SDSU Template, Version 11.1

THE TOXICITY OF COPPER AND ZINC UNDER PULSED EXPOSURE

REGIMES TO PURPLE SEA URCHINS, STRONGYLOCENTROTUS

PURPURATUS, AND MYSID SHRIMP, AMERICAMYSIS BAHIA

______

A Thesis

Presented to the

Faculty of

San Diego State University

______

In Partial Fulfillment

of the Requirements for the Degree

Master of Science in Public Health

with a Concentration in

Environmental Health Sciences

______

Jacob Paul Munson-Decker

Spring 2017

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ABSTRACT OF THE THESIS

The Toxicity of Copper and Zinc under Pulsed Exposure Regimes to Purple Sea Urchins, Strongylocentrotus purpuratus, and Mysid Shrimp, Americamysis bahia by Jacob Paul Munson-Decker Master of Science in Public Health with a Concentration in Environmental Health Sciences San Diego State University, 2017

In San Diego Bay, marine organisms are exposed to non-point source stormwater effluent containing the contaminants of concern copper and zinc. United States Environmental Protection Agency (U. S. EPA) Whole Effluent Toxicity (WET) test methods were developed to evaluate the toxicity of continuous point source discharges. These tests are now applied to episodic stormwater discharges. There is concern that static or static-renewal WET methods are not representative the episodic, short-term nature of stormwater discharges. This study modified WET test methods to assess the relative toxicity of copper and zinc under three discharge scenarios, corresponding to the 50th, 75th, and 95th percentile of historical rainfall durations in San Diego. Pulsed contaminant studies were initiated with two common WET test species, the purple sea urchin (Strongylocentrotus purpuratus), and mysid shrimp (Americamysis bahia), representing chronic and acute toxicity test endpoints. To mimic San Diego rain events, laboratory assays were performed for copper and zinc individually and as mixtures at time durations of 3, 6, and 12-hours, followed by transfer to uncontaminated filtered seawater from San Diego Bay for the remainder of the traditional 96- hour WET testing period. LC50 and EC50 values were calculated for copper and zinc exposures at each time-point. Traditional 96-hour static reference toxicant tests were performed concurrently as a means to compare static exposure LC/EC50 values to those found from pulsed toxicity tests. For both sea urchin and mysid shrimp, an increase in contaminant exposure time corresponded to increased toxicity, resulting in LC/EC50 values up to two orders of magnitude greater than standard 96-hour tests. Copper and zinc mixtures elicited less than additive toxicity for all exposure regimes, suggesting that mixtures were less toxic than when exposed as single metal exposures for these contaminants. Exposure times used in standard WET testing likely overestimate the toxic effects of short-term episodic stormwater discharges influencing San Diego Bay. The results suggest that current WET testing protocols may be unnecessarily conservative for estimating episodic discharges toxicity.

TABLE OF CONTENTS

PAGE

ABSTRACT ...... iv LIST OF TABLES ...... vii LIST OF FIGURES ...... ix ACKNOWLEDGEMENTS ...... x CHAPTER 1 INTRODUCTION ...... 1 1.1 Statement of the Problem ...... 1 1.2 Purpose of the Study ...... 2 1.3 Theoretical Bases and Organization ...... 3 2 LITERATURE REVIEW ...... 4 2.1 Background on California Water Quality Criteria ...... 4 2.2 Whole Effluent Toxicity Testing ...... 6 2.3 Stormwater Toxicity ...... 7 2.4 Contaminants of Concern ...... 9 2.5 Contaminant Mixtures ...... 10 2.6 Pulsed Exposures ...... 12 3 METHODOLOGY ...... 15 3.1 Study Design ...... 15 3.2 Pulsed Exposure Durations ...... 16 3.3 Test Organisms ...... 17 3.4 Test Materials...... 17 3.5 Acute Toxicity Test Methods...... 18 3.6 Chronic Toxicity Test Methods ...... 19 3.7 Pulsed Copper and Zinc Exposure Tests ...... 21 3.8 Pulsed Copper and Zinc Mixture Exposure Test ...... 23

3.9 Statistical Analysis ...... 23 4 RESULTS ...... 26 4.1 QA/QC: Single Metal Copper Exposure Study ...... 26 4.2 Copper Exposure Study ...... 26 4.3 Zinc Exposure Study ...... 29 4.4 Binary Metal Mixture Pulsed Exposure ...... 33 5 DISCUSSION ...... 37 5.1 Implications of Pulsed Exposure Toxicity and Risk Characterization ...... 37 5.2 Integration of Mixed Metal Exposures in Pulsed Toxicity Tests ...... 40 5.3 Implications of Pulsed Exposure Testing Modifications on Organism Response and Regulatory Considerations ...... 41 6 CONCLUSION ...... 44 REFERENCES ...... 46 APPENDIX DEFINITION OF TERMS ...... 53

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LIST OF TABLES

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Table 2.1. Water Quality Objective Criteria Limitations for Metals Recognizing Effluent Treatment Variations ...... 5 Table 3.1. Analytical Methods, Detection (MDL), and Reporting Limits (RL) for Copper and Zinc ...... 18 Table 3.2. Test Methodology and QA/QC Requirements for 3, 6 & 12-hr Pulsed Toxicity Test Using Americamysis bahia (mysid shrimp) Exposure to Copper and Zinc Individually ...... 20 Table 3.3. Test Methodology and QA/QC Requirements for 3, 6 & 12-hrs Pulsed Toxicity Test Using Strongylocentrotus purpuratus (purple sea urchin) Exposure to Copper and Zinc Individually ...... 22 Table 3.4. Test Methodology and QA/QC Requirements for 3, 6 & 12, 96-hr Exposure to Copper and Zinc Mixtures Using Americamysis bahia (Mysid Shrimp) ...... 24 Table 4.1. Static and Pulsed Exposure Nominal and Verified Copper Concentrations ...... 27

Table 4.3. Summary of Median Lethal (LC50) Copper Concentrations for Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) ...... 29 Table 4.4. Summary of Nominal and Verified Zinc Concentrations from Static and Pulsed Exposures ...... 30

Table 4.5. Summary of Median Effective (EC50) Zinc Concentrations for Purple Sea Urchin Embryo-Larval Development Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL)...... 31

Table 4.6. Summary of Median Lethal (LC50) Zinc Concentrations for Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL)...... 32 Table 4.7. Summary of Nominal and Verified Copper and Zinc Concentrations for Mysid Survival Mixed Metal Static and Pulsed Exposures ...... 33

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Table 4.8. Summary of Median Lethal (LC50) Copper and Zinc Concentrations for the Single Metal Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) ...... 34 Table 4.9. Summary of Median Lethal Copper and Zinc Concentrations for the Mixed Metal Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) ...... 34 Table 4.10. Combined Toxic Unit Calculations for Mixed Metal Tests for Each Pulsed Time Exposure ...... 35 Table 5.1. Area under the Curve (Concentration x Duration = AUC) Calculations for Single Metal Exposures ...... 38

LIST OF FIGURES

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Figure 3.1. Aggregate incidence of rainfall intervals over 24-hr periods recorded from 1951 to 2006 (N= 2,284). Laboratory based pulsed exposure experiments were based on the 50th, 75th and 95th percentiles rainfall durations...... 16 Figure 3.2. Pulsed exposure experimental design and testing times...... 17 Figure 3.3. Purple sea urchin exposure in 25µm Nitex screen polycarbonate tubes placed in 400mL HDPE tri-corner beakers...... 19 Figure 4.1. Sea urchin embryo-larval development test median effective concentrations (EC50) for copper at each exposure interval. 96-hr static* data was included from Arnold et al., 2010. The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario...... 28

Figure 4.2. Mysid shrimp Survival test median lethal concentrations (LC50) for copper at each exposure interval. The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario...... 29

Figure 4.3. Median effective concentrations (EC50) for zinc derived from the purple sea urchin embryo-larval development test. The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario...... 31

Figure 4.4. Median lethal concentrations (LC50) for zinc derived from the mysid survival test. The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario...... 32

Figure 4.5. Median lethal (LC50) concentrations for copper and zinc derived from the mysid survival test with mixed metals. The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario...... 35 Figure 4.6. Toxic Units (TU) for copper and zinc derived from the mysid survival test with mixed metals...... 36

ACKNOWLEDGEMENTS

I would like to thank to my committee: Dr. Rick Gersberg for his guidance over the last two and a half years, Dr. Eunha Hoh, and Dr. Natalie Mladenov. Moreover, I have to give a huge thank you to Gunther Rosen and Molly Colvin of SPAWAR Systems Center Pacific, for providing me with all the resources, funding, equipment, and expert guidance allowing me to complete this research. My time working under Gunther and Molly has been, and continues to be, invaluable. In addition, a huge thank you to Nick Hayman for help on all the late night pulsed testing. Naval Base San Diego supported this research, and the Space and Naval Warfare Systems Center (SSC) Pacific granted funding.

CHAPTER 1

INTRODUCTION

Anthropogenic contaminants from urban areas enter marine environments via stormwater runoff. Urban runoff can contain a multitude of contaminants that may be toxic to marine organisms, making bays and shorelines at risk environments to stormwater contamination. As a result, facilities releasing stormwater effluent into marine environments are required to comply with increasingly rigorous National Pollution Discharge Elimination System (NPDES) permits. Due to differing sampling and toxicity test methodologies, there is much debate over the actual contribution stormwater runoff has on contaminant loading in the aquatic environment. Toxicity as a function of dose, frequency, degree, and exposure time can be difficult to predict in the environment. Tidal flushing, currents, bioavailability of contaminants, and the duration and amount of the inflowing stormwater all play an essential role in determining the toxicity of urban runoff. Current regulatory testing requirements do not incorporate storm duration, intensity, or pollutant load over time when assessing the toxicity of stormwater runoff. By performing acute and chronic pulsed bioassays on sensitive marine species, this study investigated the toxicity of common metal contaminants present in stormwater samples over exposure regimes representative of storm events adjacent to San Diego Bay, California.

