Effects of Copper Sulfate Application on Zooplankton and Macroinvertebrate Communities in Upground Reservoirs

EFFECTS OF COPPER SULFATE APPLICATION ON ZOOPLANKTON AND MACROINVERTEBRATE COMMUNITIES IN UPGROUND RESERVOIRS

Meghan C. Weaver

A Thesis

Submitted to the Graduate College of Bowling Green State University in partial fulfillment of the requirements for the degree of

MASTER OF SCIENCE

August 2012

Committee: Dr. Jeffrey Miner, advisor Dr. John Farver Dr. Helen Michaels Dr. Joseph Conroy

ABSTRACT

Dr. Jeffrey Miner, advisor

Copper sulfate (25.2% Cu by weight) has been extensively used to control nuisance algae in drinking water storage reservoirs; in upground reservoirs in northwest Ohio, CuSO4 application regimens vary from no application to 600 µg Cu/L/year. While CuSO4 is effective in suppressing algae, it also has documented toxicities to zooplankton and chironomids, which are food resources for stocked sport fish. Between May-August 2010, water, sediment, sediment trap, zooplankton, and macroinvertebrate samples were collected at four upground reservoirs with varying CuSO4 application regimens in order to track the fate of copper and enumerate zooplankton and macroinvertebrate community changes before and after CuSO4 application.

Additionally, to quantify the combined effects of pulsed copper-laden food resources and contaminated sediment on chironomids of different instars, an experiment was conducted with

Chironomus riparius in the laboratory. In the reservoirs, water copper concentration was 2-4 times higher post-application than before; correspondingly, zooplankton biomass and density were depressed by as much as 93% and 87% for at least one week after application.

Furthermore, post-application zooplankton communities were dominated by copepod nauplii, which are an unsuitable food source for fish stocked into these reservoirs. Chironomid density changes appeared to reflect adult emergence rather than CuSO4 application, although standard sampling protocol prevented tracking all instars through time, so the effects of CuSO4 application on chironomid communities were not evident. The sedimentation rate of detritus and algae increased by 36% post-application and contained more than 5000 µg Cu/g dry weight. In the experiment, I found that organisms receiving a pulse of copper-laden algae in the first instar experienced 85% mortality, while organisms fed unspiked algae (controls) experienced 40% iii mortality. Field observations underscore the need for communication between reservoir managers and fisheries managers. It may be possible to time the application of CuSO4 such that both water quality and fishery needs are met. Experimental results suggest that entire chironomid cohorts may be affected by the timing of CuSO4 application. Furthermore, toxicity studies in which chironomids are fed ecologically-unrealistic food may substantively underestimate the effects of contaminated sediment. iv

ACKNOWLEDGEMENTS

Field aspects of this study were supported by the Federal Aid in Sport Fish Restoration

Program (F-69-P, Sport Fish Management in Ohio) administered jointly by the United States

Fish and Wildlife Service and the Division of Wildlife, Ohio Department of Natural Resources.

This project was also funded in part by Bowling Green State University. An especially big thank you goes to both Jeremiah Davis and Randy Williams, without whom I would not have been able to conduct my field work or experiments. Thank you also to Chris Boehler and Ryan Crouch, who not only acted as ballast in the field but also provided feedback and support in the lab, to

Audrey Maran for helping process dozens of zooplankton and macroinvertebrate samples, to

Katie Barlak for assisting in the implementation of summer experiments, and to Summer

Harshbarger for the design of the sediment trap diagram. I am indebted to my advisors for their guidance and feedback throughout this whole process. Finally, I would like to thank my family:

Pat Weaver, Jeanne and Brad Bloomster, Kevin and Brooke Weaver, and Erin Weaver, who have tolerated my less-than-prompt responses to emails, texts, and phone calls during many field seasons and who have always been supportive of my love of nature. v

TABLE OF CONTENTS Page INTRODUCTION ...... 1

METHODS ...... 5

Field Sampling ...... 5

Effects of Copper on Chironomid Instars – Laboratory Experiment ...... 9

RESULTS ...... 14

Field Sampling ...... 14

Laboratory Experiment ...... 19

DISCUSSION ...... 22

Effects of Copper Sulfate Application on Zooplankton Communities ...... 22

Effects of Copper Sulfate Application on Chironomid Communities ...... 24

Effects of Copper Sulfate Application Timing on Chironomid Instars ...... 25

Conclusions and Future Research ...... 27

BIBLIOGRAPHY ...... 29

LIST OF TABLES/FIGURES

Table

1. Wavelengths (nm) and Minimum Detection Levels used in elemental analysis of

sediment and water samples from upground reservoirs using ICP-OES ...... 37

2. Algal community composition (cells/milliliter) at Paulding Reservoir during the months

of May – July 2010 ...... 38

3. Feeding regimen of copper-pulse experiment...... 39

Figure

1. Typical copper sulfate application regimens at 13 treated reservoirs based on reservoir

manager interviews ...... 40

2. Cluster analysis of reservoir-specific seasonal CuSO4 application rates (µg Cu/L applied

as CuSO4) in 13 upground reservoirs...... 41

3. Diagram of a sediment trap ...... 42

4. Total concentration of copper, separated into dissolved and particulate-bound forms, in

the epilimnion and hypolimnion at Paulding Reservoir and Veterans Memorial Reservoir

before and after two CuSO4 applications at Paulding Reservoir ...... 43

5. Total concentration of copper, separated into dissolved and particulate-bound forms, in

the epilimnion and hypolimnion at Findlay #2 Reservoir and Bresler Reservoir before and

after CuSO4 application at Findlay #2 Reservoir ...... 44

6. Log10-transformed crustacean zooplankton dry biomass at all reservoirs ...... 45

7. Log10-transformed crustacean zooplankton density at Veterans Memorial Reservoir and

Paulding Reservoir before and after two CuSO4 treatments at Paulding Reservoir ...... 46 vii

8. Log10-transformed crustacean zooplankton density at Bresler Reservoir and Findlay #2

Reservoir before and after CuSO4 treatment at Findlay #2 Reservoir ...... 47

9. Copper concentration in the top 1 cm of sediment at Paulding Reservoir and Veterans

Memorial Reservoir before and after two CuSO4 applications at Paulding Reservoir ....48

10. Copper concentration in the top 1 cm of sediment at Findlay #2 Reservoir and Bresler

Reservoir before and after CuSO4 application at Findlay #2 Reservoir ...... 49

11. Densities of third and fourth instar chironomids at all reservoirs ...... 50

12. Composition and copper concentration of material collected in sediment traps before and

after one CuSO4 application at Paulding Reservoir ...... 51

13. Proportion Chironomus riparius surviving following exposure to copper pulses at

different instars ...... 52

INTRODUCTION

Upground reservoirs are man-made water bodies into which water is pumped from a nearby river or creek. This type of reservoir is different from an on-stream reservoir, where a dam is built and a section of streambed becomes the reservoir (Hale et al. 2006). The primary function of upground reservoirs is as drinking water storage facilities, but management strategies for water use sometimes conflict with their auxiliary function as a resource for sport fisheries.

That is, the management practices put in place to ensure clean drinking water can potentially conflict with practices that will maintain and foster a complex, diverse aquatic ecosystem

(Duvall et al. 2001, Mischke et al. 2009). One such practice is the extensive application of copper sulfate (CuSO4) to control nuisance algae.

Copper sulfate, which is 25.2% copper by weight, is effective in controlling algal blooms and has been used for over 100 years as an algicide in hatcheries, reservoirs, and private ponds

(Moore and Kellerman 1905). However, one of the potentially negative effects of CuSO4 is the affinity of copper for particulates, and thus copper from CuSO4 applications has the potential to exist in high concentrations in reservoir sediments (Haughey et al. 2000), as well as remain in the water column attached to suspended particulates (Florence 1977). In addition to being lethal to algae, CuSO4 has documented toxicity for vascular plants (Muller et al. 2001, Mal et al. 2002), zooplankton (Havens 1994a), and macroinvertebrates, including chironomids (Kosalwat and

Knight 1987, Warrin et al. 2009) and oligochaetes (Meller et al. 1998).

