spring. The freshet began in late April but, unlike 1999, its effect lasted only into early July as river flow declined rapidly. A study of the historical discharge rates of Seymour River (Fig 2.6) indicate that the river flow during the 1998 and 2000 freshets was similar to the 10-year 43 40 1998 1999 Month Fig. 2.6. Spring and summer daily discharge rates for the Seymour River, BC, 1998-2000, including the 10-year average. Data from Environment Canada (2005). 44 mean and, therefore, that the thermohaline profiles during these years were far more likely to be representative of the vertical structure typically observed in P M A than those experienced in 1999. Overall, P M A can be considered a stratified system, with pycnocline depths generally falling between 3-7 m in the eastern portion of the Arm and a slight thickening of the surface layer (-7-10 m) towards the mouth of the estuary. 2.3.3 Nutrients Three macronutrients were measured throughout this study: nitrate (NO3), orthophosphate (PO4), and silicic acid (Si(OH)4). Two seasonal trends were apparent; the presence of elevated nutrients during autumn and winter and the rapid non-conservative removal of macronutrients during spring and summer. Each trend was related directly to the stratification regime and magnitude of phytoplankton biomass present at the time. Winter concentrations of NO3 and Si(OH) 4 were typically well distributed throughout the water column, attaining seasonal maxima of >30 u.M (Fig. 2.7) and >50 u.M (Fig. 2.8), respectively. Due to the proximity of the P M A mudflats at S5, Si(OH)4 levels approached 80 u.M during March 1999. NO3 and Si(OH) 4 concentrations were quickly drawn down during the spring and summer seasons due to the spring diatom bloom, as seasonal irradiances increased and P M A waters became progressively more stratified (due to the spring freshet). During all three annual stratification events, NO3 concentrations in the upper 2 to 5 m dropped to levels that were limiting to phytoplankton growth (Appendix C). Deeper water NO3 concentrations were also much lower during these periods than during autumn and winter. The most striking feature during these stratification events was the 2 Nutrient addition assays were used to test for nutrient limitation. See Appendix C for results. 45 Fig. 2.7. Vertical profiles of N O 3 concentrations (u,M) for the five sampled sites in P M A (Sl-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 46 A S O N D J F M A M J J A S O N D J F M A M J J A S 1998 1999 2000 0 2 5 10 15 20 25 30 35 40 45 50 55 60 Fig. 2.8. Vertical profiles of Si(OH) 4 concentrations (u,M) for the five sampled sites in P M A (S1-S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. 47 continuous long term NO3 limitation observed during the summer of 1999. Surface NO3 limitation lasted for nearly five months (April to August) in the easternmost sites of P M A and Si(OH)4 concentrations during this time periodically reached levels shown to reduce diatom growth (-1-2 u.M; Guillard and Kilham 1973; Conway et al. 1976; Brzezinski et al. 1998). Nutrient limitation was more transient during spring and summer of 2000 than in 1999, with N O 3 concentrations falling below 2 u,M for nearly a month following the spring diatom bloom. After a brief input of N O 3 during May, limiting concentrations lasted from early June to mid-August. Si(OH) 4 limitation was not present during this period, as summer concentrations typically exceeded 10 u.M. Overall, the effects of N O 3 and Si(OH) 4 limitation were most pronounced near the mudflat region (S5) and the central basin of P M A (S3 and S4), and slightly less so near the mouth of the inlet and into BI (SO and SI). Unfortunately, PO4 concentrations measured during this study were frequently erratic, thus complicating their interpretation. However, samples from the deep-waters (~11 m) of P M A obtained from Burrard Generating Station showed that N03:P04 in P M A was 8.0±2.1 (u,M:u.M 1 SD; n=72). Since the cellular demands of phytoplankton are typically 16N:1P (molar ratio; Redfield 1958), the deep water entrained into the surface layer of P M A is deficient in nitrogen with respect to phytoplankton growth. N:P ratios for P M A were similar to those in other British Columbia coastal waters (see Haigh et al. 1992). Finally, in simple bioassay experiments, PO4 addition to field samples never resulted in the stimulation of phytoplankton growth (Appendix C). Based on these findings, the phytoplankton in P M A appear to be largely N03-deficient during spring and summer, and low levels of Si(OH) 4 could have potentially impaired diatom growth during certain periods. 48 2.3.4 Light P M A is a moderately turbid inlet with extinction coefficients (k) ranging from 0.1 to >1.0 m"1 (Fig. 2.9A). Annual mean extinction coefficients increased towards the head of the estuary with significant increases (p<0.05) occurring in the shoaling and subtidal regions (Fig. 2.10). Light attenuation was generally lowest during autumn and winter (Fig. 2.9A) and increased as P M A waters became stratified in the spring and phytoplankton growth was initiated. Not surprisingly, euphotic zone depths (Z e u), or the 1% light level (Z e u = [-ln(0.01)*/c"']), showed the opposite trend (Fig. 2.9B). As shown in Fig. 2.9B, the depth of the euphotic zone in the subtidal zone (S5) consistently reached the sediment interface as depths in this area generally ranged from 4-6 m. Therefore, the littoral zone extends over 1 km into P M A and the benthic region likely makes an important contribution to the annual primary production in P M A . Favourable light regimes for phytoplankton growth continued into the central basin of P M A , as Z e u was greater than the pycnocline depth in the majority of cases (Fig. 2.9C). Extinction coefficients were significantly correlated (p<0.05) with chl a concentrations at each site (Figure 2.11). The Y-intercepts from the linear regressions are representative of algal-free water or more specifically, the extinction of light due to absorbance by water, and dissolved organic and inorganic substances. Upon inspection of these intercepts (see Fig. 2.11), there was a step-wise increase in background turbidity from BI and the entrance of P M A (SO and SI) into the central section (S2 and S3) with the highest non-algal extinction coefficients located within the proximity of the subtidal areas (S4 and S5). 49 2 -4 -§ 6-> 8 - 10 J= 12 OJj 14 16 18 20 2 4 6 (UI) 8 JZ 10 a Q 12 14 16 18 20 B Site 3 co 6 0 '\ u o O O —•— pycnocline (m) - O - 1% light level O-O O \o 6 \9P-6 i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — J A S O N D J F M A M J J A S O N D J F M A M J J 1998 1999 2000 Fig. 2.9. Spatial and temporal variability of: A) light extinction coefficients (k; m" ), B) depth of 1 % light level at the entrance (Site 1) and intertidal region (Site 5) of P M A , and C) euphotic zone (1% light level) and pycnocline depth at Site 3 in P MA . 50 1.2 -1 1.0 -0.8 -0.6 -0.4 -0.2 -0.0 -0 2 3 Site Fig. 2.10. Boxplots of mean extinction coefficients (k) at the five sites sampled in P M A (Sl -S5) and one site (SO) in BI, 1998-2000. Solid line in box is the median and dashed line is the average. Boxes indicate the 25 and 75 t h percentiles, respectively, error bars represent thelO and 90 t h percentiles, and dark circles the 5 t h and 95 t h percentiles. Numbers below individual boxes are mean values (n=43). Different letters associated with each mean value imply significant differences at p<0.05 (Kruskal-Wallis One-Way A N O V A on Ranks with Dunn's pair-wise multiple comparison test). 51 1.0 0.8 0.2 E Site 4 y=0.0056x + 0.4891 R2=0.56, p<0.001 n r Site 5 y=0.0091x +0.4979 R2=0.63, p<0.001 0 5 10 15 20 25 30 0 5 10 15 20 25 30 Chla fug l 1 ) Fig. 2.11. Least squares linear regressions relating light extinction coefficients (k) to chl concentrations through the single sampling site in Burrard Inlet (A; SO) and the five sampling sites in Port Moody Arm (B-F; S1-S5), July 17, 1998 to July 8, 2000. Significance of regression coefficient (R) at p<0.05. 2.3.5 Phytoplankton biomass (chl a) and community composition Initial sampling (July 17, 1998) began in the midst of a bloom of the potentially ichthyotoxic raphidophyte, Heterosigma akashiwo (Figs. 2.12 and 2.13). At SI and S3, 10 m integrated chl a concentrations were >150 and 64 mg chl a m 2 , respectively, and cell densities of H. akashiwo reached 3x l0 6 cells l " 1 at S3, comprising nearly 70% of the phytoplankton numerical abundance in P M A (Fig. 2.13A). However, cell densities in this sample were collected throughout the 10 m water column (see Section 2.2.6), and since >90% of this biomass was in the top 2 m, cell densities in the surface layer may have exceeded 107 cells 1"'. Due to the relatively small size (10-15 u.m) of H. akashiwo, however, dinoflagellates were photosynthetic co-dominants in terms of carbon-based biomass, and made up 52% of the community biomass on this date (Fig. 2.13B). Prorocentrum gracile, Heterocapsa triquetra, and Katodinium rotundatum were the primary constituents of this assemblage. As summer progressed, the flagellate community was replaced by a mixed diatom assemblage, which continued to be the most abundant phytoplankton group into autumn. Biomass levels exhibited weekly oscillations between 20-150 mg chl a m" and the dominant diatoms during this time were Skeletonema costatum, Chaetoceros spp., and Thalassiosira spp., with Ditylum brightwellii and Leptocylindrus danicus making important contributions to the late summer phytoplankton assemblage. Standing stocks underwent dramatic declines into late autumn and winter as photoperiod and water temperature decreased sharply. Integrated biomass concentrations <2 mg chl a m"2 during this period were dominated by small (<5 u,m) nanoflagellates, which consisted of cryptomonads such as Rhodomonas minuta. Fig. 2.12. Vertical profiles of chlorophyll a (u.g 1"') for the five sampled sites in P M A (SI S5) and one site (SO) in BI from July 17, 1998 to September 29, 2000. % Phytoplankton w % Phytoplankton w Carbon Composition Abundance Composition 0 70.f 9n Vo X c 9 ? 0 % % ^  % % ^ 9n % V / •?9 0 0 0 0 0 0 0 0 0 0 0 J I I I I I 1 I L 0 0 0 0 0 0 0 0 0 0 0 S 3 • % ^ %> %> 9g *** 29 J I I L J I I L During 1999, the spring diatom bloom in P M A developed slowly, as the major biomass peak (220 mg chl a m" at S3) did not occur until mid-April. This assemblage tended to form subsurface chlorophyll maxima near 5 m (Fig. 2.12). Thalassiosira spp., Chaetoceros spp., and S. costatum successively had the highest abundances during early spring, with Thalassiosira spp. dominating the biomass fraction. After a brief decline in diatom standing stocks during early May, high levels of diatom biomass returned and lasted into mid-June. During this time, diatoms always comprised >92% of the phytoplankton composition in P M A . S. costatum was consistently the predominant alga in the late spring assemblage with cell density often exceeding 3x l0 6 cells l " 1 . The high abundance of the potentially neurotoxic diatom Pseudo-nitzschia spp. was particularly noteworthy, reaching 2.3x105 and l . l x l O 5 cells l " 1 on June 4 and 18, respectively. During mid-summer of 1999, Heterosigma akashiwo was the clear photosynthetic dominant, reaching chl a levels of > 1,050 mg m~2 at S4 during the first bloom on June 30, while the second bloom on July 15 was substantially smaller at -250 mg chl a m"2 (Fig. 2.12). Between these two dates, the most abundant phytoplankton were the dinoflagellates, Katodinium rotundatum and Gymnodinium spp. After a period of low phytoplankton biomass (<25 mg chl a m ) in late July, H. akashiwo and Prorocentrum gracile formed a prodigious co-dominant bloom on August 11, 1999, which existed throughout the length of P M A (Fig. 2.12). More impressively, biomass levels approached 1,200 mg chl a m~ , with concentrations of 150 and 100 mg chl a m occurring at a depth of 10 m at Sites 3 and 4 respectively, despite Heterosigma akashiwo being a surface-blooming alga. These may be the highest integrated phytoplankton biomass levels recorded along the west coast of North America. 56 Each of the Heterosigma akashiwo blooms observed in 1999 occurred when surface stratification was >0.4 (Fig. 2.5). As the effects of the spring freshet began to subside and water column stratification was reduced, diatom populations once again became the dominant group of phytoplankton in P M A . The peak biomass (290 mg chl a m"2) during the autumn bloom occurred in mid-October, which is unusually late for British Columbia coastal waters. These populations were once again mostly comprised of Skeletonema costatum and Thalassiosira spp. For the most part, the trends exhibited during 1999 were repeated during 2000. Again, nanoflagellates dominated in winter and were replaced by diatoms in spring. Whereas the spring bloom in 1999 did not get underway until mid-April, the 2000 spring bloom commenced a full month earlier (in early March), although concentrations (240 mg chl a m" at S3) peaked in mid-April. Thalassiosira spp. were again the most conspicuous diatom during this time, but high densities of Pseudo-nitzschia spp. (3.8xl0 5 cells l " 1 on April 13) were present as well. In early May, diatom dominance was temporarily interrupted by mixed populations of dinoflagellates (primarily Heterocapsa triquetra), but diatoms returned and persisted for virtually the duration of the summer. Exceptions occurred during two very brief intervals during June 22 at S5 and July 25-27 at Sites 1 and 3 when Heterosigma akashiwo was present. Over the course of the summer and autumn, diatom biomass was high, reaching a maximum biomass in late September (200 mg chl a m"2), and the species composition was quite diverse. The three highest densities were the diatoms Skeletonema costatum, Chaetoceros spp., and Thalassiosira spp., although Cylindrotheca closterium, Thalassionema nitzschioides, Ditylum brightwellii, Eucampia zoodiacus and Dactyliosolen spp. also made significant contributions to the late summer assemblage. Of 57 note were high densities of Leptocylindrus minimus (9.2xl0 6 cells l"1) in late August and L. danicus (1.9xl0 6 cells l"1) in early September; Overall, temporal trends in phytoplankton standing stocks and composition were similar from year to year with the exception of the alternating diatom/flagellate summer assemblages. Spatial trends show that average phytoplankton biomass levels increased steadily from Burrard Inlet (SO) into P M A (S1-S5) and peaked at the shoaling area (S4) and mudflat region (S5; Table 2.1). In fact, average chl a concentrations at the head of the inlet were nearly twice that observed in Burrard Inlet as well as being much more variable. 2.3.6 Influence of hydrography and fortnightly tidal cycles Tidal ranges, river discharge and wind speeds were plotted with respect to vertical stratification and chl a through the years 1998-2000 (Fig. 2.14) in order to understand the potential role that the spring/neap tidal cycle played on the annual physical environment of P M A and, consequently, on phytoplankton standing stocks. Riverine discharge during 1998 was slightly lower than the 10-year average (Fig. 2.6). At the beginning of the study, however, a large run-off event during mid-July (Fig. 2.14A, top) strongly stratified the water column (Fig. 2.14A, middle), which led to an environment conducive for the blooming of Heterosigma akashiwo (Fig. 2.14A, bottom). During the remainder of the summer, stratification in P M A declined steadily due to low riverine inputs, with the greatest rates of destratification occurring during two successive spring tides (Fig. 2.14). Each increase in phytoplankton biomass during this period occurred during neap tidal events, while spring tides resulted in a decrease in biomass. 58 Table 2.1. Seasonal average chl a concentrations (ug 1"') at one sampled site in Burrard Inlet (SO) and five sampled sites in Port Moody Arm (S1-S5) from July 1998 to September 2000. Values are integrated levels (mg chl a m"2) averaged over depth (10 m at Site 0 through Site 4 and 5 m at Site 5). Site Year Season n 0 1 2 3 4 5 1998 Surnrner mean 6 7.6 6.5 6.3 5.9 3.1 SD 5.3 3.7 3.2 3.8 1.8 Fall mean 3 3.6 5.1 3.9 8.9 4.6 7.0 SD 2.0 5.1 2.4 8.3 6.0 8.7 1998-99 Winter mean 3 0.2 0.3 0.3 0.3 0.3 0.5 SD 0.1 0.1 0.1 0.1 0.1 0.1 1999 Spring mean 12 5.9 7.0 7.5 8.6 10.1 8.6 SD 4.4 3.8 4.3 5.9 6.3 6.7 Summer mean 11 8.9 10.2 17.3 20.0 29.2 22.7 SD 6.9 7.9 27.7 33.3 42.4 25.7 Fall mean 3 4.8 6.8 6.4 14.0 13.3 18.4 SD 3.0 5.0 4.8 13.8 14.3 - 13.2 1999-2000 Winter mean 3 0.8 0.6 0.7 0.7 2.0 1.3 SD 0.9 0.4 0.6 0.6 2.8 1.4 2000 Spring mean 11 7.3 8.2 9.5 9.1 11.3 10.9 SD 3.6 3.4 5.5 5.3 6.0 5.5 Summer1 mean 20 8.2 9.0 10.9 9.3 12.2 10.8 SD 4.2 4.4 4.9 5.1 4.9 5.8 Fall 1 7.0 15.0 13.6 11.5 19.1 21.0 Total mean SD 73 6.5 4.9 7.8 5.3 9.3 12.6 10.0 14.2 12.8 19.7 11.2 12.6 a - n=l l for Sites 0, 2, and 4 59 50 •a 01 DC ed JS A)1998 River Discharge Wind Speed B) 1999 C)2000 3/1 3/31 4/30 5/30 6/29 7/29 8/28 9/27 3/1 3/31 4/30 5/30 6/29 7/29 8/28 9/27 Date 3/1 3/31 4/30 5/30 6/29 7/29 8/28 9/27 Fig. 2.14. 1998-2000 time series (A-C) of: Top) river discharge (Seymour River; see text for explanation) and wind speeds; Middle) density stratification index (Aa t) at S3 (surface to 10 m normalized to depth); and Bottom) tidal range and chl a (at S3) in the P M A region. Scale on.y-axis is the same for all graphs except chl a. Dashed line in middle graphs represents stratification index of 0.4. 60 This tidal trend differed during 1999. Spring chl a concentrations peaked during a period of high stratification that accompanied a neap tide (Fig. 2.14B). Through the next two months, there were no apparent hydrodynamic or tidal effects on the standing stocks. The influence of the tidal regime on the vertical stability of P M A was evident as declines in stratification always followed new moon spring tides (tidal range: >4 m; Fig. 2.14B; middle, bottom). During late June through August 1999, with the waters still strongly stratified, peaks in chl a were observed during three of the four spring tides, while minimum concentrations occurred in phase with the neap tidal cycle. A l l three of the spring tidal increases in chl a coincided with blooms of Heterosigma akashiwo, with the first bloom taking place during the highest period of stratification of the summer (Stratification index = 0.56), while the later two occurred during destratification events associated with new moon spring tides. Note that despite the tidal regime, each H. akashiwo bloom during the summer of 1999 occurred with stratification indices of at least 0.4, as was the case in 1998. The last bloom in August 1999 was the largest bloom of the study. Until mid-June, the 2000 freshet (Fig. 2.14C, top) was comparable to that of 1999, which is reflected in the stratification index (Fig. 2.14C, middle). Thereafter, water column stability underwent a substantial decline as freshwater input slowed, accompanied with the largest spring tide observed during the study (tidal range: 4.8 m). This period seemed to divide the influence of neap/spring tides on the chl a content of P M A . During the highly stratified period in June, phytoplankton biomass increased during spring tides and decreased with neap tides (Figs. 2.14C, bottom). In contrast, it is clear that after late June (i.e. as the waters became increasingly mixed), chl a levels increased during neap tides, and underwent dramatic declines during spring tides into mid-August. During this period, chl a concentrations changed by nearly an order of magnitude over a single week. The negative 61 effect of the spring tide on phytoplankton biomass was less in late summer as the tidal range of spring tides progressively declined, but biomass levels continued to increase during neap tides. Overall, the vertical stability of P M A is influenced by the stratifying effects of buoyant inputs of freshwater together with the destratifying influence of new moon spring tides. The accumulation of phytoplankton biomass occurred during spring tides when the SI was > 0.4, while biomass increases during neap tides occurred when SI<0.4. 62 2.4 Discussion 2.4.1 Overview Throughout this investigation, chl a levels in P M A averaged 95 mg m"2, ranging from 2 mg m"2 during winter to 1,200 mg m"2 during blooms of Heterosigma akashiwo. These latter concentrations are the highest ever measured along the west coast of British Columbia. Stockner and Cliff (1979) attributed this high biomass largely to the presence of a stable mixed layer and nutrient inputs from sewage discharges. To place this in a regional context, Stockner et al. (1977; 1979) found that annual mean chl a levels for the nearby Strait of Georgia and Howe Sound ranged between 29-65 (over 25 m) and 7.8-45 mg m"2 (over 20 m) respectively, demonstrating that chl a concentrations in P M A are on average 3-10 times higher than in contiguous waters. The seasonal trends in phytoplankton composition within P M A were generally typical of northern temperate coastal waters (Margalef 1958; Guillard and Kilham 1977), and specifically those previously reported for BC (Stockner et al. 1977; Takahashi et al. 1978; Stockner and Cliff 1979; Harrison et al. 1983; Haigh and Taylor 1991; Haigh et al. 1992; Henry and Harrison 2002). Diatom blooms occurred during the spring and autumn, and were dominated by Skeletonema costatum, Thalassiossira spp., and Chaetoceros spp.. S. costatum was the most abundant alga over the course of this study, a feature that is common in coastal BC waters (Stockner and Cliff 1979; Harrison et al. 1983; Haigh et al. 1992) and in temperate regions worldwide (Guillard and Kilham 1977). Winter assemblages were characterised by low concentrations of nanoflagellates, similar to Saanich Inlet (Takahashi et al. 1978). 63 The most notable feature during this 2V4 year survey was the interannual difference between the summer assemblages. Summer phytoplankton communities in P M A varied from flagellate dominance in 1999 to diatom dominance in 2000, with episodic blooms of the potentially ichthyotoxic Heterosigma akashiwo occurring more frequently (and with increased cell densities) in 1999. Because sampling began in mid-summer in 1998, it is uncertain which group of phytoplankton was dominant during the entire 1998 summer sampling period, but diatoms were the principal group from August through October, with H. akashiwo present in July. H. akashiwo has been implicated in multi-million dollar losses of farmed salmon in BC (Black et al. 1991; Taylor 1993), but was not recorded by Stockner and Cliff (1979) during their monthly sampling regime. This underscores the importance of high frequency sampling when attempting to understand the temporal dynamics of aquatic habitats. 2.4.2 Onset of the spring bloom The low phytoplankton biomass measured during winter in P M A is a consequence of the low daily irradiance levels experienced during this period. The low winter standing stocks in BC waters, which are typically dominated by low densities of nanoflagellates (Stockner et al. 1977; Takahashi et al. 1978; Harrison et al. 1983; Henry and Harrison 2002), have been linked directly to light limitation caused by the short photoperiods and increased cloud cover that are characteristic of coastal BC winters (Takahashi et al. 1978). Gilmartin (1964) also proposed that the low primary production observed in Indian Arm (adjacent to PMA) during the mid-winter season may have resulted from the rapid horizontal transport of the surface waters out of the inlet due to the high precipitation and associated run-off. 64 Phytoplankton biomass increased rapidly in P M A from March to mid-April during both years of this study. Spring phytoplankton blooms are annual events in temperate coastal regions and the growth potential of this biomass is constrained by light and nutrient availability within the surface layer (Cloern 1996; 1999). This is controlled largely by the hydrodynamics and the resulting vertical stability within the water column (Legendre and Demers 1984). P M A presents a favourable environment for phytoplankton growth due to the light transmission characteristics of the water column, coupled with the persistent stratification present in the inlet. The light attenuation measured in P M A (Stockner and Cliff 1979; this study refer to Table 1) is less than half that recorded in the turbid waters of coastal plain and river-dominated estuaries (Cloern 1987; Mallin and Paerl 1992), due to its hydrography and surrounding catchment characteristics. Since P M A is a tidally-dominated system with low intermittent freshwater inputs (Waldichuk 1965; Taylor and Fissel 1999), and the network of creeks drains a small, relatively pristine watershed comprised mainly of erosion-resistant bedrock, the sediment inputs from the adjacent mountains are small. Furthermore, P M A is a flood-dominated estuary (D. Fissel, A S L Environmental Ltd., Sidney, BC, pers. comm.) that is strongly depositional, with a short fetch length and persistent stratification. Collectively, these traits can limit the suspension of sediments into the surface layer during destabilizing events (May et al. 2003), and thus reduce the potential for high concentrations of suspended particulate matter (SPM) within the water column. As SPM concentrations are tightly correlated with water column extinction coefficients (Cloern 1987), the transmission of light within P M A is such that the euphotic zone depth (Z e u) is nearly always greater than the mixing depth of the upper layer (Zm), despite the presence of a vast intertidal area (Fig. 2.9). Increasing Z e u : Z m values will position a phytoplankton community away from its compensation irradiance (Ec) and closer to its half-saturation 65 constant (EO or maximum photosynthetic rate, provided that irradiances do not become photoinhibitory. Therefore during the spring season, the phytoplankton populations in P M A respond favourably to daily increases in PAR since ample nutrients are present during this time. The spring bloom in BC waters is further enhanced by the increased surface stratification (decreased Z m ) that occurs from increased solar insolation and buoyant additions of freshwater from direct runoff and riverine sources (Gilmartin 1964; Stockner et al. 1979; Haigh and Taylor 1991; Haigh et al. 1992). Since the surface waters in P M A were consistently stratified from November to March (due to the high winter precipitation), the presence of higher daily irradiance levels was likely the primary factor leading to the rapid phytoplankton growth during the spring season. This scenario differs from river-dominated estuaries such as the Delaware Estuary (Pennock 1985; Sharp et al. 1986), coastal plain estuaries with high SPM concentrations (decreased Zeu:Zm) such as San Francisco Bay and the Neuse Estuary (Cloern 1987; Mallin and Paerl 1992; Jassby et al. 2002; May et al. 2003), and energetic systems with deep mixed layers like the English Channel (Ragueneau et al. 1996), all of which require buoyant freshwater pulses to reduce the depth of mixing and increase the light availability within the upper layer before large spring blooms can develop. Cloern (1987) showed that San Francisco Bay became a net source of phytoplankton production only when Zeii:Zm> 0.2 with the largest annual spring blooms consistently occurring during periods of peak riverine discharge coupled with low energy neap tides (Cloern 1984; 1991). It therefore follows that the annual Z e u : Z m > 1 determined for P M A is one of the major factors contributing to its hypertrophic nature. In response to the rapid accumulation of biomass in P M A , surface N O 3 was drawn down to growth-limiting concentrations (<2 u,M; Appendix C, Eppley et al. 1969). 