1.1 STATEMENT OF THE PROBLEM Stormwater runoff from urban areas is an element of non-point source contamination of marine environments (Kayhanian, Stransky, Bay, Lau, & Stenstrom, 2008; Schiff, Bay, & Diehl, 2003; United States Environmental Protection Agency [U.S. EPA], 1995; Whipple, Hunter, & Yu, 1974). To reduce the impact of stormwater on marine environments, regulatory entities have advocated the use of NPDES permits and static or static-renewal toxicity tests on stormwater effluent (U.S. EPA, 1995, 2002). This involves the continuous

(i.e. static) exposure of sensitive marine organisms to first-flush stormwater samples, collected from the end-of-pipe (EOP), for periods of 48-hrs or more (U.S. EPA, 2000, 2002). Using standardized procedures, these bioassays expose organisms to effluent samples under a fixed duration regardless of in situ exposure dynamics, species, or toxic mode of action (Diamond, Klaine, & Butcher, 2006; U.S. EPA, 1995). There is concern that static or static- renewal Whole Effluent Toxicity (WET) test methods are not representative of effects associated with the episodic, short-term nature of stormwater discharges. Although there have been extensive studies on stormwater toxicity under static conditions, few have measured it based on real-world episodic discharge intervals (Butcher et al., 2006; Diamond et al., 2006; Handy, 1994; Hoang, Gallagher, Tomasso, & Klaine, 2007; Hoang, Tomasso, & Klaine, 2007).

1.2 PURPOSE OF THE STUDY This study modified WET testing procedures to investigate the relative impact of exposures times on the acute and chronic toxicity of the metals copper and zinc, individually and as mixtures. This research represents a novel contribution in assessing the relative toxicity of copper and zinc individually and in mixtures under pulsed exposure regimes, likely to occur during storm events. To determine if laboratory based bioassays can effectively predict in situ exposure dynamics, the median lethal (LC50) and median effective

(EC50) concentrations of copper and zinc were determined under pulsed exposure intervals. This research supports previous findings that pulsed bioassays representing real-world contaminant exposure times can potentially provide an accurate assessment of the toxicity dynamics occurring in the environment (Angel, Simpson, Chariton, Stauber, & Jolley, 2015; Angel, Simpson, & Jolley, 2010; Butcher et al., 2006; Diamond et al., 2006; Handy, 1994; Hoang, Gallagher, et al., 2007; Hoang, Tomasso, et al., 2007). Since there is an increased need to develop methodology to quantitatively predict metal mixture toxicity (Meyer, Farley, & Garman, 2015), this research also provides a preliminary assessment of mixed metal toxicity under pulsed exposure scenarios. Filling gaps in current data available, this study will aid in understanding how stormwater effluent exposure dynamics affect San Diego Bay, California. The objectives of this study are described below.

1. Quantify the 50th, 75th, and 95th percentile of historical rainfall durations at San Diego International Airport from data previously collected over a 55-year period. 2. Conduct individual acute and chronic copper and zinc toxicity tests for 96-hr static exposures, and pulsed exposure regimes defining the 50th, 75th, and 95th percentile of historical rainfall durations adjacent to San Diego Bay, on 2 test species commonly used in Southern California industry permitting bioassays. 3. Determine the relative antagonistic, synergistic, and additive toxicity of copper and zinc mixtures under pulsed exposure regimes, on one test species used in Southern California industry permitting bioassays. Determine if reduced exposure times change metal interactions and toxicity dynamics. 4. Calculate the proportion difference between static WET test LC/EC50 values and those LC/EC50 values found under pulsed exposure regimes.

1.3 THEORETICAL BASES AND ORGANIZATION Stormwater entering the marine environment can contain many chemicals, known to be toxic to marine organisms. To protect the marine ecosystems from non-point source contamination by stormwater discharge, the Clean Water Act (CWA) mandates the enforcement of NPDES permits for such discharges. Standardized testing protocols utilize criteria, defined by a fixed duration and frequency regardless of the chemical, toxic mode of action, species tested, and unique environmental exposure dynamics. In turn, toxicity testing protocols utilizing continuous exposures do not reflect variable or episodic discharge regimes (Diamond et al., 2006; Reinert, Giddings, & Judd, 2002; U.S. EPA, 1995). Providing a margin of safety, permits often assess stormwater toxicity over periods much longer than discharge events. Additionally, water quality criteria for metals are determined on an individual basis, negating any of the relative interactions contaminants may have in the environment (Diamond et al., 2006; Meyer et al., 2015). All previous studies have either examined the acute and chronic toxicity of pulsed exposure regimes or mixed toxicant exposures, but not both. Therefore, it is important to determine the toxicity of metal mixtures over varying exposure intervals. In this study, we investigated the acute and chronic toxicity of copper and zinc, individually and in mixture, over varying pulsed intervals representative of San Diego storm events to purple sea urchins (Strongylocentrotus purpuratus), and opossum (mysid) shrimp (Americamysis bahia). A list of technical terms is found in the Appendix.

CHAPTER 2

LITERATURE REVIEW

2.1 BACKGROUND ON CALIFORNIA WATER QUALITY CRITERIA The California State Water Resource Control Board (SWRCB) adopted the Water Quality Control Plan, Ocean Waters of California in 1972 to conform to CWA water quality criteria (California Environmental Protection Agency, 2012). As defined by the U.S. EPA (1991c), water quality criteria are defined by a narrative describing ideal water quality goals, and limits on contaminant concentrations aimed at protecting human health and aquatic life. Numeric contaminant concentrations are obtained from a variety of acute and/or chronic toxic endpoints, and the potential for contaminant bioaccumulation (U.S. EPA, 1987b, 1991a). The SWRCB revised the California Ocean Plan narrative in 2012, declaring that there will no longer be degradation of marine communities due to waste discharges from coastal outfalls, or exceedance of water quality objectives for all near coastal ocean waters. If a discharge contributes to the degradation of marine communities, or water quality objectives are exceeded, the SWRCB mandated that limitations and controls be placed on dischargers (California Environmental Protection Agency, 2012). The CWA requires water quality reports every two years identifying water quality trends, prioritizing polluted waters, and establishing targets for contaminant thresholds defined by Total Maximum Daily Load (TMDL) criteria (U.S. EPA, 2016b). As a part of the CWA Section 303(d), TMDLs limit the magnitude of contaminants entering bodies of water by determining the maximum amount of a pollutant a water body can receive while still maintaining water quality standards (SWRCB, 2014; U.S. EPA, 2016a). TMDLs are characterized by point source waste load allocations (WLA), and non-point source load allocations (LA). The TMDL also incorporates seasonal variations in water quality, and a margin of safety (MOS) accounting for uncertainty in pollution reductions meeting water

5 quality standards. TMDLs are calculated using the equation represented below (U.S. EPA, 2015a).

(1.1)

To meet TMDL criteria, effluent can be subject to regulation under the NPDES permit program (California Environmental Protection Agency, 2012; U.S. EPA, 2014, 2015b). As a means of maintaining water quality objectives, three criteria limitations are enforceable on effluent metal concentrations. Recognizing normally occurring variation in effluent treatment, analytical techniques, and sampling methodology TMDL criteria are defined by the 6-month median, daily maximum, and instantaneous maximum (Table 2.1; California Environmental Protection Agency, 2012).

Table 2.1. Water Quality Objective Criteria Limitations for Metals Recognizing Effluent Treatment Variations Metal 6-Month Median Daily Maximum Instantaneous μg/L μg/L Maximum μg/L Arsenic 8 32 80 Cadmium 1 4 10 Chromium 2 8 20 Copper 3 12 30 Lead 2 8 20 Mercury 0.04 0.16 0.4 Nickel 5 20 50 Selenium 15 60 150 Silver 0.7 2.8 7 Zinc 20 80 200 Source: California Environmental Protection Agency. (2012). Water quality control plan ocean waters of california. Sacramento, CA: State Water Resources Control Board.

The SWRCB has determined according to 33 U.S.C. §§ 1311, 1342 of the CWA, that certain stormwater discharges containing regulated pollutants are subject to NPDES permitting obligations. To comply with NPDES permits, U.S. EPA (1991a) WET test methods evaluate the toxicity of continuous point source discharges (Chapman, 2000). The permit acts as the explicit requirement customized for individual effluent sources to limit their impact on bodies of water, and the ability of the body of water to maintain TMDL water quality criteria (California Environmental Protection Agency, 2012; U.S. EPA, 2014, 2015a, 2015b). Thus, mandating what contaminants may be discharged, contaminant concentration

6 discharge limits, monitoring and reporting requirements, and other provisions to ensure that discharges do not impair water quality or human health (U.S. EPA, 2015a). The U.S. EPA requires that NPDES permits maintain the most stringent technology and water quality based controls as required by the CWA (California Environmental Protection Agency, 2012). The CWA Section 301(b) and 40 Code of Federal Regulations mandates that NPDES stormwater discharge permits include the implementation of the Best Available Technology (BAT) and Best Conventional Pollutant Control Technology (BCT) necessary for receiving waters to meet applicable water quality standards (SWRCB, 2014). These standards are based on a variety of criteria including biological techniques, chemical-specific analysis, and WET testing (SWRCB 2014; U.S. EPA, 1991c). The same protocols and toxicity test methods are used to evaluate episodic non-point source stormwater discharges from facilities along San Diego Bay (Katz, Rosen, & Arias, 2006). Among the metals categorized in (Table 2.1), a previous study identified copper and zinc as the primary inorganic toxicants of concern entering San Diego Bay via stormwater runoff (Katz et al., 2006).

2.2 WHOLE EFFLUENT TOXICITY TESTING WET testing methods involve both acute and chronic toxicity tests to regionally acceptable sensitive test species. Acute toxicity tests examine the lethality of an effluent sample for a period of 48-hrs or more, while chronic toxicity tests typically involve exposures of seven day or more. In this study, short-term estimates of chronic toxicity were used to examine sublethal responses (normal/abnormal embryonic development) for a period of 96-hrs or less (U.S. EPA, 1995). U.S. EPA methods, test effluent samples on a gradient dilution series defined by effluent concentrations percentages of 100, 50, 25, 12.5, and 6.25 %, in addition to a lab control. However, stormwater samples are tested at 100% only (Swietlik et al., 1991; U.S. EPA, 1995; Wang, 1990). To meet acceptability criteria laboratory controls, in both acute and chronic tests, must have ≥ 90 % normal organism development or survival (U.S. EPA, 1995). Additionally, laboratory controls and reference toxicant tests are conducted as a quality assurance/quality control (QA/QC) measure.

Reference toxicant tests yield a LC50 or EC50 for each species tested. Results are compared with the laboratory’s historical test means (± two standard deviations) to ensure laboratory test precision and as a measure of organism health and sensitivity (U.S. EPA, 1991a, 1991b).