Toxic effects of CuSO4 pose no problem to the maintenance of upground reservoirs as drinking water storage facilities. The US Environmental Protection Agency has determined that copper below 1000 µg/L is not a threat to humans, and reservoir managers are required to test 2 water for two weeks post-application to ensure copper stays below this concentration (USEPA

2002). The potential problem occurs when added-value uses are applied to these reservoirs, such as recreational fisheries.

Yellow perch (Perca flavescens), a commonly stocked sport fish, is a good example of one species that may be indirectly affected by CuSO4 application. Yellow perch undergo an ontogenetic diet shift, initially feeding on zooplankton, then switching to benthic macroinvertebrates when total length (TL) is > 40 mm, and finally feeding on fish when TL is >

150 mm (Graeb et al. 2006). The Ohio Division of Wildlife stocks yellow perch at 25-30 mm

TL (Crouch 2011); therefore, stocking success may depend on the presence of both zooplankton and benthic macroinvertebrates, particularly chironomid species (see below), in abundances that will support yellow perch growth.

Although benthic chironomids make up a substantive part of the diets of commonly stocked sport fish in many systems (Lott et al. 1996, Wilkens et al. 2002, Galarowicz et al.

2006), benthivorous fish in upground reservoirs are almost exclusively limited to chironomids

(Paxton et al. 1981, Day et al. 1983). This is partially because of chironomids’ ability to survive in the hypoxic conditions common to hypolimnetic regions of these reservoirs in summer, and partially because these reservoirs are constructed with steep sides to maximize volume, thus providing few littoral areas for other benthic macroinvertebrates to exist (Stevenson and Day

1985). Furthermore, chironomids are tolerant of pollution compared to other benthic macroinvertebrates (Winner et al. 1980). As part of a broad study of upground reservoirs in

Ohio, we found that in systems that have received CuSO4 treatments for decades, sediment copper concentrations often exceeded the Probable Effects Level (the level above which there is 3 a high probability that sediment-dwelling organisms will experience adverse effects, Buchman

2008), yet still support chironomid populations (Crouch 2011).

Although chironomids are able to live in contaminated sediments (Kosalwat and Knight

1987), their presence does not necessarily indicate that populations are sustaining positive growth. Chironomids living in copper-contaminated sediments can bioaccumulate copper

(Bervoets et al. 2004, Giguere et al. 2005) and experience delayed growth (Besser et al. 1995,

Milani et al. 2003) and development (Timmermans et al. 1992, Al-Shami et al. 2010). However, the aforementioned adverse effects may also be, in part, food-mediated. In reservoirs and lakes, precipitating algal bodies represent a high-quality food source for chironomids (Kajak and

Warda 1968) and algae are known to absorb copper (Fayed and Abdel-Shafy 1986, Guanzon et al. 1995, Mehta and Gaur 2005). While the effects of food-borne contaminants and pulsed exposure to toxicants have been studied separately (Forbes and Cold 2005, Roberts et al. 2006,

Zhao and Newman 2006, Kolts et al. 2009, Angel et al. 2010), it is unknown how chironomids will be affected by the synthesis of these two types of exposure (i.e. the effects of pulsed dietborne copper).

Given that chironomids exist in contaminated reservoir sediments and that chironomid instars are differentially sensitive to environmental contaminants (Powlesland and George 1986,

Williams et al. 1986), I hypothesized that chironomid community impairment would be exhibited in the form of sublethal effects and would be dependent on the ontogenetic timing of exposure. I also hypothesized that, in the days following CuSO4 application, copper would exist first in dissolved form before binding to particulate matter (Haughey et al. 2000, Hullebusch et al.

2002), and that dissolved copper would depress zooplankton densities and biomass. Because zooplankton species are known to be differentially sensitive to metals (Yan et al. 2004, Wong et 4 al. 2009), zooplankton communities from treated and untreated reservoirs would be expected to diverge in composition and richness through time.

The objectives of this study were 1) to quantify the fractionation of copper in the water column, precipitating detritus, and sediment before and after CuSO4 application; 2) to enumerate zooplankton and macroinvertebrate community changes in reservoirs with varying CuSO4 application regimens through the application period (May – August); and 3) to quantify the effects of pulsed copper-laden food resources on chironomids of different instars. To accomplish these objectives, I collected water, sediment, sediment trap, zooplankton, and macroinvertebrate samples at four upground reservoirs and conducted experiments with Chironomus riparius in the laboratory.

METHODS Field Sampling

In January 2010, interviews were conducted with reservoir managers from 20 upground reservoirs located throughout Northwest Ohio. During these interviews, information concerning historic CuSO4 application regimens was obtained (Fig. 1). A post-hoc cluster analysis of the reported total amount of copper applied per season as CuSO4 (25.5% Cu by weight) in the 13 typically treated reservoirs supported the division of reservoirs into two application rate categories. Low application-rate reservoirs were those whose yearly total application was < 130

µg Cu/L, while high application-rate reservoirs had a yearly total application rate of > 130 µg

Cu/L (Fig. 2). The seven reservoirs that had never received CuSO4 were designated as “no application.”

To determine effects on the zooplankton community and chironomid populations due to

CuSO4 application, a Before-After-Control-Impact (BACI) Assessment (Stewart-Oaten et al.

1986) was followed. One reservoir from each application-rate category was paired with a reservoir that was not treated with CuSO4 during the study period (May – August 2010); these reservoirs were paired based upon size (capacity and area) and feasibility of access. Paulding

Reservoir (high application rate: 73 µg Cu/L/application) was paired with Veterans Memorial

Reservoir, which had received CuSO4 in previous years but did not during the study period.

Findlay #2 Reservoir (low application rate: 21 µg Cu/L/application) was paired with Bresler

Reservoir, which had never received CuSO4. The initial sampling strategy was to sample three

BACI pairs (six reservoirs total), with Veterans Memorial Reservoir representing an impacted reservoir. However, management personnel decided not to treat Veterans Memorial Reservoir during 2010 and our sampling schedule was too full to sample at six reservoirs. Thus, samples that were taken at Veterans Memorial and Paulding reservoirs concurrently were compared using 6 the BACI approach. However, because Veterans Memorial Reservoir had previously received

CuSO4, this approach could not be used to compare the chronic effects of copper-contaminated sediments on macroinvertebrate communities. Therefore, the BACI approach was used to address the immediate effects of copper in solution on zooplankton, and on chironomids ingesting copper-laden algae.

Copper sulfate was applied at Paulding Reservoir twice during the sampling period (2

June and 28 June) and once at Findlay #2 Reservoir. Because Findlay #2 Reservoir is so large

(260 ha), CuSO4 was applied on 19-21 July. Field sampling was conducted at Paulding and

Veterans Memorial reservoirs 25 May – 13 July and at Findlay #2 and Bresler reservoirs 15 July

– 4 August 2010. Sediment and water samples were obtained in order to determine the availability of copper to chironomids and zooplankton. Macroinvertebrate and zooplankton samples were collected to quantify population changes (density and biomass) through time. An effort was made to collect samples less than one week before and approximately one and two weeks after CuSO4 application events, although it was not always possible (due to applications at other reservoirs or weather).

Water

Water samples were obtained in the epilimnion (< 1 m below water’s surface) and hypolimnion (2 m off the bottom) using an acid-washed Van Dorn sampler. At each depth, two water samples were taken from the same Van Dorn sample with sterile 60 mL syringes: one sample was filtered through a 0.45-µm nylon filter and the other remained unfiltered. All water samples were stabilized with 2% by volume trace-metals grade HNO3. Unfiltered water samples were digested according to USEPA Method 3015 (Microwave-Assisted Acid Digestion of

Aqueous Samples and Extracts) to extract bioavailable metals. Filtered water and post-digestion 7 unfiltered water samples were analyzed using a ThermoElectron iCAP 6500 Inductively Coupled

Plasma Optical Emission Spectrophotometer (ICP-OES, Geology Department, Bowling Green

State University). Water was analyzed for seventeen metals, while sediments (see next section) were analyzed for nineteen metals (Table 1); however only copper was found at concentrations above the Probable Effect Levels (PEL, Buchman 2008). Comparing the amount of copper in unfiltered versus filtered water samples allowed for determination of the fraction of copper that was particulate-bound versus dissolved.