66 Following the 1999 spring bloom, Si(OH) 4 concentrations also reached levels that have been found to be potentially limiting to diatom growth (-1-2 u,M; Guillard and Kilham 1973; Conway et al. 1976; Brzezinski et al. 1998). A notable difference between the two years surveyed in this study was the delayed development of the spring bloom in 1999 as compared to 2000 (Fig. 2.12). Since irradiance levels and surface stratification did not differ between the years (Figs. 2.1 and 2.5), it is possible that the high winds during February and March 1999 increased the wind-driven horizontal advection of surface biomass from P M A into BI, resulting in a late bloom. The high Si(OH)4 concentrations observed in the subtidal region (S5; >80 u,M) during this time may be evidence of wind-induced suspension of the sediments, although no other macronutrient showed this increase. Whereas light is the primary factor for the initiation and development of the spring bloom in BC coastal waters, winds may play a strong role in determining its timing and magnitude of the biomass peak (Haigh and Taylor 1991; St. John et al. 1993; Yin et al. 1995a, 1996, 1997a, b). This may explain why blooms in sheltered inlets along the southern coast of mainland BC commonly occur a month prior to those in the Strait of Georgia (Stockner et al. 1979; Harrison et al. 1983) and two months prior to those along the west coast of Vancouver Island (Taylor and Haigh 1996). 2.4.3 Role of the spring freshet Peak seasonal riverine inputs have been shown to promote spring diatom production along the coastlines of North America, primarily owing to the shallowing of the mixed layer (Yin et al. 1995b; Cloern and Dufford 2005), the enhanced entrainment of subsurface nutrients into the buoyant surface layer (Harrison et al. 1991; Yin et al. 1995b; 1997a, b), and the direct input of terrestrially-derived nutrients from the fresh water source (Malone 67 1991). This was not the case in P M A , particularly during 1999, as pre-freshet stratification and surface nutrient concentrations were suitable for optimal algal growth, and chl a concentrations quickly declined with the onset of the freshet. Chl a concentrations also declined in 2000, although the results were less clear due to the appearance of Heterocapsa triquetra in early May. There are several explanations for the reduction of the spring biomass during this period. First, the depletion of surface nutrients undoubtedly slowed the high rate of diatom growth observed during the spring bloom. Riverine inputs into P M A and the surrounding region are nutrient-poor (Harrison et al. 1991; Harrison et al. 1994 and references therein; Kiffney and Bull 2000), and flow rates are low (Waldichuk 1965), which reduces the entrainment potential of nutrient-rich subsurface water into the buoyant seaward surface flow. Thus, even when riverine inputs into P M A are higher than normal, nutrient inputs into the surface layer are quite small and incapable of sustaining the spring diatom populations. Second, the enhanced freshwater discharge into P M A , although low, would have increased the seaward residual flow of the surface waters, and hence the horizontal export of phytoplankton out of P M A and into BI. Freshets have been shown to transport neritic phytoplankton species (de Madariaga et al. 1992) as well as erodable fine sediments from estuaries into the coastal zone during high discharge events (Uncles et al. 2002). This surface export would be particularly significant for the diatom communities that were present at this time since they are largely confined to the surface outflowing layer. Finally, zooplankton grazing likely contributed to the decline of phytoplankton biomass during the spring. The collection of zooplankton was attempted during this study; however, due to the dense populations of chain-forming diatoms, the bongo nets consistently became clogged with phytoplankton making accurate measurements of zooplankton biomass impossible. Previous studies have confirmed that zooplankton biomass 68 in P M A and its contiguous waters is highest following the spring bloom (Parsons et al. 1979; Stockner and Cliff 1979; Stockner et al. 1979; Harrison et al. 1983; Bornhold 2000), and the ability of large copepods such as Neocalanus plumchrus to graze down the diatom spring bloom has been demonstrated in the Strait of Georgia (Harrison et al. 1983; Yin et al. 1996, 1997b; Bornhold 2000). Thus, grazing probably played some role in the biomass decrease. Additionally, there can be large annual fluctuations in the zooplankton biomass in the region during this period (Stockner et al. 1979; Bornhold 2000), which may explain the discrepancy between the phytoplankton biomass during the early stages of the freshets in 1999 and 2000. Each of these three factors likely contributed to the end of the spring bloom to some degree, although within the hierarchy of factors, NO3 limitation was most likely the dominant cause. The most interesting finding of this study was how the differing durations of the 1999 and 2000 freshets resulted in the development of either diatom or flagellate-dominated assemblages in the summer. During both 1999 and 2000, peak freshet discharges occurred in mid-June; river flow and surface stratification were comparable from April to mid-June, and diatoms dominated the phytoplankton assemblages in P M A . After this congruent period, the 1999 freshet persisted for approximately 6 weeks longer, while the 2000 freshet began to wane. The extended freshwater input into P M A during 1999 provided long-term vertical stability to the surface waters, such that surface stratification was largely unaffected by the fortnightly tidal cycles. As a result of this strong stratification, the surface N O 3 and Si(OH) 4 concentrations that were depleted through diatom uptake could not be replenished by nutrient-rich deep water. Consequently, diatoms were replaced by Heterosigma akashiwo and various dinoflagellates species from late June to late August 1999, most likely due to their ability to vertically migrate into the bottom waters to sequester the required nutrients. 69 From late June until mid-August 2000, the freshet discharge was <50% of 1999 levels (Fig. 2.6). This rapid decrease occurred simultaneously with the largest spring tide of the phytoplankton growing season, which dramatically lowered the stratification index by three-fold over the fortnight (Fig. 2.14C). Throughout the remainder of the summer, this created an environment that sustained diatom growth in two ways: (1) by reducing the stratification (and thus the energy) required to inject essential nutrients into the nutrient-limited surface waters during subsequent destratification events, and (2) through the decline of river flow that reduced lateral transport out of P M A and increased residence times in the inlet, allowing net diatom growth through decreased loss rates (Cloern et al. 1983). The surface waters of P M A remained N03-limited throughout the later summer months of 2000 despite the tidal events, indicating that diatom growth was being sustained by this tidal mixing supply of NO3 into the surface layer. It is worth noting that the operation of the thermoelectric generating station also ejected substantial amounts of nutrients into the surface waters of P M A during the summer of 2000 via a thermally-elevated buoyant discharge plume. There is evidence that this discharge also contributed substantially to the diatom growth during this season (see Chapter 4). 2.4.4 Tidal Influences The mechanical energy provided by tidal activity in coastal areas leads to high-frequency variations in phytoplankton and nutrient dynamics (Webb and D'El ia 1980; Legendre and Demers 1984; Cloern 1991; Harrison et al. 1991; Yin et al. 1995c; Ragueneau et al. 1996). Tidal currents are the dominant energetic source in P M A and are therefore the primary destabilizing force on the hydrodynamics of P M A (Waldichuk 1965; Taylor and Fissel 1999; Jiang et al. 2003). Furthermore, the large tidal prism relative to the total inlet 70 volume creates an environment that is rapidly flushed (Waldichuk 1965). Consequently, tidal advection is an important factor determining the residence time and horizontal distribution of phytoplankton within P M A . The result of this can be observed in the eastward increase in phytoplankton biomass from Burrard Inlet into P M A . Since large tidal excursions (-1.5-3.5 km) are capable of removing water from much of the western section of P M A (SI-S3) into Burrard Inlet during ebb tides, a large portion of the phytoplankton community will be flushed downstream towards the Strait of Georgia due to the increased effect of Indian Arm outflow into Burrard Inlet. The subsequent movement of low biomass water from Burrard Inlet into P M A on the flood tide then contributes to the lower biomass seen over the western portion of P M A (S1-S3) relative to that near the subtidal section (S4 and S5), which is beyond the influence of the Burrard Inlet. Therefore, tidal advection plays a major role in the daily removal of phytoplankton biomass from P M A . In addition to horizontal advection caused by tidal activity, the interplay between the destratifying influence of the fortnightly tidal cycle and the buoyancy imparted by the annual freshet were vitally important during this study since they facilitated the seasonal dominance of particular groups of phytoplankton. The strongest destabilizing tidal force during this time occurred during the new moon spring tide, which led to large decreases in stratification during the absence of the freshet. These tides attained macrotidal status (tidal range >4 m) 4-5 times from May through August in P M A (Appendix A), a period encompassing both the phytoplankton growing season and the transitional phase of the annual freshet. Since the magnitude of the tidal amplitude is related to the vertical mixing potential within an estuary (Dyer 1997), these new moon spring tides were responsible for the largest disruptions in stratification during the summer period in P M A . Coupled with the 71 hydrological dynamics, the tides were a principle factor in shaping the phytoplankton assemblages between the summer seasons. Shorter, weekly fluctuations in the P M A phytoplankton stocks were also related to the spring-neap tidal cycle and were dependent on the vertical stability present within the inlet. When stratification was low (SI<0.4), biomass tended to increase during the neap phase of the fortnightly cycle and decrease during the spring phase. This was particularly evident during late summer of 1998 and 2000, when stratification was at an annual minimum and diatoms were dominant. This phenomenon has been observed in several coastal temperate studies (Sinclair 1978; de Madariaga et al. 1989; Cloern 1991; Roden 1994; Ragueneau et al. 1996; Crawford et al. 1997) where the light and nutrient dynamics were altered by the fortnightly tides. In stratified nutrient-limited inlets, the turbulent kinetic energy (TKE) provided by spring tides can be sufficient to break down moderate vertical density gradients, thus lowering surface chl a concentrations through dilution and redistributing new and regenerated nutrients into the surface waters (Webb and D'El ia 1980; Harrison et al. 1991; Yin et al. 1995c). When waters become increasingly stratified during the less energetic neap tides, nutrients are utilized under more favourable light regimes (higher Zeu:Zm), resulting in increased diatom production. In regions such as P M A , this infusion of nutrients into the euphotic zone is an important means of sustaining diatom growth during periods of nutrient limitation. In contrast to 1998 and 2000, phytoplankton biomass increased during spring tides during the summer of 1999 when strong vertical stability (SI>0.4) existed over a 4 month period. This also occurred during the period of peak stratification in May-June 2000. Several observations have shown that spring tides may promote higher diatom growth in areas where vertical gradients are more pronounced and water transparency is high, such as 72 in the Gulf of Maine (Balch 1981), the Gernika estuary (de Madariaga et al. 1989), and the English Channel (Ragueneau et al. 1996). This may explain the high diatom biomass that occurred during two spring tides in June 2000 when stratification was at a seasonal high and P M A waters were relatively clear. The influence of spring tides was most apparent when flagellates were the dominant phytoplankton. Both flagellate blooms during 2000 (Heterocapsa triquetra on May 3; Heterosigma akashiwo on July 27), and all three during 1999 (H. akashiwo; June 30, July 15, and Aug 11; co-dominant with Prorocentrum gracile) showed dramatic increases in biomass during spring tides. The increase in flagellate biomass during spring tides may be linked to the rates of cyst resuspension, since Heterosigma akashiwo and Heterocapsa triquetra form benthic cysts as part of their life histories (Imai and Itakura 1999; Olli 2004). Tidally-induced turbulence is a bottom-generated process (Mann and Lazier 1996), where bed-shearing stress is directly related to current velocities, among other parameters. Accordingly, elevated concentrations of suspended particulate material and microphytobenthos are characteristically observed in the water column during spring tides, while the decrease in tidal straining during neap tides leads to lower suspended particulate material levels (Weekes et al. 1993; Grabemann and Krause 2001). de Jonge and van der Berg (1987) suggested that current speeds of -10 cm s"1 may be sufficient to resuspend various sediment types and benthic diatoms. Since the bottom currents measured in P M A during spring tides (-20 cm s"1) are twice that during neap tides (-10 cm s"1; Taylor and Fissel 1999), a greater input of cysts should theoretically be delivered to the overlying waters during spring than neap tides. When these resuspension events coincide with conditions that are conducive for germination and subsequent vegetative growth, the potential for the rapid accumulation of biomass exists. Experimental evidence indicates 73 germination rates of H. akashiwo cysts greatly increase when water temperatures exceed 15°C (Imia and Itakura 1999). Field observations within BC waters corroborate this as Taylor and Haigh (1993) have shown that H. akashiwo appears in English Bay when temperatures exceed 15°C and surface salinity drops below 15. Each H. akashiwo bloom during this study also occurred when surface waters were above 15°C. Therefore, it appears that the liberation of benthic cysts from sediments during spring tides, coupled with the appropriate hydrodynamic stability and ambient temperatures for successful germination, could be important contributors to the presence of H. akashiwo blooms in P M A . One of the key determinants in phytoplankton bloom dynamics is the ability of the organisms to persist within an inlet. In BC waters, Heterosigma akashiwo blooms can exist for several months when the duration of stratification is sufficiently long (Taylor and Haigh 1993). During the prolonged stratification of 1999, the appearance of H. akashiwo was a consistent feature in P M A waters, although the residence time of the flagellate appeared to be ephemeral (blooms seldom occurred on consecutive weeks; Aug 3-11 1999), despite conditions that were considered optimal for growth (Taylor and Haigh 1993; Smayda 1998). This is likely due to the rapid flushing rate of P M A where the incoming daily tidal prism has been estimated to be -10-30% of the total P M A volume depending on the tidal regime (Waldichuk 1965; Taylor et al. 2001). There is evidence that certain organisms possess the ability to withstand the effects of tidal flushing (Horstmann 1980; Garcon et al. 1986; Crawford and Purdie 1992; Lauria et al. 1999), however the tidal excursions in P M A are -2-3.5 km and can transport roughly half of the surface waters into BI on a major ebb tide. This would lead to a large net advective loss of phytoplankton into BI over the tidal cycle since the return flood waters into P M A are highly diluted because of the enhanced estuarine circulation of the Indian Arm-BI complex (Waldichuk 1965). If P M A was to maintain long-74 term populations of H. akashiwo, it is likely that these blooms would have to be present within BI, or at least be advected from the adjacent Indian Arm. 2.4.5 Harmful Algal Bloom (HAB) species Whether due to cultural eutrophication (Smayda 1997) or simply reflecting an increased social and scientific interest, greater frequencies of H A B s are being reported worldwide (Anderson 1997). This trend holds true along coastal BC and the surrounding region where, in recent years, an increasing number of bloom events from an increasing number of bloom species have been reported along the entire coastline (Taylor 1993; Taylor and Horner 1994; Taylor and Harrison 2002). References to HABs in the B C area can be found in Harrison et al. (1983), Gaines and Taylor (1985), Haigh and Taylor (1990), Taylor (1993), Taylor et al. (1994), Taylor and Horner (1994), Taylor and Haigh (1996), Whyte et al. (2001), and Taylor and Harrison (2002). 2.4.5.1 Heterosigma akashiwo Of the bloom species reported in BC, Heterosigma akashiwo is the most conspicuous algae within P M A with five 'outbreaks' (>106 cells l"1) occurring during the course of this study. H. akashiwo exists throughout the temperate regions of the world and has been responsible for multi-million dollar losses to the fish farming industry (Taylor 1993). Although the causative agent has not yet been confirmed, it is thought that breve-like neurotoxins (Khan et al. 1997) and the production of reactive oxygen species (ROS) such as superoxide radicals and hydrogen peroxide may be involved in the fish mortality (Yang et al. 1995; Oda et al. 1997; although see Twiner et al. 2001). There were no fish kills attributed to H. akashiwo during this study, despite very high cell densities. On June 30, 1999, these densities may have reached 2x l0 8 cells l" 1 in the surface layer (calculated from 75 the date-specific surface chl a concentration and a chl axell ratio of 5 pg chl a cell" ; Tomas 1980), which is similar to densities observed during fish-kills in Sechelt Inlet, B C (Taylor and Horner 1994). This suggests that the magnitude of the bloom is not the determining factor that dictates fish kills in P M A . Although negative impacts of HABs on fish communities were not documented in these waters during this investigation, it is apparent that P M A is a suitable habitat for recurrent blooms of Heterosigma akashiwo. There are several reasons for this including: the semi-permanence of vertical stratification in the inlet, optimal temperature and salinity levels for germination and growth, as well as the potential for nutrient limitation that leads to the annual collapse of the spring diatom bloom (Taylor and Haigh 1993). Furthermore, the sediment layer in P M A is a silty clay layer (T. Pedersen, University of Victoria, pers. comm.), which protects benthic resting cysts from being scoured during transport along the sediment bed. With respect to H. akashiwo dynamics, perhaps one of the most important features of P M A is the tidal current asymmetries and positive estuarine circulation that exists in the inlet. P M A is a flood-dominant estuary where the duration of the flood tide is much shorter than the outgoing ebb tide (Taylor and Fissel 1999). The result is a system that is highly depositional, which in turn leads to the formation of large subtidal and intertidal mudflat regions. In addition, the landward transport of bottom water created by positive estuarine circulation causes P M A to act as a sink for a substantial amount of sediment and pollution originating from BI (BIEAP 1998). It is therefore possible that a substantial portion of H. akashiwo cysts in P M A may originate from allochthonous sources due to its estuarine circulation and remain within P M A because of its depositional characteristics. P M A should thus function as a net annual and localized source of H. akashiwo despite the large rates of tidal flushing that exist. 76 2.4.5.2 Pseudo-nitzschia spp. Pseudo-nitzschia spp. are pennate diatoms with a broad marine distribution (Hallegraeff 1993). Several species within this genus are capable of producing domoic acid (DA), a neurotoxic amino acid linked to Amnesic Shellfish Poisoning (ASP), which can cause abdominal distress, memory loss, and occasionally coma and death in humans (Horner and Postel 1993). While ASP has been implicated in human fatalities along the eastern coast of Canada (Bates et al. 1989), no such reports have occurred along the west coast of North America. The primary Pseudo-nitzschia species responsible for toxin production along the west coast include P. australis, P. pungens, P. multiseries, and P. pseudodelicatissima, which have been implicated in avian and marine mammal mortalities along the California coast (Fritz et al. 1992; Beltran et al. 1997; Scholin et al. 2000) as well as multi-million dollar losses through shellfish closures along Oregon/Washington (Adams et al. 2000; Horner et al. 2000; Trainer et al. 2001). In the Pacific northwest, the highest abundances of Pseudo-nitzschia spp. are consistently found in shelf waters (Forbes and Denman 1991; Taylor and Haigh 1996) and persistent populations may exist within the Juan de Fuca eddy situated at the head of the Juan de Fuca Strait on the Canadian-American border (Trainer et al. 2002; Marchetti et al. 2004). Densities exceeding 105 cells l " 1 are also common within the estuarine waters of Puget Sound, Washington (Horner and Postel 1993) and inlets in the Strait of Georgia (Taylor et al. 1994). During this study, Pseudo-nitzschia spp. was an infrequent component of the temporal phytoplankton distribution within P M A (37% of samples), however there were three occasions where Pseudo-nitzschia spp. densities exceeded 105 cells l " 1 ; June 4, 1999 (2.3xl0 5 cells l"1); June 18, 1999 ( l . l x l O 5 cells l"1); and April 13, 2000 (3.8xl0 5 cells 77 I"1). These densities are consistent with cell concentrations observed during seasonal peaks in Puget Sound (Horner and Postel 1993), Sechelt Inlet, BC (Taylor et al. 1994), Barkley Sound, BC (Taylor and Haigh (1996), as well as within the Juan de Fuca eddy (Trainer et al. 2002). Pseudo-nitzschia spp. enumerated during this study were only identified to the genus level, since species-level identification requires the use of scanning electron microscopy (SEM) to resolve intricate taxonomic features. Therefore, it is impossible to know whether the species within P M A were capable of toxin production. The high cell densities observed in P M A do not necessarily imply the presence of elevated D A concentrations, though the proximity of P M A to known hotspots of D A production makes it prudent to be aware of this potential threat. Although the shellfish harvest has been closed in BI (W. Hajen, Environment Canada, pers. comm.), Dungeness crabs (Cancer magister) are collected recreationally from P M A , and these organisms have been shown to be potential vectors for the DA toxin (Wekell et al. 1994; Lund et al. 1997). Steps should be taken within PMA/BI to identify the species of Pseudo-nitzschia and whether any species are potential DA producers. 