The U.S. EPA (1991c) conducted the Complex Effluent Toxicity Testing Program (CETTP) to determine the ability of WET tests to predict receiving water conditions. For marine environments, this program examined ambient water toxicity by releasing dye into wastewater discharges, and then measured the relative concentrations of dyed effluent in receiving waters. Toxicity tests were performed to assess effluent dilution toxicity, in comparison to the ambient receiving water toxicity. For marine ecosystems, the CETTP found that WET tests were able to predict receiving water health with 94% accuracy. Accordingly, in only 6% of the tests predicting EOP toxicity was there no toxicity observed in the receiving environment (Dickson, Waller, Kennedy, & Ammann, 1992; Schimmel, Morrison, & Heber, 1989; Schimmel, Thursby, Heber, & Chammas, 1988; U.S. EPA, 1989). However, the ability of the CETTP methodology to determine WET testing accuracy is questionable. The U.S. EPA (1989) report acknowledges that locations used in the studies were not randomly selected, and did not represent a statistically defensible sample size over time. Moreover, the U.S. EPA acknowledged that impact correlations between effluent and receiving water toxicity are lower where minimal impacts were expected and higher where greater toxicity occurred. It should be noted that variable in situ factors such as flow, dilution, chemical bioavailability, dissolved oxygen, temperature, channelization, flooding and weather cycles can all impact the relative correlations between WET test results and ambient receiving water health (U.S. EPA, 1991c).

2.3 STORMWATER TOXICITY Urban stormwater contributes a significant portion of pollution to many receiving water bodies (Lee, Lau, Kayhanian, & Stenstrom, 2004). The toxicity of this discharge is variable and often dependent on the season and stage of the storm that the effluent is sampled (Taebi & Droste, 2004). Extended dry conditions increase contaminant concentrations at urban outfalls during seasonal first flush stormwater events (Lee et al., 2004; Stenstrom & Kayhanian, 2005). The U. S. EPA (1993) collected stormwater grab samples throughout a storm event and generated a pollutograph demonstrating that the contaminant load in stormwater effluent decreases over the duration of a storm. A study conducted as part of an interagency agreement between the California Department of Transportation (CalTrans), Division of Environmental Analysis, and the

University of California, Los Angeles, identified the “first flush phenomenon” as the first portion of the storm, where a majority of the contaminants loaded during a dry period flush into stormwater effluent (Stenstrom & Kayhanian, 2005). This study noted that 30-50% of pollutants observed in highway runoff were contained in the first 10-20% of the flow. Additional studies have shown that up to 80% of the contaminants of concern are contained in the first 30% of the “first flush” stormwater discharge (Kayhanian et al., 2008). Accordingly, Kayhanian et al. (2008) showed that the first 30% of the flow is responsible for 80% of the toxicity, and the other 70% of the flow is only potentially responsible for 20% of toxicity. As a result, bioassays utilizing first flush EOP samples may be more toxic than composite effluent samples taken over the length of the storm. Southern California storm events show first flush contaminant loads are often only representative of the first few hours of runoff sampled (Kayhanian et al., 2008; Stenstrom & Kayhanian, 2005). The associated toxicity of the identified grab samples corresponds to these first flush events. Accordingly, the 96-hr (or less) acute exposure tests used in bioassays for compliance may not represent first-flush exposure dynamics in receiving waters. Moreover, Best Management Practices (BMPs) dependent on volume of the water treated could potentially focus on contaminant concentrations released during the first-flush stage of the storm, rather than allocating resources to mitigate portions of the storm representing a less significant percentage of the toxicity (Bertrand-Krajewski, Chebbo, & Saget, 1998; Granato, Zenone, & Cazenas, 2002). An NPDES permit written for Naval Submarine Base San Diego requires that undiluted stormwater runoff adjoining industrial processes, cannot produce less than 90% survival 50% of the time and no less than 70% survival 10% of the time (Katz et al., 2006). A study by Katz et al. (2006) utilizing a multiple lines of evidence approach, showed that first- flush storm water samples failed NPDES permit criteria 58% of the time, while composite samples representing the total storm discharge failed permit criteria 25% of the time. The observed toxicity was primarily due to copper, zinc, and surfactants. This line of evidence demonstrated that the first-flush was responsible for a majority of the failed tests at these facilities. Although a majority of first-flush discharges failed permit requirements, Katz et al. (2006) found that in 202 receiving water tests, samples were statistically indistinguishable from lab controls (p< 0.05). The magnitude of this study, assessing 51 stormwater outfalls

9 and 85 corresponding bay water samples, demonstrates that testing criteria may not represent water quality dynamics occurring in San Diego Bay (Katz et al., 2006).

2.4 CONTAMINANTS OF CONCERN Copper (Cu) is a naturally occurring metal found in rocks, soil, water, and air (Agency for Toxic Substances and Disease Registry [ATSDR], 2004). Various anthropogenic sources have contributed to increases in ambient copper levels in the marine environment, leading to a decline in overall ecosystem health (Irwin, Van Mouwerik, Stevens, Seese, & Basham, 1997a). A report prepared by TDC Environmental (2004), for the Clean Estuary Partnership cites brake pads, architectural copper, pesticides, industrial copper use, soil erosion, domestic discharges, vehicle fluid leaks, and antifouling paints as the primary contributors of copper to non-point source urban runoff. Copper can form extremely stable oxides (Cu2+) in the marine environment, while ionic Cu+1 is only stable if part of a complex ion in aqueous solution (Irwin et al., 1997a). Once in the marine environment total colloidal particles, organismal bioaccumulation, dissolved complexes, particulate moieties, and bioavailable free aqueous copper ions (Cu2+; Campbell, 1995; Eriksen, Mackey, van Dam, & Nowak, 2001; Rivera-Duarte et al., 2005; Sunda & Ferguson, 1983) represent the total copper load. Aqueous Cu2+ is indicative of the portion of copper available for uptake by organisms, and thus reflects its potential toxicity in aquatic ecosystems (Moffett & Brand, 1996; Rivera-Duarte et al., 2005). Studies have shown that a majority of the total dissolved copper load in the environment is not in this bioavailable form (Eriksen et al., 2001; Rivera- Duarte et al., 2005). While copper is a necessary dietary element, elevated levels can create a disruption in the ion balance of marine organisms affecting growth, reproduction, enzyme activity, system chemistry, and metabolism (Eisler, 1985; Jenkins, 1981). Excessive exposure to copper in the marine environment is toxic to various species of algae, annelids, crustaceans, cyprinids, and salmonids (Irwin et al., 1997a). Section 307 (a) (1) of the CWA, lists copper as one of 126 priority pollutants (previously 129), subject to effluent limitations (Keith & Telliard, 1979; U.S. EPA, 2014). For San Diego Bay, research has indicated copper as one of the primary contaminants of concern in urban stormwater effluent (Katz et al., 2006).

Comprising roughly 0.0005-0.02% of the Earth’s crust, zinc (Zn) is a metal found in approximately 55 mineralized forms (ATSDR, 2005; Irwin, Van Mouwerik, Stevens, Seese, & Basham, 1997b). Although zinc is a naturally occurring, anthropogenic uses such as rubber, paints, galvanized steel, and numerous metal alloys have contributed to an increase in the ambient zinc content of marine ecosystems (Irwin et al., 1997b; National Center for Biotechnology Information, 2016). Zinc naturally combines with various compounds to form salts, the most common of which is zinc sulfate. This results in the formation of complexes with organic and inorganic ligands, which can increase the solubility and bioavailability of zinc in the marine environment (ATSDR, 2005; Irwin et al., 1997b). In aqueous solution, zinc exists only as the Zn2+ oxide (U.S. EPA, 1987a). Moreover, zinc is an essential macronutrient responsible for protein synthesis, metabolizing nucleic acids, carbohydrate metabolism, and stabilizing enzyme membranes in humans and animals (Viarengo et al., 1985; World Health Organization [WHO], 2001). Essential for plant and animal life as a trace element, at low concentrations zinc is a precursor of DNA and RNA polymerases and some metalloenzymes (Irwin et al., 1997b; Keen & Hurley, 1989). Additionally, increased zinc intake can act as a protective mechanism against copper and cadmium toxicity, as zinc competes for the same binding sites (Les & Walker, 1984). However, in the marine environment elevated zinc concentrations can be especially toxic to molluscs, algae, crustaceans, and salmonids (Gore & Bryant, 1986). The relative toxicity of zinc to these species is often associated with a disruption of the internal ion balance (U.S. EPA, 1993). Similar to copper, zinc has been listed as one of 126 priority pollutants subject to effluent controls under 40 CFR Part 423, Appendix A of the CWA (Keith & Telliard, 1979; U.S. EPA, 2014).

2.5 CONTAMINANT MIXTURES Urban stormwaters and ambient receiving waters contain an assortment of inorganic and organic contaminant mixtures (Masnado, Geis, & Sonzogni, 1995). The interactions and effects of these mixtures on marine organisms are not clearly defined (Phillips et al., 2003). Depending on numerous factors, contaminant combinations can have varying toxic effects (Kraak et al., 1994). Concerning trace metal contaminants, monitoring programs and water quality criteria TMDLs focus on single metal toxicity. Many researchers have recognized the

