Monitoring Report #24 Summer 2015

PROJECT QUINTE ANNUAL REPORT 2013

Bay of Quinte RAP Restoration Council / Project Quinte

BAY OF QUINTE REMEDIAL ACTION PLAN MONITORING REPORT #24

PROJECT QUINTE ANNUAL REPORT 2013

prepared by

Project Quinte members

in support of the Bay of Quinte Remedial Action Plan

Bay of Quinte Remedial Action Plan Kingston, Ontario, Canada.

Summer 2015

Editors Note: This report does not constitute publication. Many of the results are preliminary findings. The information has been provided to assist and guide the Bay of Quinte Remedial Action Plan. The information and findings cannot be used in any manner or quoted without the consent of the individual authors. Individual authors should be contacted prior to any other proposed application of the data herein.

PREFACE

BAY OF QUINTE REMEDIAL ACTION PLAN MONITORING REPORT #24

2013 PROJECT QUINTE ANNUAL REPORT (Summer 2015)

In 1985, the Great Lakes Water Quality Board of the International Joint Commission (IJC) identified 42 Areas of Concern in the Great Lakes basin where the beneficial uses were impaired. The Board recommended that the appropriate jurisdictions and government agencies prepare, submit and implement a Remedial Action Plan (RAP) in each area to restore the water uses.

The Bay of Quinte was designated as one of the Areas of concern. Ten of 14 beneficial uses described in Annex 2 of the Great lakes Water Quality Agreement (revised 1987) are impaired. The impaired uses include beach postings, eutrophication or undesirable algae, restrictions on fish consumption, taste and odour problems in drinking water, etc. The contributing factors are excessive phosphorus loadings, persistent toxic contaminants, bacteriological contamination, as well as alterations and destruction of shorelines, wetland and fish habitat.

Project Quinte is a long-term, multi-agency research and monitoring project. The project's original objectives included studying, comparing, and evaluating the Bay of Quinte limnological attributes (biological, chemical and physical) before and after phosphorus control was implemented at municipal sewage treatment plants thus reducing phosphorus loads to the bay. Project Quinte was launched over 40 years ago in 1972 and is still operating (Minns et al., 2011, 2012). The time between 1972 and 1977 has been referred to as the “pre-phosphorus control” period, while post-1977 is called “post-phosphorus control” period. Changes in water quality, as well as aquatic communities (e.g. phytoplankton, zooplankton, benthic and fish) between pre- and post-control periods have been compared. Phosphorus control strategies were assessed. Finally, long-term ecosystem responses within the Bay of Quinte were examined and described.

Another focus of Project Quinte is to assess changes in lower trophic level productivity in response to the invasion of the bay by Zebra Mussels that began in 1994. Since 2000, microbial loop assessment, consisting of the enumeration of bacteria, autotrophic picoplankton and ciliates, has been included in the program. Also, size fractionated primary productivity was included in the monitoring to assess the contributions of various size fractions of phytoplankton. The response to invasion of the bay by both Zebra Mussels and other non-native species of zooplankton and fish and how changes are transferred up the food chain is currently being addressed. Clearly, managing both the fishery as well as phosphorus loadings during alterations in the food chain presents a challenge that requires the yearly productivity assessment provided by Project Quinte.

This report is the 24th in the Bay of Quinte RAP monitoring report series. It is a window looking into the status of the Bay by providing information about projects and results currently being conducted in the Bay in support of the Bay of Quinte Remedial Action Plan. The report presents preliminary findings and data. It does not necessarily represent the views or policies of the

sponsoring agencies. The information and data contained in the report, more or less, serves as background reference material. The information has been compiled and reported so that:

(1) the Bay of Quinte RAP Restoration Council and QWC can formulate abatement and remedial options,

(2) remedial actions can be monitored and progress reported, and

(3) researchers can be informed about the current status of the health of the Bay of Quinte

The report has been prepared as part of the Bay of Quinte RAP under the auspices of the Canada- Ontario Great Lakes Remedial Action Plan. Financial assistance and technical sponsorship for the investigations and research was provided by Fisheries and Oceans Canada, Environment Canada, the Ontario Ministry of the Environment and the Ontario Ministry of Natural Resources. Through this preface the Bay of Quinte RAP Restoration Council notifies potential users that this report does not constitute publication.

References

Minns, C.K., Munawar, M., Koops, M.A., Millard, E.S., 2011. Long-term ecosystem studies in the Bay of Quinte, Lake Ontario, 1972-2008. A prospectus. Aquat. Ecosyst. Health Mgmt. 14(1), 3-8.

Minns, C.K., Munawar, M., Koops, M.A., 2012. Preface. Aquat. Ecosyst. Health Mgmt. 14(1), 369.

Bay of Quinte Information

Symposium

“Ecosystem Health and Recovery of the Bay of Quinte, Lake Ontario: Past, Present and Future”, May 9, 2010. Organized by: Aquatic Ecosystem Health & Management Society as a special session at the International Association for Great Lakes Research’s 53rd annual conference on Great Lakes Research. Toronto, Ontario.

Primary Publications

Besides the series of annual reports, of which this is the 24th, the following publications have been printed on the Bay of Quinte:

Minns, C.K., Hurley, D.A., Nicholls, K.H. (Eds.), 1986. Special publication: “Project Quinte: Point-Source Phosphorus Control and Ecosystem Response in the Bay of Quinte, Lake Ontario”. Canadian Journal of Fisheries and Aquatic Sciences, 86.

Aquatic Ecosystem Health & Management Society, 2011. Special issue: “Ecosystem Health and Recovery of the Bay of Quinte, Lake Ontario”. Aquat. Ecosyst. Health Mgmt. 14(1).

Aquatic Ecosystem Health & Management Society, 2012. Special issue: “Ecosystem Health and Recovery of the Bay of Quinte, Lake Ontario: Part II”. Aquat. Ecosyst. Health Mgmt. 15(2).

4 TABLE OF CONTENTS

INTRODUCTION 6

WATER TEMPERATURE MONITORING IN THE BAY OF QUINTE, 2013 8 K.E. Leisti, S. Avlijas, and S.E. Doka

POINT SOURCE PHOSPHORUS LOADINGS 1965 TO 2013 21 P. Kinstler and A. Morley

THE MICROBIAL AND PLANKTONIC COMMUNITIES OF THE BAY OF 24 QUINTE: OBSERVATIONS FROM THE BELLEVILLE LONG TERM MONITORING STATION DURING 2013 M. Munawar, M. Fitzpatrick, R. Rozon, and H. Niblock

ZOOPLANKTON IN THE BAY OF QUINTE – 2013 37 K.L. Bowen and R. Rozon

FISH POPULATIONS IN THE BAY OF QUINTE, 2013 58 J. A. Hoyle

2013 BAY OF QUINTE ALGAE WATCH PROGRAM 65 C. McClure, S. Watson, B. Keene, D. Eastcott, and L. Lambert

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INTRODUCTION

Project Quinte is a co-operative, multi-agency, research and monitoring project between the federal (Department of Fisheries and Oceans) and provincial governments (Ontario Ministry of Natural Resources, Ontario Ministry of the Environment) that has investigated the long-term effects of the reduction in point-source phosphorus (P) loadings (Minns et al., 1986), food-chain influences (Nicholls and Hurley, 1989), and more recently Zebra Mussel colonization (RAP monitoring reports Nos.7-12) on trophic dynamics of the entire bay ecosystem. The Bay of Quinte is one of 42 severely impaired ecosystems (Area of Concern - AOC) on the Great Lakes. Annex 2 of the Revised Great Lakes Water Quality Agreement of 1978 outlined a three stage Remedial Action Plan process that called for the identification of impaired beneficial uses (Stage I), their causes and a plan to be implemented (stage II) to restore these uses. The third and final stage of the RAP process requires monitoring to measure the success of RAP implementation that should ultimately lead to delisting the Bay of Quinte as an AOC. Project Quinte has been invaluable to stage three of the Remedial Action Plan (RAP), supporting an essential, continuing program of research and monitoring in the bay. Project Quinte has contributed long-term data that was used to produce comprehensive assessments of the status of two impaired beneficial uses, phytoplankton (Nicholls, Monitoring Report #11) and zooplankton (Johannsson and Nicholls, Monitoring Report #12). Project Quinte presented the first evidence of the impact of Zebra Mussels on the Bay of Quinte ecosystem in the 1995 monitoring report. Since that time we have observed variable impact, both spatially and inter-annually, of this invader on physical, chemical and biological properties of the bay (e.g. transparency, phosphorus, and lower trophic level biomass and production). There are no other Great Lakes studies that have the benefit of the extensive multi- trophic level database that exists for the Bay of Quinte to determine ecosystem level impacts of Dreissenid Mussels. We stand to learn a great deal about the impacts of this exotic species on fisheries productivity in the Bay of Quinte that may be of use in other ecosystems where background data is not as extensive. The scope of the work in Project Quinte is rare in freshwater ecosystem research, now encompassing multi-year, multi-trophic level data from bacteria to fish. Long-term data sets for

6 biomass, species composition and production for all trophic levels provide a unique opportunity to model trophic interactions in the bay and determine how various factors are impacting important fish populations. This report provides a detailed summary of the results and highlights for the 2013 field season, prepared by various scientists studying the bay. Please contact these individuals directly if further explanation or detail is required.

Link to the Remedial Action Plan

Progress towards restoration goals must be measured in order to delist impaired beneficial uses. The ongoing monitoring provided by Project Quinte is invaluable in assessing restoration progress as the federal government attempts to delist this AOC. The restoration goals originally set as part of the stage II development were considered to be the interim. Project Quinte has been instrumental in developing more realistic goals for phytoplankton, zooplankton, benthos and other related beneficial uses. Information on listing criteria and delisting goals has been summarized in the report, “Bay of Quinte RAP Monitoring and Delisting Strategy IBU Assessment Statements 2003” prepared for the Bay of Quinte Restoration Council under contract to Murray German Consulting and Fred Stride Environmental.

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WATER TEMPERATURE MONITORING IN THE BAY OF QUINTE, 2013

K.E. Leisti, S. Avlijas and S.E. Doka

Fish Habitat Science Section, Great Lakes Laboratory for Fisheries and Aquatic Sciences Fisheries and Oceans Canada 867 Lakeshore Road, P.O. Box 5050, Burlington, Ontario, L7R 4A6

Since 2001, GLLFAS has monitored water temperatures in a maximum of 16 locations in the Bay of Quinte area (Leisti and Doka, 2007). In 2013, data was collected at the same six offshore (i.e. mid-channel) and three nearshore stations monitored in 2012 (Figure 1). Offshore stations have been grouped into Upper Bay (Belleville and Napanee), Middle Bay (Hay Bay and Glenora) and Lower Bay (Conway and Lennox). At offshore stations temperature is monitored throughout the water column. Nearshore stations were set up in shallow locations to capture coastal thermal dynamics.

Napanee Belleville Hay Bay

Conway Lennox

Hay Bay Carnachan Bay

Glenora

Figure 1. Temperature logger locations in the Bay of Quinte in 2013.

At all offshore stations, temperature loggers were fixed onto spar buoys at approximately 1-m below the water surface (Figure 2); the buoys provide some shading for the logger so

8 temperatures should be reflective of the ambient conditions at this depth. For sites that are equal to or shallower than 12-m deep (e.g. Belleville, Napanee and Hay Bay), all loggers were attached to the spar buoy line at depths listed in Table 1 using the set-up illustrated in Figure 2A. At sites deeper than 12 m (e.g. Glenora, Conway and Lennox), U-moorings were used to ensure that the loggers were consistently positioned at a set water depth (Figure 2B). On U-moorings, loggers deployed between 1- and 9-m deep were attached to the spar buoy line, which was suspended vertically in the water column by the tension between the floating spar buoy and a weight on the lake bottom. An additional weight was placed below the deepest logger on the spar buoy line to fix the loggers at a consistent position in the water column. Loggers, deeper than 9 m, were deployed on a second line and suspended in the water column by a submersed buoy.

Figure 2. Offshore mooring methods used at (A) sites less than 12-m deep and (B) the U mooring deployed at offshore sites deeper than 12 m.

Nearshore loggers were deployed approximately 30 cm from the lake bottom on a large weight resting on the bottom, in water depths ranging from 1.5 to 2 m during spring deployment. Actual water depths varied throughout the season as water levels fluctuated but the logger typically remained above the sediments to avoid burial and 1.5-m deployment depths avoided complete dewatering during low seasonal water levels. Water temperatures were recorded at ½-hour intervals for the full year at nearshore sites and from early May to early November at offshore sites. Onset HOBO (U22-001) loggers were used at all stations except at Glenora, Conway and Lennox where the U mooring fittings only accommodated the smaller Onset Tidbit (UTBI-001) loggers at depths of 9 m and greater (Figure 2B). Additional loggers were deployed by the GLLFAS Freshwater Ecosystems Research section at the Belleville and Hay Bay stations in 2013 as part of the Lake Ontario Cooperative Science and Monitoring Initiative (CSMI). A listing of the offshore loggers deployed in the Bay of Quinte in 2013 can be found in Table 1.

Table 1. Relative depth of temperature loggers at all offshore (mid-channel) stations in the Bay of Quinte.

Offshore Station Temperature monitoring depth (m) in 2013 Belleville 0* 1 2* 4* 5 Napanee 1 5 Hay Bay 0* 1 2* 3 4* 5*” 6 8* 9 10*” 12 Glenora 1 3 6 9 12” 15 18 Conway 1 3 6 10 12 15 18 21 24 27 30 Lennox 1 5 10 15” 20” 25 30 35 40 45 50 55 60 * Loggers deployed by Warren Currie, GLLFAS Freshwater Ecosystems Research. “ Loggers malfunctioned, no data or partial year recorded.

It should be noted that air temperatures in 2013 were substantially cooler during all seasons than in preceding years (Figure 3). Comparing the 2013 mean monthly air temperatures against the 1981 to 2010 climate normals for the Trenton A weather station, winter (Dec-Mar) ranged from 2.4 to 6.6 °C cooler in 2013; spring (Apr-May) was between 5.9 and 4.8 °C; summer (Jun-

Aug) ranged from 4.4 to 6.0 °C and fall temperatures (Sep-Nov) were between 3.7 and 6.6 °C cooler in 2013.

Offshore Temperature Results

At all offshore stations, the 1-m mean daily temperatures varied more in the summer (Jun- Aug) and fall (Sep-Nov), with standard deviation values between 2.3 and 4.5, than in the spring (Apr-May) when the recorded temperatures fell within 1.5 to 2.9 standard deviations (Table 2). The minimum temperatures during the ice-free deployment ranged between 5.6 °C and 8.2 °C and occurred either within the first 13 days after deployment or the last few days before retrieval of the loggers. The maximum temperatures at most locations were recorded in mid-July; except for Glenora and Conway where this occurred in late August (Table 2).

Air Air Temperature, C 0

-5

-10

-15 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

2010 2011 2012 2013 1981-2010 Normals

Figure 3. Mean monthly air temperatures from the Trenton A weather station from 2010 to 2013, including the 1981-2010 climate normals.

Table 2. Summary of mean daily temperatures at 1-m depth at offshore stations in 2013. Minimum and maximum mean daily temperatures are listed with the date when they occurred during the deployment. The mean and standard deviation are reported for data available in the spring (Apr-May), summer (Jun-Aug) and fall (Sep-Nov).