2.4.5.3 Dinophysis and Alexandrium spp. There were periods when Dinophysis spp. concentrations may have been much higher than those recorded elsewhere in BC. On one date (Aug 19, 1999), concentrations of Dinophysis spp. were >104 cells l " 1 , which may be the highest density recorded in BC waters. This genus is commonly recorded in BC waters at low concentrations (Haigh and Taylor 1990; Taylor et al. 1994; Taylor and Haigh 1996) and is capable of producing okadiac acid, which is responsible for Diarrheic Shellfish Poisoning (DSP). Reports of DSP-78 related symptoms can occur at cell densities as low as 200 cells l " 1 (Lassus et al. 1985). Since shellfish cannot currently be harvested in BI, this likely poses little threat to human health in the area. If these symptoms are reported, however, it is possible that the cause may be misidentified as being related to outbreaks of the enteric pathogen Vibrio parahaemolyticus (DePaola et al. 1990; Taylor and Harrison 2002). There were few incidences of Alexandrium spp. recorded during this study. This is surprising since this dinoflagellate has been frequently reported just downstream of P M A in English Bay (Taylor and Haigh 1993), and especially considering the dominance of various flagellates in 1999. The BC coastline has one of the longest records of HABs in the world (Taylor 1993). Reports of outbreaks are frequent and have involved novel species (Whyte et al. 2001) as well as those contributing to multi-million dollar losses to the aquaculture industry through shellfish harvest closures and finfish kills (see reviews in Taylor 1993; Taylor and Horner 1994; Taylor and Harrison 2002). In the most extreme cases, human fatalities have occurred (Taylor and Horner 1994). The greatest threat within P M A is related to blooms of the potentially ichthyotoxic chloromonad, Heterosigma akashiwo. The appearance of Pseudo-nitzschia spp. within the BI record is also cause for awareness given its potential impact on human health, especially during summers of low run-off. 2.4.6 Role of pollution in P M A Stockner and Cliff (1979) claimed that the high level of production in P M A was due to "sufficient nutrients from sewage discharges and a relatively stable mixed-layer". This claim could not be supported by this study since no evidence of sewage sources was found. It is unlikely that sewage inputs contribute significantly to the phytoplankton production in 79 P M A since the direct discharge of domestic wastes into P M A or Burrard Inlet has not been practiced for the past 30 years (J. Boyd, Environment Canada, pers. comm.). Furthermore, combined sewer overflows (CSOs) occur further upstream in the Inner and Central Vancouver Harbour (Hall et al. 1998), and any influence of this overflow would contribute little to phytoplankton growth in P M A since the primary months of discharge occur during the high precipitation periods of winter, when ambient nutrient concentrations are already high due to insufficient light and low phytoplankton growth. Finally, there is no evidence within the chemical profiles in P M A indicating anomalously high nutrient concentrations, skewed nutrient ratios, or anoxia, all of which would be symptomatic of eutrophication within P M A . It cannot be discounted however that other anthropogenic influences may partially contribute to the high primary production seen in P M A , since it is situated within a heavily industrialized region. A primary factor contributing to the high production rates and biomass in P M A may be the pollution-mediated decoupling between the benthic community and the overlying pelagic waters. As stated previously, P M A can be classified as the most polluted area within the BI complex (BIEAP 1998; Belan 2003) with the highest recorded concentrations of heavy metals, organochlorides, and polyaromatic hydrocarbons within the Vancouver Harbour sediments (BIEAP 1998). This is likely due to the estuarine flow in BI and the flood-dominance in P M A that causes the inlet to become an allochthonous sink for industrial waste discharged into the Vancouver Harbour system. The result of this contamination is a largely depauperate benthic community that contains the lowest faunal richness, abundance, and diversity in BI (Belan 2003; Je et al. 2003). Accordingly, P M A is dominated by pollution-tolerant polychaete species (primarily Tharyx multifilis; Belan 2003), and is nearly devoid of suspension-feeding bivalves (Je et al. 2003). This has 80 substantial implications, as bivalves are capable of controlling the phytoplankton standing stocks in coastal estuaries, particularly in shallow tidal inlets (Cloern 1982; Officer et al. 1982; Koseff et al. 1993; Lucas and Cloern 2002). Cloern (1982) estimated that 1.2-1.8 times the volume of south San Francisco Bay (SSFB) was cycled through the bivalve community on a daily basis and was the primary sink for the phytoplankton production within the inlet. In addition, Alpine and Cloern (1992) reported that summer blooms in SSFB were effectively eliminated due to the appearance of an exotic Asian clam species, Potamocorbula amurensis. Thus, it is apparent that alterations in the trophic cascade can lead to drastic changes in the pelagic phytoplankton community. Since the mollusc density in P M A is <10% of any other site in BI (Je et al. 2003), and given the shallow nature of P M A relative to the rest of BI, the potential for phytoplankton biomass accumulation due to the decreased bivalve grazing pressure is substantial (assuming that higher mollusc densities result in higher grazing rates; Koseff et al. 1993). This grazing removal process has been overlooked in discussions of the phytoplankton ecology of P M A . Finally, BI and P M A are the receiving waters for large a number of industrial discharges (BIEAP 1997a). Of these, Burrard Generating Station (BGS) discharges 4 times more effluent volume into P M A than the remaining 32 industries discharge into Vancouver Harbour combined. Since BGS discharges massive volumes of thermally-elevated cooling water into P M A , it is important to include this anthropogenic source when describing the major factors contributing to the phytoplankton dynamics within P M A . This will be discussed in the subsequent chapters. 81 Chapter 3 - Effects of Entrainment Through the Cooling Waters of Burrard Generating Station on the Phytoplankton of Port Moody Arm, BC 3.1 Introduction Once-through cooled power plants are situated near coastal locations since they require enormous volumes of water to remove waste heat from their condenser systems. For a 2,000 M W generating station, this daily requirement may exceed 4 x l 0 6 m 3 (Roberts et al. 1990; Bamber 1995). The withdrawal and discharge of cooling water from thermoelectric generation has two major impacts on the planktonic biota in aquatic ecosystems: 1) the entrainment of organisms through the condenser system (intake entrainment) where the plankton are subjected to rapid increases in temperature, elevated absolute temperatures, mechanical shearing stresses, and the exposure to disinfectants (chlorination), and 2) the subsequent discharge and incorporation of thermal effluent into the receiving water body (plume entrainment). It is generally accepted that the effect of plume entrainment on planktonic communities is transient and localized (GESAMP 1984). In contrast, intake entrainment has negatively impacted many planktonic forms including bacterioplankton (Choi et al. 2002), phytoplankton (Briand 1975; Jordan et al. 1983; Ahamed et al. 1993), zooplankton (Tunowski 1988; Guseva and Chebotina 2000), and ichthyoplankton (Hergenrader et al. 1982; Patterson and Smith 1982; Mayhem et al. 2000). In effect, the process of intake entrainment acts as a large indiscriminate 'grazer' in aquatic systems, and therefore has the potential to shape trophic energy flows within the affected ecosystem. 82 Phytoplankton are the predominant source of organic material into aquatic food webs. Accordingly, changes in their stocks or photosynthetic capabilities could lead to substantial alterations in trophic dynamics, particularly in small shallow inlets where power plant withdrawals may comprise a significant fraction of the total inlet volume. Since BGS is permitted to withdraw nearly 2 x l 0 6 m 3 d"1 from the shallow environment of P M A , a major goal of this thesis was to determine the influence of this intake entrainment on the total phytoplankton stocks of P M A . Specifically, the objectives of this chapter were: 1) to determine the effects of intake entrainment on the phytoplankton biomass drawn into BGS, 2) to determine the specific effects on individual groups of phytoplankton (diatoms and flagellates) during passage through the BGS condenser system, and 3) to address the potential impact any cellular destruction could have on the total phytoplankton biomass within P M A . 83 3.2 Materials and Methods 3.2.1 Intake entrainment To assess the effect of intake entrainment on the phytoplankton standing stocks of P M A , chl a concentrations in the incoming and discharged water from the BGS cooling-water system were collected twice weekly from June 1999 to the end of October 1999, and again from March 2000 to September 2000. Phytoplankton taxonomic samples were also collected biweekly during the same time period. In all, 30 dates were sampled in 1999 and 43 in 2000. Samples were collected from the pre-intake environment (before the entrained water was subjected to heating or chlorination) via a faucet system connected directly to the intake pipes. Discharge samples were collected from a small pump system located on the discharge platform of BGS, immediately prior to the heated effluent being discharged into the receiving environment of P M A . Outfall samples were collected approximately 2-8 min after the intake samples were collected (depending on discharge rate) to account for the time a single parcel of water travels through the intake-discharge system. On one date (June 14, 1999), outfall samples were collected every two minutes over a 10 minute period to determine time-related chl a variability during pipe passage. Results showed that all outfall samples collected between two to ten minutes were statistically equal (Appendix D). However, outfall samples were collected at calculated times according to their respective discharge rates to ensure the sampled discharge water represented as much as possible, the same populations of phytoplankton drawn into the BGS intake system on each sampling occasion. Although some cell destruction might have occurred during transport of the entrained water through the faucet/pump systems, it is considered that this sampling procedure was still the most representative of the changes affecting the entrained 84 phytoplankton, as it assured that the communities sampled and analyzed were actually being entrained and moved through the discharge system. 3.2.1.1 Phytoplankton biomass (chl a) analysis Duplicate water samples were collected twice weekly from the intake and discharge system and analyzed for chl a biomass (fig l"1) as described in Section 2.2.5. Briefly, 60 ml plastic syringes were rinsed three times with sample-specific water (i.e. intake or discharge water) before the intake/discharge water was purged through a pre-combusted 0.7 |im glass-fibre filter held in a Millipore Swinnex filtering apparatus. Flow rate through the faucet systems was minimal to decrease the likelihood of cell destruction during transport. Filters were wrapped in aluminium foil and quickly frozen at -20°C until further analysis. Chl a concentrations were determined fluorometrically according to Parsons et al. (1984a). Declines in phytoplankton biomass during intake entrainment were calculated as: [100 - (biomass of discharge sample/biomass of intake sample) x 100] and are presented as percent mortality. 3.2.1.2 Enumeration and identification of phytoplankton community Samples were collected, prepared, and identified as described in Section 2.2.6. Briefly, duplicate water samples from both the intake and discharge pipes were collected under reduced faucet flow (to reduce cellular destruction) into 250 ml amber glass bottles. Samples were immediately preserved with acidic Lugol's solution and returned to the 85 laboratory where they were placed in the dark at room temperature until identification commenced (always within 3 months of collection). Phytoplankton cell counts and identifications were performed by settling 10-25 ml subsamples for 24 h in settling chambers and counting on a Zeiss inverted microscope. Random transects were used to count (300-500 individual cells counted per sample) and identify (to genus and species where possible) the phytoplankton. Sample counts were then converted to phytoplankton densities and the species/genus densities were combined into their appropriate groups (i.e., diatoms, dinoflagellates, nanoflagellates, Heterosigma akashiwo, and other phytoplankton) to provide the percent composition of each group in each sample collected. The effect of intake entrainment on the different groups of phytoplankton was determined by: [100 - (phytoplankton density of discharge sample/phytoplankton density of intake sample) x 100] and are expressed as percent declines of the particular phytoplankton group. A l l intact cells viewed under the microscope were considered to be living at the time of collection. 3.2.2 Plume entrainment To study the effect of the thermal discharge plume on the phytoplankton community of P M A , 15 sites within a 3x5 grid (Fig. 3.1) were each sampled 12 times from July 21, 1999 to September 29, 2000, during high, medium, and low discharge regimes. From April 2, 1999 to June 22, 1999 a 3x4 grid was sampled and excluded the easternmost sites near the IOCO shipping dock. The sampling grid was designed to incorporate the near-field receiving environment of BGS, the far-field receiving environment, the pre-discharge environment, and the 'mid-channel' of P M A . The sampling was designed to provide a 86 Fig. 3.1 Sampling grid design for the BGS discharge plume. Refer to Fig. 1.4 for grid location. Arrows at site 4 represent direction of discharge jet. 87 spatial evaluation of the discharge dynamics within the BGS discharge plume. The 3x5 grid began approximately 100 m to the west of the discharge platform (Site G l ; see Fig. 3.1), extended 400 m eastward to the western end of the IOCO shipping dock (Site G13), and approximately 200 m into the main channel of PMA. The grid contained five sites in an east-west direction and three sites in a north-south direction, forming three transects which encompassed the near-shore (Transect 1; Sites G l , G4, G7, G10, and G13), mid-shore (Transect 2; Sites G2, G5, G8, G i l , G14) and mid-channel (Transect 3: Sites G3, G6, G9, G12, G15) environments of P M A . Sites G l , G2, and G3 made up the pre-discharge habitat and acted as a reference to the 'downstream' sites influenced by the discharge effluent. An effort was made to sample during flood tides, to ensure that no discharge effects were measured within the immediate reference sites. Sampling occurred monthly from April 1999 to July 2000, and proceeded at a biweekly pace thereafter until the end of September 2000. Due to extremely low biological activity, only one winter date was sampled (February 24, 2000). Unfortunately, due to equipment failure, physical data were not collected during this date. Most of the hydrographic parameters that were applied in the more widespread investigation of the phytoplankton ecology of Port Moody Arm (see Section 2.2) were used to describe the potential influence of discharge waters on P M A phytoplankton assemblages. The conservative properties of temperature and salinity were used to assess the extent of the plume in a horizontal and vertical direction, while chl a concentrations were measured to determine the immediate effect of BGS discharge on phytoplankton biomass (chl a). Continuous vertical temperature and salinity profiles were performed at each grid site, and samples for the determination of chl a concentrations were collected at 0, 2, and 5 m at each 88 site using a Van Dorn water sampling device. The analytical procedures for each parameter are the same as those described in Section 2.2. Overall, the discharge plume was sampled 19 times, as follows: four times during low discharge events (i.e., <105 m 3 d"1), eight times during medium discharge events (i.e., 105 to 106 m 3 d"1), and seven times during high discharge events (>106 m 3 d"1). To best illustrate the gradual changes in the selected parameters, one date for each discharge event was chosen. Detailed results for all 19 dates are reported in Henry and Harrison (2002). 89 3.3 Results 3.3.1 BGS operation BGS operation showed strong interannual fluctuations during the course of this study (Fig. 3.2A). Electrical production peaked during the late summer and fall of 1998 as well as during two periods in the summer of 2000 (Fig. 3.2A), reaching 700 M W during both events. Between these high production periods, BGS typically operated at -40% of its maximum capacity (950 MW). The discharge of cooling water into P M A paralleled electrical production (Fig. 3.2B). During periods of high production, cooling water discharge approached the maximal allowable limit of 1.7xl0 6 m 3 d"1 (BCMELP 1995), while mean discharge levels were <50% of the allowable limit from November 1998 to May 2000, as BGS operated at less than half capacity. Overall, mean power production and discharge rates during 1999 were significantly lower (p<0.05) than in 1998 and 2000 with the most striking differences occurring between summer and fall. Temperatures of the intake water reflected the seasonal patterns, ranging between 7.5°C in winter and 16.8°C and summer (Fig. 3.2C). Effluent temperatures approached the maximum allowable levels of 27°C (BCMELP 1995) during occasions of high energy production, but were substantially less when BGS operation was reduced between the winter of 1998 and the spring of 2000 (Fig. 3.2C). Intake waters were heated by an average of 9.4±3.9°C after passing through the BGS condenser system (AT), with the greatest increase occurring in the winter as intake temperatures were at seasonal lows. Overall, electrical production, discharge rates, discharge temperatures, and AT values were much lower in 1999 than either 1998 or 2000. BGS operation during this study can therefore be 90 800 J A S O N D J F M A M J J A S O N D J F M A M J J A S Fig. 3.2. Daily values of: A) electrical production; B) discharge rates; and C) temperature characteristics (triangles - discharge temperature, circles - intake temperature, squares -temperature change during transit) during the operation of BGS from July 1998 to September 2000. 91 described by two general trends: 1) high levels of electrical production in 1998 and 2000 that produced thermally-elevated effluent, which approached both the maximum allowable discharge volumes and temperatures as dictated by municipal regulations, and 2) reduced electrical production in 1999 where discharge rates were <50% of 1998 and 2000 levels as well as being discharged at lower temperatures. 3.3.2 Effect of intake entrainment 3.3.2.1 Phytoplankton biomass To gauge the potential for phytoplankton destruction during plant passage, samples were collected on 73 dates over a two-year period for the comparison between chl a levels in the intake waters and the discharge waters. With the exception of six sampling periods (one during 1999, five during 2000), chl a concentrations were always higher within the intake waters than the discharge waters (Figs. 3.3, 3.4). Given that the sampling regimes of 1999 and 2000 were of different durations and covered different seasons, phytoplankton biomass comparisons between the two years will only be made for the summer months (June 21 to September 21) to avoid confounding seasonal differences in phytoplankton biomass. Chl a levels in the intake waters during the summer of 1999 were generally quite low, seldom exceeding 4 u,g l " 1 (Fig. 3.3). During this period, the phytoplankton biomass in the discharge effluent was significantly lower than the BGS intake waters (Table 3.1; t=4.618, df=40, p<0.001) with an average mortality of 71 ±20%. Notable dates include June 23 and 28 when communities dominated by Heterosigma akashiwo were dramatically reduced by 94 and 99%, as well as July 19 and August 16, in which communities dominated by dinoflagellates suffered 97 and 87% reductions in biomass (Fig. 3.3). Despite the dramatic reduction in phytoplankton stocks 92 Fig. 3.3. Concentrations of: A) chl a (u.g l"1) and B) phaeophytin a (u,g l"1) in the intake and discharge waters of BGS from June 14 to October 28, 1999, and C) the associated % declines in chl a during transit (i.e. discharge chl a concentrations divided by intake chl a concentrations x 100). 93 16 0 J 1 1 1 r . 1 March April May June July August Sept Fig. 3.4. Concentrations of: A) chl a (u.g l"1) and B) phaeophytin a (u,g l"1) in the intake and discharge waters of BGS from March 24 to September 29, 2000, and C) the associated % declines in chl a during transit, (i.e. discharge chl a concentrations divided by intake chl a concentrations x 100). 94 Table 3.1. Effects of BGS intake entrainment on average chlorophyll a and phaeophytin a concentrations and the predicted particulate and dissolved carbon (P/DOC) discharged into the P M A receiving waters. Standard deviations are in italics. Chlafl BGS P/DOC P/DOC Date Chlorophyll a Phaeophytin a Chl a :Phaeo a Destroyed Discharg e dischargedb discharged (M-g I') I') fogl"1) (105 m3 d ') (Ckgd') (C tonnes) Intake Outfall Intake Outfall Intake Outfall (March-Sept) 1999c 3.5 1.2 2.1 1.4 1.7 0.7 2.4 4.9 63 15.5 (n=30) 3.4 1.3 1.6 as 1.1 0.5 2.6 2.3 77 7S.7 Summer6 3.6 1.1 1.7 1.2 2.0 0.7 2.5 5.4 60 16.7 (n=21) 3.1 1.4 0.7 0.8 1.2 0.6 2.6 2.3 67 77J 2000d 4.6 2.9 2.5 2.8 2.0 1.1 1.9 10.3 100 24.5 (n=43) 3.0 1.9 1.2 7.5 1.4 0.6 2.2 3.7 112 27.5 Summer6 4.6 2.6 . 2.4 2.4 1.9 1.2 2.0 12.6 129 31.6 (n=20) 2.8 1.7 0.9 1.0 0.9 0.7 1.7 2.2 112 27.5 Total 4.2 2.2 2.3 2.2 1.8 0.9 2.1 8.1 85 20.7 (n=73) 3.2 1.9 1.4 1.4 1.3 0.6 2.4 4.2 700 24.5 a - when the intake chl a < outfall chl a, the chl a destroyed during passage through BGS was considered to be zero. b - product of daily chl a destroyed, daily BGS discharge rate, and C:chl a of 50 c - June 14 to Oct 28 d - March 24 to Sept 29 e - June 21 to Sept 21 moving through BGS condensers, phaeophytin a concentrations were significantly (t=2.172, df=40, p=0.036) lower in the outfall water than the intake waters (Table 3.1). Mean chl a concentrations in the intake waters were higher during 2000 than in 1999 (Table 3.1). As in summer 1999, mean discharge chl a concentrations during summer 2000 were significantly lower than the chl a levels (Fig. 3.4) initially drawn into BGS (t=2.484; df=38, p=0.018). The 40±19% decline in phytoplankton biomass during the summer of 2000 was significantly less than the mortality observed during the summer of 1999 (paired t-test; t=5.052, df=38, p<0.001), despite a two-fold increase in discharge rate in 2000 compared to 1999 (Table 3.1). Phaeophytin a concentrations were not significantly different between the intake and discharge waters during 2000 (p=0.53). The destruction of phytoplankton biomass due to intake entrainment during this study was considerable. Based on daily BGS discharge rates and a C:Chl a conversion of 50, this biomass destruction would have added more than 60-100 kg C d"1 of P/DOC into P M A waters during 1999 and 2000, respectively (Table 3.1). During the major period of phytoplankton growth (March-October), it is estimated that the total input of P/DOC from BGS into P M A would be on the order of 15-25 tonnes (Table 3.1). Despite the substantial destruction of phytoplankton observed in this study due to intake entrainment, BGS discharge rates appeared to have little effect on the in situ chl a levels in the main channel of P M A (Fig. 3.5). 96 1998 1999 2000 Fig. 3.5. Relationship between the integrated chl a levels (mg m"2; grey bars) at Site 3 and BGS discharge rates (open circles) from July 17, 1998 to September 29, 2000. Individual discharge rates are calculated from 5 day means±l SD. Inset is a linear regression between the date-specific integrated chl a concentrations and 5 day mean discharge rates in primary graph. Regression was non-significant (p=0.78). 97 3.3.2.2 Phytoplankton community composition Samples were collected and enumerated from the intake and discharge pipes to determine the influence of intake entrainment on individual categories of phytoplankton. Overall, total cell densities declined by >50% during passage through BGS (Table 3.2). The densities of diatoms, dinoflagellates, and nanoflagellates, the three most prominent types of phytoplankton entrained in BGS cooling water, also experienced significant declines (Table 3.2). Diatoms, with their protective siliceous outer frustules, had the lowest mean % decline in cell density (34.0±32.0%) after cooling pipe passage, while dinoflagellates experienced the highest losses (54.6±29.4%). Mean % decline in nanoflagellate density was intermediate at 48.0±32.7%. Among phytoplankton species, Heterosigma akashiwo suffered the greatest mean % declines in cell density (88.5±9.5%) due to entrainment. For each species enumerated, within-group variability was high and statistical power was low, likely the result of high temporal variability and insufficient sample size (n=12). Therefore, differences in cell density between intake and discharge pipes were deemed insignificant. When data representing each species or genus were combined into their respective groups (i.e. diatoms, dinoflagellates, etc.), thus reducing variability, proper power analysis requirements were met providing significant intake/discharge differences. In future studies, it would be beneficial to collect and count more dates for adequate statistical testing at the species level. 3.3.2.3 Effects of BGS discharge rate and temperature on phytoplankton mortality Perhaps surprisingly, there was a significant negative correlation (r= -0.66, df=40, p<0.001) between BGS discharge rates and the observed phytoplankton mortality from the 98 Table 3.2. Average cell densities of phytoplankton (±SD; italics) in the intake and discharge waters of BGS, their associated % decline of cell densities, and the significance level from the paired t-test. Bold denotes significance at p<0.05. Phytoplankton Intake Discharge % decline p value (103 cells f 1 ) (103 cells f 1 ) (n) Diatoms 1610 807 34.0 0.05 1220 635 32.0 (12) Skeletonema costatum 908 448 35.9 0.21 1110 530 35.3 (12) Chaetoceros spp. 222 135 37.4 0.47 438 227 44.7 (10) Thalassiosira spp. 46 27 31.7 0.34 61 27 26.6 (12) Dinoflagellates 49 22 54.6 0.02 29 20 29.4 (12) Nanoflagellates 311 122 48.0 0.02 238 93 32.7 (12) Heterosigma akashiwo 117 13 88.5 0.21 189 18 9.5 (6) Total cell number 2050 963 45.6 0.01 1170 625 33.0 (12) a - when intake cell densities < discharge cell densities, the % decline was assumed to be zero. 99 intake entrainment of BGS cooling waters during the summers of 1999 and 2000 (Table 3.3). In addition, mortality was negatively correlated to both the absolute discharge temperature (r=-0.51, df=40, p<0.001) and the temperature increase (r= -0.50, df=40, p<0.001) of the BGS cooling water (AT°C). There were no significant correlations between daily absolute temperature or temperature change during intake entrainment and the densities of individual categories of phytoplankton (i.e., diatoms, dinoflagellates, and nanoflagellates). Similarly, correlations between daily discharge rates for BGS cooling water and the corresponding densities of individual categories of phytoplankton were insignificant (p>0.05). The lack of any demonstrable correlation between daily temperature changes or daily discharge rates and density declines of individual categories of entrained phytoplankton was likely a consequence of low sample size; further testing using a larger database would be necessary to distinguish any such correlations, if they exist. 3.3.3 Effect of plume entrainment on temperature, salinity and phytoplankton biomass in the P M A receiving waters The effluent discharged from BGS into the P M A receiving waters can reach velocities of >1 m s"1 and temperatures of 27°C. During winter, the temperature difference between the discharge effluent and surface receiving waters can approach 20°C, while summer values are generally <5°C. To trace the effects of the discharge plume, the thermohaline characteristics (temperature and salinity) and phytoplankton standing stocks (chl a) were measured within a 200 m x 500 m grid (Fig. 3.1) over a variety of discharge regimes (range: 1.4-160xl04 m3d"'). Results from three discharge regimes, low (1.4xl0 4 m 3 d"1), medium (5.5xl0 5 m3d"'), and high (1.5xl0 6 m3d"') are presented in Figs. 3.6-3.8. 100 Table 3.3. Pearson product-moment correlation matrix of % mortality (% declines in chl a) during BGS intake entrainment and selected BGS cooling water characteristics during the summer seasons of 1999-2000 (n=42). BGS Discharge Temperature Discharge Rate Temperature Increase (AT°C) % mortality Coeff. -0.66 -0.51 -0.50 p value <0.001 <0.001 <0.001 BGS Discharge Rate Coeff. 0.83 0.81 p value <0.001 <0.001 Discharge Temperature Coeff. 0.96 p value <0.001 101 Fig. 3.6. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (jxg l"1) through the BGS discharge plume during a low discharge event, August 19, 1999. Discharge rate: 1.4xl0 4 m 3 d . Discharge temperature: 15.1°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3. 102 Fig. 3.7. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (u,g l"1) through the BGS discharge plume during a medium discharge event, August 3, 2000. Discharge rate: 5.5xl0 5 m 3 d"1. Discharge temperature: 23.6°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3. 103 13 14 15 16 17 18 19 20 24.2 24.4 24.6 24.8 25 25.2 25.4 25.6 6 7 8 9 10 11 12 13 Fig. 3.8. Vertical profiles of transect: A) temperature (°C), B) salinity, and C) chl a (jig l"1) through the BGS discharge plume during a high discharge event, September 8, 2000. Discharge rate: 1.5xl0 6 m 3 d"1. Discharge temperature: 26.2°C. Refer to Fig. 3.1 for transect locations and grid site positions. Top: Transect 1; Middle: Transect 2; Bottom: Transect 3. 104 As expected, the vertical and horizontal thermohaline gradients present during August 19, 1999 were largely unaffected during a low discharge event of 1.4xl0 4 m 3 d _ 1 (Fig. 3.6A, B). A discharge temperature of 15.1°C was similar to temperatures in the receiving area (Fig. 3.6A) and a discharge velocity of <0.04 m s"1 created no additional turbulence to the surrounding waters since this was within the daily current range naturally present within P M A . Vertical and horizontal chl a distributions were also comparable between sites (Fig. 3.6C). Differences in the physical and biological environment became evident as BGS discharge levels approached half the maximum allowable limit of 1.7xl0 6 m 3 d"1. On August 3, 2000 BGS discharge was 5.5xl0 5 m 3d"'. This thermally-elevated effluent (AT=10.1°C) was discharged at 23.6°C at a velocity of 0.3 m s'] into surface waters at 19.1°C. This discharge rate created turbulent mixing in the discharge core, as velocities were greater than typical P M A currents (0.1-0.3 m s"1; Waldichuk 1965). There were noticeable temperature and salinity increases in the surface waters along transect 1 (G4 and G7; Fig. 3.7A, B), as the discharge effluent traveled in an eastward direction along the northern shoreline. Despite the continuous discharge of the thermally-elevated effluent, initial discharge temperatures in the outfall were quickly reduced to near-ambient levels, as the warmest surface temperature measured was just 19.6°C at the point of entry. As well, the elevated temperatures near the sediment surface suggest that benthic scouring occurred within the immediate vicinity of the discharge core (Fig. 3.7A). Thereafter, the plume was quickly incorporated into the adjacent P M A waters as it moved in a southeasterly direction past the IOCO pier, 400 m downstream of BGS, with only trace temperature and salinity anomalies present at the eastern extent of transects 2 and 3 (Fig. 3.7A, B). 105 Chl a distributions within the sampling grid were consistent with the thermohaline traits of the discharge plume (Fig. 3.7C). Because the initial chl a concentrations discharged from BGS were low (1.1 u,g r1), integrated chl a levels (mg m"2; 0-5 m) at the nearshore sites, G4, G7, and G10, exhibited biomass levels that were 35, 46, and 48% lower than pre-discharge site averages (Sites G1-G3; see Fig. 3.1 for site locations), respectively. This trend continued into the eastern portion of transects 2 (Gl 1) and 3 (G12) as the plume moved into the mid-channel of P M A (Fig. 3.7C). The high discharge event (1.5xl0 6 m 3 d"1) of September 8, 2000 is shown in Fig. 3.8A-C. The BGS discharge level on this date approached the maximum allowable limit of 1.7xl0 6 m 3d"', and is therefore representative of the periods encompassing August-October 1998 and the summer of 2000 (Fig 3.2). Effluent temperatures (26.2°C) during this time also approached the permit limit of 27°C, and were nearly twice as warm (AT=12.8°C) as the intake water (13.4°C) and approximately 9°C warmer than the contiguous surface waters. The discharge velocity of -0.9 m s"' was about three times greater than the highest tidal velocities measured in P M A (Waldichuk 1965; Taylor and Fissel 1999). Similar to the medium discharge regime, both temperature and salinity were highest in the nearshore environment, although dilution of the plume to ambient surrounding levels proceeded more slowly than the medium discharge event as it moved into the central portion of P M A (Fig. 3.8A, B). Despite the high initial discharge temperature, the maximum temperature measured in the plume was 20.5°C, which was 3.9°C higher than the surface temperature in P M A adjacent to the power plant. The initial chl a concentration in the discharge effluent was 2.6 u.g T1, and hence, biomass levels in the discharge plume were substantially reduced from the >12 u,g l" 1 levels 106 present in the surrounding surface waters (Fig. 3.8C). Nearshore sites of G4, G7, and G10 showed 31, 48, and 37% declines in integrated phytoplankton biomass from the pre-discharge sites (G1-G3), while sites G l 1 and G15 had 44 and 40% declines as plume waters flowed into P M A . The high surface temperatures and low chl a levels at the pre-discharge site of G l indicates that some of the discharge effluent was moving in a westerly direction along the shoreline. The overall trends for the surface temperature, chl a, and phaeophytin a in the discharge plume are presented in Fig. 3.9 and include 12 sampling dates comprising a variety of discharge regimes. On average, surface temperatures in the discharge core were ~2°C warmer than the contiguous waters of P M A (Fig 3.9A). These waters were quickly entrained into the surrounding waters such that the temperatures near the most distant sampling site (S15) were <1°C higher than ambient conditions. Average chl a (Fig. 3.9B) and phaeophytin a (Fig. 3.9C) concentrations followed an opposite trend, with low levels in the discharge core progressively reaching higher concentrations as surrounding waters were entrained into the plume. At the furthest distance from the discharge core, chl a levels were 85-90% of the pre-discharge reference site. 107 West to East Transect Fig. 3.9. Surface contours within the BGS discharge plume of: A) average temperature increase (°C) above minimum sampling grid temperature, B) average site-specific fraction of integrated chl a grid maximum, and C) average site-specific fraction of integrated phaeophytin a grid maximum. 'Discharge' at the top of the page indicates location of BGS discharge platform and the direction of the discharge. Refer to Fig. 3.1 for surface sites and transect locations within the sampling grid. Contours are averages of 12 sampling dates covering a variety of discharge regimes. 108 3.4 Discussion 3.4.1 Effects of intake entrainment on phytoplankton stocks and species composition This study was designed to assess the potential influence of the BGS cooling water system on the entrained phytoplankton stocks of P M A , and to determine what effects the process of intake and plume entrainment had on the phytoplankton ecology of P M A . The data collected during this two-year study demonstrated that the entrainment of P M A waters through BGS caused significant declines in both biomass and cell density during passage through the condenser pipes. Biomass losses resulting from intake entrainment reached 70% during the summer of 1999 and 40% during the same period in 2000. Algal groups were differentially affected by circulation through the BGS cooling system, with dinoflagellates being the most susceptible to damage (55% decline in cell densities on average), followed by nanoflagellates (48% decline) and diatoms (34% decline). In this respect, the operation of BGS effectively acts as an indiscriminate 'grazer' within the P M A ecosystem capable of removing significant amounts of phytoplankton on a continuous basis. The entrainment losses in 1999 were higher than those reported at other power plants, whereas the losses during 2000 were more typical. Briand (1975) reported 42 and 34% reductions in phytoplankton cell numbers and biovolume during passage through a southern Californian power plant, while Socal et al. (1999) observed a 46% decline in average biomass between the intake and discharge canals of a thermoelectric plant located on the shore of the Venice lagoon. Historically though, the effects of intake entrainment on phytoplankton biomass and densities have been ambiguous. For example, Flemer and Sherk (1977) also found a 46% decrease in chl a concentration at Chalk Point Power Plant on the 109 western shore of Chesapeake Bay, but observed no differences during the same time period at an adjacent power plant (Vienna) on the eastern shore. Bienfang and Johnson (1980) and Chang and Rossmann (1985) have also reported no significant changes between intake and discharge biomass levels. The primary effects of cooling water on the standing stock and productivity of entrained algae are due to one, or potentially all of the following: the absolute temperature (T°C) of the discharge water, the temperature increase during transit (AT°C), mechanical stress from increased turbulence, and the exposure to biocide treatment usually in the form of chlorination (GESAMP 1984; Langford 1990; Bamber 1995). During this investigation, phytoplankton mortality was negatively and significantly correlated with the daily discharge rates of BGS, the absolute discharge temperature, and temperature increase (AT°C) associated with plant passage. It is unlikely that increases in cooling water heating from BGS would have enhanced the survival rates of the entrained stocks per se. Several studies have shown that the primary production of entrained assemblages increases when cooler intake temperatures are heated to levels that approach their photosynthetic optima. On the contrary, productivity tends to decline when this optima is surpassed, largely at discharge temperatures >30°C (Morgan and Stross 1969; Peck and Warren 1978; Kozasa and Anraku 1981; Keskitalo 1987). The BGS discharge temperatures between 1998 and 2000 were below the lethal temperature for growth of most temperate estuarine algae (Eppley 1972). Furthermore, transit times of cooling water through BGS are very short, ranging from two minutes during high electrical output (3-6 units in operation) to eight minutes during low output periods (Duval 1998) and algae may be able to withstand temperatures that are higher than their lethal temperatures for several minutes (i.e. brief temperature shock). This 110 rapid transport should rule out any meaningful temperature-induced increases or decreases in biomass. Although not experimentally determined, the mortality of BGS-entrained phytoplankton was probably due to biocide treatment since chlorination was routinely practiced during the summer sampling periods of this study. Numerous lab and field studies have demonstrated that chlorine and its by-products are the principal causes of phytoplankton mortality during power plant operation, leading to the inhibition of phytoplankton photosynthesis (Ho and Roberts 1984), productivity (Carpenter et al. 1972; Fox and.Mover 1975; Flemer and Sherk 1977; Sanders et al. 1981; Ahamed et al. 1993), and reductions in biomass (Brooks and Liptak 1979). Consequently, the decreased mortality during the high discharge, high temperature events at BGS was most likely the result of the shorter transit times that effectively limited the exposure time of the phytoplankton to the biocide. Hirayama and Hirano (1970) showed even a short exposure time of 5 minutes to chlorination could significantly reduce growth (<50%) in the coastal diatom Skeletonema costatum, a dominant species in the P M A waters. Hence, the rapid movement of phytoplankton populations through any intake system that treats its effluent with biocides or subjects entrained organisms to excessive thermal loads would appear to be critically important to the survival of the entrained phytoplankton (provided that the turbulent nature of the discharge water does not lead to cellular damage). This may partially explain the large disparity between the declines in phytoplankton biomass between the summers of 2000 (40%) and 1999 (70%), when the 2000 discharge rates were twice that of 1999. It is also possible that the differences in species composition in the intake waters and their differing tolerances to intake entrainment played a supplementary role in the discrepancy between the mortality rates of 1999 and 2000. In previous studies that reported 111 reductions in cellular biomass due to intake entrainment, flagellate abundance is generally the most affected (Jordan et al. 1983; Sellner et al. 1984 and references therein), although Briand (1975) observed lower declines (33%) in dinoflagellate densities than diatoms (46%) during passage through a southern Californian power plant. In this study, dinoflagellate, nanoflagellate and Heterosigma akashiwo populations were most deleteriously affected by intake entrainment (55, 48, and 89% density declines, respectively), and were the most abundant groups present in P M A during the summer of 1999 (Fig. 2.13). Their increased presence would have contributed to increased mortality rates observed during 1999. Conversely, diatoms (34% decline during passage) were the least affected group of algae and were the most abundant group during the summer of 2000, thus contributing to the decreased mortality rates observed during 2000. Due to poor statistical resolution, species-specific losses due to chlorination or to temperature effects within the cooling water were not detected during this study. Many studies, however, have reported that dinoflagellates are adversely affected by excessive turbulence (Thomas and Gibson 1990; Berdalet 1992; Thomas et al. 1995; Juhl et al. 2000), which has a negative affect on mitosis, and thus cell division (Berdalet 1992). Diatoms, on the other hand, usually dominate turbulent environments (Margalef 1978). It is likely Heterosigma akashiwo densities declined by up to 90% during intake entrainment because this alga is a naked flagellate with no cell wall (Smayda 1998) and is susceptible to even moderate levels of turbulence (M. Henry, personal observation). Consequently, the differential response of the flagellate-dominated communities of 1999 and the diatom-dominated communities of 2000 to turbulence may have contributed to the discrepancy in biomass declines between the two years in this study. Benda and Gulvas (1976) reported 33% declines in productivity in the heated effluent from a Lake Michigan nuclear power 112 plant and estimated that more than half of this decline was due to mechanical stress. However, other studies that have attempted to isolate the mechanical effects of entrained waters on algae have suggested that there is little mechanical influence on entrained phytoplankton populations (Kreh and Derwott 1976; Miller et al. 1976; Peck and Warren 1978; Reetz 1982). In any event, it is reasonable to assume that chlorination, temperature increases, and shearing stress all play some interactive role on phytoplankton mortality during the entrainment into BGS cooling waters, although within the hierarchy of factors, chlorination is most likely the primary causative agent. 3.4.2 Effects of plume entrainment on phytoplankton stocks The effect of plume entrainment on planktonic organisms has been described as extremely localized and of minimal ecological importance, since the heated discharge plumes are quickly incorporated into the surrounding water bodies beyond their discharge areas (GESAMP 1984; Langford 1990). The data collected during this study corroborates this previous suggestion. Temperature, chl a, and phaeophytin a surface contours (Fig. 3.9) all clearly showed that the discharge plume moves from the near-field sites in an easterly direction along the north shore of P M A , and then southeasterly past the IOCO shipping pier into the main channel. Chl a and phaeophytin a concentrations were both substantially reduced within the discharge core, but increased as the discharge waters were mixed into the overlying waters as the plume moved into the main channel. In contrast, temperatures were predictably higher within the discharge core before declining as the plume was entrained into the adjoining waters of P M A . By the time the discharge plume was 400 m downstream of the outfall, phytoplankton biomass and temperature levels were largely restored to that of pre-discharge sites. Duval (1998) estimated that the spatial influence of the discharge plume 113 would range between 1-17% of P M A , depending on the tidal regime, with an average of 10%. The decrease in phytoplankton biomass observed in the discharge plume was probably unrelated to plume entrainment since the discharge temperatures in the immediate outfall region were immediately cooled to temperatures within a few degrees of ambient surface temperatures. Even during the summer period when the surface waters of P M A were at seasonal highs and BGS discharge was near 27°C, discharge temperatures were quickly reduced to levels considered favourable for temperate estuarine algae (Eppley 1972; Suzuki and Takahashi 1995). Furthermore, it is likely that the residual chlorine within the discharge plume had negligible effects on the entrained phytoplankton since the effluent is dechlorinated (<0.02 mg l"1) before being ejected and undergoes rapid dilution and decay within the discharge plume (Eppley et al. 1976). Reductions of >50% in cell densities (Sanders et al. 1981) and productivity (Eppley et al. 1976) have been observed at comparable chlorine concentrations, but it is generally thought that the gross impact of the residual chlorine on phytoplankton is minimal due its the swift dilution into the receiving waters (Eppley et al. 1976; Goldman and Davidson 1977; Sanders et al. 1981). The biomass reduction observed in the discharge plume during this study is directly related to the low biomass that is initially drawn into BGS coupled with the roughly 50% mortality during plant passage. Since the intake waters are drawn from below the pycnocline where biomass is always low (see Fig. 2.12) and discharged into the surface euphotic zone where biomass levels are predictably higher, the average biomass discharged into P M A (1.8±1.6 u,g chl a 1"'; averaged summer 1999-2000 values) was generally <20% of the average biomass present in the upper 5 m (10.9±9.7 u,g chl a T1; averaged summer 1999-2000 values) causing a dilution of biomass within the discharge core. Perhaps 114 surprisingly, there was no accompanying increase in phaeophytin a concentrations in the discharge core which is a feature commonly associated with cellular destruction during grazing events (Parsons et al. 1984b). This may be because the cellular destruction within the condenser pipes was complete and the cell contents entered the dissolved phase, thereby passing through the filter during collection. Any permanent influence of the discharge waters in P M A would likely be seen in the benthic community situated in the immediate outfall region, where flow rates are high and variable, and the outflow is thermally elevated and chlorinated. Deleterious biological effects resulting from power plant effluents are consistently reported in areas inhabited by sessile organisms such as periphyton (Hein and Koppen 1979; Snoeijs and Prentice 1989), corals (Jokiel and Coles 1974), macroalgae (Keskitalo and Ilus 1987) and sea grass beds (Roessler 1971). This area in P M A would be similar to lotic habitats such as streams and rivers where alterations in flow regimes are frequent. In this instance, the best suited inhabitants for this benthic environment would be eurythermal and chlorine-tolerant organisms adapted to high flow regimes and capable of rapid recolonization after spate conditions. Because of the small area and the unique physical environment, this locale could be considered a discrete semi-permanent microhabitat within P M A . This study has shown that BGS is capable of destroying a significant portion of the phytoplankton drawn through its condenser system, ultimately ejecting several tonnes of organic carbon into the surrounding waters of P M A during the course of the phytoplankton growing season (March-October). To properly assess the potential effects that the intake entrainment had on the phytoplankton within P M A , the influence that this entrainment had on the contiguous waters of P M A must be put into an ecological perspective. The maximum allowable volume of effluent that BGS is permitted to release daily into P M A is 1.7xl0 6 m 3 . 115 In contrast, the twice-daily tidal prism moving into P M A , depending on tidal cycle, ranges from 5.1xl0 6 to 14.5xl0 6 m 3 (Waldichuk 1965). Therefore, under maximum operating conditions BGS can entrain only approximately 12-33% of the incoming tidal prism on a daily basis. However, based on average total P M A volumes, only -3-4% of the P M A waters can be maximally entrained on any given day, within which -50% (1999-2000 average) of the phytoplankton biomass is killed. This would suggest that approximately 1.5-2% of the P M A phytoplankton biomass could be killed daily during intake entrainment when BGS is operating at maximum loads. However, intake waters are drawn from depth (<5 m), which typically contains less than 20% of the average phytoplankton biomass present in surface waters of P M A . Therefore, it is likely that much less than 1% of the phytoplankton community in P M A is killed during a day of maximum discharge from BGS. An important (though previously unaddressed) issue is that a large percentage of the phytoplankton entrained into BGS is likely comprised of diatoms that had previously sunk below the pycnocline due to their lack of vertical motility and high sinking rates (Smayda 1970; Smetacek 1985: Waite 1992), particularly during N 0 3 limitation (Waite et al. 1992). Thus, the entrained diatoms would have otherwise been lost from the system via sedimentation or consumption by 'deep-water' secondary consumers. This suggests that power plants may possibly reseed the upper euphotic layer, leading to surface gains in biomass as opposed to the losses which are generally believed to occur in this field of research. Phytoplankton in coastal temperate areas have high growth rates, commonly exceeding one doubling per day (Parsons et al. 1984b). This allows the phytoplankton community to recover rapidly following exposure to the biocidal, thermal, and shearing stresses during passage through the power plant condenser system (Hirayama and Hirano 1970; Fox and Moyer 1975; Eppley et al. 1976; Goldman and Davidson 1977). Therefore, 116 phytoplankton mortality due to intake entrainment is a small and localized effect, as populations are restored by natural regeneration and the dilution of the discharge plume. It has been suggested that a daily kill rate of 20% for entrained phytoplankton would not affect net population sizes to any appreciable degree (Glasstone and Jordan 1980). Given the aforementioned factors, it is very unlikely that the effects of BGS intake entrainment-related mortality on phytoplankton, whether due to chlorination, temperature, or turbulence, had any appreciable effects on the total P M A phytoplankton biomass, productivity, or species diversity. One important element relating to the BGS cooling water and its potential effects on the phytoplankton ecology of P M A has not been discussed to this point. As mentioned previously, BGS draws its intake water from below the pycnocline and expels it into the surface waters of P M A . Because the water below the pycnocline in P M A is nutrient-rich (relative to the surface layer), this water may act as a source for new primary production in the estuary, thus potentially enhancing phytoplankton standing stocks within P M A . The chl a concentrations indicated this was not the case (Fig. 3.5), as BGS discharge rates were unrelated to the chl a standing stocks of P M A . The most important influence of the effluent may take place during the summer season when the surface waters of P M A are chronically limited by low levels of NO3 (Appendix C). Since flagellates were the dominant phytoplankton in P M A during the summer of 1999 and diatoms during 2000, it is possible that the level of BGS operation may ultimately dictate the dominant phytoplankton assemblages present during the summer season in P M A based on the cycling of nutrients into the surface waters. This phenomenon will be discussed in the next chapter. 