11 shortcomings of this approach, offering evidence of variable toxic responses for multiple contaminant exposures (Enserink, Maas-Diepeveen, & Van Leeuwen, 1991; Kraak et al., 1994; Masnado et al., 1995; Phillips et al., 2003). The toxicity of contaminant mixtures is dependent on organism sensitivity, the contribution of each toxicant in mixture, toxicant interactions, and the response of individual toxicants to varying water quality parameters (Altenburger, Boedeker, Faust, Grimme, 1996; Verslycke et al., 2003). The scope of work needed to determine the toxicity of contaminant mixtures among varying species and fluctuating water quality parameters is vast (Bellas, 2008). When multiple metals compete for binding sites, the combination can potentially produce a broad spectrum of joint effects such as synergism, antagonism, and additivity (Ahsanullah, Negilski, & Mobley, 1981; Bellas, 2006; Verslycke et al., 2003). Essential metals are necessary for proper cellular function. For example, copper acts as a catalyst for many enzyme systems, while zinc is essential for the proper function of cell membranes (Viarengo et al., 1985). However, at elevated concentrations in a mixture, copper and zinc compete for binding sites in cell membranes transport, metalloenzymes, metallothioneins, or sensitive molecules thus increasing metallothionein production and irregularities in lysosomal activity (Phillips et al., 2003; Sharma, Schat, Vooijs, & Van Heerwaarden, 1999). Phillips et al. (2003) demonstrated variable toxicological effects of metal mixtures using sea urchin embryo larval development tests. When compared to the toxicity of the individual metals, copper and zinc mixtures expressed antagonistic toxicity. Comparatively, when assessing the filtration rates of the freshwater mussel Dressena polymorpha, Kraak et al. (1994) found that copper and zinc mixtures had less of an impact than the individual metal exposures. Conversely, Lynch, Hoang, and O’Brien (2016) found copper and zinc exhibited a more than additive acute toxic effect on survival of Pimephales promelas. Moreover, a study by Meyer et al. (2015) found Daphnia magna exposed to copper and zinc mixtures resulted in additive toxicity, or a synergistic response depending variations in the respective dilution series. While these studies demonstrate variable toxic effects of copper and zinc in mixture, some studies have indicated that mixtures of a large number of metals typically results in additive toxicity, while binary metal mixtures can result in unpredictable or even contradictory toxic endpoints (Enserink et al., 1991; Kraak et al., 1994; Sharma et al., 1999; Verslycke et al., 2003). Although researchers have been unable to predict mixed metal

12 toxicity, the results of the aforementioned studies can be used to estimate the contribution of individual metals to the observed toxicity of effluent samples. This can in turn influence Toxicity Identification Evaluations (TIEs) used in NPDES permitting (Meyer et al., 2015; Phillips et al., 2003).

2.6 PULSED EXPOSURES Industrial and municipal stormwater discharges from San Diego coastlines occur during intermittent and episodic rain events. This results in fluctuating and irregular inputs of toxicants into receiving waters (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015; Katz et al., 2006). The concentrations of toxicants in stormwater effluent changes greatly as flow rates vary, and pollutants change and degrade in the receiving environment (Butcher et al., 2006; Reinert et al., 2002). Research has been performed describing both the acute and chronic toxicity under pulsed exposure regimes. These studies utilize different variables such as contaminant type, concentrations, length of pulsed exposures, effects of multiple exposures over time, and toxicant mixtures (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015; Angel et al., 2015; Angel et al., 2010; Butcher et al., 2006; Diamond et al., 2006; Hoang, Gallagher, et al., 2007; Hoang, Tomasso, et al., 2007; Hosmer, Warren, & Ward, 1998). Dependent on the array of variables, different pulsed methodology has elicited latent, greater, and lesser toxicity than traditional testing methods (Brent & Herricks, 1998; Diamond et al., 2006; Hosmer et al., 1998). In recent years, there has been a move to modify water quality standards using pulsed exposure bioassays reflecting episodic and variable contaminant exposures (Diamond et al., 2006). A study using D. magna and P. promelas measured the pulsed toxicity of copper, zinc, or ammonia under a range of concentrations, durations, frequencies, and recovery times. This research found no effects associated with 24-hr ammonia and copper pulses after the contaminant pulse was removed, while zinc pulses exhibited latent toxicity following organism transfer to contaminant-free test solutions. The results of the study did suggest that multiple exposures to zinc over a short time, followed by a period of no contaminant exposure, could result in greater toxicity than continuous zinc exposures (Diamond et al., 2006).

Angel et al. (2010) reported a significant interaction between inhibition of Phaeodactylum tricornutum biomass and the exposure concentration and duration of copper under pulsed exposure treatments. This study demonstrated a relationship between the Time- Averaged Concentration (TAC) of copper among 72-hr static exposures, and 1, 4, and 8-hr pulsed exposures, followed by transfer to uncontaminated water for the remainder of the 72- hr static exposure period. These results suggest that continuous exposures are marginally more toxic to P. tricornutum than pulsed copper exposure at equivalent TACs, since pulsed exposures at high copper concentrations have slow internalization rates because membrane transport proteins become saturated (Angel et al., 2010). In the case of P. tricornutum, this study suggested water quality criteria using continuous exposures is not representative of real-world toxicity dynamics when applied to short-term discharges (Angel et al., 2010). A technical report prepared for the City of San Diego assessed the pulsed toxicity of copper using locally representative marine organisms Holmesimysis costata, Strongylocentrotus purpuratus, and Mytilus galloprovincialis (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015). Pulsed exposure durations defined as 30-minutes, 1, 2, and 3-hrs covered a range of likely exposure scenarios for species inhabiting upper intertidal tide pools near Devil’s Slide, La Jolla. As a means of assessing the relative difference between traditional 48-hr/96-hr tests and pulsed exposure tests, this study compared relative LC50 and EC50 values for the varying exposure times. Results showed that static tests elicited a toxic response in all organisms at lower concentrations than those found under pulsed exposure regimes. During pulsed exposures, the 3-hr pulses generated the greatest toxic response in all three species, with decreasing toxicity associated with shorter exposure regimes (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015). The U.S. EPA (1992) defines ecological risk as the likelihood that an adverse effect may occur as the result of exposure to ecological stressors. Traditionally ecological risk assessment to aquatic ecosystems utilizes a deterministic approach defined by a hazard effect concentration, LC/EC50, or by a no-observed effect concentration (Barnthouse & Suter, 1986). However, ecological risk can also be defined in terms of probabilistic estimates of an adverse effect (Morton et al., 2000). A common method described by Hoang, Gallagher, et al. (2007), utilizes the Area Under the Curve (AUC) as an intuitive approach to characterize contaminant toxicity and risk to biota. As a function of exposure time and contaminant

14 concentration (AUC = concentration x duration), organisms with the same AUC would be subject to the same risk. However, research has shown that exposure time and contaminant concentrations are not always able predict toxic endpoints on a linear plane (Angel et al., 2010; Hickie, McCarty, & Dixon, 1995; Hoang, Gallagher, et al., 2007). The CWA narrative to eliminate discharges of toxics in toxic amounts has led to regulatory concern over stormwater discharges (U.S. EPA, 1991a). Although NPDES permits have traditionally regulated point-source pollutants, they now regulate EOP monitoring at stormwater outfalls. Additionally, WET test criteria may overestimate the toxicity of episodic contaminant discharges. As a result, modifications to traditional testing protocols are being researched and evaluated (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015; Angel et al., 2010; Brent & Herricks, 1998; Diamond et al., 2006; Hosmer et al., 1998; U.S. EPA, 1995, 2002). This study investigates the potential for pulsed exposure testing to better simulate stormwater runoff dynamics, than current static exposure methods, for an arid region such as San Diego Bay.

CHAPTER 3

METHODOLOGY

Considered contaminants of concern in stormwater runoff and ambient receiving waters, the chronic and acute endpoints of copper and zinc were measured at varying exposure times, individually and as mixtures. To simulate the potential toxicity of stormwater, which is often intermittent, this study modified standard U.S. EPA WET testing methods (U.S. EPA, 1995, 2002). Exposure intervals were changed from traditional 96-hr continuous contaminant exposure to 3, 6, or12-hrs, followed by transfer to uncontaminated seawater for 93, 90, or 84-hr periods, respectively. Acute and chronic toxicity endpoints were evaluated for copper and zinc individually with two standard marine test organisms, the purple sea urchin (Strongylocentrotus purpuratus) and the opossum (mysid) shrimp (Americamysis bahia). Additionally, a single test for mysid shrimp was conducted to assess the acute toxicity of copper and zinc in mixture under pulsed exposure scenarios.

3.1 STUDY DESIGN Pulsed toxicity tests on marine organisms were implemented under three distinct exposure scenarios: 1. Juvenile mysid shrimp and sea urchin embryos were exposed to seawater spiked with copper sulfate (CuSO4) for 3, 6, or 12-hr times periods, followed by a transfer to uncontaminated seawater for the remainder of the 96-hr test period. 2. Juvenile mysid shrimp and sea urchin embryos were exposed to seawater spiked with zinc sulfate (ZnSO4) for 3, 6, or 12-hr times periods, followed by a transfer to uncontaminated seawater for the remainder of the 96-hr test period.

3. Juvenile mysid shrimp were exposed to seawater spiked with CuSO4 and ZnSO4 in mixture for 3, 6, and 12-hr times periods, followed by a transfer to uncontaminated seawater for the remainder of the 96-hr test period. A static 96-hr metal mixture test was conducted concurrently.

4. Static 96-hr reference toxicant tests were conducted concurrently for CuSO4 and ZnSO4 individually, as a quality control measure to assess organism health, and as

a reference of comparison between standard method toxicity and pulsed exposure method toxicity.

3.2 PULSED EXPOSURE DURATIONS A study by Katz et al. (2006) showed a strong correlation between rainfall and EOP outfall discharge durations at Naval Submarine Base San Diego. Accordingly, experimental pulsed exposure times were determined using a database of 2,284 measureable precipitation events recorded at San Diego International Airport over a 55-year period (1951-2006). The dataset, generated from NOAA Satellite and Information Service (2016), was provided to the Navy (SSC Pacific) by Dr. Robert Pitt. The database quantified rainfall (≥ 0.1 inch) on an hourly basis for a given 24-hr period. To calculate the cumulative probable duration of any given San Diego storm event, the total number of hours of measureable precipitation received in a given 24-hour event was summed. To calculate relevant San Diego Bay stormwater pulse regimes, for use in the laboratory-based bioassays, the 50th, 75th, and 95th percentile of all 24-hour rainfall durations were determined as 3, 6, and 12-hrs, respectively (Figure 3.1).

The experimental design used in the laboratory-based pulsed exposure tests were modified from standard U.S. EPA (1995, 2002) methods and is shown in Figure 3.2. Standard methods expose organisms to undiluted samples for static periods up to 96-hrs. This experiment modified traditional testing regimes by exposing organisms to reference toxicants

(copper and zinc) for the aforementioned pulsed durations 3, 6, and 12-hrs. Following contaminant exposures, organisms were transferred to uncontaminated 0.45µm filtered seawater (FSW), collected near the mouth of San Diego Bay for the remainder of the 96-hr test thus ensuring the fulfillment of the standard WET test duration.