Belleville Napanee Hay Bay Glenora Conway Lennox Minimum 7.5 7.9 8.2 6.8 5.6 7.3 Maximum 27.6 27.3 27.0 24.1 23.6 23.6 Date of Min Nov 5 Nov 6 Nov 6 Nov 14 May 12 May 13 Date of Max Jul 18 Jul 20 Jul 19 Aug 29 Aug 28 Jul 16 Spring Mean 16.8 15.3 16.2 13.7 12.0 11.7 Summer Mean 23.2 20.7 22.3 20.1 19.6 19.8 Fall Mean 16.5 16.6 17.0 16.5 17.0 17.0 Spring SD 1.8 1.8 1.5 2.4 2.9 2.6 Summer SD 2.3 3.4 2.3 3.1 3.7 3.6 Fall SD 4.5 4.0 4.1 4.0 3.3 3.2

Typically, temperatures across the Bay of Quinte increased from early May, peaked between late July and early August and then generally declined through September until logger retrieval in November (Figure 4); although depending on their location, loggers experienced differing levels of temperature variation. The 1-m depth daily means in Belleville differed significantly from those at Napanee (N=189, p=0.002); Hay Bay was significantly different from Glenora (N=189, p=0.0008); but there was no significant difference between Conway and Lennox (N=197, p=0.93). On average, the shallower Upper Bay sites warm faster in the spring and summer, but cool faster in the fall than the Middle and Lower Bay stations (Figure 4). The 1-m temperatures in the Upper and Middle Bay are more similar than to the more exposed and deeper Lower Bay. There is an average difference of 1.1 °C and a maximum of 6.3 °C between Upper/Middle sites and Lower Bay sites. Between the Upper and Lower Bay only, the temperature difference was 2.3 °C on average and 9.7 °C at most. The mean difference between Middle and Lower bay temperatures was 1.4 °C with a maximum of 6.7 °C. The differences in temperature between the bays are principally due to thermal inertia and illustrate the heat storage capacity of waters with varying depths and exposures to the main Lake dynamics.

10 Temperature, ˚C Temperature, 5

0 1 Jul 1 Jan 1 Jun 1 Oct 1 Feb 1 Sep 1 Apr 1 Dec 1 Mar 1 Aug 1 Nov 1 May Upper Middle Lower

Figure 4. Mean daily offshore water temperatures measured at 1 m from surface in 2013. Upper Bay is the mean of Belleville and Napanee, Middle Bay is the mean of Hay Bay and Glenora, and Lower Bay is the mean of Conway and Lennox.

10 Temperature, ˚C Temperature, 5

0 1 Jul 1 Jan 1 Jun 1 Oct 1 Feb 1 Sep 1 Apr 1 Dec 1 Mar 1 Aug 1 Nov 1 May 0m 1m 2m 3m 4m 6m 8m 9m 10m 12m

Figure 5. Mean daily offshore temperatures measured at different depths at the Hay Bay Station (at the confluence between Hay Bay and Middle Bay) in 2013. The 0-m depth is the air temperature.

The offshore stations in Middle and Lower Bay are sufficiently deep (12 m or deeper) to observe stratification of the water column (Figures 5 to 8). Hay Bay began to stratify prior to logger deployment at the beginning of May and was isothermal by the third week of September with a temperature range of generally less than 1 °C at all depths until logger retrieval (Figure 5). The 12-m temperature on the date of retrieval, 6 Nov 2013, was still 8.1°C. Stratification at this site progressed between 3 and 12 m throughout the year until turnover. The maximum difference between the 1-m and 12-m temperatures (a proxy of epi- to hypolimnion) was 15.7 °C and occurred on 21 July 2013. The water column at Glenora began stratifying prior to logger deployment and by Oct 9th, the temperature range across all depths was less than 1.7 °C until the loggers were retrieved (Figure 6). At logger retrieval, the temperature at 18 m was 6.6 °C. A maximum difference in temperature of 12.8 °C occurred on July 15th between the 1-m and 18-m depth loggers at this site. In the Lower Bay, stratification had already occurred at Conway by the time of logger deployment (Figure 7). By October 23rd and until logger retrieval, there was less than 1 °C temperature difference across all depths. Logger retrieval temperature was 9.6 °C at 30 m. The max difference between 1- and 30-m temperatures was 13.8 °C and occurred on 15 July 2013. The Lennox station located in the Adolphus Reach is twice as deep as Conway and closer to Lake Ontario proper than the other stations (Figure 1). The day after logger deployment, the temperature difference between the 1-m and 60-m depth was 3.7 °C and rose to 11.2 °C by May 9th (Figure 8). Unfortunately, logger malfunction at the intermediate depths of 15 and 20 m reduced the ability to track stratification through the season. The pattern of turnover in early November was similar to that of 2012, with a sharp rise in bottom temperatures to equal the steadily declining surface temperatures as the lake became isothermal. The maximum difference between 1- and 60-m depth temperatures was 15.1 °C and occurred on 16 July 2013.

10 Temperature, ˚C Temperature,

0 1 Jul 1 Jan 1 Jun 1 Oct 1 Feb 1 Sep 1 Apr 1 Dec 1 Mar 1 Aug 1 Nov 1 May 1m 3m 6m 9m 15m 18m

Figure 6. Mean daily offshore temperatures measured at different depth at Glenora Station (Lower - Middle Bay transition) in 2013.

Temperature, ˚C Temperature, 5

0 1 Jul 1 Jan 1 Jun 1 Oct 1 Feb 1 Sep 1 Apr 1 Dec 1 Mar 1 Aug 1 Nov 1 May 1m 3m 6m 10m 12m 15m 18m 21m 24m 27m 30m

Figure 7. Mean daily offshore temperatures measured at different depths at Conway Station (Lower Bay) in 2013.

Temperature, ˚C Temperature, 5

0 1 Jul 1 Jan 1 Jun 1 Oct 1 Feb 1 Sep 1 Apr 1 Dec 1 Mar 1 Aug 1 Nov 1 May 1m 5m 10m 25m 30m 35m 40m 45m 50m 55m 60m

Figure 8. Mean daily offshore temperatures measured at different depths at Lennox Station (Lower Bay) in 2013.

Nearshore Temperature Results

Nearshore temperatures were logged throughout the year in the Upper Bay Napanee station and in the Middle Bay at both the Hay Bay and Carnachan Bay stations (Figure1). All sites were deployed in approximately 1.5 m of water but these depths then varied throughout the season. At the beginning of the year, water temperatures were consistently above 1 °C by March 15th in Carnachan Bay, March 22nd at Hay Bay and March 23rd at Napanee. Temperatures peaked in mid-July and generally followed air temperatures, albeit with a dampened response (Figure 9). Water temperatures were consistently below 1 °C to the end of December beginning on November 23rd at Carnachan Bay, November 25th at Napanee and December 14th at Hay Bay.

C 20 , ˚ ,

10 Temperature

0 1 Jul 1 1 Jan 1 1 Jun 1 1 Oct 1 1 Feb 1 Sep 1 1 Apr 1 1 Dec 1 1 Mar 1 1 Aug 1 Nov 1 1 May 1 Napanee Hay Bay Carnachan Bay Trenton Air

Figure 9. 2013 mean daily water temperatures at nearshore stations of Bay of Quinte.

Table 3. Summary of mean daily and seasonal temperatures at nearshore sites of Bay of Quinte in 2013.

Napanee Hay Bay Carnachan Bay Minimum 0.0 0.1 0.0 Maximum 29.1 29.2 28.6 Date of Max Jul 18 Jul 18 Jul 19 Spring Mean 8.5 8.5 7.9 Summer Mean 23.0 23.1 21.4 Fall Mean 12.8 12.0 13.0 Winter Mean 0.6 1.1 0.8 Spring SD 6.9 6.8 5.9 Summer SD 2.7 2.6 3.6 Fall SD 6.9 6.7 6.7 Winter SD 0.11 0.34 0.29

Temperatures in Carnachan Bay were generally cooler than at Napanee or Hay Bay in the spring and summer and then were typically less than 1°C warmer in the fall (Table 3). The maximum difference of daily mean temperatures between Napanee and Carnachan Bay was 7.7 °C and occurred on June 6th, while on June 28th Hay Bay was 8.7 °C warmer than Carnachan Bay. The cooler temperatures in Carnachan Bay in the late spring and summer of 2013 are consistent with previous years and are due to upwelling and mixing with the generally colder Lake Ontario waters by seiches and diffusion between the Lower bay and the Middle bay areas. Bay of Quinte temperatures and oxygen have been analyzed both in the past (Minns and Johnson, 1986; Minns, 2009) and more recently (Minns et al., 2011). In the 2011 paper, the offshore/nearshore temperature dataset from 2001 to 2008 and earlier data was used to assess the influence that climate, morphometry, nutrient loading and dreissenids have had on Quinte’s temperature and oxygen temporal trends and spatial patterns. The analysis revealed a slight climate warming signature, with maximum summer surface waters increasing by almost 1 °C on average from 1972 to 2008. A detailed analysis of thermal structure in major embayments and the whole Bay of Quinte are underway and coarse statistical temperature models have been used for various assessments including habitat analyses for the Remedial Action Plan and for potential delisting of this Area of Concern (Gertzen et al., 2012a,b). Fish are bioenergetically and behaviourally affected by temperature at various stages of their life history, from spawning and egg survival to growth and migration (Hanson et al., 1997). Thermal dynamics were modelled in the bay, and incorporated into a spatiotemporal habitat suitability assessment for different thermal guilds (Gertzen et al., 2012a). Specifically, thermal habitat has been included in fish habitat mapping and population modelling activities and delisting has been recommended, based on trends and various spatial comparisons (Gertzen et al., 2012b). Inter-annual variability in temperature will influence variation in production from year to year for different fish species. We will be incorporating this information into population models for some key fisheries in the eastern end of Lake Ontario so that the potential relative contribution of the Bay is quantified. Lastly, these data can contribute to other long-term datasets at Belleville Water Treatment Facility. DFO field sampling datasets can help identify trends in warming that may be occurring spatially across the Bay. These data can also contribute

18 to local and regional, hydrodynamic and climate modelling underway that can assist future management of restoration activities, coastal areas and fisheries.

Acknowledgements

Thanks go to Janet Mossman, Dave Reddick and Stephen James for logger preparation and nearshore deployment. Offshore logger support was provided by C. Talbot and R. McFadyen of Environment Canada aboard the CCG and DFO Science vessel, the Kelso. Funding for the Bay of Quinte Fish Habitat Assessment project, of which this is a subproject, was provided by the Great Lakes Action Plan, administered and led by Environment Canada for funding work in Areas of Concern towards delisting.

References

Gertzen, E.L., Doka, S.E., Minns, C.K., Moore, J.E., Bakelaar, C.N., 2012a. Effects of water levels and water level regulation on the supply of suitable spawning habitat for eight fish guilds in the Bay of Quinte, Lake Ontario. Aquat. Ecosyst. Health Mgmt. 15(4), 397-409.

Gertzen, E.L., MacEachern, J.T., Doka, S.E., 2012b. BUI #14-4: Loss of Fish and Wildlife Habitat (FWH-4). Bay of Quinte AOC BUI Status Report. Informal Report to Bay of Quinte RAP Council. Oct 2012.

Hanson, P. C., Johnson, T. B., Schindler, D. E., Kitchell, J. F., 1997. Bioenergetics model 3.0 for Windows. University of Wisconsin, Sea Grant Institute, Technical Report WISCU-T-97- 001, Madison, USA.

Leisti, K.E., Doka, S.E., 2007. Water Temperature Monitoring in the Bay of Quinte. In: Monitoring Report #16. Project Quinte Annual Report 2005, pp. 7-11. Bay of Quinte Remedial Action Plan, Kingston, Ontario, Canada.

Minns, C.K., Johnson, M.G., 1986. Temperature and Oxygen Conditions and Oxygen Depletion in the Bay of Quinte, Lake Ontario. In: Minns, C.K., Hurley, D.A., Nicholls, K.H. (Ed.) Project Quinte: point-source phosphorus control and ecosystem response in the Bay of Quinte, Lake Ontario, pp. 40-49. Can. Spec. Publ. Fish. Aquat. Sci. 86.

Minns, K., 2009. Temperature and oxygen regimes in the Bay of Quinte 1972 – 2008. In: Monitoring Report #18, Project Quinte Annual Report 2007, pp. 8-12. Bay of Quinte Remedial Action Plan, Kingston, Ontario, Canada.

Minns, C.K., Moore, J.E., Doka, S.E., St. John, M.A., 2011. Temporal trends and spatial patterns in the temperature and oxygen regimes in the Bay of Quinte, Lake Ontario, 1972- 2008. Aquat. Ecosyst. Health Mgmt. 14(1), 9-20.

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POINT SOURCE PHOSPHORUS LOADINGS 1965 TO 2013

Peter Kinstler and Andrew Morley

Ontario Ministry of the Environment and Climate Change 1259 Gardiners Road, Unit 3, P.O. Box 22032, Kingston, Ontario K7M 8S5

Daily average total phosphorus loads to the Bay of Quinte between 1965 and 2013 from six municipal sewage treatment plants (STP) bordering the Bay of Quinte were derived from routine Ontario Ministry of the Environment and Climate Change (MOECC) monitoring data (Table 1). Daily average phosphorus loadings are reported as kilograms per day (kg d-1). Loadings are calculated as the monthly average effluent phosphorus concentration multiplied by the monthly average effluent flow from the STP for the period of record. In 2013, the estimated daily average total phosphorus loads from STPs bordering the Bay of Quinte were 4.9 kg d -1 for the year and 4.2 kg d-1 for the May to October period. Since 1991, low total phosphorus inputs have been maintained by controlling variability in wastewater flow, enhancing phosphorus removal capabilities and expanding treatment capacity. To manage long term phosphorus loading, in the 1990s the MOECC advised municipalities bordering the Bay of Quinte that a Bay of Quinte RAP “phosphorus load cap” would be applied to effluent from STPs. Initially, the “phosphorus load cap” was to be calculated as the product of a total phosphorus effluent concentration of 0.3 mg L-1 and the respective hydraulic capacity for STPs bordering the Bay of Quite. This “phosphorus load cap” has been used post 1995 to measure and assess performance. Since 1995, most of the operators of municipal STPs have voluntarily reduced total phosphorus inputs or have made upgrades and the “phosphorus load cap” has been incorporated into the STP’s operating Environmental Compliance Approval. Similarly, since 1995 the MOECC has advised municipalities in the Bay of Quinte watershed that a Bay of Quinte RAP “phosphorus load cap” applies to STPs in the Bay of Quinte watershed. The watershed “phosphorus load cap” has been initially established as the product of a total phosphorus effluent concentration of 0.5 mg L-1 and the STP approved hydraulic capacity. Most operators of STPs in the watershed have voluntarily reduced total phosphorus inputs or

21 have made upgrades and the “phosphorus load cap” has been incorporated into the STP’s operating Environmental Compliance Approval. Phosphorus loading from bypassing events is not specifically reported in the data as presented. STP operators have taken steps to reduce bypassing, however bypassing continues to occur at some STPs adding to the phosphorus loading to the Bay of Quinte. Continued efforts are required to control wastewater influent flow variability and reduce bypassing to effectively “cap” phosphorus inputs to the Bay of Quinte.