117 Chapter 4 - Effects of Nutrient Inputs from Once-through Power Plant Generation and 'Natural' Entrainment on the Summer Phytoplankton Community of Port Moody Arm 4.1 Introduction Over the last 30 years, the role of nutrients in the coastal zone has come under increased scrutiny as it is evident that anthropogenic nutrient inputs are being disproportionately added to the coastal margins (Nixon 1995; Nixon 1998). These inputs have been implicated in coastal eutrophication (Nixon 1995; Cloern 2001; Pickney et al. 2001), shifts in the plankton community and trophic structure (Elmgren 1989;.Sellner 1997; McClelland and Valiela 1998; Turner et al. 1998; Turner 2002), increased presence of H A B species and bloom frequency (Smayda 1990; 1997), and the overall decline of water quality in estuarine habitats (Nixon 1998; National Research Council 2000). To assess whether any alteration to the marine environment is due to 'natural' or human causes, a proper assessment of each cause is necessary before appropriate remedial action can take place (if required). In the case of nutrients, delivery into the coastal zone comes from two primary sources: 1) non-point source inputs, which include riverine inputs, atmospheric deposition (dry and wet), groundwater, and deepwater entrainment into the surface layer; and 2) point source inputs, which encompass primarily anthropogenic sources such as sewage and wastewater treatment effluents, industrial discharge, and stormwater overflows (Pickney et al. 2001). The operation of thermal and nuclear power plants has been identified as a potential threat to estuaries and their surrounding regions due to the abundant discharge of thermal effluents (GESAMP 1984; Langford 1990), the emission of substantial quantities of 118 greenhouse gases and airborne contaminants into the atmosphere (Mohnen and Wang 1992), as well as the entrainment and impingement of marine organisms as cooling waters are drawn through the power plant condenser system (Kennish 1997). Due to inefficiencies in the conversion of thermal to electrical energy (-40%; G E S A M P 1984), 'once-through' thermal and nuclear power plants require vast quantities of cooling water to remove the excess heat from their condenser systems. Accordingly, these plants are located adjacent to bodies of water that can fulfill this requirement. In order to maximize its heat absorbing capacity, power plant cooling water is generally drawn from the cold deep waters, which are typically nutrient-rich relative to the surface layer. From a nutrient and eutrophication perspective, this is important as these intake waters are usually returned to the surface layer as a thermally-elevated plume (Sobey et al. 1988) where the nutrients become available for phytoplankton and macroalgae acquisition. Despite the recognition of this phenomenon (GESAMP 1984; Langford 1990), there have been no attempts to quantify the effects of the nutrient loads contributed to the affected water bodies during electrical production, although circumstantial evidence does exist (Jahn et al. 1998). Since electrical generation is the largest single industrial user of water (Nalewajko and Dunstall 1994), the release of this cooling water effluent is another major anthropogenic source of nutrients into the coastal zone and therefore must be considered with respect to potential regional eutrophication or trophic alterations within the aquatic habitat. Burrard Generating Station (BGS) is a gas-fired thermal electric generating station located on the north shore of Port Moody Arm. At peak performance, BGS is capable of supplying 1.7xl0 6 m 3 d~' of cooling water to the upper layer of P M A and thus it represents a prominent nutrient source to the inlet, particularly due to its shallow depth (average: -9 m) and small surface area (4.8 km 2 at mean tidal height). Furthermore, P M A is a tidally-119 dominated estuary with a positive estuarine circulation (Waldichuk 1965), and consequently, nutrient-rich subsurface water is mixed into the surface layer during seaward transport as well as over diel and fortnightly tidal cycles. Because summer precipitation is low, and riverine inputs into P M A are seasonal and low in nutrients, power plant discharges and vertical mixing processes are likely to be principle factors influencing the phytoplankton standing stock, production, and composition in P M A . Nutrient additions may have substantial impacts when light levels are sufficient, and phytoplankton growth is limited by a particular nutrient (Cloern 1999). In P M A , NO3 often limits phytoplankton biomass production during the summer season. During this study, summer phytoplankton assemblages responded differently to this NO3 limitation, as flagellates dominated during 1999 and diatoms during 2000 (see Chapter 2). The goal of this chapter was to: 1) determine the environmental factors that contributed to the differences between the summer phytoplankton assemblages in P M A , and 2) determine whether direct inputs from BGS operation or from 'natural' mixing events were responsible for the morphotypic alterations observed in the summer P M A phytoplankton assemblages. This is the first study to address the influence of injecting subsurface nutrients into the surface layer via power plant operation in a coastal inlet and its associated effects on the local phytoplankton community. 120 4.2 Materials and Methods 4.2.1 Phytoplankton composition and environmental gradients To assess the environmental factors that contributed to alternating years of flagellate and diatom dominance in the summer phytoplankton assemblages of P M A , phytoplankton abundance and environmental data collected from June to September in both 1999 and 2000 were used from the general 2-year survey of P M A (Chapter 2). The methods of collection, processing, and analysis of these data have been described in Section 2.2. Methods for phytoplankton enumeration and identification are detailed in Section 2.2.6. Environmental data including thermohaline (temperature, salinity, and stratification) and nutrient data (NO3 and Si(OH)4) are summarized in Section 2.3.2 and 2.3.3, respectively. BGS discharge and temperature data were provided by BC Hydro, and meteorological data (irradiance and wind speed) for P M A were obtained from the Greater Vancouver Regional District (GVRD). 4.2.2 Data Analysis General relationships between the summer phytoplankton taxa and the environmental variables listed above were assessed using canonical ordination; a combination of ordination and multiple regression techniques, which link one set of variables to one or more variable sets. The efficiency of a particular ordination technique is related to species' responses (linear or unimodal) to environmental gradients (Jongman et al. 1987). To establish the direct gradient canonical technique3 best-suited to evaluate P M A species-environmental relationships, Detrended Correspondence Analysis (DCA) was first 3 Direct gradient analysis expresses species patterns through environmental data unlike indirect gradient analysis, where only species data are used to calculate ordinations (i.e. D C A , Multidimensional Scaling). 121 applied to the summer phytoplankton data. D C A is typically used to remove 'arch' effects4 within ordinations (Hill and Gauch 1980), however, it is also useful for determining the gradient length of data, therefore establishing whether a linear or unimodal direct gradient technique should be applied. Short environmental gradients (<1.5 standard deviation, SD) suggest a linear species response model (Redundancy Analysis, RDA), while long gradients require a unimodal response model such as Canonical Correspondence Analysis (CCA; ter Braak 1995). D C A showed gradient lengths >2.5 SD for the first two axes (see Section 4.3.1), and the relationships between the summer phytoplankton assemblages and environmental gradients were examined using C C A ; a direct gradient technique where the ordinal axes are linear combinations of environmental variables and species responses to environmental gradients are unimodal. Statistical routines involving D C A and C C A were run on the statistical package CANOCO Ver. 4.0 (ter Braak and Smilauer 1998). Initially, phytoplankton taxa that comprised at least 1 % of the total abundance on any one summer date were used in the C C A . To reduce potential distortion in the ordination, analyses were run with the down-weighting of rare taxa (ter Braak and Smilauer 1998), and environmental variables that had high Variable Inflation Factors (VIF>20)5 and multiple significant relationships with other environmental variables were removed form the analysis (ter Braak and Smilauer 1998). Selected variables were subsequently tested for significance (p<0.05) against the first canonical axis and those that contributed significantly to the model were maintained within the analysis. The significance of the canonical axes were tested using Monte Carlo tests with 199 permutations (p<0.05). 4 distortion within an ordination caused by unimodal species responses to environmental gradients 5 Environmental variables with high VIFs (>20) are highly correlated with other environmental variables (multicollinear). Consequently, the variable is largely redundant and offers no distinct interpretive value in the analysis (ter Braak 1986). 122 4.2.3 Nutrients in BGS cooling water discharge To gauge the nutrient loading capability of BGS, NO3, PO4, and Si(OH) 4 samples were collected from the BGS cooling-water system twice weekly from June 1999 to the end of October 1999, and again from March 2000 to September 2000. In all, 30 dates were sampled in 1999 and 43 in 2000. The discharge samples were collected from a small pump system located on the discharge platform of BGS immediately prior to the heated cooling water being discharged into the receiving environment of P M A . Samples were drawn using acid-rinsed 60 ml plastic syringes, and were purged through combusted (460°C for 4 h) 0.7 u.m glass-fibre filters into acid-cleaned 30 ml Nalgene polyethylene bottles. The filters were held in 25 mm acid-cleaned Millipore Swinnex® filter holders. Once completed, the bottles were maintained at 0°C until they were returned to the laboratory where they were frozen at -20°C until further analysis. Standard nutrient analysis techniques were performed for the three macronutrients using a Technicon AutoAnalyzer® II (Hager et al. 1968). Daily BGS nutrient loading rates were calculated as the product of the daily nutrient concentration and the daily discharge rate of BGS. The daily discharge rates were provided by BC Hydro. NO3 samples were also collected within the receiving environment of the BGS discharge to understand the entrainment of this nutrient-laden discharge into the contiguous waters of P M A during different discharge regimes. The temporal and spatial sampling design is fully described in Section 3.2.2 and the sample grid is illustrated in Fig. 3.1. Briefly, NO3 samples were collected from each site of the 3x5 grid at 0, 2, and 5 m. Processing and analysis are described in the preceding paragraph. Overall, from the 19 dates sampled, three dates involving a low discharge, medium discharge, and high discharge event were chosen to best illustrate the gradual changes in NO3 concentrations as the 123 discharge plume moved into the central channel of P M A . Only the nearshore transect is presented for each discharge event, and each event occurred during periods of NO3 limitation (<2 u,M; Eppley et al. 1969; Appendix C). 4.2.4 Vertical Entrainment Model An inlet-scale 2D model was developed to estimate the degree of vertical mixing and the flux of nutrient-rich subsurface waters into the surface waters of P M A . This model was applied to 12 dates between June and September, 1999 and 14 dates from June-September, 2000 and was designed to encompass the period of the summer flagellate/diatom transition in P M A . Because of the shallow depth of the intertidal area at Site 5, only entrainment fluxes between sites 1, 2, 3, and 4 were calculated. P M A waters have been confirmed to have a positive estuarine circulation (Seaconsult 1995; Taylor and Fissel 1999; Jiang et al. 2003). The presumption of the 2D model is that the difference in surface flow rate between two sites should reflect the rate of vertical entrainment flux incorporated into the outflowing surface layer. For example, if the estimated surface layer outflow at two adjacent sites is l x l O 6 and 2 x l 0 6 m 3 d"1, respectively, then the vertical entrainment rate between the two sites is assumed to be approximately 1x10 m d" . Therefore, the nutrient flux into the surface layer can be estimated as the product of nutrient concentration of the bottom water and the vertical entrainment flux. Vertical entrainment rates, or estuarine amplification, can be estimated by the differences in water volumes that exist along spatial gradients. Conservation of mass necessitates that the volume of water leaving P M A in the surface waters must be balanced by the incoming bottom water flow (Eq. 4.1; Fig. 4.1). 124 Fig. 4.1. Two-layer estuarine flow within Port Moody Arm and the influence of the operation of Burrard Generating Station (BGS). Q = volume; T = temperature; out = outflowing; in = inflowing; d = discharge; int = intake. P 0 = initial pycnocline depth; Pf = final pycnocline depth. 125 where: Q o u t = Q i , (Eq. 4.1) Qout = volume of water leaving P M A in the surface layer Q i n = volume of water from the incoming bottom water flow However, because BGS draws water from depth and discharges the cooling water into the surface layer of P M A , this equation must be modified to include this vertical cycling (Eq. 4.2). where: Q o u t + Q d = Qin + Qim (Eq. 4.2) Q d = volume of BGS cooling water discharged into the surface layer Qim = volume of water drawn into BGS Eq. 4.2 indicates that more water is leaving P M A in the surface waters as a result of the additional BGS discharge, but this is counteracted by an increase in the flow of bottom waters into the Arm to replace the water drawn into BGS. This suggestion has been substantiated by Jiang et al. (2003). Tracking the volume of water masses does little to quantify the effects of BGS on the vertical entrainment within P M A other than to suggest that it increases surface and subsurface flow rates (and that entrainment rates should therefore increase). The conservative features of salinity and temperature, in combination with water volume estimates, can be explored to track the water properties originating from the discharge of once-through BGS cooling water. Because cooling water salinities and volumes remain 126 unaltered during passage through B G S , it is not a useful tracer as the salinity-based water volumes (QdSd and QjmSint) from B G S cancel, leaving a total salt balance for the Arm as follows (Eq. 4.3): Where: QoutSout + QdSd = QinS i n + QintSint (Eq. 4.3) S o u t = salinity of surface water Sin = salinity of incoming bottom waters Sd = salinity of discharge waters S i m = salinity of intake waters Because Q d S d = QintSint, by elimination: where: QoutS o ut = Q i ntSint (Eq. 4.4) Since BGS alters the temperature of the intake waters (AT), temperature makes a much better tracer for the influence of BGS cooling water discharge and estimates of surface layer outflow (Fig. 4.1), which is required to estimate the vertical entrainment rates in P M A (Eq. 4.5): 127 where: (Eq. 4.5) T o u t = depth-averaged temperature of surface water (°C) T^ = depth-averaged temperature of bottom water (°C) T d = temperature of cooling-water discharge (°C) T^t = temperature of cooling-water intake (°C) Because Qd=Qim and Q 0 Ut=Qin (Eq. 4.1), rearranging the above equation gives an estimate of the surface layer volume at a particular site within P M A as follows: Thus site-specific volumes of outgoing water can be estimated directly from the following: BGS discharge volumes, the difference between BGS intake and discharge temperatures, and the depth-averaged difference between the surface and subsurface layers of P M A . The degree of entrainment between two sites is therefore the difference between their estimated outflow rate, which is a function of the incorporation of colder and presumably nutrient-rich deep water into the surface layer in the seaward direction (i.e. towards Site 1 and into Burrard Inlet). The average temperatures of the surface and subsurface layer at each site (Eq. 4.5) were calculated by first determining the pycnocline depth (Pawlowicz and Farmer 1998) using density data (ot) collected during the standard sampling survey (see Section 2.2.3). Vertical density profiles are presented in Appendix E. The pycnocline depth was assumed to represent the interface between the seaward and landward flowing masses as no current Qout - Qd LTdlTin tL [T o ut"Tin] (Eq. 4.6) 128 measurements were made during this study. Using thermohaline data from September 1998, calculations of pycnocline depth at SI using this method (6.2 m) were very similar to those generated during the same period from current profiles (-6.5 m) by Taylor et al. (2001) using an Acoustic Doppler Current Profiler (ADCP). Therefore, it is assumed that this method accurately estimates the depth of the opposite flowing layers. Once pycnocline depths were determined, depth-averaged temperatures were calculated for each layer using trapezoidal integration (Ichimura et al. 1980). To avoid spatial biases, the depth of integration for the bottom layer temperatures at each site over one sampling date was taken as the shallowest depth of the four sampled sites. This would ensure that deeper sites did not have colder average bottom water temperatures due to the deeper cold water profile. Since S4 was normally the shallowest site, all bottom depths (i.e. for SI-S3) were taken as the depth between the pycnocline and the sediment depth of S4. 4.2.5 Estimation of vertical nutrient fluxes The quantity of a particular nutrient added to the surface layer of P M A is a function of its rate of addition and its concentration within the original water mass. For BGS, this quantity was accurately determined through known discharge rates and discharge nutrient concentrations. To assess the nutrient load that is 'naturally' entrained into the P M A surface layer, the estimated vertical entrainment rates were multiplied by the most representative bottom water nutrient concentration collected in this study. Because of the increased sampling frequency of the BGS intake waters relative to the general survey (Chapter 2), and the fact that the intake depth of this cooling water occurs at -11 m (which represents true bottom water within PMA), nutrient concentrations in the intake waters were used to estimate the vertical nutrient entrainment into the P M A surface waters. However, the intake 129 nutrient samples were not collected at similar periods of the tidal cycle during each sampling date. Since the BGS intake pipes are at a fixed depth, these samples had lower nutrient concentrations during low low spring tides when surface nutrient limitation was present than those samples collected at high tides due to the greater deep-water influence (further from the pycnocline; see Jiang and Fissel (2004) for description of cooling water recirculation due to diel tidal fluctuations). To minimize these daily tidal discrepancies, monthly averaged intake nutrient concentrations were used to determine the vertical nutrient entrainment flux in P M A . By using deep water samples (-11 m) together with the entrainment calculations, estimates of nutrient fluxes into the surface waters may be slightly overestimated since the upper section of the landward travelling bottom water (the portion of water that would be incorporated into the seaward flow) is measurably diluted by the downward movement of the overlying seaward-flowing surface layer (Taylor and Fissel 1999). Accordingly, this section of the water column should be lower in nutrients during the periods of nutrient limitation observed during the summer season. Thus, using the higher nutrient concentrations of the true bottom water will slightly overestimate the nutrient load naturally mixed into the surface layer. 130 4.3 Results 4.3.1 Relationship between environmental factors and phytoplankton community composition during P M A summers A major finding during this study was the interannual variation in phytoplankton composition, where diatoms largely dominated the summer assemblages in 1998 and 2000, and flagellate species were dominant in 1999 (Fig. 4.2). Skeletonema costatum was the most prevalent diatom during 1998 and 2000, while Heterosigma akashiwo, Prorocentrum gracile, Heterocapsa triquerta, and Katodinium rotundatum made important contributions to flagellate densities during the summer of 1999. A full account is described in Section 2.3.5 and the data are presented in Henry and Harrison (2002). The relationship between the summer phytoplankton assemblages and their potential causative environmental factors (both naturally occurring and those resulting from BGS operation) were assessed using D C A and C C A . D C A revealed that the lengths of the environmental gradients for the first two ordination axes were sufficiently large (>1.5; Table 4.1) to test for species-environmental interactions using C C A (ter Braak 1995; see Section 4.2.2 for explanation). To best describe the species data, eight environmental factors were initially selected for analysis. Of these, temperature explained the least variance in the species data (Table 4.2) and was subsequently removed from further analysis. Salinity was also removed from the C C A due to high variable inflation factors (VIF>20) as it was highly correlated with stratification, Si(OH) 4, and BGS operation (Table 4.3), and thus yielded little additional interpretative value. 131 Diatoms |;"^ H[ Dinoflagellates | | Nanoflagellates H. akashiwo Others 2 100 80 60 40 20 0 ' I W July Aug June July Aug 1998 1999 (6.4±3.6 ug chl a l"1) (18.2+18.2 ug chl a l"1) TTTT June July Aug 2000 (8.9±3.7 iig chl a I'1) Fig. 4.2. Summer phytoplankton community composition (carbon-based biomass) at Site 3 in Port Moody Arm, 1998-2000. Values in parentheses on x-axis are average chl a concentrations for that particular summer (1998, n=6; 1999, n=12; 2000, n=21). 132 Table 4.1. Summary of Detrended Correspondence Analysis (DCA) involving the summer phytoplankton of P M A , 1998-2000. Total Axes 1 2 3 4 inertia Eigenvalues 0.908 0.629 0.028 0.015 Lengths of gradient 3.717 2.500 0.604 1.099 Species-environment correlations 0.973 0.962 0.890 0.809 Cumulative percentage variance of species data 45.7 77.3 78.7 83.1 of species-environment relation 53.9 89.6 133 Table 4.2. Summary of variance explained by environmental variables within Canonical Correspondence Analysis (CCA) and interset correlations between six forward-selected environmental variables and the first two canonical axes. Environmental Variance Interset correlations variable explained3 Axis 1 Axis 2 Stratification 0.64 0.85 0.12 Si(OH) 4 0.63 0.80 0.30 BGS Discharge 0.48 -0.61 -0.45 N 0 3 0.31 -0.15 -0.67 Wind 0.25 0.47 0.23 Light 0.17 0.27 -0.46 Salinity b 0.60 Temperature c 0.03 - testing for variance in species data explained by individual environmental variables using manual forward-selection in C C A (ter Braak and Smilauer 1998) b - removed from C C A due to high correlation other environmental variables (Table 4.3) c - removed from C C A due to low variance explained factor 134 Table 4.3. Pearsons correlation matrix of eight selected environmental variables used in C C A . * indicates significance at p<0.05. Env. Variable Temp Salinity Strat N 0 3 Si(OH) 4 BGS Light Salinity 0.27 Stratification -0.23 -0.96 * N 0 3 -0.18 -0.04 0.10 Si(OFf)4 0.10 -0.81 * 0.78 * -0.22 BGS discharge 0.19 0.80 * -0.78 * 0.17 -0.78 Light 0.43 -0.03 0.13 0.29 -0.01 0.11 Wind 0.40 -0.24 0.43 -0.25 0.42 -0.21 135 The six remaining environmental variables explained a large portion of the phytoplankton variation within the C C A (>65%; Table 4.4), yielding large eigenvalues for the first (0.77) and second (0.52) canonical axes, with both canonical axes being significant (p=0.015; 199 Monte Carlo permutations). Subsequent axes had extremely low eigenvalues and thus provided very little additional information and their contribution will not be discussed further. Interset correlations revealed that (in descending order) vertical stratification, Si(OH) 4 concentrations, BGS discharge rates, and NO3 concentrations were most highly correlated with the first two canonical axes (Table 4.2). Stratification and Si(OH) 4 levels had strong positive correlations with the first axis, while BGS discharge was negatively correlated. NO3 concentrations had the highest association with the second axis and were also negatively correlated. To best interpret the species-environmental relationship provided by the CCA, an ordination-regression biplot diagram was employed (Fig. 4.3). Species abbreviations presented in the biplot are listed in Table 4.5. The summer phytoplankton assemblages of P M A were clearly separated into two distinct morphotypes along the environmental gradients. A group of diatom species, including the dominant species, Skeletonema costatum, was located in the lower left-hand quadrant of the biplot and was therefore associated with periods of high BGS discharge and high surface NO3 concentrations that occurred during the summers of 1998 and 2000. These diatom species also occurred during periods of diminished stratification, wind speeds, and Si(OH) 4 concentrations. Chaetoceros spp., Leptocylindrus danicus, L. minimus, and Thalassionema nitzschoides were most abundant during a period of low irradiance and NO3 levels, and are accordingly clustered in the upper left quadrant of the biplot. 136 Table 4.4. Results of Canonical Correspondence Analysis (CCA) on common summer P M A phytoplankton and selected environmental variables. Total Axes 1 2 3 4 inertia Eigenvalues Species-environment correlations Cumulative percentage variance of species data of species-environment relation • p value of canonical axes* * Monte Carlo tested with 199 permutations 0.778 0.523 0.060 0.017 0.934 0.864 0.568 0.663 39.2 65.5 68.5 69.4 55.8 93.2 97.5 98.7 0.015 0.015 1.986 137 Axis 2 (X=0.52) r Lmin Ldan T n e m a • Chaet Axis 1 (X=0.78) 1 Amph <$> Gyro Gymno Scrip <>Dphy < > K a t 0 c : r r » u \ < & y p w i n d e r : s , ( O H ) 4 ^•Nano ^ Haka ^ ^ ^ ^ _ J ^ - * Strat /i5enn / / * / Cylin / / T s i r a ^ / • T t y B G S „ . S k e l / Pnitz 7 • / Rhiz • N 0 3 \ D Eugl P P d m " Light ° H c a p Fig. 4.3. Canonical Correspondence Analysis (CCA) biplot showing relationships between phytoplankton taxa (points) and environmental factors (arrows) during the summers of 1998-2000 in Port Moody Arm. Groups of algae are diatoms (dark circles), dinoflagellates (grey diamonds), Heterosigma akashiwo (dark diamond), unidentified nanoflagellates (dark square), and euglenoids (white square). Phytoplankton taxon abbreviations are listed in Table 4.5. BGS=Burrard Generating Station discharge rate, Strat=stratification. 138 Table 4.5. Phytoplankton taxon abbreviations used in C C A biplot in Fig. 4.3. Phytoplankton Group C C A Abbreviation S pecies/Genera/Unidentified Diatoms Skel Skeletonema costatum Chaet Chaetoceros spp. Tsira Thalassiossira spp. Tnema Thalassionema nitzschiodes Dity Ditylum brightwellii Ldan Leptocylindrus danicus Lmin Leptocylindrus minimus Cylin Cylindrotheca closterium Pnitz Pseudo-nitzschia spp. Perm Pennate diatoms Rhiz Rhizosolenia spp. Proro Prorocentrum spp. Dphy Dinophysis spp. Oxy Oxyphysis Heap Heterocapsa spp. Kato Katodinium spp. Gymno Gymnodinium spp. Gyro Gyrodinium spp. Cera Ceratium fusus Ppdm Protoperidinium spp. Scrip Scrippsiella spp. Amph Amphidinium spp. Dinoflagellates Dflag Unidentified dinoflagellates Others Haka Nano Heterosigma akashiwo Nanoflagellates Euglenoids Eugl 139 Contrary to the diatoms, the flagellate species were located in the upper right-hand quadrant, indicating that they were more common in highly stratified water with low NO3 and high Si(OH)4 concentrations. Species within this quadrant also occurred during periods of reduced BGS operation. The occurrence of the dominant flagellates, Heterosigma akashiwo and Prorocentrum gracile was related primarily to the high surface stability and reduced BGS operation present during the summer of 1999 (Fig. 4.3), whereas Katodinium spp., Dinophysis spp., Oxyphysis spp., Ceratiumfusus, and unidentified nanoflagellates were present in increasingly N03-limited waters. Gymnodinium spp., Amphidinium spp., Gyrodinium spp., and Scrippsiella spp. were most abundant during a period of severe NO3 limitation and BGS inactivity (low discharge), and formed an independent group in the top right-side in the biplot. Overall, it is apparent that the presence of a particular phytoplankton morphotype within the summer waters of P M A is dictated by the factors that influence the hydrodynamic environment (i.e. stratification and BGS output) and the nutrient content of the P M A surface water. The dominant nutrient source advected into the summer surface waters of P M A will be the focus of the next section. 