3 Hour 93 Hour 0.45µm Filtered Seawater Exposure Exposure

6 Hour 90 Hour 0.45µm Filtered Seawater Exposure Exposure

12 Hour 84 Hour 0.45µm Filtered Seawater Exposure Exposure

96 Hour Exposure

Figure 3.2. Pulsed exposure experimental design and testing times.

3.3 TEST ORGANISMS As a means of representing both acute and chronic toxicity endpoints, test species were selected based on their regional relevance and use in standard toxicity tests (U.S. EPA, 1995, 2002). Due to NPDES monitoring requirements, it is essential that quality test organisms are available year round. Accordingly, U.S. EPA (2002) approved species are characterized as sensitive to toxicants, easily cultured in a laboratory setting, and readily commercially available. Opossum (mysid) shrimp (Americamysis bahia) were used established the acute toxicity testing endpoint, while embryos of the purple sea urchin (Strongylocentrotus purpuratus) were used to determine the chronic toxicity testing endpoint.

3.4 TEST MATERIALS To minimize the variation of metal bioavailability altered by ionic binding to particulate matter, all reference toxicant dilutions were made using 0.45µm FSW drawn from the SPAWAR Systems Center Pacific (SSC Pac) seawater intake near the mouth of San Diego Bay, where dissolved organic carbon content is also low (Rosen et al., 2008). Toxicant

18 stock solutions were prepared at SSC Pac using reagent grade CuSO4 and ZnSO4 salts volumetrically added into FSW. All stock solutions and test dilutions were analyzed and verified by State of California Environmental Laboratory Accreditation Program (ELAP) certified laboratories; Enviromatrix Analytical, Inc. (EMA) using Inductively coupled plasma atomic emission spectroscopy (ICP-AES) U.S. EPA method 6010, or Weck Laboratories (Weck), using Inductively coupled plasma mass spectrometry (ICP-MS) U.S. EPA method 1640, respectively (U.S. EPA, 1997). Analytical methods and reporting limits for each test is shown in Table 3.1. Daily water quality measurements were recorded for all tests with an Oakton hand-held pH/mv/temperature/RS 232 meter, and Hach HQ40d portable conductivity, dissolved oxygen and temperature meter.

Table 3.1. Analytical Methods, Detection (MDL), and Reporting Limits (RL) for Copper and Zinc MDL RL Study Type Metal Test Method Laboratory (µg/L) (µg/L) Single & ICP-MS; EPA Zinc 0.04 0.20 Weck Mixed Metal 1640 Copper and ICP-AES; Single Metal 2 100 EMA Zinc EPA 6010 ICP-MS; EPA Mixed Metal Copper 0.004 0.01 Weck 1640

3.5 ACUTE TOXICITY TEST METHODS Acute toxicity test methods (U.S. EPA, 2002) were modified to reflect pulsed exposure regimes defined in Figure 3.2. Juvenile mysid shrimp were 4 days old when received from a commercial supplier, Aquatic Research Organisms Inc., Hampton, New Hampshire. To ensure organism health, mysids were fed roughly 40 newly hatched Artemia nauplii each twice daily. Following U.S. EPA protocols, organisms were acclimated to the test conditions via a slow drip of 0.45µm 34 ppt FSW during an acclimation period of 24-hr prior to test initiation. Organisms were kept at 20 ± 1°C, for a 16-hr light/8-hr dark photoperiod. Water quality (dissolved oxygen, temperature, pH, and salinity/conductivity) was assessed daily. A static reference toxicant test was performed concurrently as a QA/QC measure of organism health. Static reference toxicant tests received a 50% water renewal at

48-hrs. All mysid exposures were conducted in 500mL plastic drinking cups. Four replicates were prepared for each test concentration. At test initiation, 5 healthy organisms were randomly placed into each replicate. Following pulsed exposures, organisms were transferred from toxicant containing solutions into 0.45µm FSW by pouring mysids from exposure chambers onto an 80µm polycarbonate Nitex screen tube. Organisms were gently rinsed with FSW from the screen into a separate plastic cup containing FSW, where they remained for the rest 96-hr exposure period. As a quality control measure to ensure transfer methods did not damage or kill organisms, lab controls underwent a mock transfer at 3-hrs. Light tables were used to enumerate mysid shrimp mortalities on a daily basis. Mortalities found were removed from the test as a means of preserving water quality. Test specifications can be found in Table 3.2.

Figure 3.3. Purple sea urchin exposure in 25µm Nitex screen polycarbonate tubes placed in 400mL HDPE tri-corner beakers.

3.6 CHRONIC TOXICITY TEST METHODS Sea urchins were field collected off Point Loma in San Diego, California. Embryo development tests were conducted in 400mL High-density polyethylene (HDPE) tri-corner beakers kept at 15 ± 1°C, for a 16-hr light/ 8-hr dark photoperiod. Water quality parameters (dissolved oxygen, temperature, pH, and salinity/conductivity) were measured and recorded daily to ensure appropriate testing conditions. Sea urchin eggs were fertilized with the appropriate sperm density to provide > 95% fertilization success. At time zero, roughly 250

Table 3.2. Test Methodology and QA/QC Requirements for 3, 6 & 12-hr Pulsed Toxicity Test Using Americamysis bahia (mysid shrimp) Exposure to Copper and Zinc Individually Test organism Americamysis bahia (mysid shrimp) Test organism source Aquatic Research Organisms, NH Test type Pulsed Test endpoints Survival Test solution renewal Once at 48-hr Feed 40 newly hatched Artemia nauplii per larvae twice Feeding daily, morning and evening Test Chamber 500mL Plastic Cups size/type Test solution volume 250 ml Test temperature 20 ± 1°C Test salinity 34 ± 2 ppt None, unless DO concentrations fall below 4.0 mg/L, then Aeration aerate all chambers. Light quality Ambient laboratory illumination Light intensity 10-20 µE/m2/s (Ambient laboratory levels) Photoperiod 16-hrs light/ 8-hrs dark No. of organisms per 5 chamber Age of test organism 5 days; 24-h range in size No. of replicates 4 Filtered (0.45µm) natural seawater collected from near the Dilution water mouth of San Diego Bay at SSC Pacific Laboratory Nominal test CuSO : 0, 200, 400, 800, 1600, 3200 μg/L concentrations (Cu) 4 Nominal 96-hr static test concentrations ZnSO4 : 0, 500, 1000, 2000, 4000, 8000, 16000μg/L (Zn) Exposure for 3,6, & 12-hrs followed by exposure to Test duration contaminant free 34 ppt sea water for the remained of the 96- hr test period Test acceptability ≥ 90% survival in controls criteria Nominal reference Copper sulfate: 0, 50, 100, 200, 400, 800 µg/L toxicant concentrations Zinc Sulfate: 0, 125, 250, 500, 1000, 2000 µg/L Test protocol EPA 821/R-02/012 (U.S. EPA, 2002)

21 sea urchin embryos, developed to the 2-4 cell stage, were inoculated into pre-cleaned 25µm Nitex screen polycarbonate tubes (Figure 3.3) set in toxicant containing tri-corner beakers. Four replicates were initiated for each concentration tested. Following the pulsed exposure duration, screen tubes were removed from toxicant containing solutions, thoroughly rinsed with FSW, and placed in to tri-corner beakers containing FSW for the remainder of the 96-hr period. As a quality control measure ensuring transfer methods did not damage embryos, lab controls underwent a mock transfer at the 3- hr pulse. At test termination, larvae were rinsed with FSW from screen tube bottoms into 30mL scintillation vials. Organism developmental stage was preserved using 1mL of 10% formalin acetate solution in 10mL of exposure water. Larvae were scored as normal or abnormal depending on the presence or absence of a 4- armed pluteus with well developed skeletal rods and a well differentiated gut (U.S. EPA, 1995). Larval development was assessed by examining 100 individuals under an inverted Olympus microscope at 100x magnification. Test specifications can be found in Table 3.3.

3.7 PULSED COPPER AND ZINC EXPOSURE TESTS

Dose response determinations (LC50 and EC50 endpoints) for both copper and zinc were generated for the purple sea urchin and mysid shrimp at each pulsed exposure duration.

To evenly bracket the expected LC50 and EC50 values, six concentrations (four replicates each) were prepared for each respective pulsed period of each metal. As essential metals, copper and zinc exist in a homeostatic balance where excess amounts of the metals entering cells can cause increased metallothionein production and irregularities in lysosomal activity, resulting in similar mechanisms of toxicity (Langston, 1990; Phillips et al., 2003). Moreover, San Diego Bay stormwater studies performed by Katz et al. (2006) showed the dose response curves for copper and zinc individually and in mixtures are similar. It was hypothesized that the ratio between copper and zinc static 96-hr exposure LC/EC50’s could be applied to copper pulsed exposure LC/EC50 concentrations to yield the potential LC/EC50’s of zinc pulsed exposures. Therefore, the baseline dilution series used in the zinc reference test was based on the historical upper bound EC50 (120 µg/l) and

LC50 (550 µg/l) concentrations generated from various studies (Cripe, 1994; Katz et al., 2006; Lussier, Gentile, & Walker, 1985).

Table 3.3. Test Methodology and QA/QC Requirements for 3, 6 & 12-hrs Pulsed Toxicity Test Using Strongylocentrotus purpuratus (purple sea urchin) Exposure to Copper and Zinc Individually Test organism Strongylocentrotus purpuratus (purple sea urchin) Test organism source Field collected in San Diego, CA Test type Pulsed Test endpoints Embryo Development Rate (Proportion Normal) Test solution renewal None Feeding None Test Chamber 400 ml HDPE tri-corner containers /polycarbonate screen size/type tubes with 25µm mesh Test solution volume 250ml Test temperature 15 ± 1°C Test salinity 34 ± 2 ppt Light quality Ambient laboratory illumination Light intensity 10-20 µE/m2/s (Ambient laboratory levels) Photoperiod 16-hrs light/ 8-hrs dark No. of organisms per 250 eggs, appropriate sperm density to provide > 90% chamber fertilization success (determined in a pre-test trial). No. of replicates 4 Filtered (0.45µm) natural seawater collected from near the Dilution water mouth of San Diego Bay at SSC Pacific Laboratory Nominal Test CuSO : 0, 31.3,62.5, 125, 250, 500μg/L concentrations (Cu) 4 Nominal Test ZnSO : 0, 1280, 2560, 5120, 7680, 10240μg/L concentrations (Zn) 4 Exposure for 3, 6, & 12-hrs followed by exposure to Test duration contaminant free 34 ppt sea water for the remained of the 96-hr test period Test acceptability ≥ 80% normal development in surviving controls; criteria < 25% Minimum Significant Difference (MSD) Nominal 96-hr static Copper sulfate: 0, 5.8, 8.4, 12, 17.2, 24, 35µg/L reference toxicant Zinc Sulfate: 0, 20, 40, 80, 160, 320μg/L concentrations Test protocol EPA 600/R-95/136 (U.S. EPA, 1995)

Both copper and zinc dilution series employed are outlined in Table 3.1 and Table 3.2 for purple sea urchins and mysid shrimp, respectively. As a measure of organism health and means of quantifying the relative ratio of toxicity between static exposures and pulsed exposures, static reference tests for both copper and zinc were performed concurrently using standard methods (U.S. EPA, 1995, 2002), for both purple sea urchins and mysid shrimp.