Sediment

Sediment samples were collected using an acid-washed stainless steel petite Ponar

(approximately 0.02 m2/sample). Two Ponar grabs were taken at two randomly-determined sites within each reservoir. Two 1-cm deep samples from each grab were collected using an acid washed plastic scoop to ensure enough material was collected for analysis (N = 2 sediment samples from each reservoir/sampling day). Sampling was restricted to the top 1 cm in an effort to collect material that represented the environment in which chironomids were likely to occur

(Oliver 1971).

With the knowledge that collection of the top 1 cm of sediment would include historic as well as recently-deposited material, sediment traps were deployed at Paulding Reservoir in order to collect precipitating detritus, including algal deposition. Three sediment traps (12 collection containers total, see below) were placed at Paulding Reservoir before CuSO4 application and retrieved before application, then re-deployed and retrieved less than one week after application, and, again, re-deployed and retrieved two weeks after application. Each sediment trap consisted of a plastic support basket with ballast into which four acid-washed 3.2-L plastic containers (609 cm2 surface area at the top), were inserted for collection of precipitating material (Figure 3). 8

Sediment and sediment trap samples were digested according to USEPA Method 3051A

(Microwave Assisted Acid Digestion of Sediments, Sludges, Soils, and Oils) to extract bioavailable metals and were analyzed using the ICP-OES in the BGSU Geology Department.

Percent organic matter of each sample was also determined. To obtain percent organic matter, samples were dried, pulverized, and dried to constant weight in pre-weighed crucibles in a drying oven at 60˚C. They were then heated in a muffle furnace for 16 hours at 375˚C, a method that has been recommended for sediments that are composed mostly of clays and have low organic content (Beaudoin 2003, Boyle 2004), cooled in a dessicator, and reweighed.

Aquatic Organisms

Benthic macroinvertebrate (i.e., chironomids) and zooplankton samples were also collected before and after application to understand how their populations were affected by

CuSO4. Zooplankton samples were collected during the day in triplicate using a vertical zooplankton tow (63-µm mesh, 1 m in length, 30 cm diameter) fitted with a flow meter (General

Oceanics Inc., Model 2030 with Standard Rotor). The net was lowered to 1 m above the bottom of the reservoir and pulled upward through the water column. To determine community composition and density, at least 100 organisms in at least two 5-mL subsamples were counted under 30x magnification and identified to genus using a Zeiss Stereoscope (model 47 50 52

9901). The first 25 organisms encountered from the most abundant taxa and the first 10 organisms encountered from all other taxa were measured in each subsample using an ocular micrometer (30X, 1 ocular unit = 0.031 mm). Using length-weight regression equations

(Dumont et al. 1975, Bottrell et al. 1976, Rosen 1981, Culver et al. 1985), biomass per taxon was then estimated. ANOVAs were used to determine overall within- and between-reservoir differences in both biomass and density. If significant overall differences were detected, a 9

Tukey-Kramer HSD test was used to determine on which dates crustacean zooplankton populations differed.

To assess chironomid community changes through time, three sediment samples were taken in random locations throughout each reservoir on each sampling day using a stainless steel petite Ponar. Complete samples were preserved in 10% buffered formalin and then organisms were elutriated from sediment using 500-µm mesh filter according to USEPA SOP LG406

(Benthic Invertebrate Field Sampling, 2003) upon arrival at the lab. After elutriation, macroinvertebrates were stored in 10% buffered formalin until identification. All chironomids were removed from each sample and identified to subfamily. Head widths were measured for all chironomids (McCauley 1974) and biomass was determined using regression equations (Smock

1980, Baumgartner and Rothhaupt 2003).

Effects of Copper on Chironomid Instars - Laboratory Experiment

Sediment

Sediment from a local drinking water reservoir (Paulding Reservoir) was used in this experiment because its copper concentration was closest to the concentration at which detrimental effects on the growth of larval Chironomus spp. have been demonstrated (Kosalwat and Knight 1987). Sediment was collected in November 2011 using an acid washed stainless steel petite Ponar. The top 5 cm of sediment was collected, homogenized, and then frozen at

-20˚ C to kill in situ organisms (Carins et al. 1984, de Haas et al. 2004); sediments were then rehomogenized and stored at 4˚C until the start of the experiment.

Test Organisms

Several Chironomus riparius egg strands of similar age were obtained from

Environmental Consulting and Testing, Inc. (Superior, WI). Upon arrival, egg strands were 10 separated into groups of approximately 26 ± 2.5 eggs (mean ± SD) using a sterile stainless steel scalpel; groups were placed into individual Petri dishes, and covered with the same water in which they were transported. Eggs were monitored daily for hatching; once all larvae left the egg mass, any un-hatched eggs were counted and discarded. Known numbers of first-instar individuals were then transferred to test aquaria and were allowed to acclimate for 24 h before algal resources were added.

Experimental Setup

Trials were conducted in walk-in environmental chambers kept at 22 ± 1˚C under a

16L:8D cycle. Six rectangular Rubbermaid® containers (16 x 24 x 8 cm deep) per treatment were used as experimental vessels (hereafter aquaria). Prior to the addition of sediment, aquaria were cleaned by soaking in a 10% HNO3 bath and then a Micro-90® detergent bath. Containers were thoroughly rinsed with Milli-Q water before and after each bath.

Each aquarium received 570 cm3 of homogenized wetted sediment to create a sediment depth of 1.5 cm and dechlorinated tap water (11.7 µg Cu/L) was added to generate a total sediment/water depth of 4 cm. Water and sediment were allowed to equilibrate for 7 days prior to the addition of organisms. Water quality measures (D.O., pH, conductivity, temperature) were recorded every other day during equilibration with a HACH sensION 156 in order to ensure common starting conditions. Water quality measures were also recorded every experimental day and it was not necessary to add overlying water for the duration of the experiment. Preliminary experiments demonstrated that the water was shallow enough for adequate oxygen exchange to take place without the addition of an air stone (always > 2 mg/L).

The cyanobacterium Spirulina platensis was used as a food source for C. riparius larvae in this experiment. S. platensis was obtained in powder form from Pure Bulk, Inc. (Roseburg, 11

OR). Although S. platensis is not a naturally-occurring algae in Paulding Reservoir waters (M.

Winners, Paulding Reservoir Manager, personal communication, 6 October 2011, Table 2), it has several characteristics that make it ideal for use as a food source: 1) it contains a low concentration of copper ( 3 µg Cu/g dry wt, Tokusoglu and Unal 2003), a high percentage of both protein and lipids, and is easily digested (Richmond 1988); 2) its suitability as a high- quality food source for C. riparius larvae has been demonstrated (Goedkoop et al. 2006); and 3) the ability of both live and dead S. platensis to absorb copper has been well-documented (Doshi et al. 2006, Solisio et al. 2006, Doshi et al. 2009, Fang et al. 2011).