4.3.2 Nutrient concentrations in BGS cooling water discharge BGS cooling water is drawn from depth (-11 m) and, after travelling through the condenser system, its effluent is discharged as a buoyant thermally-elevated plume into the surface waters of P M A (-3-5 m). This cooling water discharge was seasonally and annually variable during 1998-2000 (Fig. 3.2), and during the phytoplankton growing season, it was enriched with nutrients (with respect to the upper water column). During the sampling period of 1999 (June-September), NO3 concentrations discharged from BGS into the P M A receiving waters averaged 11.3±4.4 u,M. Early to mid-summer NO3 discharge 140 concentrations were typically between 5-10 uM, while late summer and fall levels were always >10 \iM (Fig. 4.4A), giving an average summer NO3 concentration of 9.3±3.4 u.M (June 21-September 21). Mean discharge concentrations of PO4 during the summer of 1999 were 1.6±0.4xiM, resulting in a relatively low N 0 3 : P 0 4 ratio of 6.4±2.2 (Table 4.6). Si(OH) 4 discharge concentrations were generally >35 \iM, but exhibited high variability, frequently oscillating between 10-30 uM (Fig. 4.4C). Summer Si(OH)4 concentrations in the BGS discharge water averaged 31.0±11.6 uM, and Si:N was 3.7±2.3 (Table 4.6). During the summer of 2000, NO3 levels within the discharge effluent were significantly higher than those observed within the 1999 discharge (14.2±3.7 \iM; t=4.497, df=40, p<0.001; Table 4.6). Differences between 1999 and 2000 N 0 3 concentrations were most obvious during the mid-summer months, when 1999 levels were relatively low at 5-10 fxM, while 2000 NO3 discharge levels ranged from 10-20 u,M. As with NO3, summer PO4 concentrations were also significantly higher (1.9±0.3 u.M; t=2.757, df=33, p=0.009) within the cooling water during 2000 than 1999. The N 0 3 increase in the 2000 discharge however was greater than that measured for PO4, which resulted in significantly higher N 0 3 : P 0 4 ratios of 7.8+1.7 in 2000 than 1999 (p<0.05). During both 1999 and 2000, N03:P04 ratios were lower during summer than spring or autumn. The average deep water N03:P04 ratio during the course of this study was 8.0±2.1 (n=73). Si(OH) 4 concentrations in 2000 were similar to that discharged in 1999, with a summer average of 33.6±6.2 u,M. The Si:N ratio (2.5±0.7) in 2000 was significantly lower than during the summer of 1999 (Table 4.6). 141 25 J J A S O M A M J J A S 1999 2000 Fig. 4.4. Concentrations of: A) N 0 3 , B) P 0 4 , and C) Si(OH) 4 in the cooling water discharge of Burrard Generating Station from June-October, 1999 and March-September, 2000. 142 Table 4.6. BGS discharge rates and the nutrient concentrations and ratios within the cooling water discharge during 1999 and 2000. Student's t-tests were used to determine annual summer differences. Significance at p<0.05. ns- not significant. Summer is June 21 to September 21. Mean BGS discharge rates were calculated from daily discharge rates. Date Duration BGS discharge N O , PO, n (105 m 3 d"1) mean Si(OH) 4 (pM) N 0 3 : P 0 4 S i ( O H ) 4 ^ r 0 3 (uM) (uM) SD mean SD mean SD mean SD mean SD mean SD 1999 June-Oct. 30 4.5 2.7 11.3 4.4 1.8 Summer 21 5.4 2.4 9.3 3.4 1.6 0.4 32.2 10.2 6.8 1.9 3.3 2.0 0.4 31.0 11.6 6.4 2.2 3.7 2.3 2000 March-Sept. 43 9.2 4.0 15.8 4.1 1.8 0.3 33.8 8.7 8.7 1.9 2.2 0.6 Surnmer 21 11.8 2.9 14.2 3.7 1.9 0.3 33.6 6.2 7.8 1.7 2.5 0.7 t-test p<0.001 p<0.001 p=0.009 ns p=0.042 p=0.037* * normality failed; analyzed using Mann-Whitney Rank Sum test 143 4.3.3 Nutrient load discharged into PMA during BGS operation The nutrient load discharged into the surface waters of P M A is a function of the ambient nutrient concentration within the intake water and the daily discharge rate of BGS. It is therefore a measure of the total mass of a specific nutrient that is cycled from the deep-water of P M A into the surface water (i.e. above the pycnocline) during one day (24 h) of plant operation. The nutrient cycling capabilities of BGS (in terms of nutrient loading rates into the surface waters during BGS operation) were clearly distinguishable between years of high and low electrical generation (Fig. 4.5). Mean daily BGS discharge rates were significantly higher (Mann-Whitney test; T=234.5; p<0.001; Table 4.7) during the summer of 2000 (11.8xl0 5 ±2.9xl0 5 m 3 d"1) than during the summer of 1999 (5.4xl0 5 ±2.4xl0 5 m 3 d"1). Consequently, the levels of macronutrients (NO3, PO4, and Si(OH) 4) that were cycled into the surface waters of P M A were substantially higher in 2000 than during 1999 (Fig. 4.5). Overall, the mean NO3 load cycled daily into the summer P M A surface waters during 2000 was significantly different (t=8.690, df=40, p<0.001; Table 4.7) and approximately three to four times higher (1120±369 kg d"1) than that discharged in 1999 (319±206 kg d"1). This difference was mainly a result of the -220% increase in the mean daily BGS discharge rate during the summer of 2000, and to a lesser degree, the higher N 0 3 concentrations that were present within the 2000 discharge waters as compared to 1999 (Fig. 4.4A). The quantities of PO4 and Si(OH) 4 cycled daily through BGS to the surface waters of P M A followed a similar trend to those of N 0 3 , as each was significantly greater in 2000 144 2000 1600 1 M 1200 c « O 800 H o 400 H 0 1 1—•—i 1 1 1 r i 1 1 r 350 Fig. 4.5. Loading rates of: A) NO3, B) P 0 4 , and C) Si(OH) 4 discharged from Burrard Generating Station into the surface waters of Port Moody Arm from June-October, 1999 and March-September, 2000. 145 Table 4.7. Nutrient loading into the P M A surface waters from BGS cooling water discharge during the 1999-2000 sampling regime. Student's t-tests were used to determine annual summer differences. Significance is at p<0.05. Summer for 1999 and 2000 is June 21 to September 21. BGS discharge means are calculated from daily rates (June 21-Sept 21) while nutrient loads are calculated from nutrient concentrations and BGS discharge rates for sampled dates. Nutrient loads ejected into P M A during BGS operation BGS discharge N 0 3 P 0 4 Si(OH) 4 Date Duration n (105 m 3 d"1) (kgd"1) (kgd 1 ) (kgd"1) mean SD mean SD mean SD mean SD 1999 June-Oct. 30 4.5 2.7 343 192 84 34 1560 884 Summer 21 5.4 2.4 319 206 86 37 1650 998 2000 March-Sept. 43 9.2 4.0 1010 387 186 77 3400 1270 Summer 21 11.8 2.9 1120 369 227 50 4060 925 t- test p<0.001 p<0.001 p<0.001 p<0.001 146 than in 1999 (Figs. 4.5B, C; Table 4.7). Daily amounts of Si(OH) 4 cycled by BGS in 1999 (1,650±998 kg d"1) were less than half those discharged in 2000 (4,060+925 kg d"1). Monthly means for each daily macronutrient load during 2000 were significantly higher (p<0.05) than the corresponding months (June-September) in 1999. The influence of the discharge rates on the NO3 content within the receiving area of the discharge waters can be seen in Fig. 4.6. During low discharge events (Fig. 4.6A; August 19, 1999, BGS discharge 1.4xl0 4 m 3 d"1), elevated NO3 concentrations were not evident within the receiving core sampling grid (see Fig. 3.1). In contrast, medium discharge (Fig. 4.6B; August 3, 2000, BGS discharge 5.5xl0 5 m 3 d"1), and high discharge events (Fig. 4.6C; September 8, 2000; BGS discharge 1.5xl0 6 m 3 d"1) showed dramatic increases in surface NO3 at the point of discharge and >400 m beyond as waters flowed past the IOCO shipping pier (Site 13; Fig. 4.6). The evidence of surface nutrient loading is particularly apparent since each sampled date shown in Fig. 4.6 occurred during periods when P M A surface waters were N03-limited, which is characteristic of the summer season in P M A . 4.3.4 Comparison of nutrient inputs due to 'natural' entrainment and BGS operation 4.3.4.1 Pycnocline depths To determine whether subsurface nutrient entrainment into the surface layer was occurring during this study, the surface layer depths throughout P M A were approximated in order to calculate the average surface and bottom water temperatures required for the entrainment model (Eq. 4.6). 147 1 4 7 10 13 Nearshore Sites (West to East) Fig. 4.6. Vertical profiles of NO3 concentrations (u,M) along the nearshore discharge plume (see Fig. 3.1 for transect sites) during: A) low (1.4xl0 4 m 3 d"1; August 19, 1999), B) medium (5.5xl0 5 m 3 d"1; August 3, 2000), and high BGS discharge events (1.5xl0 6 m 3 d" ; September 8, 2000). The 'Discharge' at top of diagram refers to the location and direction of discharge. Site 4 is the immediate discharge area and Site 1 receives no discharge and acts as a control. Sites 7, 10, and 13 are directly downstream of the discharge outfall. Each profile was obtained during periods of NO3 limitation. The distance between Site 1 and Site 13 is 500 m. 148 Between June and September, pycnocline depths ranged from 3.6-6.9 m during 1999 and 3.2-10.2 m during 2000 (Fig. 4.7A, B). Average site-specific pycnocline depths showed very little variability between 1999 and 2000 and the intraseasonal variability at each site was low as well (Appendix F. l ) . Spatially, mean pycnocline depths during the summer of 1999 increased in a seaward direction, as mean depths at SI (6.3±0.8 m) and S2 (5.7±0.7) were significantly deeper (one-way A N O V A , df=2, F=8.335, p=0.001; Tukey's multiple comparison test, p<0.05)6 than at the central section (S3; 5.0±0.8 m). During the same period in 2000, mean pycnocline depths at SI (7.3+1.9 m) were slightly deeper than in 1999, and significantly deeper than at both S3 (5.3±1.0 m) and S4 (5.3±0.8 m; one-way A N O V A , df=3, F=8.752, p<0.001; Tukey's multiple comparison test, p<0.05). Pycnocline depths at S2 (6.2±1.1 m) were intermediate between the mouth (SI) and central regions (S3 and S4) during this time. 4.3.4.2 Average temperatures for surface and bottom water layers In 1999, surface temperatures peaked during mid-summer (19.4°C) before declining with the onset of fall (Fig. 4.7C). Bottom water temperatures exhibited the same trend, although the temperature increase (Fig. 4.7E) and variability (Appendix F. l ) was considerably lower than that in the surface waters. This led to temperature stratification (AT) that peaked in mid-summer before declining into autumn (Fig. 4.7G). On average, summer surface and bottom water temperatures were lowest at SI and successively higher towards the easternmost sites (Fig. 4.7). Similarly, AT was also progressively higher at the 6 Only S1-S3 were used in One-way A N O V A because of missing temperature data for S4 during 1999 149 1999 2000 "i r June July Aug Sept 12 10 8 6 4 B V v 12 15 14 13 12 11 F V 1 Site 1 • ••3 • o Site 2 - - • — Site 3 -v— Site 4 June July Aug Sept Fig. 4.7 Physical features of P M A during the summer of 1999 and 2000 that were used in the entrainment model (Section 4.2.4); pycnocline depth (A, B), average surface layer temperature (C, D), average bottom layer temperature (E, F), and temperature difference (AT) between surface and bottom layer (G, H). Symbol key in Fig. 4.7F applies to each graph. 150 east-most stations and was lower at the most western sites, ranging from 0.4°C at SI to 5.7°C at S3. During the summer of 2000, the average surface temperature peaked twice at 19.7°C (Fig. 4.7D) following prolonged periods of seasonally elevated atmospheric temperatures. The collapse of both temperature peaks coincided with periods of destratification brought about by spring tidal cycles and reductions to riverine inputs into P M A (Fig. 2.14C). Surface temperatures continued to decline into fall following atmospheric trends. As in 1999, surface temperatures in 2000 were highest at the eastern sites (S3 and S4) and coolest towards the mouth of P M A (SI). Bottom water temperatures increased gradually throughout the summer and peaked two weeks later in 2000 than in 1999 (Fig. 4.7F). Similar to 1999, the average temperatures increased eastward from the mouth towards the head of PMA. The greatest temperature stratification of the surface and bottom waters occurred early in the summer (June 29; AT=7.0°C) following the largest spring tide of the season, as well as the termination of the spring freshet. After this period, AT was generally <5°C at the easternmost sites and <2 °C near the entrance of P M A (Fig. 4.7H). Overall, surface layer depths and site-specific average temperatures in the surface and bottom layers of P M A were remarkably similar between the summers of 1999 and 2000 (Appendix E. l ) despite the intrinsic hydrodynamic differences between the years. 4.3.4.3 Estimates of site-specific surface layer outflow and estuarine entrainment Estimates for the seaward upper layer transport in P M A were calculated for specific sites using Eq. 4.6. The extent of entrainment within P M A was approximated as the 151 difference between the horizontal volume fluxes between the respective sites. Positive results indicate an expansion of the upper layer, and hence entrainment of subsurface water into the seaward flowing upper layer, while a negative result indicates the increased stratification in the outflowing water. As shown in Fig. 4.8, the surface volume transport, entrainment, and subsequent vertical nutrient flux in P M A is seasonally and annually variable. During the summer of 1999, seaward flow at each site (mean: <2.0xl0 6 m 3 d"1) was approximately three-fold lower than the mean flow at comparable sites during 2000 (Table 4.8). The largest discrepancy between the 1999 and 2000 seaward transport was primarily a result of the reduced outflow during the mid-summer in 1999 (Fig. 4.8A) and pronounced outflow observed during late August and September 2000 (Fig. 4.8B). The seaward transport during September 2000 was the highest of the study, and consistently exceeded 107 m 3 d"1 at the westernmost site (SI). Notwithstanding this seasonality, the surface volume transport in 2000 was still twice that seen in 1999 during mid-summer at all sites. Early in the summer of 1999, there were slight increases in horizontal transport as the surface waters moved westward out of P M A , but from mid-July to late August there was virtually no difference between the sites (Fig. 4.8A). Accordingly, entrainment rates between SI and S3 were exceptionally low during this period (Fig. 4.8C), exceeding l x l O 6 m 3 d"1 only during the commencement of the fall season (September 24, 1999). Notably, -60% of the vertical flux occurred at the mouth of the inlet (S1-S2), whereas entrainment was minimal between S3 and S4 (Table 4.8). In fact, the surface waters appeared to be more prone to stratification during seaward flow as 4 of the 7 dates had negative 'entrainment' rates (Fig. 4.8C). 152 1999 2000 c B I a — v © «5 w S_ 3 18 16 14 12 10 8 6 4 2 0 —•— Site 1 A o Site 2 — T — Site 3 v — Site 4 D c 7 - E n 8 = S t | g 3 = O J* -12 10 8 6 4 2 0 -2 15000 12000 •-c Z 9000 i | 6000 C 3000 B B J • 57.. — • S 2 t o S l O S3 to S2 - - • S3 to SI - V S4toS3 - E . : — • — N O 3 -: o P O 4 -— S i ( O H ) 4 • June July August September F . . . . . - 0 - - 0 ° - o June July August September V 30000 40000 20000 •§ u IOOOO «L o s o Date Fig 4.8. Volume of seaward flowing surface water at each respective site within Port Moody Arm during the summer of: A) 1999, and B) 2000; vertical entrainment volumes between sites during the summer of C) 1999, and D) 2000; and the vertical flux of NO3, PO4, and Si(OH) 4 into the surface waters of P M A during the summers of E) 1999, and F) 2000. Surface entrainment rates are the difference between outflowing volumes at respective sites. Vertical nutrient fluxes were calculated as the product of the entrainment rates and the respective bottom water nutrient concentration. 153 Table 4.8. Summary of summer outflow surface volumes and associated vertical entrainment rates during 1999 and 2000 and contribution of BGS and natural entrainment to the total P M A surface inputs. Note: Total input is the sum of the BGS discharge volume and the entrainment volume between sites 1 and 3. Date Outflow Volume (10 m d ) 5 3 - 1 Entrainment Volume (10 m d ) Site Sites 4* 1-2 2-3 3-4 1-3 1-4* BGS Total BGS:Total Discharge (10 5 m 3 d _ 1 ) ( 1 0 5 m V ) (%) 1999 mean (n=12) SD 18.7 13.5 15.1 9.6 13.6 14.0 12.6 3.5 4.7 1.6 2.2 0.3 3.3 5.1 5.3 5.6 5.7 5.8 2.1 10.9 6.8 64.4 23.1 Summer mean (n=10) SD 14.6 10.8 12.5 9.1 10.7 7.5 7.8 8.4 2.2 1.8 1.7 2.0 0.9 1.7 3.9 3.6 4.7 6.9 5.4 2.6 9.3 5.6 66.5 18.1 2000 mean (n=14) SD 62.7 43.5 46.8 34.6 34.3 16.3 34.5 15.9 15.9 15.9 12.6 22.5 -0.3 5.8 28.5 30.3 28.2 30.1 11.6 2.5 40.0 31.9 43.7 28.4 Summer mean (n=ll) SD 39.0 18.0 30.0 11.2 25.1 8.7 27.4 12.2 9.1 7.1 4.9 3.8 -2.4 4.1 14.0 10.6 11.6 9.2 10.9 2.7 24.8 11.7 52.9 29.5 * - Site 4 during 1999 (n=7); Summer 1999 (n=6) a - average BGS discharge calculated from days actually sampled. Sample size is same as listed in column 154 In contrast to 1999, average entrainment rates between SI and S3 were nearly 6-fold higher during the summer of 2000 (Fig. 4.8D; t-test, df=23, t=2.746, p=0.012). More importantly, during the critical period when flagellate species dominated the phytoplankton assemblages in 1999 (June 30 - August 19), the entrainment rates were 3-4 times higher over the same time frame in 2000. These rates were essentially stable from June to late August, but a dramatic increase occurred as autumn approached (Fig. 4.8D). As in 1999, the water between Sites 3 and 4 alternated between periods of increased stratification and entrainment to the upper layer (Fig. 4.8D). On average, this section of P M A appears to be a net source of stratification (Table 4.8). Conversely, the dominant source of entrainment in 2000 occurred in the narrows of P M A between Sites 1 and 2, which produced 65.0±23.1% of the total entrainment in P M A measured by the model. This effect becomes even more apparent considering that the distance between SI and S2 (0.9 km) is approximately 25% of that between S2 and S4 (3.4 km). 4.3.5 Relationship between vertical nutrient entrainment and direct BGS enrichment The vertical flux of nutrients into the surface layer of P M A was substantially lower in 1999 than 2000 for two reasons: 1) reduced entrainment rates observed in 1999 compared to 2000 (Figs. 4.8C-F) and, 2) lower NO3 concentrations in the bottom waters during the summer of 1999 relative to 2000 (Section 4.3.2). On average, during June to September 1999, 0.4±0.5 tonnes N 0 3 d' 1, 0.09±0.10 tonnes P 0 4 d"1, and 1.5+1.7 tonnes Si(OH) 4 d"1 were mixed into the surface layer compared to 2.7±3.1 tonnes NO3 d"1, 0.5±0.6 tonnes P 0 4 d"1, and 8.5±9.2 tonnes Si(OH) 4 d"1 during 2000. During the important period of the diatom/flagellate transition (June 30-August 19), the vertical flux of nutrients was still 4-5 155 times higher in 2000 than during 1999. During the summer of 2000, there were indications the entrainment rates were influenced by spring-neap tidal oscillations as three successive peaks in mixing coincided with the spring tidal cycle during June and early July. The largest early summer peak (July 6) coincided with the largest new moon spring tide of 2000 (Fig. 4.8F) and reduction of the spring freshet. These tidal peaks were not evident during the summer of 1999 despite identical fortnightly tidal trends. The decline in summer stratification in 2000, and the associated increase in vertical mixing, resulted in tremendous nutrient inputs during September reaching a peak of over 11 tonnes of NO3 and 33 tonnes of Si(OH) 4 on September 8 (Fig. 4.8F). The low levels of entrainment observed during 1999 were supplemented with low nutrient contributions from BGS due to low discharge rates and the low seasonal nutrient concentrations within the cooling water. Despite the reduction in the power plant operation, BGS was still the dominant source of nutrients for the surface waters of P M A from June to September 1999 (64±23%; Fig. 4.9A). In 2000, BGS operation accounted for 30-99% of the surface nutrients into P M A during early and mid-summer (Fig. 4.9A) before the contribution of BGS was overwhelmed by the substantial volumes of water mixed into the outgoing waters as stratification levels declined into late summer and early fall (Fig. 4.8F). Between the in situ and anthropogenic sources, BGS contributed an estimated 44±28% of nutrients delivered into the surface waters of P M A during 2000, but was 53±30% when only the summer season (June 21 and September 21) was considered (Table 4.8). During 1999, P M A was dominated by various flagellate species over a two month period between late June and late August. During this time, BGS and in situ mixing processes supplied 9.3xl0 5 ±5.6xl0 5 m 3 d"1 of nutrient-enriched water into the surface 156 C5 s-s c 13 © o u c .22 °E s 120 100 80 60 -| 40 20 0 A • / V \ . L / > \ \ \ /jk <N—1 • • 1999 2000 • • • % 6000 O Z 4000 June July August September Month Fig. 4.9. A) The relative contribution of BGS nutrient inputs compared to the total nutrient flux (BGS + 'natural' entrainment) for 1999 and 2000, and B) the total N 0 3 load contributed to Port Moody Arm from BGS and entrainment (PMA) sources during the summer of 1999 and 2000, respectively. 157 layer. By comparison, BGS alone discharged I l x l 0 5 ± 2 . 7 x l 0 5 m 3 d"1 within this time frame during 2000, while in situ mixing added another 1 4 x l 0 5 ± l l x l 0 5 m 3 d"1. Therefore, BGS and natural mixing events individually contributed higher nutrient loads to P M A surface waters during 2000 than the combination of the two sources did in 1999, and when combined, the two sources nearly tripled the surface nutrient fluxes of 1999 (Fig. 4.9B). 158 4.4 Discussion 4.4.1 Relationship between the summer phytoplankton community composition and environmental factors Estuarine phytoplankton are exposed to a variety of biotic, abiotic, and anthropogenic factors that constrain growth, primary production, biomass, and community composition (Cloern 1996; Cloern and Dufford 2005). In P M A , the most important environmental gradients responsible for partitioning the summer phytoplankton composition appear to be vertical stratification, the nutrient content of the surface layer, as well as the operation of the thermoelectric power plant, BGS. Overall, summer assemblages were structured such that flagellate species, particularly the raphidophyte Heterosigma akashiwo and the dinoflagellate Prorocentrum gracile, were dominant during the highly stratified N03-limited period in 1999 when BGS operated at less than half capacity. Diatoms, specifically Skeletonema costatum, were the principal component of the summer phytoplankton community during 2000 when vertical stability was moderate, BGS operated at near capacity, and surface NO3 concentrations were higher than in 1999. Due to their high growth rates, diatoms are typically the dominant phytoplankton in nutrient-replete temperate systems, whereas flagellate species tend to inhabit stable nutrient-limited waters by exploiting their ability to migrate freely between the surface and subsurface layers (Margalef 1978). In BC coastal waters, diatoms are normally replaced by various flagellate species from late June/early July into August, which coincides with peak levels of stratification due to the spring freshet and surface NO3 exhaustion following the spring bloom (Harrison et al. 1983; Harrison et al. 1994). Flagellate dominance during this period has been reported previously for P M A (Stockner and Cliff 1979) as well as nearby 159 Sechelt Inlet (Haigh et al. 1992), and sections of the Strait of Georgia (Stockner et al. 1979; Haigh and Taylor 1991). Despite this evidence, summer diatom blooms can occasionally occur in the Georgia Basin when pulses of nutrient-rich water are supplied to the surface layer (Takahashi et al. 1977: Parsons et al. 1983; McQuoid and Hobson 1997). During this study, surface NO3 concentrations were at levels considered to be limiting to. phytoplankton growth (<2 u.M) throughout each summer season (Appendix C; Eppley et al. 1969). Consequently, there must have been a continuous, or semi-continuous, elevated supply of nutrients to the surface layer of P M A to sustain the high diatom growth observed during summer 2000 relative to that in 1999. Because the summer season in coastal BC is characterised by low precipitation, and the watershed surrounding P M A is generally pristine, with only intermittent summer freshwater inputs, the major sources of nutrients supplied to the surface waters are expected to be: 1) vertical mixing from estuarine circulation and/or tidal activity, and 2) the upward cycling of bottom water by direct BGS cooling water discharges into the surface layer of P M A . 4.4.2 Sources of nutrients for summer surface waters: BGS operation and estuarine circulation The combined nutrient contribution from direct BGS discharges and vertical mixing was insufficient to support diatom growth between July and August 1999. This was primarily related to the pronounced snowmelt following the winter of 1998-99 as river discharges in the vicinity of the Fraser Delta were far above seasonal means and approached historical maxima through late June to September (see Fig. 2.6). This would have suppressed the delivery of nutrients into the surface layer in three ways. First, due to the long-lasting freshet in 1999, surface stratification in P M A was strong and sustained 160 throughout the summer period, and accordingly, mixing rates and vertical nutrient fluxes were the lowest of the study. Since coastal run-off in BC is typically nutrient-poor during the summer (Harrison et al. 1994; Kiffney and Bull 2000), this created a stable nutrient-poor upper layer where nutrient replenishment was severely inhibited; an environment that was optimal for flagellate growth. Secondly, BGS is predominantly used as a supplementary electrical source in BC (-12% of the provincial requirement) and normally operates during periods of reduced hydroelectric production (BC Hydro 2005). Thus, the peak period of operation generally occurs in the summer months following the spring freshet and preceding the winter rains (i.e. when run-off is low). Because riverine discharge in the BC region reached near historic highs during 1999, the electrical needs of BC were largely fulfilled through hydroelectric means. As a result, BGS operated at <50% capacity during this time and its direct nutrient input into P M A was correspondingly low. Despite the decreased operation, BGS remained the largest source of surface nutrients to P M A over the summer period (-65%), indicating the importance of this point source input. Finally, the stronger stratification in 1999 also likely contributed to the reduced NO3 and PO4 concentrations found in the bottom waters (which would influence the nutrient flux into the surface layer), and is related to the positive estuarine flow in P M A and its downstream topography. During spring tides, the flow at the Second Narrows in Burrard Inlet just west of P M A reaches supercritical velocities as stratified flows pass over a section of complex topography, which includes a lateral constriction, a sill, and a deep hole just east of the Narrows (Isachsen and Pond 2000). The result is that the surface and bottom waters are mixed throughout the water column in this area. Therefore, during periods of sustained nutrient limitation, the continuous incorporation of nutrient depleted surface waters into the 161 landward flowing bottom waters (i.e. towards PMA) leads to progressively lower nutrient levels in the bottom waters throughout the summer. Since the surface NO3 concentrations were severely depleted due to phytoplankton uptake and low vertical replenishment in 1999 (resulting from the intense stratification), the average bottom water NO3 concentrations (~9 \iM) were -35% lower during this period than the summer of 2000 (-14 \iM). Consequently, the nutrient inputs delivered to the surface layer of P M A from BGS and mixing processes would have reflected this decrease. It is also probable that the preponderance of flagellate species during 1999 contributed to the reduced bottom water NO3 and PO4 concentrations due to their deep water nutrient retrieval mechanisms. 