3.8 PULSED COPPER AND ZINC MIXTURE EXPOSURE TEST

The binary metal mixture experiment used the LC50 values generated in the two previously described single metal tests. As defined by Kraak et al. (1994), this study utilized a Toxic Unit Fixed Ratio Design. Toxic units (TUs) are calculated by dividing metal concentrations by their corresponding LC50’s (TU = concentration / LC50). This results in single metal LC50 values equaling 1 TU. Assuming that copper and zinc contribute equally to the total toxicity, the relative additive, antagonistic, or synergistic effects of copper and zinc mixtures were determined using methods outlined by Phillips et al. (2003) and Bellas (2008).

Accordingly, a median lethal response is expected if half the copper LC50 is added to half the zinc LC50. This is defined by the Phillips et al. (2003) formula: 0.5 (TU-Cu) + 0.5 (TU-Zn). To bracket the appropriate mixed metal dose response, a six concentration dilution series was developed for each time point by halving and doubling the respective single metal TU 0.5 values two times as outlined in Table 3.4 If the binary metal mixture toxicity was additive, the sum of the individual metal TUs was equal to approximately 1. If mixture TUs were found to be less than 1, lower than expected concentrations of the individual metals would have caused a 50% adverse effect in mixture, suggesting synergy (more than additive toxicity). Finally, if mixture TUs were greater than one, an antagonistic (less than additive) interaction between copper and zinc was concluded. By establishing the relative proportion difference between mixture TUs at each time point, a comparative analysis of time-varying exposure of metal mixtures was performed. If mixed metal TU values were different at varying exposure regimes, mixed metal toxicokinetic dynamics may be influenced by varying exposure times.

3.9 STATISTICAL ANALYSIS Multi-concentration juvenile mysid survival and embryo-larval development tests represented the acute and chronic endpoints, respectively. Median Lethal (LC50) or median effective (EC50) concentration values and associated 95 % confidence limits were calculated for all tests that exhibited a dose-response curve. These endpoints were determined using U. S. EPA approved decisions trees automated with the statistical software package Comprehensive Environmental Toxicity Information System (CETIS), version 1.025b.

Table 3.4. Test Methodology and QA/QC Requirements for 3, 6 & 12, 96-hr Exposure to Copper and Zinc Mixtures Using Americamysis bahia (Mysid Shrimp) Test organism Americamysis bahia (mysid shrimp) Test organism source Aquatic Research Organisms, NH Test type Pulsed Test endpoints Survival Feed 40 newly hatched Artemia nauplii per larvae twice daily, Feeding morning and evening Test Chamber 500mL Plastic Cups size/type Test solution volume 250mL Test temperature 20 ± 1°C Test salinity 34 ± 2 ppt None, unless DO concentrations fall below 4.0 mg/L, then Aeration aerate all chambers. Light quality Ambient laboratory illumination Light intensity 10-20 µE/m2/s (Ambient laboratory levels) Photoperiod 16-hrs light/ 8-hrs dark No. of organisms per 5 chamber Age of test organism 5 days; 24-h range in size No. of replicates 4 Filtered (0.45µm) natural seawater collected from near the Dilution water mouth of San Diego Bay at SSC Pacific Laboratory Nominal 3-hr test CuSO4 : 0, 184, 369, 737, 1475, 2950 μg/L concentrations ZnSO4 : 0, 1305, 2610, 5220, 10440, 20880 μg/L Nominal 6-hr test CuSO4 : 0, 99, 199, 398, 795, 1590 μg/L concentrations ZnSO4: 0, 489.3, 978.5, 1957, 3914, 7828 μg/L Nominal 12-hr test CuSO4 : 0, 30, 60, 120, 240, 479 μg/L concentrations ZnSO4: 0, 257, 515, 1030, 2060, 4120 μg/L Nominal 96-hr test CuSO4 : 0, 17, 34, 68, 135, 270 μg/L concentrations ZnSO4: 0, 63.8, 127.5, 255.1, 510.1, 1020.2 μg/L Exposure for 3-hrs, 6-hrs, & 12- hrs followed by exposure to Test duration contaminant free 34 ppt sea water for the remained of the 96- hr test period Test acceptability ≥ 90% survival in controls criteria Nominal 96-hr static Copper sulfate: 0, 50, 100, 200, 400, 800 µg/L reference toxicant test Zinc Sulfate: 0, 125, 250, 500, 1000, 2000 µg/L Test protocol Phillips et al., 2003, U.S. EPA, 1995, 2002

CETIS normalized data calculating endpoints using a linear regression model probit analysis, identifying the maximum likelihood estimate (MLE) for approximating the LC/EC50 values and associated 95 % confidence limits (Finney, 1978; U.S. EPA, 1995, 2002). This analysis consists of adjusting the data for mortality in the control then using the MLE to estimate the parameter of the underlying log tolerance distribution, assumed to have a particular shape. If there was an incomplete range of responses the Trimmed Spearman-Karber Method was used to calculate LC/EC50 values and associated upper and lower 95% confidence limits.

CHAPTER 4

RESULTS

All laboratory controls met standard acceptability criteria of ≥ 90% survival for mysid shrimp tests, and ≥ 80% normal larval development for sea urchin embryo-larval development tests. Water quality parameters were within recommended ranges for all tests.

4.1 QA/QC: SINGLE METAL COPPER EXPOSURE STUDY The embryo-larval development static reference toxicant test for the single-metal pulsed copper study did not produce an EC50 due to the improper preparation of the dilution series. As a result, the static 96-hr copper EC50 value (14.8 µg/L) reported by Arnold, Cotsifas, Ogle, DePalma, and Smith (2010) was used instead. This concentration was within the confidence limits of the EC50 14.3 µg/L (13.8-14.9 µg/L) reported by Rosen et al. (2008), as well as within 2 standard deviations of the mean SSCPac bioassay laboratory values.

4.2 COPPER EXPOSURE STUDY Target copper concentrations identified in Tables 3.2 and 3.3 were verified using ICP- AES following U.S. EPA method 6010c. Table 4.1 summarizes the target and actual copper dilution series used in the single metal sea urchin embryo-larval development and mysid shrimp survival tests. Increased exposure time to copper corresponded to a decrease in the proportion of normally developed sea urchin pluteus larvae (Figure 4.1). Larvae exhibited the greatest sensitivity (the lowest EC50 value) at the longest pulse (12-hr) exposure duration, with decreasing sensitivity as the pulsed period was shortened. The copper EC50 s and the corresponding upper and lower confidence limits for sea urchin embryo-larval development are shown in Table 4.2. The EC50 values for pulsed exposure were 8-20 times greater (less sensitive) than that of standard 96-hr exposures.

Table 4.1. Static and Pulsed Exposure Nominal and Verified Copper Concentrations Nominal Copper Verified Copper Exposure Time Test Species Concentration Concentrationa (hr) (µg/L) (µg/L) 3, 6, & 12 0 ND 3, 6, & 12 31.3 15.0 3, 6, & 12 62.5 36.0 S. purpuratus 3, 6, & 12 125 81.0 3, 6, & 12 250 179 3, 6, & 12 500 367 3, 6, & 12 0 ND 96 (Static) 96 (Static) 50 44.4 3, 6, & 12 100 88.8 96 (Static) 3, 6, & 12 200 178 A. bahia 96 (Static) 3, 6, & 12 400 355 96 (Static) 3, 6, & 12 800 693 96 (Static) 3, 6, & 12 1600 1390 3, 6, & 12 3200 2970 a U.S. EPA method 6010 (EMA) ND=below method detection limit.

Table 4.2. Summary of Median Effective (EC50) Copper Concentrations for the Purple Sea Urchin Embryo-Larval Development Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) Factor Difference Pulse Duration EC 95% LCL 95% UCL 50 from 96-hr Static (hr) (µg/L copper) (µg/L) (µg/L) Exposure 3 296 241 335 20 6 224 216 232 15 12 114 106 122 8 96 (Static) 14.8 14.6 15.1 NA

Figure 4.1. Sea urchin embryo-larval development test median effective concentrations (EC50) for copper at each exposure interval. 96-hr static* data was included from Arnold et al. (2010). The associated error bars represent the upper and lower 95% confidence intervals for each exposure scenario. Mysid shrimp tests yielded a similar response, exhibiting progressively greater toxicity corresponding to increased exposure time (Figure 4.2). During pulsed exposures mysid shrimp exhibited the greatest sensitivity (the lowest LC50 value) at the longest (12-hr) pulse. The dose response of mysid shrimp to pulsed copper is shown in Figure 4.2. The LC50 values and the corresponding upper and lower confidence limits for mysid shrimp are summarized in Table 4.3. The LC50 values for pulsed exposures were 2-11 times greater (less sensitive) than standard 96-hr exposures. However, the confidence limits between the 3 and 6-hr pulsed durations overlapped, in addition to the confidence limits between 12-hr and 96- hr exposures.

Table 4.3. Summary of Median Lethal (LC50) Copper Concentrations for Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) Factor Pulse Duration LC50 95% LCL 95% UCL Difference from (hr) (µg/L copper) (µg/L) (µg/L) 96-hr Static Exposure 3 1475 968 2248 11 6 795 438 1202 6 12 240 104 367 2 96 (Static) 135 91.3 172 NA

4.3 ZINC EXPOSURE STUDY Target and verified zinc concentrations are shown in Tables 3.2 and 3.3 were verified using ICP-AES following U.S. EPA method 6010c, and ICP-MS following U.S. EPA method 1640. Table 4.4 summarizes the nominal and verified zinc dilution series used in the single metal sea urchin embryo-larval development and mysid shrimp survival tests.