Table 1. Daily average for phosphorus loadings from municipal sewage treatment plants bordering the Bay of Quinte (total P kg d-1) from 1965 to 2013. Period T CFB B D N P TOTAL 1965-72 53.0 13.0 110.0 2.7 20.0 15.0 214.0 1973-75 47.0 5.9 73.0 2.3 17.0 11.0 156.0 1976-77 43.0 2.1 39.0 2.3 63.0 9.1 158.0 1978-86 Jan-Dec 7.5 1.9 31.0 0.9 24.0 2.4 68.0 May-Oct 6.3 1.7 25.0 0.7 24.0 1.9 60.0 1987 Jan –Dec 5.9 2.7 16.0 0.6 11.0 2.1 38.0 May-Oct 6.4 2.0 15.0 0.5 6.2 1.8 32.0 1988 Jan-Dec 6.9 4.7 15.0 0.7 9.2 2.2 42.0 May-Oct 6.3 1.7 11.0 0.5 12.4 1.9 34.0 1989 Jan-Dec 6.0 * 13.0 0.7 5.6 1.4 26.7 May-Oct 5.3 * 11.0 0.6 5.9 1.0 23.8 1990 Jan-Dec 5.9 * 14.0 1.0 4.1 1.8 26.8 May-Oct 6.5 * 9.0 0.6 3.8 1.5 21.4 1991 Jan-Dec 6.8 * 9.2 1.0 2.6 1.7 21.3 May-Oct 5.2 * 4.6 0.8 2.3 0.9 13.8 1992 Jan-Dec 5.9 * 12.3 1.4 2.5 1.5 23.6 May-Oct 5.1 * 9.6 0.9 2.3 0.8 18.7 1993 Jan-Dec 6.1 * 12.1 0.9 3.0 1.0 23.1 May-Oct 4.6 * 10.1 0.3 2.8 0.7 18.5 1994 Jan-Dec 4.2 * 18.5 0.5 2.0 1.3 26.5 My-Oct 3.3 * 12.1 0.4 1.4 0.9 18.1 1995 Jan-Dec **4.2 * 14.3 0.3 1.5 1.2 21.5 May-Oct **3.3 * 14.2 0.2 1.1 0.9 19.7 1996 Jan-Dec 3.4 * 18.4 0.4 1.4 1.7 25.3 May-Oct 4.4 * 14.9 0.3 1.2 1.1 21.9 1997 Jan-Dec 4.0 * 16.7 0.5 1.8 1.4 24.3 May-Oct 3.2 * 13.2 0.4 1.4 0.6 18.8

Period T CFB B D N P TOTAL 1998 Jan-Dec 2.3 0.7 11.6 0.4 1.5 1.1 16.9 May-Oct 2.3 0.6 11.8 0.4 1.3 0.6 16.4 1999 Jan-Dec 3.3 0.6 6.8 0.7 1.6 1.3 13.6 May-Oct 3.5 0.5 7.0 0.6 1.3 0.8 13.2 2000 Jan-Dec 2.0 0.7 6.1 0.6 1.5 1.5 12.4 May-Oct 2.5 0.8 7.3 0.5 1.4 0.8 13.4 2001 Jan-Dec 1.5 0.5 4.8 0.2 1.3 1.8 10.1 May-Oct 1.4 0.5 3.5 0.1 1.3 1.0 7.8 2002 Jan-Dec 1.7 0.5 7.0 0.9 1.3 1.2 12.6 May-Oct 1.9 0.4 9.2 0.4 1.4 1.2 14.5 2003 Jan-Dec 2.2 0.7 13.0 0.2 4.0 2.2 22.3 May-Oct 2.8 0.8 4.0 0.2 1.0 0.5 9.3 2004 Jan-Dec 3.3 0.4 7.0 0.2 1.4 0.9 13.2 May-Oct 3.2 0.4 2.0 0.2 0.9 0.7 7.4 2005 Jan-Dec 2.9 0.4 5.5 0.2 1.3 0.8 11.1 May-Oct 2.8 0.3 3.8 0.1 1.1 0.5 8.6 2006 Jan-Dec 4.8 0.8 8.9 0.3 1.7 2.0 18.5 May-Oct 4.0 0.3 5.7 0.3 1.4 1.2 12.9 2007 Jan-Dec 3.6 0.3 1.6 0.3 1.4 1.4 8.6 May-Oct 3.4 0.2 1.2 0.2 1.1 0.9 7.0 2008 Jan-Dec 4.1 0.5 1.4 0.3 1.6 1.2 9.1 May-Oct 3.8 0.3 1.5 0.1 1.3 0.8 7.8 2009 Jan-Dec 2.0 0.4 21.5* 0.2 1.3 0.9 26.3 May-Oct 2.1 0.3 27.9* 0.1 1.0 0.9 32.3 2010 Jan-Dec 1.3 0.2 10.6 0.1 1.0 0.7 13.9 May-Oct 1.4 0.3 9.9 0.2 0.7 0.5 13.0 2011 Jan-Dec 1.8 0.3 2.6 0.1 1.2 0.6 6.2 May-Oct 2.1 0.3 2.3 0.1 1.0 0.2 6.0 2012 Jan-Dec 1.1 0.2 1.3 0.1 1.1 0.2 4.0 May-Oct 1.5 0.3 1.5 0.1 1.0 0.2 4.6 2013 Jan-Dec 1.5 0.3 2.0 0.1 0.9 0.1 4.9 May-Oct 1.4 0.3 1.5 0.1 0.8 0.1 4.2

T=Trenton, CFB=Canadian Forces Base Trenton, B=Belleville, D=Deseronto, N=Napanee, P=Picton. * CFB Trenton loadings not available. ** No 1995 records available for Trenton, 1994 loadings substituted to estimate total loading. * In October 2009, Belleville reported bulking as a result of a hydraulic overload caused by heavy rain.

23 Click here to return to the Table of Contents.

THE MICROBIAL AND PLANKTONIC COMMUNITIES OF THE BAY OF QUINTE: OBSERVATIONS FROM THE BELLEVILLE LONG TERM MONITORING STATION DURING 2013

M. Munawar, M. Fitzpatrick¹, R. Rozon, H. Niblock,

Great Lakes Laboratory for Fisheries and Aquatic Sciences, Fisheries and Oceans Canada, Burlington, ON L7S 1A1 ¹Correspondence: [email protected]

Introduction

This annual report is based on the data collected under the auspices of Project Quinte, the long term research and monitoring program maintained by Fisheries & Oceans Canada in conjunction with Environment Canada, the Ontario Ministry of the Environment and the Ontario Ministry of Natural Resources. Project Quinte began in 1972 in response to widespread concerns over eutrophication and focused on monitoring the response of phytoplankton and zooplankton communities to phosphorus load reductions. Beginning in 2000, the microbial food web, including bacteria, autotrophic picoplankton, heterotrophic nanoflagellates and ciliates was added to the regular suite of parameters in order to provide a more holistic assessment of the lower trophic levels. This is an interim report for the 2013 sampling period (May – October) which provides details of the structure and function of the microbial-planktonic food web at the Belleville monitoring site. In addition, long term changes in total phytoplankton biomass as well as total phosphorus and chlorophyll a as they relate to Remedial Action Plan (RAP) targets are also discussed.

Methods

A total of 13 sampling events were conducted bi-weekly from May 9 – October 30, 2013 at the long term monitoring site near Belleville (Fig. 1). Integrated (0 – 4 m) whole water samples

24 were collected from our research vessel, the Leslie J, and transported on ice to our laboratory in Burlington, Ontario for further analysis. Various physical measurements were performed on site including temperature (YSI ExoSonde) and water column irradiance (Li-Cor Quantum sensor). Total phosphorus and nitrate + nitrite were analyzed in accordance with the standard protocols of the National Laboratory for Environmental Testing (NLET). Chlorophyll a concentrations were determined by filtering up to 1 L of water through Whatman GF/C filters followed by cold acetone pigment extraction and spectrophotometric analysis (Strickland & Parsons, 1968)

Figure 1. Map of the Bay of Quinte showing long term monitoring sites at Belleville (B), Napanee (N), Hay Bay (HB) and Conway (C). Only the Belleville data is reported on here.

Microbial loop (bacteria, heterotrophic nanoflagellates and autotrophic picoplankton) samples were preserved in 1.6% formaldehyde and processed using DAPI staining and epifluorescent microscopy. Details of this technique are given in Munawar et al. (1994). Ciliate samples were preserved with acidified Lugol’s iodine and prior to analysis post-fixed with Bouin’s fluid. Enumeration and identification followed the Quantitative Protargal Staining technique (Montagnes and Lynn, 1993).

Phytoplankton samples were fixed with acidified Lugol’s iodine upon collection. Enumeration and measurement followed the HPMA (2-hydroxypropyl methacrylate) technique described by Crumpton (1987) and is broadly compatible with the Utermöhl (1958) inverted microscope technique. Size fractionated primary productivity was estimated by 14-Carbon uptake for three size categories of phytoplankton (<2 µm, 2-20 µm, >20 µm) following the standard technique of Munawar and Munawar (1996).

Results and Discussion

General characteristics and RAP Targets

Phosphorus abatement was first implemented in the Bay of Quinte in 1978 in response to concerns over eutrophication. Project Quinte has included long term monitoring of several physical and chemical parameters including temperature, water column irradiance (light penetration) and nutrients in addition to measures of primary productivity and algal standing crop. Surface water temperatures at the Belleville site ranged from 8.0 ºC to 27.0 ºC between

May and October 2013. Likewise, the vertical attenuation coefficient (Ԑpar), a measure of light penetration through the water column, was between 0.7 and 1.3 m-1 indicating relatively low water clarity, especially during the summer months. When the Remedial Action Plan was implemented, a target concentration for total phosphorus (TP) of 30 μg l-1 was established. During 2013, we observed that TP concentrations ranged from 20 – 81 μg l-1 (Fig. 2a) with a mean of 34 μg l-1 and exceeded the target on 7 of 13 occasions. While no formal targets were established for nitrogen, we observed that nitrate + nitrite concentrations in the spring and fall were generally between 20 and 125 μg l-1 (Fig. 2b). However, during the summer the nitrate + nitrite concentration of 2.5 – 8.0 μg l-1 was near or below the analytical detection limit (5 μg l-1) indicating a significant drawdown of nitrogen by photosynthesizing algae which is characteristic of a eutrophic environment.

a) Total Phosphorus (μg l-1)

b) Nitrate + Nitrite (μg l-1)

c) Chlorophyll a (μg l-1)

Figure 2. Total phosphorus, nitrate + nitrite and chlorophyll a concentrations observed at Belleville during the 2013 monitoring season. All units are μg l-1.

a) Size fractionated primary productivity (mg C m-3 h-1) and percent size composition

b) Phytoplankton biomass (g m-3) and percent taxonomic composition

c) Microbial Loop biomass (g m-3) and percent composition

Figure 3. a) Primary productivity (mg C m-3 h-1) and percent size composition; b) phytoplankton biomass (g m-3) and percent taxonomic composition, and c) microbial loop biomass (g m-3) and percent composition observed at the Belleville site during 2013.

Chlorophyll a is often used as an indicator of the algal standing crop, and a target concentration of 12 – 15 μg l-1 was established for the Bay of Quinte. In 2013, we observed that chlorophyll a ranged from 3.5 – 18.9 μg l-1 (Fig. 2c) exceeding the (upper) target only 3 times. Unique to the Bay of Quinte RAP, a target for phytoplankton biomass (4 – 5 g m-3) was also established to provide a more reliable measurement of the algal standing crop. Phytoplankton biomass at the Belleville site ranged from 1.4 – 4.9 g m-3 (Fig. 3b) falling within the target range or lower. While both targets for the algal standing crop were met during 2013, it is worth noting that the average chlorophyll a concentration of 10.5 μg l-1 is considered indicative of a eutrophic environment and the mean phytoplankton biomass of 2.8 g m-3 is considered to be on the threshold of mesotrophy and eutrophy (e.g. Munawar et al., 2012).

Phytoplankton dynamics and food web interactions

Primary productivity ranged from 7 to 85 mg C m-3 h-1 during the 2013 field season (Fig. 3a) with the lowest level reported in late May and the highest level in early July. Our size fractionation experiments revealed that larger sized net plankton (> 20 µm) was dominant from early June until early October (43 – 64%), with the only exception being a peak in nanoplankton in August. By way of comparison, primary productivity in the offshore, oligotrophic waters of eastern Lake Ontario (station 81, near Kingston, ON) ranged from 1.2 – 10.8 mg C m-3 h-1 during the same time frame with relatively equal contributions from the net (> 20 µm), nano- (2 – 20 µm) and pico- (< 2 µm) size classes (M. Munawar, unpubl.). The much higher rates of production by larger sized algae observed in the Bay of Quinte is characteristic of a nutrient enriched environment. As noted in the previous section, total phytoplankton biomass was estimated to be 1.4 – 4.9 g m-3 during the sampling season (Fig. 3b). While these observations fell within or below the RAP target, the target itself does not take into account the presence or absence of algal blooms. We observed the presence of algal blooms, defined where phytoplankton biomass exceeds 3.0 g m-3, on 5 of 13 sampling events during 2013 (May 9, June 4, July 3, July 31 and October 9). Four of these blooms (all except July 3) were overwhelmingly composed of filamentous diatoms (75 – 86% of total biomass), particularly species of Aulacoseira including: A. ambigua, A.

29 crenulata, A. granulate and A. muzzanensis and to a lesser extent Fragilaria capucina and F. crotonensis. The July 3 bloom contained a mixture of diatoms (54% A. granulata and A. muzzanensis) and dinoflagellates (34% Peridinium umbonatum). Throughout 2013, diatoms including species of Aulacoseira, Cyclostephanos, Cyclotella, Fragilaria, Stephanodiscus and Synedra composed > 50% of the biomass in 12 of 13 observations. Cyanophyta (blue-green algae) were most prevalent from the mid-summer to early fall contributing 13 – 34% of the total biomass and containing several filamentous and colonial forms including species of Dolichospermum (Anabaena), Aphanocapsa and Microcystis. The phytoplankton assemblage contains a mixture of diatoms, blue-greens and other algae that reflect cultural eutrophication. Microbial loop biomass (composed of bacteria, autotrophic picoplankton, heterotrophic nanoflagellates, and ciliates) ranged from a low of 1.1 g m-3 in late May to a high of 2.9 g m-3 in late July (Fig. 3c). Bacteria were the major components (0.5 to 1.4 g m-3) contributing from 41 to 67% of the microbial loop biomass followed by autotrophic picoplankton (0.06 to 0.98 mg m-3) accounting for 6 to 35% of the biomass. Considering the combined phytoplankton and microbial communities, total autotrophic biomass (phytoplankton + autotrophic picoplankton) was 1.4 – 5.0 g m-3 compared to 0.9 – 2.2 g m-3 for heterotrophs (bacteria, heterotrophic nanoflagellates, ciliates). On average, autotrophs accounted for 69% of the microbial food web biomass compared to 31% for heterotrophs. While the problem of eutrophication is generally regarded as an excess of autotrophic production driven by nutrient enrichment, our results show that heterotrophic microbes also have an important role in determining the fate of autochthonous production.

Long Term Trends

Project Quinte began in 1972 in response to widespread concerns over eutrophication, both within the bay and throughout the Great Lakes basin. At the time, phosphorus loads into the bay exceeded 200 kg d-1 and controlling point source phosphorus loadings was deemed to be the best course of action to alleviate eutrophication (e.g. Vollenweider et al., 1974; Minns et al., 1986). Project Quinte was unique in that it not only informed the discussion and policy over nutrient controls under Annex 4 of the Great Lakes Water Quality Agreement, but it also was intended to

30 be the case study for measuring long-term ecosystemic response to phosphorus abatement. In 1978, point source phosphorus loads were cut to < 80 kg d-1 and have been steadily reduced since. The current recommended cap is 15 kg d-1 (Minns and Moore, 2004) and in 2013, P loadings into the bay were < 5 kg d-1. Similarly, the total phosphorus concentration (seasonal average) was reduced from ≈ 90 μg l-1 in the early 1970s to > 50 μg l-1 following P abatement to current levels which were near the target of 30 μg l-1, however TP concentrations appear to have been on an upswing over the past 6 years (Fig. 4a). While point source phosphorus loads continue to be reduced, TP concentrations have not followed the same trend. This is because the TP concentration in the water column is influenced by many other non-point factors including tributary loadings, storm sewer overflows and sediment re-suspension (e.g. Munawar et al., 2012). Reductions in phosphorus loadings were predicted to reduce the algal standing crop. Chlorophyll a concentrations have declined from ≈ 40 μg l-1 (on average) prior to phosphorus abatement to current levels that are at or near the target of 12 – 15 μg l-1 although there has been considerable variability (Fig. 4b). Likewise, phytoplankton biomass has declined from an annual average > 10 g m-3 in the early 1970s to levels that have generally been consistent with the RAP target of 4 – 5 g m-3. It is worth noting however, that meeting the RAP targets for both chlorophyll a and phytoplankton biomass still results in persistently eutrophic conditions (see Munawar et al., 2012 for a full discussion). There is not enough room in the current report to provide a full discussion of the composition of the phytoplankton community however, such a discussion is critical to assessing and understanding the Beneficial Use Impairments of “Eutrophication” and “Degradation of phytoplankton and zooplankton communities”. Nicholls and Carney (2011) considered the long term data from 1972 – 2008 and observed that while most of the dominant taxa declined since P abatement, the blue-green Microcystis actually showed an increase in biomass following the establishment of dreissenid mussels in the bay during the mid 1990s. Having said that, diatoms notably Aulacosiera spp., have consistently been reported as the largest contributor to total phytoplankton biomass followed by species of Cyanophyta (Dolichospermum spp.) as shown in Fig. 5. Both Diatomeae and Cyanophyta blooms have contributed to the eutrophication

31 impairment of the Bay of Quinte over the years and further research into the complexity and variety of individual bloom events is warranted. a) Total phosphorus (μg l-1)

b) Chlorophyll a (μg l-1)

Figure 4. Total phosphorus and chlorophyll a trends in the Bay of Quinte at the Belleville long term monitoring station. Values are the seasonal weighted mean and units are μg l-1. The RAP target is indicated by the gray bar.