4.4.3 Influence of vertical mixing in P M A during 2000 During the summer of 2000, 'natural' mixing processes and BGS operation each contributed higher nutrient fluxes into the surface waters of P M A than the combined sources of 1999. It is not clear whether vertical entrainment rates alone could have supported the diatom growth observed in 2000, however due to their high nutrient affinities (low Ks values; Smayda 1997), particularly for NO3, diatoms are thought to have a selective advantage in low nutrient waters provided that Si(OH) 4 concentrations are adequate (Reigman 1995). During 2000, stratification was low and nutrients entrained into the surface layer were 4-5 times higher than in 1999, and so it is possible that mixing processes alone could have supported summer diatom populations. However, precipitation and riverine discharge in the Georgia Basin during summer 2000 were similar to the 10-year average (see Fig. 2.6). This leads to the assumption that the climatic and hydrodynamic conditions during this period were characteristic of P M A , and under such conditions, surface nutrient limitation and summer flagellate dominance have been previously reported 162 for P M A (Stockner and Cliff 1979). Furthermore, the vast majority of the estimated total mixing volume within P M A (65%) occurred over a very short distance (~1 km) between SI and S2 at the constriction at the mouth of P M A . Therefore, this potential nutrient source would have only been available to phytoplankton over short spatial scales, except possibly during large spring tides. In fact, much of the horizontal length of the central channel was a net source of stratification during 2000 (Table 4.8) as the tidally mixed waters passed over the shoaling ledge at S4 and restratified as the surface waters move towards the mouth of P M A . Hence the mixing factors in 2000 may not have been able to support summer diatom growth over much of the central channel. One of the weaknesses of the model presented here is its inability to identify the source of mixing within P M A or to distinguish between vertical mixing and horizontal advective processes. Tidal flows dominate in P M A (Waldichuk 1965; Taylor and Fissel 1999), ranging between 20 and 30 cm s"1 during neap and spring tides, respectively, and tidal velocities are 3-4 times higher during the passage through the narrows at SI than in the eastern sections of P M A (Seaconsult 1995). In contrast, residual surface and bottom water flow rates due to positive estuarine circulation are extremely slow at ~1 cm s"1 when BGS is not in operation (Jiang et al. 2002; 2003). Therefore, it appears that tidal flow drives the substantial mixing that occurs in P M A , particularly when BGS is not in operation, and that a majority of the mixing is confined to the western section of the inlet. However, the daily tidal prism in P M A represents -33% of the total P M A volume (Taylor et al. 2001), and the tidal excursions into P M A range between 1.5 km during neap tides and 3.5 km during new moon spring tides. During flood tides, these intrusions bring colder and more brackish water from Burrard Inlet into P M A giving the inlet its estuarine character (Waldichuk 1965). Thus, the colder waters at the mouth of P M A used to calculate outflowing surface volumes 163 in the 2D model may have resulted from horizontal advection and mixing events involving tidally advected Burrard Inlet waters and seaward flowing P M A waters instead of from intrinsic vertical mixing of cold nutrient-rich bottom water into the surface layer through tidal mixing or from estuarine two-layer flow. As a result, outflowing volumes at the mouth of P M A and, in turn, mixing volumes calculated using SI and to a lesser extent S2, may have been overestimated by the model. Since the surface waters of Burrard Inlet were similarly NOvlimited (Figure 2.7), any advection or horizontal mixing would have yielded little additional surface nutrients to the system for diatom growth. At the very least, the potential for vertical mixing at the mouth of P M A would contribute nutrients over a limited spatial scale within the inlet, which could restrict their availability to much of the phytoplankton community in the central and eastern portion of P M A . The overestimation of vertical mixing volumes in P M A , on the other hand would indicate the growing importance of BGS's contribution as the prominent nutrient source that facilitated summer diatom dominance during the summer of 2000. 4.4.4 Influence of BGS on the P M A ecosystem during 2000 When aquatic systems are supplemented with anthropogenically-derived nutrients, the greatest potential for eutrophication or community modification occurs during periods when primary producers are limited by a specific nutrient (Mackas and Harrison 1997). In order for these inputs to be transferred into autotrophic biomass, the supply of nutrients must be freely accessible to phytoplankton stocks within the surface light field. In this regard, one of the most important characteristics of the BGS is its ability to alter the conservative properties of the intake water during transit through the condenser system. As the cooling water temperatures increase by ~10°C during passage, the thermally-elevated 164 effluent is discharged as a nutrient-laden buoyant jet, and because of its central position in P M A , this discharge is traceable over much of the central channel as flood and ebb tides move the plume back and forth over the length of P M A (Hodgins and Webb 1991; Seaconsult 1995; Jiang et al. 2002; 2003). Consequently, the total nutrient load produced by BGS operation covers a large surface area within P M A and is capable of being entirely accessed by the primary producers, as the surface layer is completely within the photic zone. This phenomenon is particularly important if the exogenous input of nutrients is going to be integrated into diatom production due to their lack of motility. Average BGS discharge rates during the summer of 2000 were nearly three times those of 1999 and were -75% of maximum allowable limits (1.7xl0 6 m 3 d"!). This input was substantial as the summer discharge rates were comparable to the flow rates from the two major river systems (Capilano and Seymour Rivers) that empty into the much larger Burrard Inlet (-100 km2) just west of P M A . Overall, this discharge contributed >50% of the estimated nutrient flux into the surface layer of P M A during the summer, and average daily loading rates were -1 tonne of NO3 and -3 tonnes of Si(OH) 4, respectively. This nutrient load is considerable, but its ecological relevance is realized mainly because it is discharged into the small surface area and maintained within the shallow surface layer present in P M A , therefore maximizing its loading potential. Based on a surface area of 4.8 km 2 (at mean tidal height; Waldichuk 1965) and a surface depth of 5 m, the surface layer volume of P M A is 7 3 estimated at 2.4x10 m . Assuming all BGS discharge is incorporated into the P M A surface water, a mean summer effluent NO3 concentration of 14 u,M and average summer discharge 6 3 1 rate of 1.3x10° m J d" could raise ambient surface N O 3 concentrations in P M A by -0.7 \iM during the summer. Operating at full capacity, these levels could increase nearly 1 uM. By further taking into account the mixing volume estimates during 2000, the total N O 3 165 delivered to the summer surface waters would be of order 1.5-1.7 u,M, which may have been sufficient to alleviate many of the symptoms associated with the nutrient limitation present during 2000. The capacity to compete in low nutrient environments is reflected in a species' ability to sequester and assimilate the available nutrients for growth. Furthermore, the ability of an algal species to outcompete its competitors is tied to its ability to reduce ambient nutrient concentrations to levels where the growth-loss equilibrium for the competitor becomes unfavourable (Tilman et al. 1982). In marine phytoplankton, the affinity for a particular nutrient is reflected in the half-saturation constant ( K S ) , where species with low K S values are competitively superior at reduced nutrient concentrations (i.e. high affinity) than those species with higher K S . In general, diatoms have higher (in many cases much higher) affinities for low NO3 concentrations than do flagellates (Eppley et al. 1969; Smayda 1997; Lomas and Glibert 2000). During 2000, the N 0 3 supplied to P M A via BGS discharges was well within the K s (NO3) of many marine diatoms (see Table 3 in Smayda 1997), which is further extended when the 2000 mixing estimates are included. The total NO3 delivered in 1999, in contrast, would only amount to an increase of 0.3 u,M ft ^ -1 (based on bottom water N O 3 concentration of 9 [xM and 9.3x10 m d" combined BGS and mixing volume), which is at the lowest end of diatom affinities. Hence, because of the low surface nutrient fluxes in 1999, surface N O 3 levels could not satisfy diatom species with even the most efficient uptake systems. Flagellates, on the other hand, could exploit their ability to migrate into the bottom waters where their low N O 3 affinities are suited to the high nutrient concentrations found there, and therefore became the primary autotrophs in the inlet. 166 In contrast, Skeletonema costatum, the predominant phytoplankter in P M A during the summer of 2000, has one of the most efficient NO3 uptake systems among marine algae (K s (N03) = 0.4-0.5; Eppley et al. 1969; Lomas and Glibert 2000), and its growth is maximized at high irradiance and low NO3 concentrations (Eppley et al. 1969). Given the nutrient loads contributed by BGS into P M A during the summer of 2000, the high NO3 affinity of S. costatum, and the favourable light habitat of P M A , it is possible that the BGS discharges alone could have sustained these diatom populations through nutrient-limited periods when operating near capacity. However, it is more likely that reduced stratification and elevated mixing rates in 2000 would have raised the surface NO3 concentrations near the threshold required to promote diatom growth (i.e. 0.5-1.0 u,M), and the additional contribution of BGS discharges increased the surface NO3 concentrations in P M A to levels that could sustain diatom dominance (1-2 \iM). The continuous elevated operation of BGS throughout the 2000 summer season, coupled with moderate mixing rates, created a steady state low-nutrient environment in P M A that effectively led to the temporal extension of the spring diatom bloom in which S. costatum is typically the last successional diatom in BC waters (Harrison et al. 1983). Reigman (1995) has shown that as long as surface waters are in moderately nutrient-limited steady state conditions (as opposed to pulsed delivery), smaller algae should outcompete larger species and, therefore, dominate these environments. Due to its small size, high affinity for low N O 3 concentrations (Lomas and Glibert 2000), and high growth rates at high irradiances (Eppley et al. 1969), S. costatum is well adapted to the environment created by the reduced stratification in P M A and elevated BGS operation during the summer of 2000. Accordingly, it was the dominant phytoplankter during this period. 167 Thus far I have only discussed the direct nutrient inputs contributed to P M A from the operation of BGS. In addition to the direct nutrient input, there two other major features where BGS likely contributed substantially to the nutrient budget of P M A : 1) the near-field entrainment of bottom water into the discharge jet as it rises to the surface, and 2) the subsequent increase in vertical entrainment from increased two-layer estuarine flow due to the surficial BGS discharge. In modelling the circulation patterns of the BGS discharge jet in P M A , Jiang et al. (2003) noted that there was indeed very strong entrainment of subsurface ambient water during the rising stage of the cooling water effluent. Quantification of this nutrient source was not attempted by the authors (J. Jiang and D. Fissel, pers. comm.), however, Jahn et al. (1998) stated that the discharge jet from a Californian nuclear power plant was capable of incorporating >10 times the discharge volume from the nutrient-rich subsurface zone, indicating the importance of this nutrient source. Jiang et al. (2003) also found that BGS enhances the water flow in both the seaward flowing surface water and the landward flowing bottom water. Their model results showed that BGS operation imparts the strongest control on the positive estuarine circulation in P M A as predicted seaward flows reached -10 cm s"1 during maximum operation, which was 5-fold higher than that predicted for periods of non-operation (<2 cm s"1). In stratified flows, the ability to mix subsurface water into surface flow is governed by the ratio of the stabilizing forces of density stratification compared to the destabilizing forces of velocity shear due to the velocity differential between the upper and lower layers, and is parameterized as the Richardson number, Ri (Dyer 1997). Since BGS ejects thermally elevated saline water into the surface layer, which results in an increase in the velocity differential between the two layers, the down-inlet vertical entrainment, and hence vertical nutrient flux into the surface layer should increase with respect to BGS operation. In fact, it 168 is common in estuarine systems that seaward transport and nutrient flux into the surface waters due to discharge jets and riverine flow is often several times greater than the actual riverine discharge or effluent jet driving the system (Lewis 1997; Mackas and Harrison 1997; Jahn et al. 1998). Therefore, it is very likely that the dynamic flow of the BGS discharge jet may actually exceed the nutrients contributed directly to the surface waters of P M A from the BGS effluent by several fold. For this reason, when determining the influence of the power plant operation on the nutrient budgets of coastal systems, there are three main factors that require consideration: 1) the direct nutrient inputs within the discharge jet, 2) the near-field entrainment of bottom water into the discharge plume, and 3) the subsequent vertical entrainment due to enhanced estuarine circulation. 4.4.5 Atmospheric and riverine inputs Nutrient inputs from riverine and atmospheric sources are considered to be among the greatest contributors to the cultural nutrification in coastal ecosystems (Peierls et al. 1991; Paerl et al. 2002a, b). These two nutrient sources were not considered within the 2D-model (Eq. 4.6), which was used to determine the dominant source in P M A that contributed to the summer phytoplankton assemblages. From a riverine perspective, direct fresh water inputs into P M A were not measured and discharge data were not available from any governmental, public or private source. Nonetheless, Waldichuk (1965) reported that summer run-off into P M A is intermittent, and the estuarine nature of P M A is largely due to the tidal intrusions of Indian Arm/Burrard Inlet water into the inlet. Even if the exceptional snowmelt of 1999 contributed inflow into P M A , it is likely that this water would have contributed little in the way of nutrients since freshet waters in the BC lower mainland are typically nutrient-poor in nature (Harrison et al. 1994). The fact that flagellates species were 169 the main phytoplankton constituents during the maximum run-off period of 1999 supports this suggestion. Therefore, it is highly unlikely that river inputs could have contributed significantly to diatom growth during the summer of 2000 due to the lower summer discharge rates coupled with the nutrient-poor nature of the freshwater input. It has also become clear in recent years that atmospheric deposition (wet and dryfall), is among the fastest growing sources of nutrient additions to the coastal zone (Paerl 1993; 1995; 1997; Seitzinger and Sanders 1999). This nutrient source is capable of stimulating surface primary production when supplied to a nutrient-limited system (Aguilar et al. 1999; Paerl et al. 1999), and in some cases, may contribute >50% of the new nitrogen delivered to surface waters (see Table 1 in Paerl 1997) and account for >10% of the annual new production in marine waters (Paerl 1997). It is doubtful any atmospheric input contributed significantly to the P M A nutrient budget despite the high annual precipitation within the region (ca. 1500-2200 mm between 1998 and 2000) and potentially high N concentrations within the rainfall (Mackas and Harrison 1997). Because P M A has a maritime west coast climate, rainfall is seasonal and is characterised by high winter and low summer precipitation. Consequently, summer rainfall varied between -100-300 mm during the course of this study being highest during 1999 and lowest during 2000. Considering an upper N concentration of 20 u,M for the wet deposition into P M A (from Mackas and Harrison 1997), and the maximum summer precipitation rate of 3.3 mm d"1 (based on precipitation accumulation between June 1-August 31, 1999; -300 mm), the atmospheric contribution to P M A would amount to only 5-6 kg N d"1. Based on an upper water depth of 5 m, this would increase daily N concentrations in the upper layer of P M A by only 1 nM, an inconsequential amount for phytoplankton growth. 170 There are also other nutrient sources not considered in this study. Groundwater inputs have been acknowledged as significant contributors to coastal budgets (Paerl 1997), however, much like wet deposition, 2000 levels would have been much less than in 1999 due to a -70% decrease in summer precipitation. Sewage in the area was diverted to Annacis Island in 1974, and discharge from combined sewer overflows occurs far upstream in Burrard Inlet. Moreover, any contribution from this source is proportional to precipitation levels (Hall et al. 1998), and therefore maximal inputs occur during the winter season when low irradiance levels limit phytoplankton production. Although all of the sources mentioned in this section undoubtedly contribute some nutrient load to P M A , each would have had the highest contribution during 1999 when flagellate species and abundant Si(OH)4 concentrations were present within the inlet. Furthermore, I have attempted to show that the combined total of each of these sources is negligible in comparison to that delivered by BGS cooling water discharge and that mixed into P M A through estuarine circulation and tidal mixing. In this regard, when monitoring P M A during the summer season, fluctuations in phytoplankton dynamics should be first addressed through BGS operation, tidal mixing, and to a lesser degree, estuarine circulation. 4.4.6 Ammonium (NH4) and the Role of Grazing The largest supply of nitrogen not accounted for in this study is the absence of N H 4 data. N H 4 samples were collected during three dates in March 1999 and concentrations were generally <0.5 uM in the upper 5 m and 0.5-1.0 uM at 10 m depth (Appendix B). It is difficult to speculate how this nitrogen source could have contributed to the phytoplankton growth and community dynamics in P M A , particularly considering its potential contribution to the summer phytoplankton populations. No additional N H 4 data has been published for 171 P M A , however NH4 concentrations in the Georgia Basin are typically related to the presence of elevated zooplankton grazing activity (Harrison et al. 1983; Bornhold 2000). Stockner and Cliff (1979) found that zooplankton biomass increased rapidly in P M A following the spring diatom bloom with peak zooplankton biomass being attained in May and June. Thereafter, they suggest that jellyfish and hydroid medusae maintained the zooplankton biomass at low levels. Assuming elevated zooplankton biomass represents elevated grazing rates and N H 4 excretion by the zooplankton community, it would appear that the highest potential for NH4 would occur over a two-month period following the spring diatom bloom. Therefore, the largest NH4 contribution would take place before the summer period when the potential for diatom-flagellate shifts would be most likely to occur (based on the findings of this study). Consequently, I can only speculate that it was unlikely that N H 4 was the dominant nutrient contributing to the continuation of diatom dominance through the summer of 2000. 4.4.7 Ecological implications of BGS operation The largest current pollution threat to coastal areas is the elevated input of nutrients (Howarth et al. 2002). The delivery of enriched nutrient levels and altered nutrient ratios into the coastal zone has led to the eutrophication and/or changes in community composition in several major areas around the globe (e.g. Mee 1992; Sin and Chau 1992; Elmgren 2001; Rabalais et al. 2002a; Druon et al. 2004). The increase in organic production and its export has led to bottom water hypoxia or anoxia (Rabalais et al. 2002b; Hagy et al. 2004), and the associated alteration of benthic communities and losses of important fisheries (Diaz and Rosenberg 1995), while nutrient-related community shifts have produced greater incidences of toxic or otherwise harmful algal blooms in recent years (Anderson et al. 172 2002). The consequence of these changes are far reaching as they are transferred throughout the food web (Nixon and Buckley 2002; Sommer et al. 2002; Kvitek and Bretz 2004; 2005) as well as potentially affecting human and other marine health (Van Dolah et al. 2001; Friedman and Levin 2005), and contribute to recreational and economic losses (Hoagland et al. 2002; Segerson and Walker 2002). The key features of BGS with respect to the primary producers within P M A are related to its discharge rate and the buoyancy of the nutrient-laden effluent. The cycling of Si(OH) 4 into the surface layer of P M A concurrent with the N and P inputs is critically important in terms of diatom production. In this regard, BGS is unlike other anthropogenic inputs as BGS maintains the stoichiometric nutrient ratios that are inherent to P M A . Studies have shown that diatoms favour high Si:N environments (Sommer 1989), and flagellate-dominated communities become increasingly common when Si:N ratios fall below 1:1 (Dortch et al. 2001; Turner 2002). This can lead to decreases in mesozooplankton abundance and particulate organic export flux (Turner et al. 1998). By discharging deep water with Si:N ratios of -3:1 into the upper layer, Si(OH)4 concentrations in P M A are always sufficient to support diatom growth during elevated BGS operation (as well as during mixing events). Since bottom water N:P is -8:1, the supply of N will control the potential of new primary production in P M A and hence its community composition. The greatest reason for ecological concern during this study was the presence of the potentially ichthyotoxic flagellate Heterosigma akashiwo, a H A B species has been responsible for multi-million dollar losses to farmed salmon in BC (Taylor 1993). The three largest blooms in P M A occurred during 1999 when BGS operation and mixing rates were the lowest. During this period, H. akashiwo was overwhelmingly the largest contributor to the high average summer biomass levels. In contrast, only two small, short-lived blooms 173 occurred during 2000. It has been suggested that a precondition for the blooming of H. akashiwo in BC waters is the collapse of the spring diatom bloom, strong vertical stability, and surface N 0 3 depletion (Taylor and Haigh 1993; Taylor and Harrison 2002). Thus BGS operation should fundamentally decrease the frequency and size of H. akashiwo blooms by supplying the surface layer with the required N O 3 levels to sustain the spring diatom bloom throughout the summer season. If blooms did form during peak BGS operation, then they should be advected from the inlet more rapidly due to the enhanced surface seaward transport. Outside of their potential ichthyotoxic effects, the decline of H. akashiwo blooms in P M A may also impart food web effects within the inlet as this H A B species is a poor food source to secondary producers (Tomas and Deason 1981; Verity and Stoecker 1982; Verity 1987; Taniguchi and Takeda 1988; Kamiyama 1995), in contrast to diatoms which are generally thought to be effective energetic vectors to higher trophic levels (Parsons et al. 1984b; Ware 2000; although see Ban et al. 1997; Miralto et al. 1999; Irigoien et al. 2002). While at first glace it appears summer diatom dominance is preferential to the presence of H. akashiwo in P M A , high concentrations of the pennate diatom Pseudo-nitzschia spp. have also been observed in the summer assemblages. Given the evidence that BGS operation has the potential to increase summer diatom production, it would be prudent to monitor these potentially neurotoxic diatoms during summers when BGS operation is near capacity since increased nutrient inputs in other coastal systems have led to blooms of these potentially toxic diatoms (Pan and Subba Rao 1997). In Chapter 3, it was shown that BGS operation has the ability to act as an indiscriminate grazer by drawing phytoplankton through its intake system. On average, 2 u.g chl a 1"', or -50% of the entrained phytoplankton was destroyed during passage through the 174 condenser system (Table 3.1). During maximum operation, this should have destroyed 2 x l O u fig C d"1 (based on C:Chl a of 50). Assuming a BGS NO3 discharge concentration of 14 \iM, and a C:N of 6.6 in phytoplankton (Parsons et al. 1984b), the potential 'new' production due to maximum BGS discharge could be ~2x l0 1 2 u.g C d"1, if the N is entirely incorporated into algal biomass. This is an order of magnitude higher than that destroyed through intake entrainment. In this respect, BGS is likely a net source of production into P M A as opposed to contributing to any declines in production due to intake entrainment. Importantly, any additional production of diatom biomass did not appear to contribute to the eutrophication of the inlet. Measurements of primary production were not made during this study due to the high sampling frequency and the large number of other parameters that were measured. Whether the operation of BGS, by definition, had any eutrophying effect in P M A cannot be directly addressed. However, the average