Organisms were fed on experimental days 2, 3, 5, 6, 9, 10, 13, and 14. These days were chosen because they fell in the presumed middle of each instar (Table 3). Algae were added to the aquaria over two days per instar (e.g., days 2 and 3) instead of one because preliminary experiments demonstrated that dissolved oxygen fell below 1 mg/L when all food was added in a single dose. Feeding amounts (mg dry weight algae/cm2) were based upon field observation of sedimentation rates determined using the sediment traps placed at the bottom of Paulding

Reservoir during June of 2010. Sedimentation rates differed before and after CuSO4 application: before application the rate was 0.31 ± 0.02 mg dry weight/cm2/day and in the three days after application the rate was 0.49 ± 0.09 mg dry weight/cm2/day (mean ± SE). Sediment surface area in each test aquarium was 315 cm2; thus, on each feeding day, the control treatments and the aquaria containing instars that were not experiencing a copper-laden algal pulse received 0.049 grams dry weight of algae in 5 mL of Milli-Q water. Aquaria experiencing a copper pulse received 0.077 grams of spiked algae in 5 mL of Milli-Q water. In order to replicate the copper concentration of precipitating material after a CuSO4 application, algae was soaked in a 5000 mg

Cu/L solution (7.9365 g CuSO4·5H2O dissolved in 40 mL Milli-Q water). To prevent unbound 12 copper from entering the water along with the algae, spiked algae was rinsed three times with

Milli-Q water prior to addition to aquaria.

Although traditional toxicity testing is conducted over a 10 or 28-day period (ASTM E

1706), I wanted larvae pulsed in the fourth instar to have the opportunity to forage on spiked algae, so this experiment was ended after 15 days. On day 15, samples of the entire sediment column were collected using an acid-washed plastic vial (mouth approx. 7 cm2) and frozen for later analysis. Water samples were collected using sterile syringes, filtered through a 0.45 µm nylon filter and stabilized with 2% by volume trace-metals grade HNO3. Remaining water and sediment from each aquarium was then rinsed through 1 mm and 0.25 mm mesh filters. All larvae, pupae, and any adults were stored in 70% ethanol.

Analysis

Surviving larvae, pupae, and adults were enumerated and proportional survivorship estimated. Proportion surviving was arcsine square root transformed to ensure normality and analyzed using a one-way ANOVA followed by Tukey’s HSD test with JMP (Version 9, SAS

Institute Inc., Cary, NC, 1989-2010).

Less than 24 hours after being placed into ethanol (minimizing weight loss, Donald and

Paterson 1977; Ristola et al. 1999), half of the larvae from each aquarium were weighed: following head width measurement using the ocular micrometer on a binocular microscope

(30X, 1 ocular unit = 0.031 mm, Zeiss Stereoscope 47 50 52-9901), each larva was placed into a pre-weighed propylene weighing dish, dried to constant weight at 60˚C, cooled in a dessicator, and then weighed to the nearest 0.1 mg using a Mettler AE 100 analytical balance. Individual weights were averaged for each aquarium. Head widths of all un-weighed larvae were obtained in the same manner as for weighed larvae. Head width and weight data were similarly analyzed 13 using a one-way ANOVA with Bonferroni correction (α = 0.05/2). To assess developmental differences among treatments, the proportion of C. riparius that had developed into pupae and adults was compared (arcsine square root transformed, ANOVA).

Sediment was dried to constant weight at 60˚C, allowed to cool in a dessicator, and pulverized before digestion according to USEPA Method 3051A. Both digested sediment samples and filtered water samples were analyzed for a suite of metals (Table 1) on the ICP-OES in the BGSU Geology Department. Percent organic matter was also determined using the methods described above for field-collected sediment. Samples of Cu-augmented algae were also digested, analyzed, and ashed as above to determine copper content. Copper concentration for un-augmented S. platensis was obtained from Tokusoglu and Unal (2003). 14

RESULTS Field Sampling

Water

Within each BACI Assessment pair, the concentration of copper in water at treated reservoirs reflected CuSO4 application rate and amount, while at untreated reservoirs the amount of copper in water did not change (Figs. 4 & 5). At Paulding Reservoir (application rate 73 µg

Cu/L in one day), the concentration of both epilimnetic dissolved and particulate-bound copper increased ~ 4-fold following each of the two CuSO4 treatments (Fig. 4), with dissolved copper concentration greater than twice the level determined to cause acute toxicity to zooplankton

(Buchman 2008). Two to three weeks after application, epilimnetic copper concentrations had returned to levels that were similar to the pretreatment concentration and were below acute toxicity. At Veterans Memorial Reservoir (no added Cu) dissolved and particulate copper did not change through time and remained well below toxic levels. Copper concentrations and trends in the hypolimnion of each reservoir corresponded to those observed in the epilimnion

(Fig. 4).

At Findlay #2 Reservoir, where a total of 21 µg Cu/L was applied over three days (as the reservoir is too large to treat in one day), concentration of both epilimnetic and hypolimnetic dissolved and particulate copper was 2-3 times higher than the pre-treatment concentration in the week after treatment (Fig. 5). However, unlike at Paulding Reservoir, dissolved copper concentrations did not approach levels that cause acute toxicity to zooplankton in either the epilimnion or hypolimnion (determined using equations from Buchman 2008). At Bresler

Reservoir, the comparison reservoir for Findlay #2 Reservoir, both epilimnetic and hypolimnetic copper concentrations exhibited slight fluctuations but always remained well below toxic concentrations and approximated the pre-treatment concentrations at Findlay #2 Reservoir. 15

Zooplankton

In both control reservoirs, there were no significant within-system changes in zooplankton biomass through time (Veterans Memorial Reservoir: F4,10 = 0.25, P = 0.91; Bresler

Reservoir: F2,6 = 0.60, P = 0.58; Fig. 6). At both treated reservoirs, the pre-treatment biomass was significantly greater than any subsequent sample (Paulding Reservoir: F4,10 = 7.73, P =

0.004; Findlay #2 Reservoir: F2,6 = 9.89, P = 0.013; Fig. 6), but the magnitude of the difference reflected application rate. At Paulding Reservoir, dissolved copper in the water exceeded levels toxic to zooplankton (Buchman 2008) for 1 – 2 weeks after each treatment (2-June and 28-June,

Fig. 4). Correspondingly, post-treatment zooplankton biomass decreased 93% compared to pre- treatment biomass and remained depressed for at least two weeks post-application (Fig. 6).

Within one week post-application at Findlay #2 Reservoir, the concentration of dissolved copper in the water was 20% of the amount that would cause acute toxicity to zooplankton according to

Buchman (2008), but the zooplankton biomass had decreased by 64% (Fig. 6).

Similar to biomass, in both control systems neither total zooplankton density (Veterans

Memorial Reservoir: F4,10 = 0.39, P = 0.81; Bresler Reservoir: F2,6 = 1.86, P = 0.23) nor density of cladocerans, adult copepods, and copepodites (hereafter non-nauplii) changed through time

(Veterans Memorial Reservoir: F4,10 = 0.43, P = 0.79; Bresler Reservoir: F2,6 = 2.52, P = 0.16;

Figs. 7 & 8). At Paulding Reservoir, total crustacean zooplankton density was only significantly lower than pre-treatment density on 10-June; however, non-nauplii density was significantly lower than pre-treatment density (F4,10 = 15.54, P < 0.001) for all samples taken after the initial application (Fig. 7). This was due to the fact that copepod nauplii accounted for 18% of the pre- treatment crustacean zooplankton community but comprised >75% of the density in all samples taken after the initial CuSO4 application (2 June). Furthermore, the post-treatment non-nauplii 16 zooplankton community was comprised solely of adult copepods and copepodites, remaining devoid of cladocerans (< 1%, Fig. 7).

At Findlay #2 Reservoir, total density was significantly lower two weeks after CuSO4 application than before application (F2,6 = 12.96, P = 0.007), but non-nauplii density displayed a non-significant decrease (F2,6 = 4.61, P = 0.061; Fig. 8). The pre-treatment crustacean zooplankton community at Findlay #2 Reservoir was 50% copepod nauplii and increased to 71% within the week following application. However, copepod nauplii densities decreased to 35% of the community within two weeks, and cladocerans were still present, comprising 18% and 42% of samples taken one and two weeks post-application, respectively. The fact that densities of cladocerans and adult copepods were less affected by CuSO4 application at Findlay #2 Reservoir than at Paulding Reservoir suggests that differences in the application rate (21 µg Cu/L over three days vs. 73 µg Cu/L over one day) and timing (two versus one application in four weeks) may have differentially affected zooplankton.