Figure 5. Long term trends in phytoplankton biomass (g m-3) and composition (percent biomass) at the Belleville monitoring site.

A note on the long term phytoplankton record

For 40 consecutive years, 1972 – 2011 inclusive, phytoplankton biomass and community composition was assessed by the discerning eyes of Elaine Carney who followed a modified version of the Utermöhl inverted microscope technique (see Nicholls et al., 2002) which was meticulously analyzed and interpreted by Ken Nicholls. Their work has resulted in one of the most consistent long term databases of phytoplankton in the Great Lakes (and probably the world). Here we would like to pay tribute to both Ms. Carney and Mr. Nicholls for their long service to Project Quinte which extended well beyond their formal retirements from the Ontario Ministry of the Environment. We also acknowledge their respective contributions to the science of eutrophication management which resulted from this work. Beginning in 2012, we have worked with a new contractor for the phytoplankton assessment. One noteworthy change is the use of the HPMA (2-hydroxypropyl methacrylate) technique (Crumpton, 1987) rather than the Utermöhl technique. In describing the HPMA method, Crumpton demonstrated that the results produced by this technique were consistent with the Utermöhl method. It is our professional opinion that this change will not disrupt the continuity of the long term data set.

Summary

As part of its ongoing commitments to Project Quinte, Fisheries & Oceans Canada surveyed the phytoplankton and microbial communities of the Bay of Quinte during 2013. This work included microscopic assessments of the phytoplankton, microbial loop (autotrophic picoplankton, bacteria, heterotrophic nanoflagellates) and ciliate communities in addition to radioisotope measurements of primary productivity. Regular measurements of temperature, water column irradiance, nutrient levels (total phosphorus, nitrate + nitrite), and chlorophyll a were also included. During 2014, we found that seasonal weighted averages of total phosphorus (34 μg l-1), chlorophyll a (10.5 μg l-1) and phytoplankton biomass (2.8 g m-3) were consistent with Remedial Action Plan targets, but still indicative of a eutrophic environment. Algal blooms were observed on 5 of 13 sampling events and were typically dominated by species of

Aulacoseira, a filamentous diatom strongly associated with phosphorus enrichment. We also observed that > 30% of the biomass of the microbial – planktonic food web was heterotrophic. The long term monitoring record, dating back to 1972, indicates that eutrophic conditions still persist in the Bay of Quinte.

Acknowledgements

We thank Robert Bonnell and Ashley Bedford of DFO, as well as our summer students (Sonya Oetterich, Jason Reed, Giulia Rossi, Jilian Zsolt) for the long hours in the field and the lab. We also thank Jennifer Lorimer of the AEHMS for assistance with data analysis and processing.

References

Crumpton, W.G. 1987. A simple and reliable method for making permanent mounts of phytoplankton for light and fluorescence microscopy. Limnol. Oceanogr. 32: 1154-1159.

Minns, C.K., Owen, G.E., Johnson, M.G. 1986. Nutrient loads and budgets in the Bay of Quinte, Lake Ontario, 1951-81. In: Minns, C.K., Hurley, D.A., Nicholls, K.H. (Eds), Project Quinte: point-source phosphorus control and ecosystem response in the Bay of Quinte, Lake Ontario. Can. Spec. Publ. Fish. Aquat. Sci. 86, 59-76.

Minns, C.K. and Moore, J.E. 2004. Modelling phosphorus management in the Bay of Quinte, Lake Ontario in the past, 1972 to 2001, and in the future. Can. Manuscr. Rep. Fish. Aquat. Sci. 2695: v+42p.

Montagnes, D.J.S., Lynn, D.H., 1993. A quantitative protargol stain (QPS) for ciliates and other protists. In: Kemp, P.F., Sherr, B.F., Sherr, E.B., and Cole, E.J. (Eds), Handbook of methods in aquatic microbial ecology. Lewis Publishers, Boca Raton, FL.

Munawar, M., Munawar, I.F., Weisse, T., Leppard, G.G., Legner, M., 1994. The significance and future potential of using microbes for assessing ecosystem health: The Great Lakes example. J. Aquat. Ecosyst. Health 3, 295-310.

Munawar, M., Munawar, IF. 1996. Phytoplankton dynamics in the North American Great Lakes, Vol. 1. Lakes Ontario, Erie and St. Clair. SPB, Amsterdam

Munawar, M., Fitzpatrick, M., Munawar, I.F., Niblock, H., Kane, D., 2012. Assessing ecosystem health impairments using a battery of ecological indicators: Bay of Quinte, Lake Ontario example. Aquat. Ecosyst. Health Mgmt. 15(4), 430-441.

Nicholls, K.H., Heintsch, L., Carney, E., 2002. Univariate step-trend and Multivariate assessments of the apparent effects of P loading reductions and zebra mussels on the phytoplankton of the Bay of Quinte, Lake Ontario. J. Great Lakes Res. 28(1), 15-31.

Nicholls, K.H., Carney, E.C. 2011. The phytoplankton of the Bay of Quinte, 1972 – 2008: Point source phosphorous lading control, dreissenid mussel establishment, and a proposed community reference. Aquat. Ecosyst. Health Mgmt. 14(1), 33-43.

Strickland, J.D.H., and Parsons, T.R. 1968. A Practical Handbook of Seawater Analysis. Fisheries Research Board of Canada, Bulletin 167, 71–75.

Utermöhl, H., 1958. Zur vervolkommnung der quantitativen phytoplankton-methodik. (The improvement of quantitative phytoplankton methodology. In German.) Mitt. Internat. Verein. Limnol. 9: 1-38.

Vollenweider, R.A., Munawar, M., Stadelmann, P. 1974. A comparative review of phytoplankton and primary production in the Laurentian Great Lakes. J. Fish. Res. Board Can. 31, 739-762.

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ZOOPLANKTON IN THE BAY OF QUINTE – 2013

K.L. Bowen and R. Rozon

Great Lakes Laboratory for Fisheries and Aquatic Sciences Department of Fisheries and Oceans 867 Lakeshore Road, P.O. Box 5050 Burlington, Ontario, L7S 1A1

Introduction

The zooplankton community in the Bay of Quinte is dynamic. It changes in structure down the length of the Bay from the shallow, eutrophic habitat in the upper bay, represented by the stations at Belleville (B), through eutrophic but slightly deeper conditions in the middle bay near Hay Bay (HB), to a mesotrophic to oligotrophic environment in the mouth of the bay at Conway (C). Station depth increases from 5 m at B, to 12 m at HB and 32 m at C. However, depth drops off fairly rapidly at C, and in 2013, the mooring was located slightly closer to shore, at a depth of only about 26 m. Thermal stratification occurs throughout the summer at C, sporadically at HB and not at B. Zooplankton community structure and productivity is controlled not only by the physical environment but also by the type and quantity of available food. This is determined by nutrient conditions, the structure of the fish community which is the primary source of mortality in this system, and the introduction of new species which either compete for food or alter predation on or within the zooplankton community. Monitoring of zooplankton in the Bay of Quinte tracks species composition, size structure, abundance, biomass and productivity in an effort to understand the response of the community to changes in the controlling factors and assess its ‘health’ as a component of the ecosystem. This report presents data from the 2013 field year in the context of previous data.

Methods

In 2013, zooplankton samples were collected at three primary monitoring stations, B, HB and C. Samples were collected on alternate weeks from late April until the end of October for B and HB, but only monthly at C from May to August then biweekly until early October; this was due to budgetary and logistical constraints and an attempt to capture any community changes that coincided with potential algal blooms events. As in previous years, discrete samples were collected through the water column with a 41-L Schindler-Patalas trap fitted with 64-m mesh. Three depths were sampled at B (1 m, 2 m and 3 m) and five at HB (1 m, 2 m, 3 m, 6 m and 9 m). A single composite sample was constructed for each station-date by combining 50% of the sample from each depth. On 23-May and 09-Oct, a total water column net was taken at C with a 153-µm Wisconsin- style zooplankton net. On six dates between June and September, epilimnetic (epi) and meta- hypolimnetic (MH) vertical net hauls were taken at C using a 40-cm diameter, 64-µm closing net. The MH net was taken from the bottom of the epi sample to 1-m off bottom. To compare catches in the zooplankton net to those of the Shindler Patalas traps traditionally used at C, eight Schindlers were also collected (1 m, 3 m, 5 m, 8 m, 10 m, 15 m, 20 m and 25 m) on each of six dates at C. Half of each sample was pooled to make an epi and a MH sample, corresponding to the same depths sampled with the net. For each gear type, counts from each stratum were combined to estimate densities in the total water column. For the net hauls, the densities from each stratum were weighted according to the proportion of the water column that stratum occupied. We assumed that the MH sample to a maximum depth of about 25 m was representative of the water column to 33 m, the more typical depth at this station over the duration of Project Quinte. Using 18-Jun as an example, the epi net was taken from 0-10 m and the MH net was 10-25 m. The epi net was then given a weighting of 0.30, and the MH a weighting of 0.70, again assuming the more typical depth of 33 m. The Schindler depths traditionally used at this station are potentially problematic in that they over-represent the epi and underrepresent the MH. Although the sampling design would be improved by spacing the Schindlers evenly throughout the water column, this has not been done

38 in order to preserve long-term continuity in the data set. In 2013, densities in the epi and MH Schindler samples were weighted two different ways to test this bias, again assuming a station depth of 33 m. For the above example, the four traps from 1-8 m were pooled to make an epi sample, and the remaining four traps from 10-25 m were pooled to make a MH sample. Using the “traditional” method where all traps were given an equal weighting, both the epi and MH samples were given a weighting of 4/8, or 0.5. In the “proper” method, we attempted to correct for the unequal distribution of traps in the water column, and gave the less frequent MH traps a higher weighting. In this example, the epi traps were assumed to represent 0-9 m in the water column (weighting of 0.27), and the MH traps represented 9-33 m (0.72). A minimum of 400 individual zooplankters and all loose eggs within a subsample aliquot were enumerated from each sample. The counting method is described in detail in Bowen and Johannsson (2011). Seasonally-weighted mean (SWM) abundance, SWM biomass and total production of zooplankton were calculated over the May 01 to October 06 sampling season. Cladoceran mean lengths were estimated over the June 01 to October 06 periods. SWM biomass and seasonal biomass trends were calculated by applying length-weight regression equations to measured zooplankton lengths (Bowen and Johannsson, 2011). These “measured” biomass estimates represent a departure from earlier Quinte reports (e.g., Bowen and Gerlofsma, 2010) when SWM values were based on Great Lakes fixed weights (Bowen and Johannsson, 2011). It was recently brought to our attention that the equation used to estimate veliger weights (Hillbricht-Ilkowska and Stanczykowska, 1969) should be divided by two to better represent dry weight (Sprung, 1984). Therefore, all veliger weights in this report have been halved compared to previous reports. Production was estimated by the egg-ratio method of Paloheimo (1974) for the more abundant taxa, and P/B relationships for the less dominant taxa (Johannsson and Bowen, 2012).

Rotifers

Rotifer sampling began in 2000 by collecting 1 L of water from each depth indicated above using a Van Dorn sampler. For each station-date at B and HB, water was pooled and filtered through 20-µm mesh. Rotifers were narcotized using carbonated water and preserved as above.

Each year, a seasonal composite sample for each station was made by combining 50% of the sample from each date. Enumeration and identification of rotifers in these seasonal composites were performed using a weighted counting method and limits similar to that of the zooplankton (Bowen and Johannsson, 2011). The biovolume (mm3) for each individual was calculated by using formulae of Ruttner-Kolisko (in McCaulley, 1984).

Results

Belleville

The zooplankton seasonally-weighted mean (SWM) biomass at Belleville in 2013 was only 56.4 mg m-3 (Fig. 1A). Mean biomass at this station has ranged between 127 and 176 mg m-3 since 2008, and represents the lowest SWM at Belleville since 2004. Biomass at Belleville has been highly variable over the years, but was generally greater in the 1982 to 1991 period, and again in 2001. The lowest values were seen in 1992, 1997 and 2000. This variability has been largely driven by fluctuations in cladoceran biomass, which in 2013 averaged 45 mg m-3 or 79% of the total (Tab. 1). Cyclopoid copepod SWM biomass was 4.8 mg m-3 in 2013, a slight decrease compared to the 2008 to 2012 period when it averaged 13.5 mg m-3 or 9%. Cyclopoids were more abundant prior to the arrival of dreissenids (1979-1994), when they averaged 47 mg m-3, or 18% of the total. Calanoid copepods have usually comprised <2% to 4% of the biomass at B. The contribution by dreissenid veligers has been variable since the species invaded in the mid-1990s. In 2013, veliger biomass was 5.8 mg m-3 (19%), very similar to the 1995 to 2012 mean of 6 ± 1 mg m-3. Veliger biomass peaked at 19 mg m-3 or 18% in 2010, but was only 3% of the total biomass in 2012. This is the first year that the dry weight correction of 50% has been applied to all past veliger data. Rotifers only contributed an additional 2.3 mg m-3 to SWM biomass, a value similar to the 2000 to 2012 mean of 2.9 mg m-3. The dominant rotifer taxa in 2013 in terms of biomass were Polyarthra vulgaris, Keratella cochlearis, Synchaeta pectinata and S. stylata (Fig. 2A).

400 A) Belleville 8000 D) Belleville DM 7000 DM ) 3 )

300 -

3 6000 - PC PC 5000

200 4000

3000

Biomass (mg (mg Biomassm 100 2000 Production (mg Production(mg m

1000 * * 0 0 ** 400 8000 B) Hay Bay E) Hay Bay 7000 DM )

3 6000 -

) 300 3 - DM 5000 PC PC 200 4000

3000 Biomass (mg (mg Biomassm 100 (mg m Production 2000 * * * * * * 1000

0 0

200 2000 C) Conway Rotifers F) Conway Veligers

) 150 1500 ) 3 Cyclopoids - 3 - Calanoids Cladocerans 100 1000 DM PC PC DM Biomass (mg (mg m Biomass 50 500 Produciton (mg Produciton(mg m

* * ** * * * * * *

0 0 1975 1977 1979 1981 1983 1985 1987 1989 1991 1993 1995 1997 1999 2001 2003 2005 2007 2009 2011 2013 1975 1977 1979 1981 1983 1985 1987 1989 1991 1993 1995 1997 1999 2001 2003 2005 2007 2009 2011 2013

Figure 1. Trends in seasonally-weighted mean dry biomass (A-C) and total seasonal production (D-F) of zooplankton groups at Belleville, Hay Bay and Conway. These are volumetric estimates based on measured lengths from May 1 to October 6. Time stanzas are indicated by the arrows: implementation of phosphorus control (PC) and invasion by dreissenid mussels (DM). Note the different scale used for Conway. Rotifer sampling began in 2000. * indicates no samples were collected that year.