autotrophic biomass decreased substantially during the summer of 2000 (8.9±3.7 u.g chl a l"1) compared to 1999 (18.2±18.2 u,g chl a l"1) when BGS operation was the highest. This strongly suggests that BGS does not significantly enhance eutrophication in P M A . The biomass of the 2000 diatom community was also less than the peak biomass levels observed during either the spring or fall blooms (15-20 \ig chl a l"1), indicating that the nutritional requirements of the phytoplankton standing stocks were not being met and/or summer grazing was high. The important features of P M A with respect to its ability to modulate the effects of eutrophication, are the absence of a sill at the mouth of the inlet, which is common in many mainland BC inlets and fjord systems (Pickard 1961), and the strong hydrodynamic forcing within the inlet. Consequently, any enhanced diatom production, export, and the potential for oxygen depletion, is successfully moderated by the large twice daily tidal prism that 175 moves into P M A as well as the enhanced estuarine circulation caused by increased BGS operation. The result is the removal of a substantial portion of the biomass into Burrard Inlet on ebb tides and bottom water oxygen replenishment on flood tides. Furthermore, as BGS operation approaches its maximum, the seaward surface transport in P M A also increases 4-5-fold (Jiang et al. 2003), increasing the horizontal flux of diatom biomass out of the inlet before it settles into the subsurface layer, while highly oxygenated deep water is freely moved into the Arm. Together, these two features appear to mitigate any increase in primary production caused by BGS operation as well as the potential for deep water anoxia due to increased vertical export. The results of this chapter provide evidence of the importance of the operation of BGS within P M A during the summer season. By discharging nutrients from the bottom into the surface waters at stoichiometrically relevant ratios, the operation of BGS is analogous to an upwelling system in the coastal ocean. In such systems, diatom production is typically high and production levels are controlled by a limiting nutrient (Parsons et al. 1984b). In P M A , this nutrient is NO3 and thus the ability of BGS to provide NO3 during periods of NO3 limitation during the summer period should control a substantial portion of the new production in the inlet (Dugdale and Goering 1967). In addition, BGS operation is also the dominant process that controls the positive estuarine circulation in P M A and thus is the major contributor to the vertical estuarine entrainment within P M A as well as influencing the residence time of the surface layer within P M A . In this regard, the operation of BGS is similar to the presence of a riverine system, which directly inputs nutrients into the coastal habitat and controls the subsurface entrainment and the hydraulic residence time of the surface layer within an estuary. Based on these nutrient inputs, the operation of BGS is 176 likely the dominant summer nutrient source in P M A and therefore dictates the phytoplankton community composition during periods of nutrient limitation within P M A . Lastly, the operation of BGS is mainly used as a supplementary energy source in BC, and its operation increases when hydroelectric generation is minimal due to low riverine discharges brought about by the end of the spring freshet. Therefore, increases in BGS operation, and its effects on the phytoplankton community of P M A , will accompany natural increases in intrinsic mixing rates within P M A due to the decrease in stratification brought about by the decrease in surface run-off. Accordingly, during years of low spring run-off, diatoms should be the dominant group of phytoplankton during the summer period because of concurrent increased BGS operation and intrinsic mixing rates. Conversely, the presence of flagellates, particularly Heterosigma akashiwo, should dominate the summer P M A phytoplankton assemblages during years of elevated snowmelt due to the increased vertical stratification of the inlet and reduced thermoelectric production resulting in decreased vertical mixing and reduced nutrient supply to the surface waters to support diatom growth. 177 Chapter 5 - Summary and conclusions 5.1 Phytoplankton Dynamics of Port Moody Arm This 2lA year study provides important new insights into the phytoplankton dynamics of an unusually productive British Columbia inlet, Port Moody Arm, and the operation of a thermoelectric power plant within a shallow tidal inlet. Based on the previous estimate of Stockner and Cliff (1979) and the classification of Nixon (1995), P M A is a hypereutrophic estuary. During this investigation, the average phytoplankton biomass of the inlet (95 mg chl a m2 integrated over 10 m) was more than 3 times higher than that measured in the contiguous waters of the Strait of Georgia (Stockner et al. 1979) and Howe Sound (Stockner et al. 1977). In addition, the maximum biomass of 1,200 mg chl a m2 recorded during a bloom of the ichthyotoxic flagellate, Heterosigma akashiwo is the highest concentration ever measured along the west coast of British Columbia. Therefore, P M A supports some of the highest phytoplankton standing stocks in BC coastal waters. The high phytoplankton biomass levels within P M A are a result of the physical forcings within the inlet and possibly due to the pollution-influenced reduction in benthic grazing. Because of its semi-confinement, wind-driven turbulence is low, and P M A is an essentially permanently stratified system. Furthermore, water column turbidity is relatively low due to the depositional environment of P M A and the lack of river-borne sediments flowing into the inlet. Consequently, the upper layer of P M A (Zm) is located entirely within the euphotic zone (Z e u), thus suspending the phytoplankton within an optimal light regime for growth. The Z e u : Z m > 1 measured during this study is far above the Z e u :Z m > 0.2 that Cloern (1987) predicted to support positive net primary production in San Francisco Bay, indicating the favourable light regime within P M A . It is also possible that the prominent 178 pollution within the sediments and associated reduction in bivalve populations contributes significantly to the elevated biomass in the inlet through reduced benthic grazing pressure. Seasonal phytoplankton trends in P M A were characteristic of that observed in other BC coastal waters with blooms occurring during the spring and fall seasons and the presence of low biomass during winter (Harrison et al. 1983). Low winter biomass was generally due to light limitation, and phytoplankton populations were dominated by tiny naked flagellates comprised mainly of various cryptomonad species, with Rhodomonas minuta being the most abundant. Spring and fall populations were dominated by diatoms with Skeletonema costatum being the dominant species throughout this study. Other important diatom species contributing to P M A standing stocks include Thalassiosira spp. and Chaetoceros spp., along with the potentially neurotoxin-containing pennate diatom Pseudo-nitzschia spp. An important finding during this study was the interannual variability in the composition of the summer phytoplankton assemblage in P M A . Flagellate species were dominant during the summer of 1999 when coastal waters were strongly influenced by the large 1999 freshet that resulted in extremely strong and long-lasting vertical stratification. The strength of this stratification was such that it was largely unaffected by the fortnightly tidal cycle and resulted in a N03-limited system throughout the summer that was ideal for the initiation and longevity of prodigious blooms of the potentially ichthyotoxic flagellate, Heterosigma akashiwo. During the summer of 1999, H. akashiwo blooms were the most frequent (three), had the highest biomass (1,200 mg chl a m~2), and were the longest lasting blooms during this study. Peak biomass of this raphidophyte appeared to coincide with spring tides indicating these tides may have contributed to the resuspension of H. akashiwo cysts, which excysted under favourable environmental conditions. 179 In contrast to 1999, diatoms dominated the biomass during the summers of 1998 and 2000 when vertical stratification within P M A was low. It seems likely that these diatom communities were supported by small fluxes of subsurface nutrients into the nutrient-limited upper layers from tidal additions during spring tides followed by subsequent phytoplankton growth during neap tides. The operation of BGS also played a substantial role in sustaining diatom populations throughout the summer of 2000 due to its effluent discharge (described below). 5.2 Burrard Generating Station 5.2.1 Intake and Plume Entrainment The Burrard Generating Station is the most dominant anthropogenic influence on the circulation within P M A , and thus influences the phytoplankton therein. By drawing its cooling water from 10 m in P M A , BGS subjects the entrained phytoplankton populations to physical and biocidal stress during plant passage, and dispels the thermally elevated effluent into the upper layers of P M A . Thus elevated nutrient levels are discharged into the euphotic zone where they become available to phytoplankton. The nutrient enrichment in the surface water of P M A becomes extremely important during the summer months when the surface water in P M A is typically NCMimited, and vertically migrating flagellates tend to dominate if the water column remains highly stratified. Therefore, one of the goals of this study was to understand the contrasting effects of phytoplankton destruction during intake entrainment within the BGS cooling system, versus the potential for enhanced phytoplankton production due to the injection of nutrient-rich bottom water into the surface layer of P M A . The investigation of the effect of intake entrainment on the P M A phytoplankton showed that during the summers of 1999 and 2000, the total algal biomass was reduced by 180 an average of 70 and 40%, respectively, as it moved through the BGS intake system. Declines in phytoplankton biomass during intake entrainment were inversely related to the discharge rate of BGS and positively related to the absolute temperature and temperature increase during transit. This likely reflects the quicker transit rates of the entrained algae, which reduced the duration of exposure to in-plant chlorination and elevated temperatures (2-8 min). Based on findings from past studies (see Langford 1990 for review), the greatest impact on phytoplankton mortality during intake entrainment was likely the chlorination of cooling water at the intake as opposed to the brief temperature shock. The greater percent decline in phytoplankton biomass of entrained phytoplankton during the summer of 1999 (relative to summer 2000) is thought to be due largely to the predominance of fragile dinoflagellates and nanoflagellates in the intake-entrained cooling water in 1999 as opposed to the more tolerant diatom assemblages that dominated during summer 2000. In total, the 'mortality' of the entrained algae contributed an estimated 15-25 tonnes of particulate and dissolved organic carbon (P/DOC) into P M A waters from the months of March to October, when phytoplankton growth was the greatest. Although these values for seasonal declines in phytoplankton biomass during intake entrainment and related releases of particulate and dissolved organic carbon to P M A may appear high, the estimated BGS-influenced phytoplankton mortality during this study was <1% of the total P M A phytoplankton biomass. This was a consequence of the entrained algal populations being drawn from below the pycnocline (10 m where the biomass is typically the lowest) and only -50% of the entrained phytoplankton was actually destroyed. An important aspect within this scenario, and one that has not been discussed in studies investigating the influence of intake entrainment, is that much of the entrained phytoplankton is largely comprised of diatoms that had already sunk below the pycnocline due to the summer N limitation of the surface 181 waters. Therefore, the conclusion is that the phytoplankton mortality caused by BGS intake entrainment is of little ecological importance to P M A in terms of a decrease in the total phytoplankton biomass. 5.2.2 BGS Nutrient Discharges and 'Natural' Entrainment in P M A In response to public concerns, a majority of studies on the potential effects of power plants on phytoplankton have focused on the effects of their mortality during plant passage. A far more important issue, however, may be the effects of the nutrient enriching capacity of power plants on the aquatic environment, since the nutrient-rich cooling water drawn from depth is discharged into the euphotic zone of the receiving waters as a buoyant jet where it becomes available to phytoplankton communities for growth and thus enhanced primary production. Therefore, BGS (and other similar plants that discharge cooling water drawn from depth to the surface waters) may act as an 'artificial upwelling system'. This is particularly important in areas where seasonal nutrient limitation occurs, such as in P M A during the summer. During the summers of 1999 and 2000, BGS discharged an average N 0 3 load of 275 kg d"1 (1999) or 900 kg d"1 (2000) and an average Si(OH) 4 load of 1,500 kg d"1 (1999) or 3,200 kg d"1 (2000) into the surface waters of P M A . Plume entrainment studies showed that the nutrient concentrations were the highest in the immediate receiving environment, and remained higher than ambient levels as the plume moved past the IOCO shipping dock into the central basin of P M A . Because the nutrient load discharged into P M A is directly related to the BGS operation, the quantity of nutrients discharged into P M A was substantially higher during the summer of 2000, when BGS was operating at near-peak discharge rates (> 1.2x10 m d" ). If this nutrient input was fully incorporated into phytoplankton biomass it would replace by 10-fold that which was destroyed during the 182 intake entrainment. Upon initial investigation, it would appear that this massive daily influx of surface nutrients could lead to the potential eutrophication of P M A , however, it was found that BGS discharge rates and P M A phytoplankton standing stocks were not significantly correlated, and that no cause for concern in this respect was evident. This is probably due to the strong tidal flushing in P M A , as well as the increased seaward surface water transport that accompanies elevated BGS operation (Jiang et al. 2003). The effects of BGS injecting nutrients into the surface waters of P M A were likely manifested in seasonal shifts from flagellate-dominated assemblages to diatom communities during summer periods when BGS operates at peak discharge rates. Canonical Correspondence Analysis (CCA) showed that BGS operation and the influence of surface stratification were the primary factors contributing to this summer phytoplankton shift. Flagellates, including Heterosigma akashiwo, dominated the P M A phytoplankton community during the summer of 1999, when the water column was highly stratified and BGS operated at low levels. This scenario created an environment where nutrient fluxes into the upper waters of P M A were greatly reduced, either through reduced direct BGS input due to operational declines, or from reduced 'natural' entrainment due to elevated vertical stratification. Therefore, this low nutrient, highly stratified environment selected for assemblages of flagellates that retrieve nutrients at depth (below 5 m) and migrate into the surface waters to photosynthesize during the daylight hours. Because diatoms are non-motile and rely on available nutrients supplied above the pycnocline by physical processes, flagellates (especially those that vertically migrate such as Heterosigma akashiwo) dominated the N-limited summer environment in P M A during 1999. During the summer of 2000, stratification of the water column in P M A was substantially less than that in 1999, while BGS operated at significantly higher levels and 183 diatom populations dominated the summer phytoplankton community. The effects were two-fold. First, the quantities of nutrients moved daily into the euphotic layer of P M A from the subsurface layer, via the BGS cooling-water discharge, were ~3-fold higher than 1999 and were estimated to be sufficient to permit diatom dominance. Second, the decrease in vertical stratification, coupled with the accelerated positive estuarine circulation created by BGS operation, increased the subsurface flux of nutrients into the surface layers by 3-4-fold compared to 1999 summer levels. Each of these sources contributed -50% to the surface nutrient flux during the summer of 2000 and each were individually higher than their combined effect during 1999. Therefore, BGS contributes to summer diatom dominance by both cycling the cooling water drawn from the deep waters into the surface waters of P M A , and by altering the water column circulation in P M A . Accordingly, the operation of BGS at levels near full capacity was likely the main reason for the summer diatom populations during 2000. I have established that diatom assemblages replace summer communities of flagellates when BGS discharge rates are high and the P M A water column is not strongly stratified. I have further determined that the major nutrient source delivered into the summer surface waters of P M A is from the direct input of heated cooling water when BGS discharge rates are high. Therefore, BGS acts as an 'artificial upwelling' system that supplies limiting nutrients to the upper layers of P M A , thus supporting diatom populations which are otherwise not normally abundant in P M A or other coastal BC waters during the summer months. Perhaps one of the beneficial features of BGS is that it discharges comparable levels of both Si(OH) 4 and N 0 3 assuring that diatoms will be abundant within P M A , as they consistently and effectively outcompete flagellate species for ambient nutrients within the euphotic zone. 184 The strengths of this study can be seen in the intensive temporal and spatial sampling regime. Weekly sampling during the phytoplankton growing season (March-September) permitted a detailed temporal examination of the different summer algal assemblages present within P M A , and provided an understanding of the bloom dynamics of the ichfhyotoxic Heterosigma akashiwo in P M A . This study also provided the first investigation into the nutrient loading capability of once-through thermoelectric power production (and also applies to nuclear power production utilizing the same cooling method). The summer period was the key season to determine if there was an influence of BGS on P M A phytoplankton stocks. This study indicates that small, well-flushed estuaries with no sill (e.g. PMA) may be satisfactory sites in which to construct thermal electric power plants in the future. However, if a sill is present at the mouth of the inlet (i.e. typical of many other coastal BC inlets), it would restrict bottom water exchange with adjoining waters, and would likely lead to bottom water anoxia from the increased downward flux that results from increased diatom production and the subsequent respiratory consumption of this organic material. 5.3 Future Studies Over 30 years of study into the effects of once-through power plant operation on phytoplankton ecology have led to some general conclusions: 1) that elevated discharge temperatures below the thermal tolerances of phytoplankton stimulate primary production, 2) temperatures above phytoplankton tolerances inhibit primary production, and 3) in-plant chlorination leads to substantial phytoplankton mortality during intake entrainment. This study reveals the importance of cooling water nutrient discharges into the surface layer in aquatic systems. One of the features of BGS is that its discharge limits are permitted to be 6 3 1 no greater than 1.7x10 m d" . In the future it is anticipated that BC Hydro will seek to 185 increase the energy production from BGS to meet the increased energy needs of the 6 3 1 province and will thus request an increase in the discharge limits to 2.5x10 m d" . If this is indeed the case, the following should occur: 1) a similar study to the one presented in this thesis should be conducted. Particular emphasis should again be placed on the summer season and should include primary production measurements in the in situ environment and for samples that have passed through the cooling water to determine their recovery potential (physiological assessment) after intake exposure. 2) a taxonomic inventory involving species of potentially toxic Pseudo-nitzschia should be undertaken (since this was not attempted during this study). During summers of elevated BGS operation when diatom production may be favoured, P M A should be periodically checked for domoic acid indicators, whether through isolation of Pseudo-nitzschia species, or through other vectors such as bivalves or Dungeness crabs to ensure there is no potential for a domoic acid outbreak. 3) The interactive effects of grazing (pelagic and benthic) and recycled nutrients (particularly ammonium, N H 4 ) should be undertaken to understand the contribution of this nitrogen pool to the summer phytoplankton community. 4) Finally, all future power plant studies should involve an understanding of the nutrient loading potential of 'once-through' power plant operation. One facet in this study that was not estimated was the subsurface water entrainment that was incorporated immediately into the discharge jet as it traveled through the near-field. Estimating this entrainment is difficult (J. 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Site Date Depth (m) 0 1 2 3 4 5 March 9 1999 0 mean SD 0.21 0.03 0.17 0.03 0.20 0.04 0.12 0.00 0.22 0.03 0.36 0.04 2 mean SD 0.25 0.05 0.28 0.03 0.16 0.05 0.31 0.04 0.34 0.09 0.57 0.13 5 mean SD 0.44 0.12 0.48 0.04 0.37 0.07 0.39 0.03 0.55 0.06 0.82 0.17 10 mean SD 0.77 0.03 0.69 0.10 0.82 0.03 0.76 0.15 0.97 0.18 March 24 1999 0 mean SD 0.33 0.10 0.22 0.17 0.26 0.09 0.47 0.08 0.44 0.05 0.40 0.11 2 mean SD 0.27 0.09 0.36 0.03 0.45 0.07 0.33 0.13 0.23 0.02 0.52 0.10 5 mean SD 0.35 0.07 0.38 0.10 0.42 0.14 0.49 0.15 0.49 0.07 0.92 0.26 10 mean SD 0.56 0.01 0.44 0.21 0.51 0.19 0.72 0.23 0.99 0.20 March 31 1999 0 mean SD 0.20 0.15 0.30 0.07 0.13 0.06 0.16 0.05 0.38 0.12 0.76 0.24 2 mean SD 0.25 0.06 0.35 0.04 0.36 0.11 0.28 0.13 0.48 0.10 0.88 0.16 5 mean SD 0.37 0.06 0.51 0.18 0.24 0.00 0.63 0.22 0.65 0.28 0.68 0.16 10 mean SD 0.42 0.18 0.52 0.10 0.36 0.03 0.57 0.05 0.67 0.15 217 Appendix C . l . Four-day nutrient addition experiments showing the effect of added NO3 (open circles), PO4 (dark circles), and no nutrient addition (control; dark triangle) on the phytoplankton growth (fluorescence) from samples collected from Site 3 in Port Moody Arm during 1999. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uE m"2 s"1. 1999 1 2 3 4 5 0 1 2 3 4 5 Days 218 Appendix C.2. Four-day nutrient addition experiments showing the effect of added NO3 (open circles), PO4 (dark circles), and no nutrient addition (control; dark triangle) on the phytoplankton growth (fluorescence) from samples collected from Site 3 in Port Moody Arm during 2000. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uE m"2 s"1. 219 Appendix C.3. Four-day nutrient addition experiments showing the effect of added NO3, PO4, and no nutrient addition (control) on the phytoplankton growth from samples collected from Site 3 in Port Moody Arm during 1999 and 2000. Positive sign (+) represents positive growth and's' represents stationary growth. Growth conditions at 18°C on 16h:8h light:dark cycle at 100 uE m"2 s"1. Year Date Treatment Day 2 Day 4 Control s s 1999 July 21 N 0 3 + + PO4 + s Control + s 2000 June 14 N O 3 + + P 0 4 + s Control + + July 6 N 0 3 + + PO4 + + Control + s July 20 N O 3 + + PO4 + s Control + s Aug 3 N O 3 + + P Q 4 + s 220 Appendix D . l . Chl a concentrations (ug l"1) in the BGS discharge effluent collected 2-10 minutes after intake sampling. Samples collected on June 14, 1999. A l l discharge chl a concentrations were significantly lower than in the intake water (One-way A N O V A , df=17, F= 7.328, P=0.02; Tukey's Pairwise Multiple Comparison Test) and were statistically similar to each other. Error bars represent standard deviaton from mean and n=3. 3H 2 ^ 1 •{ 0 _L ^BilflSII! T" T _T_ iillll^ llill lIBi 10 min Intake 2 min 4 min 6 min 8 min Outfall Time from Intake Sample Collection 221 Appendix E . l . Vertical profiles of density (at) for three sites in P M A (S1-S3) from June 18 to July 15, 1999. 222 Appendix E.2. Vertical profiles of density (at) for three sites in P M A (S1-S3) from July 21 to August 11, 1999. 223 Appendix E.3. Vertical profiles of density (at) for three sites in P M A (S1-S3) from August 19 to September 24, 1999. 224 Appendix E.4. Vertical profiles of density (ot) for three sites in P M A (S1-S3) from June 6 to June 29, 2000. 225 Appendix E.5. Vertical profiles of density (ot) for three sites in P M A (S1-S3) from July 6 to August 3, 2000. Appendix E.6. Vertical profiles of density (c t) for three sites in P M A (S1-S3) from August 9 to September 9, 2000. 227 Appendix E.7. Vertical profiles of density (ot) for three sites in P M A (S1-S3) from September 19 and September 29, 2000. Q 0 2 4 6 8 10 12 14 16 18 20 22 24 17 0 2 1 4 6 8 10 12 14 16 18 20 22 24 17 i 1 1 1 r 0 17.5 18.0 18.5 19.0 19.5 20.0 0 17.5 18.0 18.5 19.0 19.5 20.0 228 Appendix F. 1. Site-specific pycnocline depths (Z p y c ; mean±SD), average surface (T s u rf) and bottom water (Tb0t) temperatures, and the temperature difference between the two layers (AT) in Port Moody Arm from June 18 to September 24, 1999 and June 6 to September 29, 2000. Summer is June 21 to September 21. Temperature data were used in 2D entrainment model (Eq. 4.6) and the determination of pycnocline depth and average surface and bottom layer temperatures are described in Section 4.2.4. Date Zpyc (m) Site 2 3 4* T s u r f (°C) Site . 2 3 4* Tbot (°C) Site 2 3 AT (°C) Site 2 3 4* 1999 mean (n=12) SD 6.3 5.8 5.0 4.9 0.8 0.7 0.8 0.6 15.5 16.1 16.7 17.0 1.8 1.8 1.9 1.7 12.6 12.7 12.9 13.3 0.8 0.8 0.9 0.9 2.9 3.4 3.8 4.7 1.1 1.2 1.2 3.3 Summer mean (n=10) SD 6.8 6.1 5.1 4.8 0.6 0.6 0.8 0.7 16.0 16.7 17.5 17.7 1.7 1.7 1.7 1.7 12.8 12.9 13.1 13.8 0.8 0.8 0.9 0.8 3.3 3.8 4.4 5.9 0.9 1.1 1.0 4.0 2000 mean 6.9 6.1 5.3 5.1 (n=14) SD 1.6 0.9 0.8 0.8 15.1 15.9 16.8 16.9 1.8 1.7 1.5 1.4 12.5 12.7 12.8 13.0 1.0 1.0 1.0 1.0 2.6 3.3 3.9 3.9 1.6 1.4 1.1 1.1 Summer mean 7.3 6.2 5.3 5.3 16.1 16.8 17.6 17.5 (n=ll) SD 1.9 1.1 1.0 0.8 1.9 1.6 1.4 1.4 12.9 13.1 13.3 13.4 0.6 0.6 0.7 0.7 3.2 3.8 4.3 4.1 1.9 1.6 1.3 1.2 * - Site 4 during 1999 (n=7); Summer 1999 (n=6) 229 

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