Sediment

Historic CuSO4 application was a good predictor of the sediment copper concentration at the beginning of summer 2010. Sediment copper concentrations significantly differed among reservoirs and reflected historic (pre-2010) application rates with Paulding > Findlay #2 >

Veterans Memorial > Bresler (ANOVA: F3,4 = 66013.97, P < 0.0001, with Tukey-Kramer HSD).

Additionally, sediment at both Findlay #2 and Veterans Memorial reservoirs had copper concentrations higher than the Threshold Effects Level (TEL, minimum level at which organisms experience adverse effects) for sediment-dwelling organisms, while Paulding

Reservoir sediments contained copper at concentrations that were approximately 8 times the

Probable Effects Level (PEL) and 44 times the TEL (Buchman 2008). 17

Over the course of the summer, sediment copper concentration reflected the addition of copper during CuSO4 application. At Paulding Reservoir, sediment copper concentration increased 45% through time with the final concentration significantly greater than the first sample (ANOVA: F4,5 = 10.78, P = 0.011 with Tukey-Kramer HSD; Fig. 9). Interestingly, there was also a significant increase (53%) in sediment copper concentration at Veterans Memorial

Reservoir (the control system with no CuSO4 applied in 2010) from the first to the last sample

(F3,4 = 20.36, P = 0.007) . However, unlike the trend at Paulding Reservoir, copper concentration did not steadily increase throughout the study period, so the dynamic is not simple to interpret (Fig. 9). In the other paired reservoirs, there was significantly more copper (25%) in the top 1 cm of sediment two weeks after application at Findlay #2 Reservoir than there was before CuSO4 was applied (F2,3 = 28.05, P = 0.011, Fig. 10). As expected, sediment copper concentration at the control reservoir (Bresler), did not change significantly during the study period (F2,3 = 0.59, P = 0.61).

Chironomids

The most commonly-encountered chironomid genera in Ohio upground reservoirs are

Chironomus (subfamily Chironominae), Procladius and Coelotanypus (subfamily Tanypodinae), and Tanytarsus (subfamily Orthocladiinae) (Paxton et al. 1981; Day et al. 1983), although I only encountered individuals from the subfamilies Chironominae and Tanypodinae in this study. As such, head capsule widths of third and fourth instar larvae from Chironomus, Procladius, and

Coelotanypus genera are generally greater than 400 µm (McCauley 1974). Chironomids with head capsule widths of 400 µm are known to be retained in a 500 µm mesh (Hamilton 1965), especially if they are preserved first, but those with smaller head capsule widths are likely to be lost (Storey and Pinder 1985, Hudson and Adams 1998). Therefore, analysis conducted on 18 chironomid populations in this study excluded individuals with head widths < 400 µm (0.4 mm, first and second instar), because chironomids of this size could not be reliably enumerated.

At Paulding and Veterans Memorial reservoirs, densities of third and fourth instar chironomids exhibited a peak in mid- to late-June and then steadily decreased (Fig. 11). The decrease in density was more pronounced at Veterans Memorial Reservoir, such that Paulding

Reservoir sediment contained a significantly higher density of third and fourth instar chironomids than Veterans Memorial Reservoir in the last two samples even after two CuSO4 applications (1-July: F1,4 = 9.80, P = 0.035; 13-July: F1,4 = 64.00, P = 0.001, Fig. 11). Third and fourth instar chironomid densities at Findlay #2 and Bresler reservoirs also exhibited a steady decrease, although sediment at Bresler Reservoir consistently had significantly greater densities than at Findlay #2 Reservoir (15-July: F1,4 = 18.64, P = 0.008; 27-July: F1,4 = 46.17, P = 0.002;

3-August: F1,4 = 23.28, P = 0.008, Fig. 11).

Unlike the density shifts observed in each reservoir, there were no significant within- system changes in chironomid biomass at any reservoir during the study period. However, there were significant between-reservoir biomass differences: Veterans Memorial had significantly less biomass than Paulding in both samples taken in July (1-July: F1,4 = 10.80, P = 0.030; 13-

July: F1,4 = 10.19, P = 0.033). Chironomid biomass was significantly higher at Bresler Reservoir compared to Findlay #2 Reservoir throughout the study period (15-July: F1,4 = 23.69, P = 0.005;

27-July: F1,4 = 21.96, P = 0.009; 3-August: F1,4 = 139.36, P < 0.001).

Sediment Traps at Paulding Reservoir

In the five days preceding CuSO4 application, sediment accumulated at a rate of 0.31 ±

0.02 mg dry weight/cm2/day, contained 1678 ± 188 µg Cu/g dry weight, and was 15.8 ± 5.91% organic matter (mean ± SE, n=3, Fig. 12). During the three days following application, the 19 sedimentation rate was 0.49 ± 0.09 mg/cm2/day, 36% higher than pre-application, and copper concentration rose to 5429 ± 459 µg Cu/g dry weight, while the percentage of precipitating material that was organic matter was relatively the same at 16.0 ± 3.6%. During days 4-15 following application, the sedimentation rate was 0.55 ± 0.13 mg/cm2/day, 45% higher than pre- application, while the percent organic matter and copper concentration fell to 10.6 ± 1.75% and

3717 ± 459 µg Cu/g dry weight, respectively (Fig. 12). The pre-application sediment accumulation rate and days 4-15 post-application rate were significantly different (ANOVA: F2,6

= 8.77, P = 0.03, with Tukey-Kramer HSD test); the rate during days 1-3 post-application was intermediate. Furthermore, the copper concentration of precipitating material significantly differed through time: pre-application < days 4-15 post-application < days 1-3 post-application

(ANOVA: F2,6 = 73.34, P < 0.001, with Tukey-Kramer HSD test, Fig. 12).

Laboratory Experiment

Survival of chironomids receiving a pulse of copper-laden food in the first instar was significantly lower than survivorship of organisms that never received a pulse or that received a pulse in the third or fourth instars (ANOVA: F4,25 = 5.16, P = 0.004, with Tukey-Kramer HSD test, Fig. 13). Survivorship of individuals that received a copper pulse in the second instar was not significantly different from all other treatments.

Although there were significant differences in survivorship among treatments, differences in developmental rate were less obvious. All organisms retrieved as larvae (i.e., surviving individuals) from this experiment were in the fourth instar (HW = 0.54 ± 0.003 mm, mean ± SE,

Watts and Pascoe 2000) and there were no significant differences in the proportion of surviving individuals that were pupae/adults (F4,25 = 1.20, P = 0.33, Fig. 13). However, larvae receiving a pulse of copper in their fourth instar weighed significantly less (-36%) than larvae from the 20

control treatment (F4,125 = 4.52, P = 0.002), suggesting that they may have stopped feeding in response to the addition of copper-laden algae, thus slowing their development.

Mean temperature ± SE within all aquaria was 22.3 ± 0.04˚C. Aquaria on one of the three shelves had a consistently higher temperature (+ 0.5 ˚C) than the other two shelves, but because aquaria were distributed such that two aquaria from each treatment were on each shelf, the overall temperature of all six aquaria within a treatment did not vary significantly (F36, 65 =

0.79, P = 0.77) throughout the experiment. Dissolved oxygen levels exhibited significant time*treatment (F60, 45 = 2.55, P < 0.001) and time (F15, 11 = 228.1, P < 0.0001) effects, but no significant treatment effect (F4, 25 = 1.16, P = 0.35). This was to be expected, because the dissolved oxygen declined in the days following an algal pulse, and declined to slightly lower concentrations when aquaria received Cu-spiked algae than when they received unspiked algae because of the amount added (0.077 vs. 0.049 grams). There were no significant time*treatment

(F60, 45 = 0.95, P = 0.58) or treatment (F4, 25 = 1.24, P = 0.32) differences in pH, although pH did vary significantly through time (F15, 11 = 30.73, P < 0.0001), increasing from 7.57 ± 0.03 to 7.68

± 0.02 (mean ± SE). Conductivity decreased in the two days during which the aquaria received a copper pulse and the day after the last copper pulse. Because conductivity is related to the amount of dissolved solids in the water, it is possible that in the three days surrounding a copper pulse, the water was cleared of solids. This could be the result of several factors: a) copper binds to particulate matter, causing it to precipitate out of the water column (a similar effect was observed in the epilimnion at Paulding Reservoir) and/or b) the immediate effect of the addition of Cu-spiked algae may have been to decrease larval feeding (and bioturbation), considering that larvae experiencing a spike of Cu-spiked algae in the fourth instar weighed less than control larvae. 21

Sediment copper concentration in treated aquaria was 1356 ± 8.7 µg Cu/g dry weight

(mean ± SE, n=24), which was significantly greater (F4,25 = 9.02, P < 0.001) than the concentration in control aquaria, which contained 1243 ± 17 µg Cu/g dry weight (mean ± SE, n =

6). This matched observations at Paulding Reservoir, where the copper concentration in the upper 1 cm of sediment increased by approximately 100 µg Cu/g dry weight after the first

CuSO4 treatment (Fig. 9).