Table 1. Distribution of zooplankton SWM biomass (May 1 - Oct. 06) amongst the taxonomic groups in the Bay of Quinte at B, HB and C in 2013.

Taxon Feeding Percentage of Total Percentage of SWM Guild SWM Biomass biomass without Veliger Larvae B HB CB HB C Cladocerans 79.1 76.7 37.3 88.2 78.4 48.1 Herbivorous 79.1 76.6 35.7 88.1 78.3 46.0 Carnivorous 0.0 0.1 1.6 0.0 0.1 2.1 Cyclopoids Herb./Carn. 8.6 19.3 37.8 9.5 19.7 48.8 Calanoids Herbivorous 2.1 1.8 2.4 2.3 1.9 3.1 Veligers Herbivorous 10.3 2.1 22.5 ------

SWM Biomass (mg.m-3) 56.4 149.6 66.3 50.6 146.4 51.4

Bay of Quinte zooplankton biomass shows strong seasonal patterns, which are influenced by water temperature, phytoplankton abundance, and fish predation. Biomass at Belleville is typically low in early May (Fig. 3A). In 2013, biomass began increasing in early June with the appearance of Bosmina, with Daphnia retrocurva and D. galeata mendotae appearing throughout July. Biomass reached its highest levels first in June (84 mg m-3) and again in late July (181 mg m-3) when D. galeata mendotae, D. retrocurva, and Eubosmina coregoni peaked. Daphnia and Eubosmina abruptly crashed in early August and remained low for the rest of the season. From late August through early October, biomass averaged 43 mg m-3, and dropped to only 3.7 mg m-3 by late October. Bosmina was again the dominant taxon, with much lower levels of the cladoceran Chydorus sphaericus and veligers. Total zooplankton seasonal production has fluctuated in recent years and generally mimics biomass (Fig. 1D), although this was not the case at Belleville in 2012, it was once again on trend in 2013. 2013 total seasonal dry production was 1 321 mg m-3 for the 01 May to 06 Oct period, the lowest level since 2000. This was less than half the 2002-2012 mean of 3 267 mg m 3. The sudden decrease as compared to 2012 is due to biomass declines in all zooplankton groups, except for veligers. When rotifers were excluded, cladocerans contributed

89% of total production in 2013 (Tab. 2), and the bosminids (Bosmina + Eubosmina) together contributed 60%. Cyclopoids and calanoids each made up 3.5% or less of the total, while veligers made up 6.9% of the total production. Unlike the previous year, rotifers contributed more to production than cyclopoids and calanoids combined (65 mg m-3). Rotifers represented 4.7% of total zooplankton + rotifer production. The May-October SWM zooplankton density at Belleville was 93.5 no. L-1 in 2013. Although small in size, Bosmina sp., E. coregoni, C. sphaericus, cyclopoid nauplii larvae and veligers were numerically the most dominant taxa (Tab. 3). Although D. retrocurva and D. galeata mendotae were less abundant, they were still very important given their larger size and high productivity. Although most of the cyclopoids found at all stations were juveniles not identified to species, the most dominant adult cyclopoids at B were Eucyclops agilis in late spring and Tropocyclops extensus in the summer and early fall. The Belleville June 1 - October 6 mean cladoceran length has fluctuated widely in the last decade. The 2013 mean cladoceran length at B was 0.33 mm in 2013, the lowest since 2002 and very similar to values from the early 1980’s. Once again, the 2013 value was below the proposed target value of 0.45 to 0.50 mm, largely due to the dominance of small taxa such as Bosmina and C. sphaericus.

Table 2. Distribution of zooplankton total seasonal production (May 1 - Oct. 06) amongst the taxonomic groups in the Bay of Quinte at B, HB and C in 2013.

Taxon Feeding Percentage of Total Percentage of Production Guild Production without Veliger Larvae B HB CB HB C

Cladocerans 88.9 86.7 42.2 95.6 88.6 62.4 Herbivorous 88.9 86.6 40.9 95.6 88.4 60.5 Carnivorous 0.0 0.1 1.3 0.0 0.1 1.9 Cyclopoids Herb./Carn 3.5 10.6 24.5 3.8 10.8 36.1 Calanoids Herbivorous 0.6 0.7 1.0 0.6 0.7 1.5 Veligers Herbivorous 6.9 2.1 32.3 ------

Total Production (mg.m-3) 1321 2234 719 1229 2188 486

Table 3. Seasonal mean abundances of major zooplankton taxa at stations B, HB and C (no. L-1) in 2012 and 2013 compared to mean values from earlier time stanzas. Copepod juveniles (nauplii and copepodids) are not included in these totals. Note that C was not sampled in 2012. Post-Dreissenid Post- Pre P Control Post P Control Pre-Cercopagis Cercopagis (1975, 1976) (1979-1994) (1995-1998) (1999-2011**) Species Mean SE Mean SE Mean SE Mean SE 2012 2013

B Bosmina sp. 33.36 28.35 27.78 8.29 16.66 1.94 29.41 5.36 41.97 26.59 Eubosmina coregoni 38.94 9.79 58.43 4.99 25.46 6.70 19.42 4.54 14.95 9.21 Daphnia retrocurva 11.69 0.95 15.52 2.28 10.38 4.12 4.84 0.71 5.63 2.83 D. galeata 0.15 0.14 5.39 1.00 6.96 1.18 6.31 2.50 7.92 0.62 Ceriodaphnia sp. 3.12 2.63 4.28 1.11 1.72 0.31 5.99 1.27 4.92 2.65 Chydorus sphaericus 47.94 6.73 38.52 4.30 9.68 3.25 10.17 2.89 10.48 4.69 Diaphanosoma 2.70 0.57 1.93 0.36 2.07 0.55 1.60 0.29 0.81 0.17 Leptodora kindtii 0.37 0.02 0.10 0.02 0.04 0.02 0.06 0.01 0.07 0.00 Cercopagis pengoi ------0.00 0.00 0.00 0.00 Diacyclops thomasi 0.08 0.06 0.21 0.05 0.12 0.05 0.03 0.01 0.04 0.04 Cyclops vernalis 1.33 0.19 1.55 0.45 0.62 0.15 0.10 0.03 0.06 0.00 Tropocyclops extensus 0.88 0.83 1.49 0.54 0.46 0.13 1.72 0.34 2.20 0.56 Mesocyclops edax 0.21 0.03 1.29 0.29 1.26 0.56 0.74 0.21 0.82 0.03 Veligers -- - - 26.91 6.45 27.14 5.64 16.20 32.21

HB Bosmina sp. 43.99 3.90 65.30 18.94 20.21 1.97 39.22 6.64 29.67 39.06 Eubosmina coregoni 22.87 6.48 35.54 4.03 13.88 0.91 14.00 2.05 9.81 14.30 Daphnia retrocurva 5.98 0.40 13.57 1.63 8.36 1.17 8.00 0.99 10.12 6.76 D. galeata 0.03 0.00 3.81 0.75 9.89 2.35 5.71 0.95 3.42 3.78 Ceriodaphnia sp. 2.67 0.10 11.79 2.70 4.54 0.62 7.91 1.40 8.78 15.28 Chydorus sphaericus 24.76 4.97 36.83 4.75 16.58 4.55 14.85 3.01 24.01 18.61 Diaphanosoma 0.92 0.32 0.89 0.21 0.76 0.24 0.61 0.14 0.31 0.25 Leptodora kindtii 0.18 0.02 0.04 0.01 0.03 0.02 0.01 0.00 0.01 0.01 Cercopagis pengoi ------0.01 0.00 0.01 0.00 Diacyclops thomasi 0.30 0.21 0.75 0.09 0.72 0.12 0.17 0.04 0.19 0.12 Cyclops vernalis 0.33 0.01 0.79 0.22 0.40 0.14 0.05 0.02 0.06 0.10 Tropocyclops extensus 4.99 4.47 3.82 0.96 0.79 0.11 2.33 0.32 2.18 1.34 Mesocyclops edax 0.09 0.02 1.28 0.26 1.60 0.32 0.92 0.14 0.66 0.64 Veligers -- - - 15.23 1.44 30.21 5.66 14.58 14.01

C Bosmina sp. 13.92 0.19 27.50 4.58 8.54 1.47 5.11 0.70 - 11.84 Eubosmina coregoni 3.43 1.18 6.85 0.96 2.20 0.54 3.24 0.50 - 2.45 Daphnia retrocurva 1.59 0.54 4.58 0.49 4.08 0.64 2.15 0.35 - 1.09 D. galeata 0.03 0.03 0.31 0.08 1.07 0.49 0.91 0.32 - 0.40 Ceriodaphnia sp. 1.16 0.50 3.24 0.61 4.35 2.10 2.33 0.58 - 1.92 Chydorus sphaericus 2.65 0.29 3.01 0.57 2.89 1.43 1.87 0.37 - 5.38 Diaphanosoma 0.13 0.09 0.22 0.05 0.17 0.02 0.25 0.06 - 0.05 Leptodora kindtii 0.04 0.01 0.02 0.00 0.01 0.01 0.02 0.00 - 0.00 Cercopagis pengoi ------0.10 0.02 - 0.05 Diacyclops thomasi 1.40 0.15 2.13 0.25 1.66 0.33 0.37 0.04 - 0.65 Cyclops vernalis 0.10 0.02 0.06 0.01 0.04 0.02 0.01 0.00 - 0.02 Tropocyclops extensus 0.72 0.71 2.07 0.44 0.34 0.15 0.26 0.07 - 0.43 Mesocyclops edax 0.12 0.02 0.20 0.03 0.18 0.03 0.16 0.03 - 0.16 Veligers -- - - 21.71 5.60 16.81 4.72 - 50.53

*Other species seen in 1997-2013: Alona sp., Bythotrephes longimanus, Ceriodaphnia quadrata, Holopedium gibberum, Polyphemus pediculus, Pleuroxus sp., Scaphaloberis kingi, Eucyclops agilis, Eurytemora affini, Leptodiaptomus sicilis, L. ashlandi, L. siciloides, Skistodiaptomus oregonensis. **Average 1999 - 2008 for Conway

12 Other A) Belleville 10 Synchaeta ) 3 - Trichocerca 8 Polyarthra 6 Keratella Asplanchna 4

2 Dry Biomass Dry Biomass (mg m 0

12 B) Hay Bay ) 3 - 10

Dry Biomass Dry Biomass (mg m 2 N/A

0 12 C) Conway ) 3

- 10

Dry Biomass Dry Biomass (mg m 2 N/A 0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013

Figure 2. Seasonal mean dry biomass of dominant rotifer groups in the Bay of Quinte from 2000 to 2013 at B and HB, and 2000 to 2009 at C.

200 A) Belleville

150

100

) 500 3 - B) Hay Bay Veligers 400 Calanoids Cyclopoids 300 Other Herb Clad Bosminds Biomass (mg (mg Biomassm 200 Daphnia Predatory Clad. 100

0 03-Jul 16-Jul 31-Jul 09-Oct 30-Oct 04-Jun 19-Jun 13-Aug 27-Aug 10-Sep 24-Sep 09-May 23-May 200 C) Conway

150

100

0 15-Jul 09-Oct 18-Jun 12-Aug 26-Aug 09-Sep 23-Sep 23-May

Figure 3. May to October 2013 seasonal trends in volumetric dry biomass at Belleville, Hay Bay and Conway in the Bay of Quinte. Bosminids include Eubosmina and Bosmina and Daphnia includes D. galeata mendotae and D. retrocurva. "Other Herb. Clad." represents the remaining herbivorous cladocerans, particularly Ceriodaphnia and Chydorus sphaericus. Note the different scale for Hay Bay, and the reduced sampling frequency at Conway.

Hay Bay

In 2013, SWM zooplankton biomass at HB (150 mg m-3) was almost three times higher than at B, but was similar to the relatively stable HB 2004 to 2012 mean of 158 mg m-3 (Fig. 1B). Prior to 1991, variability was caused by changes in both cladoceran and cyclopoid biomass, whereas in recent years cyclopoids have been lower and more stable. Mean cladoceran biomass in 2013 was 115 mg m-3 representing 77% of the total biomass. This indicates a recovery compared to the previous year when cladocerans had reached their second lowest level since 1975. This group has always been dominant in the Bay of Quinte, and Lake Ontario in general, in the 1980s and much of the 1990s. Mean cyclopoid biomass was stable between 2012 (22 mg m-3) and 2013 (29 mg m-3), each representing 19% of the total biomass. Veliger biomass peaked at HB in 2008 (22 mg m-3) and has shown a declining trend since. Excluding 2008, veliger biomass has averaged 4.5 mg m-3 since 1995 when it had recently invaded. The 2013 mean of 3.2 mg m-3 (2.1%) was similar to this mean. Biomass contributions by both calanoids and rotifers were small (2.7 mg m-3 and 1.1 mg m-3, respectively). Rotifer biomass has declined steadily at HB since 2010, although the 2013 mean was similar to values in the mid-2000s (Fig. 2B). The dominant rotifers were Asplanchna sp., Trichocera cylindrica, P. vulgaris and K. cochlearis. There were several biomass peaks through the season at HB, the highest reaching 431 mg m-3 on 23-May and 383 mg m-3 in late July (Fig. 2B). Bosmina was one of the dominant taxa through most of the sampling season, except in mid-summer when it was largely replaced by the somewhat larger E. coregoni. The cladocerans D. galeata mendotae and D. retrocurva began climbing in June, with a peak on July 31st. However, their biomass peaked later and nearly matched the entire zooplankton biomass at B on that date; HB had more than double the biomass of B on July 31st. The herbivorous cladoceran Ceriodaphnia appeared in mid-July, and C. sphaericus in August and September. Veliger larvae peaked at the end of August (11 mg m-3), whereas cyclopoids remained at fairly low levels through the season.

250 Cercopagis Belleville 200 Hay Bay ) 3 - Conway 150

100 SWMDensity(No.m 50

0 5 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013

4.5 Bythotrephes 4 ) 3 - 3.5 Belleville

3 Hay Bay

2.5 Conway 2

1.5 SWMm Density(No. 1

0.5

0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013

Figure 4. Seasonal weighted mean density (SWM) of the invasive predatory cladocerans Cercopagis and Bythotrephes in the Bay of Quinte. Note that 2010 to 2012 was not sampled at C, and 2009 was not sampled at HB.

Total seasonal production has shown small fluctuations at HB since 2000, largely driven by changes in cladoceran production (Fig. 1E). In 2013, production rose to 2 234 mg m-3. This was similar to the 2002-2012 mean of 2 088 mg m-3. In 2013, cladocerans, cyclopoids and veligers contributed 87%, 11% and 2%, respectively, to the total. Unlike B, bosminids only contributed 47% of the production total. Rotifers contributed only 25 mg m-3 at HB in 2013, or 1.1% of the total zooplankton + rotifer production. The SWM density at Hay Bay in 2013 was 158.5 no. L-1. Not surprisingly, the small taxa were the most numerically dominant - Bosmina, followed by C. sphaericus, juvenile cyclopoids (nauplii and copepodids) and veligers. Other important, larger taxa included D. retrocurva, E. coregoni, Ceriodaphnia, D. galeata mendotae and the copepod T. extensus (Tab. 3). Densities of the invasive cladoceran Cercopagis pengoi remained low at HB and B in 2013, and the spiny water flea Bythotrephes was not found at either station (Fig. 4). Small cladoceran taxa dominated the Hay Bay zooplankton community in 2013, and mean cladoceran length was only 0.38 mm (Fig. 5). This was similar to other low values seen over the last decade. Values were generally higher in the late 1980s and 1990s.

0.6

0.55

0.5

0.45

0.4

0.35

0.3 Belleville Hay Bay

CladoceranLength(mm) Body 0.25 Conway 0.2

1976 1979 1984 1987 1990 1993 1996 1999 2002 2005 2008 2011 Figure 5. Trends in annual mean cladoceran length (mm) at Belleville and Hay Bay and Conway from 01 June to 06 October, 1975 to 2013. Note that 1977 and 1978, and Hay Bay in 2009 and Conway from 2010 to 2012 were not sampled. The proposed target size of 0.45 mm is shown.