Table 4.4. Summary of Nominal and Verified Zinc Concentrations from Static and Pulsed Exposures Nominal Zinc Verified Zinc Exposure Duration Test Species Concentration Concentration (hr) (µg/L) (µg/L) a 3, 6, & 12 0 6.1 96 (Static) 96 (Static) 20 37 96 (Static) 40 63 96 (Static) 80 120 96 (Static) 160 220 S. purpuratus 96 (Static) 320 480 3, 6, & 12 1280 1900 3, 6, & 12 2560 3900 3, 6, & 12 5120 7700 3, 6, & 12 10240 15000 3, 6, & 12 20180 31000 3, 6, & 12 0 13 96 (Static) 96 (Static) 125 100 96 (Static) 250 180 3, 6, & 12 500 398 96 (Static) 3, 6, & 12 A. bahia 1000 753 96 (Static) 3, 6, & 12 2000 1520 96 (Static) 3, 6, & 12 4000 3280 3, 6, & 12 8000 5430 3, 6, & 12 16000 16700 a U.S. EPA method 1640 (Weck) for S. purpuratus, U.S. EPA method 6010 (EMA) for A. bahia.

Increased exposure time to zinc beyond 6-hr corresponded to reduced normal development of purple sea urchin larvae. As the confidence limits overlapped between 3-hr

and 6-hr EC50 values this study was unable to establish a relative difference in toxicity between these exposure times. Purple sea urchin dose response to pulsed zinc exposures is

shown in Figure 4.3. The EC50 values and corresponding upper and lower confidence limits for zinc exposure to sea urchin larvae are summarized in Table 4.5 and shown in Figure 4.3.

The EC50 values for pulsed exposures were 112-186 times greater (less sensitive) than standard 96-hr exposures.

Table 4.5. Summary of Median Effective (EC50) Zinc Concentrations for Purple Sea Urchin Embryo-Larval Development Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) Factor of Pulse Duration EC50 95% LCL 95% UCL Difference from (hr) (µg/L zinc) (µg/L) (µg/L) 96-hr Static Exposure 3 27,120 22,060 32,540 185 6 27,140 23,300 32,980 186 12 16,330 12,890 20,420 112 96 (Static) 146 134 159 NA

Compared to sea urchin larvae, mysid shrimp were found to be far more sensitive to zinc under pulsed exposure scenarios (Figure 4.3 and 4.4). Having non-overlapping

32 confidence limits, mysid shrimp exhibited differences between lethal toxic concentrations and the corresponding exposure times. During pulsed exposures, larvae exhibited the greatest sensitivity (the lowest LC50 value) following the longest 12-hr exposure, with decreasing sensitivity as the exposure period was shortened. The LC50 values and the corresponding upper and lower confidence limits to mysid shrimp, are shown in Table 4.6 and shown in

Figure 4.4. When compared to standard 96-hr exposure LC50s, pulsed exposures were 4-20 times greater (less sensitive).

Table 4.6. Summary of Median Lethal (LC50) Zinc Concentrations for Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL). Factor of Pulse Duration LC50 95% LCL 95% UCL Difference from (hr) (µg/L zinc) (µg/L) (µg/L) 96 hr Static Exposure 3 10,440 7,262 15,300 20 6 3,914 3,114 4,674 8 12 2,060 1,685 2,492 4 96 (Static) 510 415 606 NA

4.4 BINARY METAL MIXTURE PULSED EXPOSURE To determine if reduced toxicity remained consistent in the presence of complex real world mixtures, a binary metal mixture of copper and zinc was assessed at the respective 3, 6, and 12-hr pulsed time points. Verified using ICP-MS following U.S. EPA method 1640, Table 4.7 summarizes the target and copper and zinc concentrations for the dilution series used in the mixed metal mysid survival test.

Table 4.7. Summary of Nominal and Verified Copper and Zinc Concentrations for Mysid Survival Mixed Metal Static and Pulsed Exposures Pulse Cu-Nominal Cu-Verified Zn- Nominal Zn- Verified Duration Concentration Concentrationa Concentration Concentrationa (hr) (µg/L) (µg/L) (µg/L) (µg/L) Laboratory 0 1.6 0 6.4 Control 184 190 1305 1800 369 350 2610 3900 3 738 880 5220 7400 1475 1500 10440 15000 2950 3100 20880 30000 99.4 99.0 489 850 199 200 979 1500 6 398 440 1957 2500 795 770 3914 5200 1590 1600 7828 11000 30 34 258 360 59.9 72.0 515 910 12 120 120 1030 1500 240 340 2060 3100 479 570 4120 5900 16.9 21.0 63.8 81.0 33.8 34.0 128 150 96 (Static) 67.6 71.0 255 320

135 140 510 860 270 280 1020 1600 a U.S. EPA 1640 method (Weck)

Mysid shrimp showed the greatest sensitivity to pulsed copper and zinc mixtures

(lowest LC50 value) following the longest 12-hr pulsed exposure. The dose response for mysid shrimp exposed to the copper and zinc mixture is shown in Figure 4.5. The LC50 values and associated 95% upper and lower confidence intervals for the single metal static

34 copper and zinc reference toxicant tests are shown in Table 4.8. For the mixed metal exposures, the LC50 values and the associated 95% upper and lower confidence intervals, calculated for each metal individually, are shown in Table 4.9. Overall, the toxicity of the copper and zinc in mixture was reduced with shorter exposure times. Per Phillips et al. (2003), Table 4.10 summarizes the results of mixed metal toxicity relative to single metal toxicity, summarized in the form of TUs. The TUs were calculated by dividing mixed metal LC50 values (Table 4.9) by the corresponding single metal LC50 values (Tables 4.3 and 4.6). The resulting TUs for both copper and zinc were summed to produce a collective TU for the mixture. The closer the combined metal TU was to 1, the more additive the relative toxicity the metal mixture was. If a combined metal TU was significantly greater than 1, more copper and zinc were required to elicit a toxic response than in a single metal exposure, suggesting an antagonistic response (Table 4.10). As observed in Table 4.10 and Figures 4.5 and 4.6, the lower TUs for copper suggests that it is driving toxicity in mixed metal exposures at each exposure interval.

Table 4.9. Summary of Median Lethal Copper and Zinc Concentrations for the Mixed Metal Mysid Survival Test with Corresponding 95% Upper (UCL) and Lower Confidence Limits (LCL) Exposure LC 95 % LCL 95% UCL Metal 50 Duration (hr) (µg/L) (µg/L) (µg/L) Cu 832 661 1029 3 Zn 7810 6271 9733 Cu 527 425 651 6 Zn 3399 2782 4236 Cu 267 209 341 12 Zn 2774 2260 3456 Cu 106 84.6 123 96 (Static) Zn 573 412 707

Table 4.10. Combined Toxic Unit Calculations for Mixed Metal Tests for Each Pulsed Time Exposure Exposure Time (hr) Metal Mixed Metal Toxic Unit (TU) Cu 0.56 3 Zn 0.75 TU Sum 1.31 Cu 0.66 6 Zn 0.87 TU Sum 1.53 Cu 1.11 12 Zn 1.35 TU Sum 2.46 Cu 0.59 96 (Static) Zn 0.96 TU Sum 1.55

Figure 4.1. Toxic Units (TU) for copper and zinc derived from the mysid survival test with mixed metals.

CHAPTER 5

DISCUSSION

The results of this study demonstrate that modified standard WET test protocols can be used to assess the toxicity of short-term episodic discharges. Examining both chronic and acute toxicity endpoints over a range of realistic exposure scenarios, a majority of these experiments consistently show that pulsed toxicant exposures (followed by replacement with uncontaminated seawater), representing San Diego Bay storm events, result in toxic responses that decreased with contaminant exposure times. As a result, standard 96-hr static test methods may be overestimating the toxicity of EOP stormwater discharges.

5.1 IMPLICATIONS OF PULSED EXPOSURE TOXICITY AND RISK CHARACTERIZATION Bioassays using pulsed contaminant exposure times have been widely publicized in peer-review literature. The toxicity of pulsed exposures is determined by the interactive effects of exposure concentration and duration, or is a result of non-steady state exposure dynamics (Hickie et al., 1995). A vast majority of research in this area shows increased exposure time corresponds to amplified toxicity (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015; Angel et al., 2010; Butcher et al., 2006; Diamond et al., 2006; Hosmer et al., 1998; Reinert et al., 2002). However, exposure times and concentrations are not always predictors of toxic endpoints. Quantifying the individual effects of contaminant concentrations and exposures durations is important for estimating risk to biota. The probabilistic AUC (AUC= concentration x duration) method for estimating the toxicity of episodic contaminant exposures, can be applied when concentration and exposure time are inversely proportional to one another, and thus equal with regards to their relative contribution to toxicity (Hoang, Gallagher, et al., 2007; Morton et al., 2000; U. S. EPA, 1992). This method assumes varying

38 contaminant concentrations under varying exposure times with the same AUC, will have the same risk. The results of this research demonstrate that when calculating for the AUC, there are factors other than concentration and duration influencing toxicokinetic dynamics of copper and zinc (Table 5.1).

Table 5.1. Area under the Curve (Concentration x Duration = AUC) Calculations for Single Metal Exposures Exposure Concentration Species Metal AUC Duration (Hrs) (µg/L) 3 296 888* Purple Sea 6 224 1344 Urchin 12 114 1368 96 14.8 1421 Copper 3 1475 4425 6 795 4770 Mysid Shrimp 12 240 2880 96 135 12960 3 27120 81360 Purple Sea 6 27140 162840 Urchin 12 16330 195960 96 146 14016 Zinc 3 10440 31320 6 3914 23484 Mysid Shrimp 12 2060 24720 96 510 48960 * Indicates significantly different values P< 0.05 by Grubbs’ outlier test

Although the sea urchin 6, 12, and 96-hr copper AUC is close, the Grubbs’ test identifies the 3-hr exposure a significant outlier (p < 0.05), underestimating the toxicity by roughly 1.5 times for 3-hr exposures (Table 5.1). Accordingly, the AUC is not an applicable method to estimate toxicity of 3-hr copper exposures. The Grubbs’ test did not identify any other significant outliers in the remaining tests. However, all tests need to be replicated several times to ensure consistency across all time points. Additionally, the relative unresponsiveness of the sea urchin to zinc under pulsed exposure regimes needs to be addressed before any conclusion can be made regarding the relative effectiveness of the AUC model to predict toxic response across varying exposure durations.