Sediment from the control aquaria had significantly less organic matter than sediment from aquaria that were pulsed in the first or second instar (F4,25 = 5.01, P = 0.004), but had similar levels of organic matter to tanks that were pulsed in the third or fourth instar. It may be that percent organic matter is related to the number of surviving organisms in each treatment: there were more organisms to consume the organic matter in aquaria with higher survivorship.

For this experiment, copper binding by the algae, S. platensis, was 29.1 mg Cu/g dry weight algae, which is less than the maximum absorption rate of 100 mg Cu/g dry weight algae at room temperature as reported by Doshi et al. (2006). In this experiment, prior to addition to aquaria, spiked algae was placed in the refrigerator between rinses with Milli-Q water to retard cellular breakdown, potentially slowing the rate of absorption of copper.

DISCUSSION

Effect of Copper Sulfate Application on Zooplankton Communities

While zooplankton biomass and community structure did not exhibit significant changes in untreated systems (Bresler and Veterans Memorial reservoirs), zooplankton biomass was significantly depressed at both Findlay #2 and Paulding reservoirs post-CuSO4 application (64% and 93%, respectively), and density also was reduced by as much as 87%. However, both the magnitude and duration of the depression, as well as which taxa were most affected, were different between treated reservoirs, corresponding to CuSO4 application rate. Dissolved copper in water samples taken at Paulding Reservoir within one week of each CuSO4 application (73 µg

Cu/L per application) exceeded levels that were toxic to zooplankton (Buchman 2008), contributing to decreases in abundance. At Findlay #2 Reservoir (21 µg Cu/L per application) the dissolved copper concentration within one week post-application was 20% of the concentration that causes acute toxicity to zooplankton (Buchman 2008). However, biomass and density of zooplankton decreased, suggesting that, at some point within the first week, dissolved copper concentrations at Findlay #2 Reservoir reached levels acutely toxic to zooplankton. At

Findlay #2 Reservoir, biomass and total density remained depressed for at least two weeks (59% and 48% lower than pre-treatment, respectively); hence the effect of CuSO4 application even in low doses had substantive effects on zooplankton communities (Havens 1994a, 1994b; Duvall et al. 2001; Mischke et al. 2009; Tew et al. 2010).

In addition to total biomass and density, zooplankton community structure also was altered by CuSO4 application: at treated reservoirs (Paulding and Findlay #2), pre-application communities were 18% and 50% copepod nauplii, respectively. In the week following application, the proportion of copepod nauplii increased to 86% at Paulding Reservoir and 71% 23 at Findlay #2 Reservoir, while percent nauplii at control reservoirs (Veterans Memorial and

Bresler) decreased. These results suggest that cladocerans and adult copepods were more sensitive to CuSO4 application; the presence and apparent tolerance of copepod nauplii further suggests that copepod communities more rapidly recover from copper pulses than do cladocerans. Cladoceran populations are more likely than copepod populations to be depressed by CuSO4 for several reasons: cladocerans are generally more sensitive to copper than copepods

(Wong et al. 2009), cladocerans rely more heavily on the algal resources depleted by applications

(Havens 1994a), and any surviving cladocerans are more likely to be consumed by fish (Yan et al. 2004).

While copepod nauplii can be an important resource for larval fish < 8 mm TL (Bremigan et al. 2003, Graeb et al. 2004), yellow perch and walleye populations in upground reservoirs are supported by stocking young-of-year (YOY) fish at 25 – 30 mm TL. Early juvenile percids have higher growth rates on a diet of cladocerans and adult copepods compared to nauplii (Mayer and

Wahl 1997, Romare 2000, Bremigan et al. 2003, Graeb et al. 2004), such that any fish stocked into these reservoirs within two weeks of CuSO4 treatment have limited planktonic food resources available to them. Copper sulfate-induced depression of zooplankton populations was similarly observed by Mischke et al. (2009) in catfish hatchery ponds, suggesting that stocking should occur at least two weeks post-CuSO4 application. While many additional factors beyond

CuSO4 application contribute to stocked sportfish success in upground reservoirs (Crouch 2011), reservoir management and fisheries management need to communicate about stocking and

CuSO4 application timing in order to promote sport fish stocking success.

Effect of Copper Sulfate on Chironomid Communities

Densities of third and fourth instar chironomids at Paulding and Veterans Memorial reservoirs peaked in June, suggesting that adult emergence and subsequent egg laying may have occurred relatively early in summer. As newly-lain eggs hatched in reservoirs with copper- contaminated sediments, it is possible that first instar chironomids suffered mortality; even at concentrations as low as 50 µg Cu/g dry weight (i.e., sediment concentrations in three of four study reservoirs), Servia et al. (2006b) found up to 60% mortality in C. riparius larvae. Since most reservoir-dwelling chironomids feed on surficial sediments (Rasmussen 1984), this mortality may actually be an underestimate. For newly-hatched chironomids living in copper- contaminated sediments, detritus and algae with concentrations of up to 5000 µg Cu/g dry weight precipitating from the water column post application may have caused additional mortality as it was ingested. Older individuals may not have been affected, given that older instars are less sensitive than younger instars to environmental contaminants (Powlesland and George 1986,

Williams et al. 1986, and this study).

There is also the possibility that the developmental rate of chironomids living in copper- contaminated sediments was impaired (Timmermans et al. 1992, Postma et al. 1995, Al-Shami et al. 2010), resulting in lower densities over time (Postma and Davids 1995). Paulding, Findlay #2, and Veterans Memorial reservoirs are all 20+ years old; because the sediments in these reservoirs act as a sink for copper, many generations of chironomids would have experienced chronic copper exposure even if CuSO4 was not applied yearly. Delayed growth over several decades may have contributed to the lower densities observed at these reservoirs compared to Bresler

Reservoir (also 20+ years old), which was never treated with CuSO4. 25

Despite among-reservoir differences in chironomid densities, all four reservoirs always had densities greater than 25 ind/m2, except at Veterans Memorial Reservoir on 13 July when no chironomids were found. Capture efficiencies of chironomids by YOY walleye and yellow perch in optimal foraging experiments were high enough to support fish growth when density was ≥ 25 ind/m2 (Graeb et al. 2005; Galarowicz et al. 2006; Graeb et al. 2006), suggesting that chironomid densities are generally adequate for YOY and juvenile fish growth in these systems.

However, although these densities may support YOY fish growth, adult fish in both treated and untreated upground reservoirs with relatively low productivity may also rely on benthos for a significant portion of their diets (Jansen and Mackay 1992, Wilkens et al. 2002), and more so in metal-contaminated systems with slow fish growth and absent metal-sensitive benthic species

(Kovecses et al. 2005, Rasmussen et al. 2008). Fisheries management agencies should integrate information about each reservoir’s existing fish population age structure and chironomid density into encounter rate models, such as those developed by Persson and Greenberg (1990), to determine whether chironomid densities are great enough to support communities of both YOY and adult fish.