Conway

After three years without seasonal sampling, C was sampled in 2013 a total of 8 times. Unless indicated otherwise, the data presented here were results from the Schindler traps calculated the “traditional” way. When Schindlers were not taken (23-May and 23-Sep), net haul data were used. This method makes the data as comparable as possible to the previous years. The SWM zooplankton biomass at C in 2013 was 66.3 mg m-3 (Fig. 1C), which was actually higher than the value at B. Between 2007 and 2008, the biomass mean at C jumped from 36 mg m-3 to 62 mg m-3 and has not returned to pre-2008 values. The 2013 biomass value was the second highest observed in the lower bay in the last 20 years. The biggest contributing factor to this increase was veliger biomass, which rose from 6% to 29% of the total biomass. In 2013, veligers contributed 15 mg m-3 or 23%. Cladocerans and cyclopoids each contributed about 25 mg m-3 (37%). Mean cyclopoid biomass dramatically increased in 2013, to levels more typically seen prior to 1997. Between 1998 and 2009, cyclopoids contributed on average 19% of total biomass at C. In contrast, calanoids remained only a small contributor at C in 2013, adding only about 1.6 mg m-3 to total biomass (2%). Rotifers were not collected at C. Biomass peaked at C on multiple occasions in 2013, with the highest reaching 143 mg m-3 on 18-Jun. June biomass was dominated by Bosmina and cyclopoids, particularly juveniles and some Diacyclops thomasi. These taxa gradually declined during the rest of the year. The dominant adult cyclopoid was Mesocyclops edax in mid-summer, and T. extensus later in the season. There were two later biomass peaks on 09-Sep (109 mg m-3) and 09-Oct (67 mg m-3), which were comprised mostly of veligers. From August until October, veligers represented between 15% and 83% of the total biomass at C in 2013. D. retrocurva and D. galeata mendotae peaked in August, representing 21% of total biomass. Calanoids and other cladocerans remained at fairly low levels throughout the season. Throughout Project Quinte, total seasonal production at C has consistently been much lower than the Upper and Middle Bay stations (Fig. 1F). Production to biomass (P:B) ratios tend to decline from the upper to lower bays. In 2013, these values were 23.4 at B, 14.9 at HB and 10.8 at C. In 2013, production at C was 719 mg m-3. Since 2008, levels have been consistently higher than most years since 1992, largely due to the increased contribution by veligers. In

2013, veligers contributed 32% of the total production, with cladocerans and cyclopoids contributing 42% and 25% respectively. Rotifer data were not available at C in 2013. The SWM density at Conway in 2013 was 85.56 no. L-1. As at the other stations, the most numerically dominant species at C were veligers, Bosmina, cyclopoid copepodids, and C. sphaericus. Other important, larger taxa included D. retrocurva, E. coregoni, Ceriodaphnia, and D. thomasi (Tab. 3). The invasive Cercopagis pengoi was found at Conway in low numbers (0.003-0.006 no. L-1), except for a spike on 13-Jul, when 0.27 no. L-1 were found. Bythotrephes reached its highest abundance since it was accurately counted starting in 2000 (Fig. 4). As at Hay Bay, small cladoceran taxa dominated the Conway zooplankton community in 2013, and mean cladoceran length was only 0.35 mm (Fig. 5), similar to values from the late 1980’s. Zooplankton densities calculated using the net samples were often comparable to those values obtained from the Schindler traps (Fig. 6). However on some dates, the net estimates were considerably lower, for example bosminids on 18-Jun and 15-Jul, Daphnia on 15-July and 12-Aug, calanoids on 09-Oct and veligers on 90-Sep. Veliger and nauplii comparisons between the gear types are not valid on 09-Oct due to the use of a larger mesh net. Even when the same mesh size was used on the remaining dates, nauplii densities were usually greater using the Schindler trap. Cyclopoid densities also tended to be higher in the Schindlers than the nets later in the season. The only example where the net densities were substantially higher occurred with veligers on 12-Aug. Reassuringly, there were often few differences between the Shindler estimates calculated using the “traditional” method, which may over-represent epilimnetic taxa, and the “proper” weighting method. Two exceptions occurred when bosminids were abundant (18-Jun and 15- Jul). Since bosminids tend to be epilimnetic, the “traditional” method appears to have overestimated their abundance compared to the “proper” method. The traditional method may also have overestimated Daphnia on 12-Aug, and calanoids on 09-Oct. It is interesting that the two estimates were generally similar for cyclopoids, which are often thought to reside deeper in the water column, and therefore may be under-represented by the traditional method. The only time this appeared to occur was on 15-Jul, when the “proper” cyclopoid densities were higher. It also appears that the “traditional” method may under-represent veligers later in the season, the time

51 when they are most abundant. In 2013, fall veliger densities appeared to be greater deeper in the water column.

40000 Bosminids 16000 35000 14000 Cyclopoids S-traditional 30000

) 12000 3 - S- proper 25000 10000 20000 net 8000 15000 6000 Density (No.m Density 10000 4000 5000 2000 * * * * 0 0 7000 8000 Daphnia Date Date 6000 7000 Calanoids

) 5000 6000 3 - 4000 5000 4000 3000 3000

Density (No.m Density 2000 2000 1000 1000 * 0 * * 0 * 400000 16000 Date Date 350000 Veligers 14000 Nauplii 300000 ) 12000 3 - 250000 10000 200000 8000 150000 6000 Density (No.m Density 100000 4000 50000 2000 T T * T T * 0 * 0 * 23-May 18-Jun 15-Jul 12-Aug 26-Aug 09-Sep 23-Sep 09-Oct 23-May 18-Jun 15-Jul 12-Aug 26-Aug 09-Sep 23-Sep 09-Oct Date Date

Figure 6. Comparison of 2013 dominant zooplankton taxa in the total water column at C using Shindler traps and net hauls. “Net” represents densities derived from 64-µm net hauls taken of both the epi and MH, except for 23-May and 09-Oct when only a total water column haul was taken with a 153-µm net (T). This coarser mesh net catches fewer veligers and nauplii. “S-traditional” represents results from a Shindler Patalas trap fitted with a 64-µm codend, where the epi and MH samples were weighted using the traditional Schindler distribution. This method over-represents the epilimnion. “S-proper” estimates densities by weighting the Schindler strata samples according to a method that better represents the deeper, under-sampled, proportion of the water column. Schindlers were not taken on 23-May or 23-Sep., as shown by *.

Discussion

Overall, the zooplankton community in the Bay of Quinte has fluctuated markedly over time. The reasons for fluctuations over the past 38 years likely include changes in phosphorus loading, water temperature, water clarity, invasive species, variations in the phytoplankton community, and fish predation (Bowen and Johannsson, 2011; Bowen and Gerlofsma, 2012; Johannsson and Bowen, 2012; Nicholls and Carney, 2011). Zooplankton biomass and production dropped precipitously in 1992 at all three stations, and has generally remained lower in the last two decades than during the peak pre-dreissenid period of 1982 to 1991. Biomass and community structure was relatively stable at B between 2008 and 2012, although biomass dropped markedly in 2013 due to a decline in cladocerans. Cladoceran production has also fluctuated at this station in recent years. In 2013, planktivorous fish densities appeared to be very high at B as shown by the Ontario Ministry of Natural Resources (OMNR) bottom trawling program (OMNR, 2014). This program indicates that in 2013, counts of alewife, white perch and gizzard shad were at their highest levels in the last twelve years. Other planktivores in the upper bay include shiners and the juveniles of many species such as yellow perch and even walleye. Densities of the large cladoceran D. galeata, a preferred prey item, were about an order of magnitude below their 1999 to 2011 mean value. Conversely, very high cladoceran production at this station in 2012 was driven by high densities and egg production by D. galeata mendotae in June and July. This was likely due to a combination of a good supply of food and low fish predation during that period, which probably allowed this species to thrive. In contrast to 2012, fish planktivore densities appeared to be lower in the middle bay than in the upper bay in 2013 (OMNR, 2014), which may help explain the unusually low zooplankton biomass and production at B relative to HB. It is noteworthy that in 2013, total biomass was higher at C than at B, largely due to the higher contribution by cyclopoids and dreissenid veligers in the lower bay. When veligers were excluded, biomass values were very similar at the two stations (Tab. 2). However, total zooplankton production and the P:B ratio was still higher at B than at C. This is in part due to the warmer temperatures in the upper bay which enhance production, especially early in the summer. The upper bay is also largely comprised of herbivorous cladocerans, and production for a given biomass tends to be higher for this group

53 than for copepods and veligers. Cladoceran populations can rise and fall very quickly to take advantage of suddenly appearing food sources such as blooms of edible algae, largely due to their ability to reproduce rapidly by cloning (parthenogenesis). As a result, changes in both the abundance and composition of algal communities also influence production of herbivorous zooplankton such as Daphnia. Blooms of inedible, potentially toxic cyanophytes and other large colonial algae frequently occur in the upper and middle reaches of the Bay of Quinte in late summer and early fall (Munawar et al., 2012; Nicholls and Carney, 2011). In addition to potentially interfering with the filtering apparatus of cladocerans (e.g., de Bernardi and Giussani, 1990), these filamentous“ bloom” taxa may outcompete smaller, more edible phytoplankton and cause food limitation in zooplankton. On the other hand, large numbers of cladocerans and their competitors, dreissenid mussels, have the ability to “graze down” desired forms of phytoplankton and can contribute to improved water clarity. Bay of Quinte lower food web sampling complemented the large-scale 2013 Lake Ontario CSMI program (Cooperative Science and Monitoring Initiative, 2013). In general, cyclopoid copepod biomass, and the proportion of cyclopoids, appeared to be higher in 2013 relative to the earlier CSMI program in 2008 (Rudstam et al., 2015). This was also observed in the lower Bay of Quinte where mean cyclopoid biomass reached its highest level since the mid-1990s. However, this trend did not appear to carry through into the middle and upper reaches. Veligers continue to be a major contributor to zooplankton biomass in nearshore areas of the lake, including the lower Bay of Quinte. Their role in the planktonic food web needs further study in Lake Ontario and its embayments. Another finding of the 2013 CSMI program was that the predatory spiny water flea, Bythotrephes, was virtually absent in most Lake Ontario net hauls, although it continued to be found in the Kingston Basin. Bythotrephes reached densities of 0.024 no. L-1 at C in early October 2013, which is the highest density observed for this species in the Bay. It prefers open water and is generally not observed the middle or upper reaches. Bythotrephes are readily consumed by alewife (Mills et al., 1992), and high densities of this fish species can effectively supress this invasive zooplankton (Pothoven et al., 2007). This appeared to be the case in the open waters of Lake Ontario in 2013, but not in the Kingston Basin or the lower bay of Quinte.

Therefore, these areas may serve as a refuge for this species when faced with high predation in the main basin of the lake. In 2013, monthly sampling was resumed at C following a three year gap. It was felt that sampling this station was important given recent changes in the Lake Ontario zooplankton community. It was noteworthy that algal blooms at C were first evident in the zooplankton samples in mid-July, and continued through the remainder of the sampling period. This was supported by anecdotal field observations of blooms. Algal blooms in Quinte are usually most evident in the upper and middle reaches, and blooms of this intensity are not commonly seen in the lower bay. Depending on its edibility, this abundance of algae probably contributed to the higher than average zooplankton biomass at C in 2013. Comparison of vertical net hauls and Schindler traps was undertaken at C in 2013 to better understand biases associated with gear types and the depths sampled. At times there appeared to be under-representation of deeper taxa such as cyclopoids and over-representation of epilimnetic taxa such as Bosmina in the Schindlers. Sometimes catches in the nets were much lower than in the Schindlers, especially for abundant taxa. Reasons for this are unclear but may relate to problems in estimating net efficiency. At other times, the two gear types were fairly comparable. This shows the importance of standardizing sampling methods and consistency in long-term studies such as Project Quinte. The use of different gear types and vertical distribution of zooplankton will be investigated further in 2014. In summary, zooplankton biomass in the Bay of Quinte is likely determined by a number of factors, including both top-down and bottom-up effects. These include changes in the abundance of planktivorous fishes such as alewife, invertebrate predators, and phytoplankton community composition, such as domination by inedible filamentous cyanophytes.

References

Bowen, K.L, Gerlofsma, J., 2012. Zooplankton in the Bay of Quinte – 2009 and 2010. In: Monitoring Report #21. Project Quinte Annual Report 2010, pp. 75-98. Bay of Quinte Remedial Action Plan, Kingston, ON, Canada.

Bowen K.L., Johannsson, O.E., 2011. Changes in zooplankton biomass in the Bay of Quinte with the arrival of the mussels Dreissena polymorpha and D. rostiformis bugensis, and the predatory cladoceran Cercopagis pengoi: 1975-2008. Aquat. Ecosyst. Health Mgmt. 14(1), 44-55.

Bowen, K.L, Gerlofsma, J., 2010. Zooplankton in the Bay of Quinte -2008. In: Monitoring Report #19. Project Quinte Annual Report 2008, pp. 63-82. Bay of Quinte Remedial Action Plan, Kingston, ON, Canada. de Bernardi, R., Giussani, G., 1990. Are blue-green algae a suitable food for zooplankton? An overview. Hydrobiologia 200/201, 29-41.

Hillbricht-Ilkowska, A., Stanczykowska, A., 1969. The production and standing crop of planktonic larvae of Dreissena polymorpha Pall. in two Mazurian lakes. Pol. Arch. Hydrobiol. 16, 193-203.

Johannsson, O.E, Bowen, K.L., 2012. Zooplankton production in the Bay of Quinte 1975–2008: relationships with primary production, habitat, planktivory, and aquatic invasive species (Dreissena spp. and Cercopagis pengoi). Can. J. Fish. Aquat. Sci. 69, 2046–2063.

Cooperative Science and Monitoring Initiative, 2013. Lake Ontario 2013 Cooperative Science and Monitoring Initiative (CSMI) Progress Report, http://www.dec.ny.gov/docs/water_pdf/csmi2013progrpt.pdf, accessed June, 2015.

McCaulley, E., 1984. The estimation of the abundance and biomass of zooplankton in samples. In: Downing, J.A., Rigler, F.H. (Eds.), A Manual on Methods for the Assessment of Secondary Productivity in Fresh Waters, Second Edition, pp. 228–265. Blackwell Scientific Publications, Oxford.

Mills, E.L., O'Gorman, R. DeGisi, J., Heberger, R.F., House, R.A., 1992. Food of the alewife (Alosa pseudoharengus) in Lake Ontario before and after the establishment of Bythotrephes cederstroemi. Can. J. Fish. Aquat. Sci. 49, 2009-2019.

Munawar, M. Fitzpatrick, M., Niblock, H., Lorimer, J., 2012. The spatial distribution of microbial-planktonic communities and size fractionated primary Productivity in the Bay of Quinte, 2010: A brief overview. In: Monitoring Report #21. Project Quinte Annual Report 2011, pp. 69-74. Bay of Quinte Remedial Action Plan, Kingston, ON, Canada.

Nicholls, K.H., Carney, E.C., 2011. The phytoplankton of the Bay of Quinte, 1972–2008: point- source phosphorus loading control, dreissenid mussel establishment, and a proposed community reference. Aquat. Ecosyst. Health Mgmt. 14(1), 33–43.

Ontario Ministry of Natural Resources, 2014. Lake Ontario Fish Communities and Fisheries: 2013 Annual Report of the Lake Ontario Management Unit. Ontario Ministry of Natural Resources, Picton, Ontario, Canada.

Paloheimo, J.E., 1974. Calculation of instantaneous birth rate. Limnol. Oceanogr. 19, 692-694.

Pothoven, S.A., Vanderploeg, H.A., Cavaletto, J.F., Krueger, D.M, Mason, D.M., Brant, S.B., 2007. Alewife planktivory controls the abundance of two invasive predatory cladocerans in Lake Michigan. Freshwater Biology 52, 561–573.