Effect of Copper Sulfate Application Timing on Chironomid Instars

High mortality was experienced by C. riparius that received a pulse of copper-laden algae in their first instar (85% ± 3.4% mortality, mean ± SE). Survivorship of individuals receiving a pulse in the second instar was not statistically lower than that of any other instar, but was intermediate between survivorship of first instar larvae and older larvae. These findings support previous research indicating decreased chironomid sensitivity to toxicants with age

(Nebeker et al. 1984, Williams et al. 1986, Timmermans et al. 1992, Pery et al. 2003). If a large 26 proportion of the benthic community was made up of first or second-instar chironomids when

CuSO4 was applied, high population-level mortality is likely.

However, organisms also experienced mortality simply from living in copper- contaminated sediments. Organisms in control aquaria (natural sediment containing 1243 ± 17

µg Cu/g dry weight, fed ecologically-realistic amounts of unspiked S. platensis) experienced

40% mortality. In preliminary experiments, however, we observed as low as 5% mortality in the same sediment when C. riparius were fed ground fish food (Nutri Source® Farm Pond Diet).

This suggests that high quality food can overcome the effects of living in copper-contaminated sediment (de Haas et al. 2004). This also suggests that previous whole-sediment toxicity tests conducted using fish food, either in measured portions or ad libitum, may underestimate the effects of the contaminated sediment. This hypothesis is supported by Wentsel et al. (1977a), who found 50 to 60% survival of unfed C. tentans in sediment contaminated with Cd, Zn, and

Cr.

In addition to the acute effects of contaminated food and sediment, there may have also been subtle developmental differences among treatments. Larvae that had experienced a pulse in the fourth instar weighed less than controls, suggesting that they had ceased feeding. Once C. riparius is in the fourth instar, there are several developmental phases prior to pupation (Ineichen et al. 1983), and C. riparius can delay pupation, remaining in phase 1 of the fourth instar, if exposed to high levels of copper (Servia et al. 2006a). Delayed pupation has been observed in response to other contaminants as well: Forbes and Cold (2005) found that C. riparius have slower developmental rates when exposed to environmentally-realistic levels of the pesticide esfenvalerate. 27

Chironomids will consume high quality food sources even if the food is in highly contaminated sediment (de Haas et al. 2006), and will not actively avoid highly contaminated sediment (Dornfeld et al. 2009). Therefore, it is not difficult to understand why chironomid densities and biomass were lower at our target reservoirs with historic CuSO4 application than the reservoir that had never been treated. However, chironomid populations exposed to sediment- and food-borne copper over many generations may also experience adaptations and selection to better tolerate high concentrations of copper (Wentsel et al. 1977b).

Conclusions and Future Research

Copper sulfate application, while beneficial for algal control in drinking water reservoirs, negatively affects zooplankton and chironomid communities, two important food resources for fish. If reservoir managers work in conjunction with fisheries managers, it may be possible to time the application of CuSO4 such that both water quality and fishery needs are met. If possible, the algal community should be analyzed prior to application so that only the minimum amount of CuSO4 needed to eliminate noxious species is used. Furthermore, the timing of stocking events should take into account CuSO4 application, and vice versa. Other means of algal control can also be investigated, such as peroxygen (Harvey and Howarth 2008) and solar powered circulation (Hudnell et al. 2010).

Further research should be conducted to determine the role of adaptation on food-borne pulses of copper, since Wentsel et al. (1978) found that survival of chironomids from contaminated sediment was greater than those that had never experienced contaminated sediment. It will also be necessary to determine the species composition of the benthic community in treated reservoirs to accurately assess the risk of CuSO4 application. Finally, toxicity tests in which chironomids are fed ecologically unrealistic food (i.e. fish flakes) may 28 substantively underestimate the effects of contaminated sediments. Incorporating appropriate food sources into future toxicity testing is integral to realistic prediction of effects. 29

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TABLES

Table 1: Wavelengths (nm) and Minimum Detection Levels used in elemental analysis of sediment and water samples from upground reservoirs using ICP-OES. Because of slight changes in sensitivity, various wavelengths were used to optimize estimates.

Wavelength(s) and Minimum Detection Levels used in ICP-OES Analysis Sediment Water

MDL (µg/kg) MDL (µg/L) Wavelength(s) mean ± SD Wavelength(s) mean ± SD Al 2269 13 ± 1.1 n/a n/a As 1890 3.1 ± 0.18 1890 10 1937 5.2 ± 0.99 1937 9.2 ± 1.3 Ba 4554 0.051 ± 0.002 2304 0.55 4934 0.12 4934 0.17 ± 0.23 Ca 3706 4.2 ± 0.81 3706 9.3 ± 0.21 Cd 2144 0.062 ± 0.001 2144 0.31 ± 0.014 2265 0.086 ± 0.002 2265 0.3 Co 2286 1.3 ± 1.7 2286 1.0 ± 0.12 2307 0.44 Cr 2677 0.51 ± 0.024 2677 1.2 ± 0.21 2835 0.85 Cu 3273 1.1 ± 0.018 3247 1.2 3273 2.7 ± 0.071 Fe 2395 0.71 2599 1.0 ± 0.058 2599 0.05 ± 0.017 K 4047 860 ± 200 7664 15 ± 0.71 Mg 2790 3.2 ± 0.029 2852 0.62 ± 0.029 Mn 2576 0.088 ± 0.015 2576 0.20 Mo 2020 0.34 ± 0.008 2020 1.5 ± 0.17 Na 8183 23 ± 6.6 5889 0.11 5895 3.9 ± 0.21 Ni 2216 0.26 ± 0.01 2216 0.93 ± 0.046 Pb 2203 1.0 ± 0.16 2203 5.7 ± 0.57 Se 1960 3.2 ± 0.14 n/a n/a Sr 3464 1.0 ± 0.04 4077 0.04 4077 0.02 4215 0.07 Zn 2138 0.11 ± 0.005 2138 0.30 ± 0.006 38

Table 2: Algal community composition (cells/milliliter) at Paulding Reservoir during the months of May – July 2010. Category Taxa 20-May 27-May 3-June 10-June 17-June 24-June 1-July 8-July 15-July 22-July Cyanobacterium Aphanocapsa spp. 12800 16000 10400 7200 27000 24000 0 0 0 0 Chroococcus spp. 0 0 12880 0 0 0 0 0 0 0 Planktothrix 0 0 0 0 0 0 4980 0 0 0 Diatom Cyclotella spp. 24480 0 0 0 0 0 0 0 0 0 Nitschia spp. 0 4880 0 0 0 0 0 0 0 0 Synedra spp. 0 0 0 0 0 0 0 17500 5300 0 Green algae Chrysochromulina spp. 0 0 0 61200 0 0 0 0 0 0 Oocystis spp. 0 0 5120 0 7700 4600 0 0 0 0 Scenedesmus spp. 0 0 0 0 7500 8400 22560 10200 7000 8800 Rotifer Rhodomonas spp. 0 4240 0 0 0 0 0 0 0 0 Picoplankton 0 4640 0 0 0 0 0 0 0 5300 Microflagellates 0 0 0 7200 0 0 0 5000 6500 26500 Other 14836 5980 10189 28068 19778 18231 18340 10791 9072 12708 Total (cells/ml) 52116 35740 38589 103668 61978 55231 45880 43491 27872 53308

Table 3: Feeding regimen of copper-pulse experiment.