Rudstam, L. G., Holeck, K. T., Bowen, K. L., Watkins, J. M., Weidel, B. C., Luckey, F. J., 2015. Lake Ontario zooplankton in 2003 and 2008: Community changes and vertical redistribution. Aquat. Ecosyst. Health Mgmt. 18(1), 43-62.

Sprung, M., 1984. Physiological energetics of mussel larvae (Mytilus edulis). I. Shell growth and biomass. Mar. Ecol. Prog. Ser, 17, 283-293.

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FISH POPULATIONS IN THE BAY OF QUINTE, 2013

J. A. Hoyle

Ontario Ministry of Natural Resources, Lake Ontario Management Unit Glenora Fisheries Station, 41 Hatchery Lane, R.R. #4, Picton, Ontario, K0K 2T0

Introduction

This report updates long-term abundance trends of Bay of Quinte fish populations to 2013. Fish community sampling programs have existed in the Bay of Quinte since the 1950s, initially using gill nets and later adding trap nets and bottom trawls. While each gear has unique bias with respect to habitat and species selectivity, collectively, these programs provide a very comprehensive data set within which the long-term dynamics of the Bay of Quinte fish community can be examined. Observed major changes in species dominance and abundance have been related to large-scale ecological events associated with nutrient levels and invasive species. Hurley and Christie (1977) described a Bay of Quinte fish community depreciated by the impacts of cultural eutrophication in the 1960s and 1970s. Hurley (1986) evaluated the impact of point-source phosphorus control implementation in municipal sewage treatment plants, during the winter of 1977-78, on Bay of Quinte fish populations. Hoyle et al. (2012) documented major changes in the fish community following dreissenid mussel invasion in the early 1990s, and related these changes to increased water clarity and the return of submerged aquatic macrophytes to vast areas of the bay. Round goby invaded the Bay of Quinte in 1999 and by 2003 provided the missing food-web link between dreissenid mussels and piscivores but have not led to major changes in fish species dominance to date.

Sampling Programs

Long-term fish community sampling programs, first initiated on the Bay of Quinte in the late 1950s (Hurley 1986), provide intensive geographic coverage from the mouth of the Trent River

58 in the upper bay to Lake Ontario in the lower bay. Gill nets and bottom trawls sampled offshore habitats while trap nets were used nearshore. The nearshore is distinguished from the offshore at approximately 5-m depth. Bottom trawls sampled all sizes of fish, particularly smaller ones, whereas gill nets and trap nets more effectively sampled large-bodied fish. Here, only summer (June-September) data are presented. Detailed sampling methods have been reported elsewhere (e.g., Hoyle, 2012).

Data Analysis and Presentation

In this summary report, only the catches of dominant species, as ranked in Hoyle et al. (2012), are included. Species-specific catches, by number, were averaged across geographic area of the bay (except for trap nets; only upper Bay of Quinte is reported here since the lower Bay is not sampled annually), and then summarized and reported as an overall mean across years prior to 1978 (phosphorus time-stanza) and from 1978-1994 (post-phosphorus time-stanza). After 1994 (dreissenid mussel time-stanza), mean annual catches were reported. These results are reported in Tables 1, 2 and 3 for gill nets, bottom trawls, and trap nets, respectively.

Walleye

Walleye is one of the most important members of the Bay of Quinte fish community. As the dominant piscivore, it plays a pivotal role in fish community trophic structure. The Bay of Quinte recreational fishery is centered on Walleye. Abundance trends in gill nets (juvenile and adult fish) and bottom trawls (YOY) are illustrated in Fig. 1 and 2.

Table 1. Catch per gill net for major Bay of Quinte fish species. Shown are mean catches for phosphorus (1972-1977) and post-phosphorus (1978-1994) time-stanzas, annual catches from 1995-2013, and the mean for 1995-2013. Yellow White Brown White Gizzard Largemouth Freshwater Round Time-stanza Perch Perch Alewife Bullhead Pumpkinseed Walleye Bluegill Sucker Shad Bass Drum Goby Phosphorus mean 55.652 134.780 222.559 2.228 0.701 0.479 0.005 1.881 40.032 0.013 0.091 - Post-phosphorus mean 108.703 51.352 143.982 2.079 0.257 16.385 - 6.504 19.518 0.002 3.617 - 1995 108.583 28.500 24.736 1.625 0.125 11.833 - 6.542 8.292 - 4.250 - 1996 106.583 12.813 8.563 1.000 0.667 9.979 0.083 5.271 0.083 - 5.979 - 1997 144.417 15.500 16.756 1.650 0.778 6.072 0.222 5.772 0.139 - 6.106 - 1998 183.493 16.872 19.650 2.306 4.806 6.217 0.556 4.850 0.206 - 5.283 - 1999 162.083 22.672 10.644 2.750 3.528 5.178 0.278 4.478 5.528 0.056 3.333 - 2000 129.083 16.350 5.392 2.458 4.333 4.267 0.583 3.333 0.417 - 4.867 - 2001 133.417 7.500 9.167 2.558 6.058 2.633 2.375 4.125 0.667 - 6.433 - 2002 124.167 13.933 3.967 1.875 3.100 4.208 1.250 4.908 2.167 - 2.433 0.417 2003 91.494 23.019 2.556 0.728 0.503 3.900 0.167 2.244 0.667 - 2.258 3.861 2004 76.272 49.217 6.872 0.928 2.153 3.653 0.111 4.169 0.072 - 4.172 10.544 2005 91.250 26.542 23.233 0.453 0.236 2.603 0.833 1.533 14.000 - 3.614 0.475 2006 80.083 75.908 6.600 1.258 1.292 4.667 2.167 2.400 0.250 0.083 9.425 0.050 2007 45.567 45.500 7.700 0.503 0.433 3.311 1.778 1.775 0.442 - 4.381 0.033 2008 62.658 49.744 24.042 0.389 0.583 3.839 1.056 2.461 1.297 - 2.428 0.083 2009 74.256 44.231 29.122 0.222 0.472 4.033 1.850 2.628 - - 2.856 - 2010 89.806 31.267 68.367 0.000 0.306 4.578 2.222 4.206 1.056 0.056 1.739 0.017 2011 55.847 32.111 62.981 0.072 0.528 3.253 2.319 2.100 29.478 - 2.822 - 2012 21.606 48.725 39.525 0.167 0.333 2.306 0.389 3.267 3.736 - 1.353 0.017 2013 19.464 67.364 30.072 0.389 0.833 6.831 3.778 2.158 1.792 - 2.681 - 1995-2013 mean 94.744 33.040 21.050 1.123 1.635 4.914 1.159 3.591 3.699 0.010 4.022 0.816

Table 2. Catch per trawl for major Bay of Quinte fish species. Shown are mean catches for phosphorus (1972-1977) and post-phosphorus (1978-1994; no trawling in 1989) time-stanzas, annual catches from 1995-2013, and the mean for 1995-2013. Yellow White Brown White Gizzard Largemouth Freshwater Round Time-stanza Perch Perch Alewife Bullhead Pumpkinseed Walleye Bluegill Sucker Shad Bass Drum Goby Phosphorus mean 15.84 318.32 388.87 8.42 0.08 0.07 - 0.20 29.11 0.01 0.01 - Post-phosphorus mean 90.95 95.30 443.15 5.80 0.58 9.28 0.00 4.80 65.50 - 1.81 - 1995 312.28 192.33 139.11 14.41 4.45 10.45 0.13 1.28 183.51 0.40 5.39 - 1996 99.36 204.96 41.29 8.42 11.56 5.04 0.15 1.74 18.56 0.08 6.24 - 1997 205.92 298.47 60.16 7.98 12.78 3.00 0.34 2.08 14.23 0.59 3.79 - 1998 263.41 44.01 30.20 18.25 10.49 2.68 0.29 2.54 7.99 0.05 8.19 - 1999 441.81 229.92 294.03 19.11 29.26 3.45 0.05 2.07 62.77 0.46 3.82 - 2000 307.43 126.28 261.93 10.89 62.81 1.63 11.31 5.69 168.93 0.19 2.39 - 2001 373.09 10.53 216.12 21.60 32.25 5.71 10.56 46.30 12.15 0.29 18.90 0.15 2002 429.37 141.05 44.84 10.05 13.88 2.20 1.23 10.06 26.20 0.97 8.97 1.69 2003 211.61 76.28 38.97 8.43 19.07 4.48 1.73 4.47 8.62 0.41 6.31 106.11 2004 253.92 713.83 63.23 10.10 11.29 4.39 0.14 5.55 53.50 0.10 6.50 35.23 2005 236.61 137.13 75.22 7.53 5.96 2.68 0.88 1.82 10.73 0.68 34.87 72.46 2006 315.96 445.28 156.00 12.07 18.46 3.92 1.06 1.55 15.18 0.53 37.57 17.42 2007 457.53 89.04 320.29 6.44 12.07 5.64 2.00 5.64 8.89 0.02 85.58 72.55 2008 263.73 190.71 319.64 1.80 16.34 10.61 1.75 1.30 47.62 0.73 5.97 35.92 2009 181.14 405.91 258.63 2.05 9.18 4.69 0.98 0.32 31.48 0.84 6.63 33.13 2010 161.74 46.39 394.99 1.96 11.04 4.75 1.02 1.70 57.32 1.77 12.64 51.06 2011 355.76 265.18 416.25 0.84 8.94 7.32 2.38 2.01 88.10 1.75 6.01 89.14 2012 89.72 96.72 224.21 1.63 4.88 3.49 0.84 1.50 142.90 2.32 6.86 72.00 2013 157.34 374.38 295.25 4.74 13.38 1.78 0.70 2.21 91.62 1.36 7.49 35.26 1995-2013 mean 269.39 215.20 192.14 8.86 16.22 4.63 1.98 5.25 53.28 0.71 14.43 32.74

Table 3. Catch per trap net (upper Bay of Quinte only) for major Bay of Quinte fish species. Shown are mean catches for phosphorus (1969-1971, 1976) and post-phosphorus (1978-1979, 1981-1983, 1985-1988) time-stanzas, annual catches from 2001-2005 and 2007-2013 and the mean for the years 2001-2005, 2007-2013. Yellow White Brown White Gizzard Largemouth Freshwater Round Time-stanza Perch Perch Alewife Bullhead Pumpkinseed Walleye Bluegill Sucker Shad Bass Drum Goby Phosphorus mean 19.95 279.74 546.11 18.08 24.40 0.48 0.83 1.93 0.12 0.02 0.02 - Post-phosphorus mean 11.06 100.19 50.25 20.33 19.94 12.24 0.64 2.07 1.27 0.07 0.55 - 2001 3.75 2.19 - 167.67 89.39 3.17 169.58 1.03 1.11 2.47 6.36 - 2002 3.42 2.89 - 95.83 73.08 2.47 142.64 1.47 1.44 6.11 3.31 - 2003 1.94 7.69 - 37.33 26.94 2.22 66.25 1.72 2.00 7.92 3.81 - 2004 0.83 3.67 - 20.83 15.33 2.56 75.19 1.25 0.06 6.08 2.14 - 2005 1.00 2.75 - 17.89 15.97 2.14 44.44 1.11 20.42 2.75 4.36 - 2007 4.72 4.61 - 7.25 18.61 1.61 63.92 0.44 0.39 4.53 1.25 - 2008 7.00 4.31 - 6.42 18.14 2.50 159.11 0.92 1.00 5.39 1.17 - 2009 2.64 3.86 - 2.56 23.42 1.75 71.75 0.64 0.06 4.33 1.89 - 2010 6.11 1.69 - 10.56 29.08 2.53 61.50 0.44 0.64 4.25 1.97 - 2011 6.25 3.75 - 13.69 37.53 2.36 136.03 0.42 0.14 10.39 1.67 - 2012 1.31 3.58 - 7.11 28.11 1.44 74.92 0.72 0.33 2.72 2.19 - 2013 2.69 19.42 15.28 14.72 7.56 53.56 0.86 0.06 4.33 0.94 2001-2013 mean 3.47 5.03 - 33.53 32.53 2.69 93.24 0.92 2.30 5.11 2.59 -

Gill net Gillcatch 10

0 1972 1973 1974 1975 1976 1977 1978 1979 1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 Phosphorus Post-phosphorus Dreissenid Gillnet Figure 1. Walleye catch in gill nets in the Bay of Quinte from 1972-2013.

40.3 20

15 (YOY only) (YOY

5 Trawl catch catch Trawl

0 trawlingno 1972 1973 1974 1975 1976 1977 1978 1979 1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 Phosphorus Post-phosphorus Dreissenid

Figure 2. Young-of-the-year (YOY) Walleye catch in bottom trawls in the Bay of Quinte, 1972- 2013 (no trawling in 1989).

References

Hurley, D.A., 1986. Fish populations in the Bay of Quinte, Lake Ontario, before and after phosphorus control. In: Minns, C.K., Hurley, D.A., Nichols, K.H. (Eds.), In: Project Quinte: point source phosphorus control and ecosystem response in the Bay of Quinte, Lake Ontario. Canadian Special Publication of Fisheries and Aquatic Science 86, 201-214.

Hurley, D.A., Christie, W.J., 1977. Depreciation of the warmwater fish community in the Bay of Quinte, Lake Ontario. Journal of the Fisheries Research Board of Canada 34, 1849-1860.

Hoyle, J.A., Bowlby, J.N., Brousseau, C., Johnson, T., Morrison, B.J., Randall, R., 2012. Fish Community Structure in the Bay of Quinte, Lake Ontario: The Influence of Nutrient Levels and Invasive Species. Aquat. Ecosyst. Health Mgmt., 15(4), 370-384.

Hoyle, J.A., 2012. Bay of Quinte Fish, 2010. In: Monitoring Report #21. Project Quinte Annual Report 2010, pp. 116-131. Bay of Quinte Remedial Action Plan. Kingston, Ontario, Canada.

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2013 BAY OF QUINTE ALGAE WATCH PROGRAM

C. McClure¹, S. Watson², B. Keene¹, D. Eastcott¹, and L. Lambert¹

¹Quinte Conservation, Belleville, ON, K8N 4Z2 ²Watershed Hydrology and Ecology Research, Environment Canada, Canada Centre for Inland Waters, Burlington, ON, L7R 4A6

In an attempt to understand the nature of algae blooms on the Bay of Quinte and conditions leading up to a bloom, Quinte Conservation, in partnership with the Bay of Quinte Restoration Council, Ministry of the Environment and Environment Canada, undertook a 4-year intensive sampling program of the Bay of Quinte and nine tributaries. An analysis of several chemical and physical parameters of the Bay of Quinte was completed to look at temporal and spatial trends that might help to establish the conditions leading to a bloom. In 2013/14, the study team also tracked nutrient input from the Napanee River system in an attempt to determine the context of high nutrient loadings into the Bay of Quinte from that river. Figures 1 and 2 show the location of the sampling sites for both open water (in the Bay of Quinte) and the tributaries to the Bay.

Figure 1. Open water sampling stations.

Figure 2. Tributary monitoring stations.

Results

The Bay of Quinte water quality sampling program was carried out for a period of four years from 2010 to 2013. An additional month of sampling was added for April of 2014 to capture a runoff event in the Napanee River. During the course of this study well over 40,000 data points were collected. There were 55 open water sampling events at nine stations and 83 sampling events at nine tributary stations. Five special sampling series were undertaken including two at Trent River, two at Potter Creek and one on the Napanee River.

The mean monthly TP concentration in the Bay of Quinte was developed from the four years of data from June to November (

Figure to Figure ). The figures highlight the spatial changes in concentration throughout the Bay. A TP concentration below 0.025 mg L-¹ will show blue, above 0.025 mg L-¹ will transition to green, 0.040 mg L-¹ will show as yellow and above 0.05 mg L-¹ appears red. Highest concentrations of TP are observed in the Napanee to Picton Harbour area peaking in August and September. High levels seem to persist in Picton Harbour through October.