Experimental Day 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

Instar 1 1 1 1 1 & 2 2 2 2 2 & 3 3 3 3 3 & 4 4 4 4

Control 0 0 0.049 0.049 0 0.049 0.049 0 0 0.049 0.049 0 0 0.049 0.049 0 1st Instar 0 0 0.077 0.077 0 0.049 0.049 0 0 0.049 0.049 0 0 0.049 0.049 0 Amount Pulse of 2nd Instar 0 0 0.049 0.049 0 0.077 0.077 0 0 0.049 0.049 0 0 0.049 0.049 0 Algae Pulse Added 3rd Instar (g) 0 0 0.049 0.049 0 0.049 0.049 0 0 0.077 0.077 0 0 0.049 0.049 0 Pulse 4th Instar 0 0 0.049 0.049 0 0.049 0.049 0 0 0.049 0.049 0 0 0.077 0.077 0 Pulse 40

FIGURES

500700 Per Season Per Application Acute Toxicity to ZP

600400

300500 µg Cu/L µg

200

100

0 North Paulding McComb Van Wert Shelby Van Wert Amick Powers Up. San. Findlay Wauseon Findlay Veterans Baltimore #2 #2 #3 #1 #2 #1 #2 #2 Memorial Figure 1: Typical copper sulfate application regimens at 13 treated reservoirs based on reservoir manager interviews. Copper sulfate is 25.2% copper by weight; gray bars indicate the amount of copper received in one CuSO4 application; black bars indicate the total amount received over the course of one season (May – September). Dashed line indicates the average concentration of copper which is acutely toxic to zooplankton according to Buchman (2008). 41

North Baltimore

> 130 µg Cu/L

< 130 µg Cu/L

Figure 2: Cluster analysis of reservoir-specific seasonal CuSO4 application rates (µg Cu/L applied as CuSO4) in 13 upground reservoirs. Two groups of reservoirs appear. Note: North Baltimore Reservoir grouped alone due to the high application rate; for analytical purposes it was included in the "high application rate" group.

Figure 3: Diagram of a sediment trap. The plastic collection containers were held in place by bungee cords (pictured) and the support basket contained bricks for ballast (not pictured). 43

90 PR, Particulate Cu Epilimnion 80 PR, Dissolved Cu 70 VMR, Particulate Cu VMR, Dissolved Cu 60 50 40 30 20 10 0 Cu/L Jul Jul Jul 90 Jul Jul Jul Jul Jul - - - Jun - - - - - Jun Jun Jun Jun Jun µg Jun Jun Jun Jun Jun Jun Jun Jun Jun May May May ------1 3 5 7 9

Hypolimnion1 3 5 7 9 11 80 13 15 11 13 15 17 19 21 23 25 27 29 26 28 30 70 60 50 40 30 20 10 0

Figure 4: Total concentration of copper, separated into dissolved and particulate-bound forms, in the epilimnion and hypolimnion at Paulding Reservoir (PR) and Veterans Memorial Reservoir (VMR) before and after two CuSO4 applications at Paulding Reservoir (indicated by arrows). Horizontal dashed line indicates the concentration of dissolved copper acutely toxic to zooplankton according to Buchman (2008). In each application, 73 µg Cu/L was added. 44

16 Epilimnion F2R, Particulate Cu 14 F2R, Dissolved Cu BR, Particulate Cu 12 BR, Dissolved Cu 10 8 6 4 2 0 16 Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Jul Cu/L

- Hypolimnion------Aug Aug Aug Aug Aug - - - - - 1 2 3 4 5 µg 14 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 12 10 8 6 4 2 0

Figure 5: Total concentration of copper, separated into dissolved and particulate-bound forms, in the epilimnion and hypolimnion at Findlay #2 Reservoir (F2R) and Bresler Reservoir (BR) before and after CuSO4 application at Findlay #2 Reservoir (indicated by arrows). On three consecutive days, a total of 21 µg Cu/L was added. 45

3.5 Veterans Memorial Reservoir 3.5 Bresler Reservoir Paulding Reservoir Findlay #2 Reservoir

3 3

2.5 2.5

2 2 µg/L 10 log 1.5 1.5

1 1

0.5 0.5

0 0

Figure 6: Log10-transformed crustacean zooplankton dry biomass (mean ± SE, n = 3) at all reservoirs. Arrows indicate CuSO4 treatment at Paulding Reservoir and Findlay #2 Reservoir.

Figure 7: Log10-transformed crustacean zooplankton density (mean ± SE, n = 3) at Veterans Memorial Reservoir (VMR) and Paulding Reservoir (PR) before and after two CuSO4 treatments at Paulding Reservoir (indicated by arrows). 47

Figure 8: Log10-transformed crustacean zooplankton density (mean ± SE, n = 3) at Bresler Reservoir (BR) and Findlay #2 Reservoir (F2R) before and after CuSO4 treatment at Findlay #2 Reservoir (indicated by arrows). 48

10000 Probable Effects Level Threshold Effects Level

B A A A A

1000

E 100 D, E C, D C sediment copper sediment µg Cu/g dry weight dry Cu/g µg

Figure 9: Copper concentration (mean ± SE, n = 2) in the top 1 cm of sediment at Paulding Reservoir (shaded bars) and Veterans Memorial Reservoir (open bars) before and after two CuSO4 applications at Paulding Reservoir (indicated by arrows). Letters denote significant within-reservoir differences as determined by an ANOVA (Paulding Reservoir: F4,5 = 10.78, P = 0.011; Veterans Memorial Reservoir: F3,4 = 20.36, P = 0.007) with post-hoc Tukey-Kramer HSD test for each reservoir. 49

10000 Probable Effects Level Threshold Effects Level

1000

B A A, B 100

C C C sediment coppersediment µg Cu/g dry weight dry Cu/g µg

Figure 10: Copper concentration (mean ± SE, n = 2) in the top 1 cm of sediment at Findlay #2 Reservoir (shaded bars) and Bresler Reservoir (open bars) before and after CuSO4 application at Findlay #2 Reservoir (indicated by arrows). Letters denote significant within-reservoir differences as determined by an ANOVA (Findlay #2 Reservoir: F2,3 = 28.05, P = 0.011; Bresler Reservoir: F2,3 = 0.59, P = 0.607) with post-hoc Tukey-Kramer HSD test for each reservoir. 50

1200 1200 Veterans Memorial Reservoir Bresler Reservoir Paulding Reservoir Findlay #2 Reservoir

1000 1000

) 800 2 800

600 600

400 400 density (chironomids/m density

200 200

0 0

Figure 11: Densities of third and fourth instar chironomids (HW > 0.4 mm, Mean ± SE, n = 3) at all reservoirs. Arrows indicate CuSO4 treatment at Paulding Reservoir and Findlay #2 Reservoir. 51

0.8 6000 Organic Matter B Inorganic Matter 0.7 Cu Concentration E 5000

0.6 D,E

10.6% C 4000

/day 0.5 2 16.0% 0.4 3000 D

0.3 sedimentation rate sedimentation

15.8% weight dry Cu/g µg A 2000 mg dry weight/cm dry mg 89.4% 0.2 84.0%

84.2% 1000 0.1

0 0 6/23 - 6/28 6/28 - 7/1 7/1 - 7/13

Pre-Application Post-Application Pre-Application Post-Application Figure 12: Composition and copper concentration of material collected in sediment traps before and after one CuSO4 application (28-June) at Paulding Reservoir (mean ± SE, n = 3). Sediment was collected for 5 days before application and 3 and 15 days after application. Letters denote significant between-day differences as determined by an ANOVA (sedimentation rate: F2,6 = 8.77, P = 0.03; Cu concentration: F2,6 = 73.34, P < 0.001) with post-hoc Tukey-Kramer HSD test for all days. 52

1.0 Proportions by Life Stage Adults 0.9 Pupae Larvae 0.8 B B 0.7 B

0.6 3.6%

A,B 16.9% 14.5% 0.5 28.7%

16.5% 15.4% 1.4% 0.4 13.1%

Proportion Surviving Proportion 0.3

A 67.7% 0.2 66.6% 70.1% 85.5% 15.7% 0.1 84.3%

0.0 1st 2nd 3rd 4th Control

Instar at Exposure Figure 13: Proportion Chironomus riparius surviving (mean ± SE, n = 6) following exposure to copper pulses at different instars. Letters denote groups with significant differences in the proportion of total individuals surviving as determined by an ANOVA (F4,25 = 5.16, P = 0.004) with post-hoc Tukey-Kramer HSD for all treatments.