Figure 3. Bay of Quinte total phosphorus concentrations, June average (mg L-¹).

Figure 4. Bay of Quinte total phosphorus concentrations, July (mg L-¹).

Figure 5. Bay of Quinte total phosphorus concentrations, August (mg L-¹).

Figure 6. Bay of Quinte total phosphorus concentrations, September (mg L-¹).

Figure 7. Bay of Quinte total phosphorus concentrations, October (mg L-¹).

Figure 8. Bay of Quinte total phosphorus concentrations, November (mg L-¹).

When the sampling seasons are compared on a monthly basis (Figure ), one observes that the 2012 year saw the highest peak concentrations in August and September (above 0.04 mg L-¹).

2010 2011

0.09 0.09 0.08 0.08 0.07 0.07 0.06 0.06 0.05 0.05 0.04 0.04

PPUT (mg/L) PPUT 0.03 (mg/L) PPUT 0.03 0.02 0.02 0.01 0.01 0.00 0.00 1 2 3 4 5 6 7 8 9 10 11 12 13 1 2 3 4 5 6 7 8 9 10 11 12 13 MONTH MONTH

2012 2013

0.09 0.09 0.08 0.08 0.07 0.07 0.06 0.06 0.05 0.05 0.04 0.04

PPUT (mg/L) PPUT 0.03 (mg/L) PPUT 0.03 0.02 0.02 0.01 0.01 0.00 0.00 1 2 3 4 5 6 7 8 9 10 11 12 13 1 2 3 4 5 6 7 8 9 10 11 12 13 MONTH MONTH

Figure 9. Total phosphorus monthly concentration trend by year.

One of the delisting criteria for the “Eutrophication and Undesirable Algae” BUI is to reduce the average total phosphorus concentration from 40 µg L-¹ to 30 µg L-¹ in the upper Bay of Quinte for the period of May to October. The average annual (May to October) total phosphorus concentrations for the upper Bay of Quinte stations are given in Table . Only one sampling season, 2012 spanned the entire May to October period. During 2012, the upper bay averaged 29.5 µg L-¹ of total phosphorus. The Moira River Mouth measured 30.9 µg L-¹ – just above the 30-µg L-¹ threshold. The Napanee River Mouth was over 50% higher than the 30 µg L-¹ target with an average concentration of 47.5 µg L-¹. For the other years of study the Napanee River Mouth continued to have an average total phosphorus concentration

71 greater than delisting criteria of 30 µg L-¹. Over the four years of study the seasonal, upper bay average total phosphorus concentration was calculated to be 26 µg L-¹.

Table 1. Seasonal total phosphorus concentrations at upper Bay of Quinte stations by year (µg L-¹). July – July – May – June – Average August October October August by 2010 2011 2012 2013 station Dead Creek Mouth 22.3 20.6 26.1 21.2 22.6 Trent River Mouth 25.3 25.1 22.0 18.5 22.7 Bayside Narrows 30.5 34.2 27.2 20.9 28.2 Moira River Mouth 19.0 26.4 30.9 18.6 23.7 Muscote Bay 24.3 25.4 25.0 22.5 24.3 Salmon River Mouth 18.3 35.0 27.5 18.0 24.7 Napanee River Mouth 33.3 34.1 47.5 36.5 37.9 Upper Bay Average 24.7 28.7 29.5 22.3 26.3

Similar to the phosphorus constituents, the various forms of nitrogen were analysed and were found to be highest in all but total Kjeldhal nitrogen (TKN) at Picton Harbour. Figure 0 shows the ammonia and ammonium as well as the nitrates and nitrites are very high compared to the other stations. This suggests there are local sources in Picton Harbour, which could indicate sewage or agricultural runoff. Long Reach appears to have much higher TKN concentrations than the other stations; however, Long Reach was only sampled in 2010. A special sampling was performed of the Napanee River during the spring melt in April 2014. Runoff is recorded at Napanee River at Camden East 02HM007. Records exist for the Napanee River at this station from 1974 to current. The peak flow in 2014 was the highest flow on record for the gauge. The Napanee River drains a large forested drainage basin with 34% classified as forest cover and 16% as wetland (Source: SOLRIS). The upper reaches of the watershed are dominated by these land types, while the lower reach is mostly agricultural and urban. Owing to the long, narrow shape of the lower reach of the watershed, the Napanee River responds rapidly to precipitation showing as a sharp rise in the hydrograph. The upper reach has substantially more storage and contributes to the hydrograph more slowly.

0.3 0.15

0.2 0.10

NNH34

NNH3F 0.1 0.05

0.0 0.00

HayBay HayBay DeadCrM MoiraRM TrentRM DeadCrM MoiraRM TrentRM LongReach SalmonRM LongReach SalmonRM BaysideNarro MuscoteBayNapaneeRMPictonHarbou BaysideNarro MuscoteBayNapaneeRMPictonHarbou STATION STATION

1.0 0.7

0.9 0.6 0.8 0.5 0.7 0.4 0.6 0.3

NNO23F NNTKUR 0.5 0.2 0.4

0.3 0.1

0.2 0.0

Figure 10. Nitrogen concentrations in mg L-¹ – 2010 to 2013 at open water stations in Bay of Quinte.

Concentrations of total phosphorus were measured 13 times over a period of 28 days at the two stations along the Napanee River. The distance between the two sites is approximately 12 km. The downstream station is an active Provincial Water Quality Monitoring Network station number 17003500202. The upstream station (17003500302) in Camden East was discontinued in 1971. Figure 3 shows the location of the two stations with respect to the watershed area for the Napanee River.

Figure 3. Location of the special flood sampling stations on Napanee River.

The time series of unfiltered total phosphorus concentration for the two stations (Figure 4) saw a peak in concentration occurring on the same date – April 8, 2014. The concentration at the upper station rose to a high of 41 µg L-¹ and the downstream station reached 71 µg L-¹. The peak flow of the Napanee River hydrograph occurred 8 days later on April 16th. However, the initial peak flow from the lower reach occured on April 8th – the same date as the peak in phosphorus concentrations. This pattern where TP concentration follows flow was also observed in the Potter Creek sampling reported in 2012. The upper watershed contribution, appearing as the second peak, coincides with the falling nutrient concentrations. This suggests a first flush type of response to nutrient contribution from the watershed, particularly in the lower, more settled, portion of the watershed. The effect of the upper watershed runoff, when it dominates the flow, is to dilute the nutrient concentrations. A further observation is that the downstream station is more impacted showing higher concentrations than the upper station. This means there are high concentration contributions of total phosphorus between the two stations.

Figure 4. Total phosphorus at two stations along Napanee River, April 2014 flood.

Total loadings were calculated for the days for which concentrations of total phosphorus were available using the single sample point and the average flow for the day. It is estimated that for 12 common days, 1651 kg of total phosphorus was transported in the Napanee River at the downstream station (NR) and 1383 kg for the upper station (NR-CE). The difference between the two stations is 268.6 kg. Possible sources of TP between the two stations include intensive agricultural operations, industrial release and septic systems. It is suggested that future management actions for the Bay of Quinte include some review of possible sources of phosphorus and that opportunities for nutrient management be investigated in this area. Of the phosphorus types analyzed, SRP showed the highest difference between the stations (Figure 5). Over the 12 common days of sampling a total of 313 kg of SRP was calculated to be discharged in the Napanee River at the downstream station vs. 122 kg at the upstream station. This is a factor of approximately 2.6 times greater and it suggests there is a significant source of SRP between the two sites. The nitrogen series shows some difference between the sites (see Figure 6). While the median concentration is higher at the downstream station for the three constituents (NH2, NO23, TKN), the TKN upper quartile is higher at the upper station. Generally, the nitrogen compounds show more impact at the downstream station.

20 80 30

70 15 60 20 50 10

SRP 40 TPF 10 5 30 20

0 10 0 NR NR CE NR NR CE NR NR CE SITE SITE SITE

Figure 5: Special Sampling of Napanee River Phosphorus Series (µg L-¹).

0.06 0.4 0.50

0.05 0.3 0.45 0.04

0.03 0.2 0.40

NH3

NO23

TKNF

0.02 0.1 0.35 0.01

0.00 0.0 0.30 NR NR CE NR NR CE NR NR CE SITE SITE SITE

Figure 6. Special sampling of Napanee River nitrogen series (mg L-¹).

The summer highest TP concentration measured in the Algae Watch program in Napanee was 0.05 mg L-¹ recorded on July 30, 2012. This is compared to 0.043 mg L-¹ high for the Provincial Water Quality Monitoring Network (PWQMN) program around the same period. It is also observed from the graph that a spike of TP was recorded during the spring of 2012. A similar spike is also observed in the flood sampling series in 2014. No corresponding spike in TP is observed, however, in the PWQMN data.

Napanee River - TP (unfiltered) 0.08 90

0.07 80 70 0.06 ¹ ) - L

0.05 50 0.04 40 0.03 Flow (cms) 30

Concentration(mg 0.02 20

0.01 10

0 0

Routine Flood Sampling PWQMN Flow (cms)

Figure 7. Comparison of special flood sampling with routine Algae Watch program sampling and PWQMN on Napanee River.

A monthly loading estimate was prepared for the period between February 2011 and April 2014 using the Algae Watch program data series to compare with loadings estimates based on PWQMN data. The Algae Watch program used a more intensive sampling schedule and contains more dense data than the PWQMN program. Over the period from February 2011 to November 2013 the study team collected 58 samples. In the same period 21 sample results were obtained in the PWQMN program. A comparison of annual load estimate is summarized for 2011 to 2013 in Table .

Table 2. Annual loading of TP (unfiltered) in Napanee River at station 17003500202 in Kg. AW Year PWQMN Routine Diff 2011* 6401 8034 26% 2012 3507 3295 -6% 2013 4187 5645 35% Total 14,095 16,974 20% * Partial year – January 2011 was not estimated

In 2011 and 2013 the total phosphorus load estimated from the PWQMN data is 26% and 35% lower respectively than estimates made from the more densely populated dataset gathered in the Algae Watch program. However, the estimate from 2012 agreed quite closely between the two programs. For the entire period the loading using the Algae Watch program exceed PWQMN by almost 2,900 Kg, or almost 20% higher. The loading estimate rises to 69% when the sampling during the 2014 flood sequence is compared for the month of April 2014 only (3,743 Kg of TP as compared to 2,218 Kg). Therefore, loading calculations based on the PWQMN data would underestimate TP. This would agree with previous findings by the Toronto & Region Conservation Authority (TRCA).

Conclusions

Generally, the study found what is not unexpected that algae abundance (measured by PCRFU and chlorophyll α) varies seasonally in proportion to the presence of total phosphorus – the limiting nutrient. Some years of higher algae abundance coincide with higher concentrations of total phosphorus, however, the pattern is not perfect. The year 2012, during which algal toxin (Microcystin) was found to be most abundant, was also the year of interest for high TP, PCRFU and chlorophyll α values. High TP, PCRFU, chlorophyll α and temperature appear to coincide with high concentrations of toxic algae. A high risk of toxins may be indicated by the coincidence of these four. The study team found that predictive trends that could help identify the onset of an algae bloom were not apparent (McClure et al., 2012). Blooms on the Bay of Quinte during the study were sporadic and not persistent. Blooms also varied in appearance and were suspected to be comprised of more than one species of algae (Watson et al., 2010). While water temperature and total phosphorus generally trend with phycocyanin and chlorophyll α, there was no clear predictor that a bloom would occur. A simple predictive water constituent or physical parameter was not found. Total microcystin, however, does appear to coincide with high TP, PCRFU, chlorophyll α and prolonged high temperature. The tributary data did not provide an adequate explanation for observations within the Bay of Quinte, with the exception of a single year, 2012. In the 2012 Algae Watch report, study members observed the Bay of Quinte was strongly influenced by the characteristics of the Trent

River. However, over the three years of tributary sampling this observation was not consistently true. The river sampling did not reveal any clear indicators of what happened in the Bay. Open water sampling showed that the Bay of Quinte is not homogeneous and that conditions at one sampling station were not reflective of conditions at other stations. Some trending is observed with toxic algae with several parameters, but none were fully predictive of a bloom or of toxic algae conditions. In summary, four parameters measured in open water stations show some correlation with algal toxins in the Bay. These include; total phosphorus, temperature, phycocyanin and chorophyll α. Two 4-parameter plots were prepared to present the data and develop potential relationships. The first is based on temperature, total phosphorus, phycocyanin and total microcystin (Figure 8). The second is developed using temperature, total phosphorus, chlorophyll α and total microcystin (Figure 9). What is apparent from this type of presentation is that the general trend of increasing values is to the upper right quadrant for each of the phycocyanin and chlorophyll α sequences. Thus, during periods of high temperature and high concentrations of total phosphorus, the Bay measures higher phycocyanin and chlorophyll α and experiences higher risk of microcystin presence. These are not hard rules, but general trends. The figures can be useful to provide some concept of risk of the presence of algal toxins in the Bay of Quinte. If temperature in the Bay is very high (say 25 °C) and concentration of total phosphorus is 40 µg L-¹, then one may anticipate a high risk of toxic algae. If researchers do not have total phosphorus values, but have phycocyanin and temperature, then the chart could be read using those two parameters and the risk is interpreted to be high for presence of microcystin. The figures can be understood as risk charts and will be tested in 2014 sampling by researchers.

Figure 8. Risk chart for potential for toxic algae bloom using chlorophyll α, temperature and total phosphorus.

Figure 9. Risk chart for potential for toxic algae bloom using phycocyanin, temperature and total phosphorus.

Water quality was the most degraded in the lower Bay, and specifically within the Picton Harbour and Napanee River Mouth area where samples were found to have among the highest concentrations of nutrients and most chemical parameters investigated. Picton Harbour showed

80 the highest values of chloride, conductivity (as reported by the lab), fluoride, potassium (filtered), magnesium (filtered), sodium (filtered), nitrogen (as ammonia and ammonium), nitrogen (as ammonia), nitrogen (as nitrate and nitrite), total phosphorus (filtered), soluble reactive phosphorus, and sulphate (filtered). Total phosphorus (unfiltered) and total particulate phosphorus were both highest at Napanee River mouth station. High values of particulate organic nitrogen and particulate organic carbon tended to coincide at Napanee River mouth, Moira River mouth and Dead Creek mouth, but the high readings did not repeat each year. The runoff sampling conducted in April 2014 on the Napanee River suggest 69% more TP loading occurred in the month than would have been estimated using only PWQMN data. As high as 35% more TP loading may be occurring on an annual basis in the Napanee River. No clear predictive trends were found that would indicate an algae bloom would occur. Instead, it is postulated that a grouping of measures including; temperature, total phosphorus, chlorophyll α and phycocyanin may provide some general understanding of higher risk of bloom development. Since total phosphorus results must be obtained through laboratory analysis and the turnaround time on results is inadequate to use this parameter for early warning of an algae bloom, the study team developed two charts of risk; one based on phycocyanin and another based on chlorophyll α, where a researcher may use any two of three independent parameters to determine risk of toxic bloom events.

References

Bastien, C., Cardin, R., Veilleux, E., Deblois, C., Warren, A., Laurion, I., 2010. Performance evaluation of phycocyanin probes for the monitoring of cyanobacteria. J Environ Monit.13(1), 110-118.

McClure, C., Watson, S.B. , Keene, B., Eastcott, D., Lambert, L., Ogunlaja, S., 2012. Bay of Quinte Algae Watch Program – Project Report, March 31, 2013. Quinte Conservation, Belleville, ON, Canada.

Quinte Conservation, 2014. 2013 Bay of Quinte Algae Watch Program Project Report. August 20, 2014. Belleville, Ontario, Canada

Watson, S., Munro, J., Carney, E., Munawar, M., 2010, Environment Canada, Bay of Quinte taxa 2010 Preliminary report. Environment Canada, Burlington, ON, Canada.

Edited by

M. Munawar Fisheries & Oceans Canada

J. Lorimer Aquatic Ecosystem Health & Management Society