In general, a majority of pulsed-exposure experiments found that episodic contaminant exposures are less toxic than static continuous exposures (Hosmer et al., 1998). Similarly, this study found that for 3-hr exposures, categorizing the 50th percentile of storm durations, the EC50 value was 20 times greater than the static exposure values. However, the

3-hr LC50 levels affecting mysid survival were only 11 times greater than static exposure values (Tables 4.2 and 4.3). These varying toxicokinetic effects may be due to observed fact that crustaceans commonly adsorb certain cation metals to the exoskeleton or bind these metals to the inner exoskeleton matrix following uptake and transport through the hemolymph (Keteles & Fleeger, 2001). Moreover, chitin serves as binding site for many cation metals resulting in a high potential for crustacean adsorption, while copper has been shown to redistribute to the tissue before ecdysis (Keteles & Fleeger, 2001). Concerning pulsed exposure studies, Angel et al. (2010) observed delayed mortality in Melita plumulosa. This was attributed to a faster copper uptake rate than depuration rate.

The present study found that 12-hr copper exposures to mysid shrimp resulted in LC50 values only two times less than standard 96-hr static tests, in addition to overlapping confidence intervals. Accordingly, copper may be bound to mysid shrimp exoskeleton and potentially redistributed to tissues following transfer to FSW. This may result greater relative toxicant exposure times and respective toxicity. However, further experiments must be performed to determine the overall toxicant uptake and depuration rates mysids are experiencing following specific pulsed exposure regimes.

Single metal zinc pulsed exposure experiments yielded LC50 values that nearly halved as exposure times doubled (Table 4.6). Keteles and Fleeger (2001) found that zinc is quickly depurated by excretion, in contrast to copper redistribution. Accordingly, 96-hr static exposures yielded LC50 values 4 times lower than 12-hr exposures. This suggests that the depuration of zinc in mysids occurs quickly, relative to copper. Purple sea urchin larvae, did not experience toxicity during pulsed exposure regimes until concentrations were several orders of magnitude greater than static 96-hr exposures.

The 6-hr exposure demonstrated a lower EC50 value than the 3-hr exposure, while the confidence interval of the 3-hr and 6-hr EC50 values overlapped. Only after the 12-hr exposure was a change in toxicity observed. A study by Timourian (1968) found that sea urchin embryos placed in ZnCl2 solutions 30 minutes following fertilization and removed

40 from solution 18-20-hrs later developed further than larvae exposed to toxicant solutions for a static duration. The same study found that larvae placed in zinc solutions for 18-20-hr after reaching the mid gastrula stage of development, abnormally developed into heavy-bottom blastulae. This demonstrates the urchin larvae must reach a certain stage of development before zinc will have deleterious effects on larvae development. Additionally, Timourian (1968) found that the zinc uptake rate by developing sea urchins was consistent after an initial lag period of 8 to 10-hrs. Therefore, in order to observe the toxic effects of zinc to sea urchin larvae, under pulsed exposure scenarios, a development period of at least 10-hrs needs to occur prior to contaminant exposure initiation. The single metal zinc pulsed exposure tests did not account for the lag period for zinc uptake demonstrated by Timourian (1968). Accordingly, the exposure periods occurring for 3, 6, and 12-hrs did not allow larvae enough time to develop to the point where zinc could have deleterious effects on development. This could account for the high environmentally unrealistic zinc concentrations needed to elicit toxic responses in urchin larvae. However, further research needs to be performed to determine toxic endpoints following certain developmental stages.

5.2 INTEGRATION OF MIXED METAL EXPOSURES IN PULSED TOXICITY TESTS As stormwater and ambient receiving water samples contain metal mixtures, this study determined how metal mixtures interact under pulsed exposure scenarios. The results of this study on mysid shrimp demonstrated a consistent trend showing that copper and zinc mixtures produced less than additive toxicity across all pulsed exposure scenarios, yielding TU values between 1.31 and 2.46 (Table 4.10; Figure 4.5; Figure 4.6). Similarly, a majority of published studies has found that copper and zinc metal mixtures produce antagonistic toxicity in aquatic organisms. A study using S. purpuratus by Phillips et al. (2003), found that copper and zinc mixtures were antagonistic when compared to individual metal toxicity producing TUs values between 1.25 and 1.42. A toxicity identification evaluation (TIE) prepared by Nautilus Environmental on behalf of Katz et al. (2006), evaluated the 96-hr acute toxicity of copper and zinc mixtures to mysid shrimp, finding TU values of 1.45 and 1.62 indicating antagonistic toxicity. The TUs values yielded in the present study for both static and pulsed

41 exposure to copper and zinc mixtures found similar results (Table 4.10; Figure 4.5; Figure 4.6). This suggests that exposure times may not have an effect on the toxicokinetics of copper and zinc in mixture. Using the TU test design, this study yielded results showing that copper may be driving toxicity in mixed metal tests. Although Katz et al. (2006) describe zinc as one of the primary toxicants of concern, its relative toxicity may potentially be undermined in the presence of copper. The TUs describing the toxicity of copper and zinc in mixture to mysid shrimp under pulsed exposure durations, showed that less copper relative to single metal tests was required in mixture to elicit a toxic response (Table 4.8; Table 4.9). Relative to single metal zinc exposure tests, more zinc was required to elicit a toxic response in mixed metal tests (Table 4.8; Table 4.9). Corresponding to those results reported in the Nautilus Environmental TIE report and Phillips et al. (2003) study, the results of the presented study show that copper may be driving toxicity in copper and zinc mixture tests. The variables assessed in this study, contaminant exposure times and mixtures, demonstrate that standard WET testing protocols used for regulatory purposes omit important and quantifiable variables influencing the overall toxicity of effluent samples. This study demonstrates that it is possible to perform bioassays adjusted for relevant exposure times and contaminate mixtures. Moreover, this study demonstrates that the primary inorganic contaminants of concern from stormwater discharges are antagonistic to one another. From this observation, it can be inferred that TMDL requirements based on single metal toxicity can potentially be developed to reflect metal mixtures occurring in the environment.

5.3 IMPLICATIONS OF PULSED EXPOSURE TESTING MODIFICATIONS ON ORGANISM RESPONSE AND REGULATORY CONSIDERATIONS The results of this research indicate, in general, that standard WET methods used to quantify the relative toxicity of stormwater effluent are conservative in protecting aquatic life. However, if latent mortality is deemed a significant factor, as Diamond et al. (2006) describe with zinc exposures to both P. promelas and D. magna, traditional WET methods may underestimate effluent toxicity, signifying that in some cases modified pulsed exposure testing protocols may be a more conservative methodology. Following, WET testing is a

42 useful tool in identifying and quantifying the toxicity of discharges in the environment, however, it is not a perfect application. As the overlying objective of WET tests is to protect the receiving environment, not the organisms used in WET tests, continually expanding the understanding of contaminant toxicity dynamics is important (Chapman, 2000). Katz et al. (2006) performed 333 toxicity tests on 136 discrete stormwater and receiving water samples in San Diego Bay, showing a disconnection between the toxicity of EOP stormwater samples and receiving environment samples. This study found that first- flush EOP samples failed 90 % survival NPDES permit requirements 58% of the time, grab samples failed the permit requirements 25% of the time, while of the 202 receiving water toxicity tests performed less than 1% were toxic (Katz et al., 2006). It was concluded the effect of stormwater discharges on the receiving environment was limited in magnitude, spatial extent, and short-lived. This data parallels the U.S. EPA (1991c) document stating “there is a less likely chance for receiving water impacts to be observed in saltwater systems as predicted by toxicity tests” (p. 9). Traditional WET testing protocols do not account for several environmental variables including storm intensity, storm duration, toxicants binding to sediment, and tidal flushing (Chadwick, Zirino, Rivera-Duarte, Katz, & Blake, 2004; Diamond et al., 2006; Rivera-Duarte et al., 2005). Our results suggest that WET tests can be modified to reflect stormwater discharge exposure durations experienced in San Diego Bay. Though WET tests are considered the best method to assess stormwater toxicity, a study by Chadwick et al. (2004) observed that stormwater from all sources only accounts for 7% of the total copper loading in San Diego Bay. More importantly, this study found that the primary source of copper in the Bay is dominated by chronic industrial discharge sources, while stormwater plumes from industrial outfalls have a limited spatial extent and often rapidly assimilate into surrounding environment within 12-hrs. This natural mixing resulting in a lack of toxicity in the receiving environment is indicative of what is occurring at some San Diego Bay stormwater outfalls (Katz et al., 2006). As WET tests were developed to assess the toxicity of continuous industrial effluent discharges, static WET protocols should be modified to reflect the episodic nature of stormwater runoff (U.S. EPA, 1991c). Concerning in situ exposure times, this study demonstrates that WET testing methods are protective and likely over estimating the effect stormwater is having in the receiving environment. The simplistic model developed for WET testing protocols leaves out crucial

43 variables such as contaminant exposure times and the intensity of the exposure, when assessing stormwater discharges. Accordingly, NPDES permits and the corresponding WET testing protocols do not reflect different dynamics of varying regions of the county. This “one size fits all” approach to assessing the toxicity of stormwater in the given environment can leave areas of low rainfall and high intensity storms, such as San Diego, at risk of adhering to protocols not representative of the local environment. As WET testing criteria were developed for continuous point-source discharges their direct application is not representative of episodic discharge events (Diamond et al., 2006; Katz et al., 2006; U.S. EPA, 1991c). Various studies have shown that pulsed test modifications to current WET testing methodology is a simplified means of measuring the toxicity of short-term exposures (AMEC Foster Wheeler Environment & Infrastructure Inc., 2015; Angel et al., 2010; Butcher et al., 2006; Diamond et al., 2006; Hosmer et al., 1998; Reinert et al., 2002). The results of the present study agree, demonstrating that bioassays representing short-term exposure events, characteristic of EOP stormwater discharges, result in toxic endpoints many times greater (less toxic) than endpoints established by static exposure tests (Butcher et al., 2006; Diamond et al., 2006; U.S. EPA, 1991c). Consequently, the application of pulsed bioassays representing stormwater exposure times may potentially be more environmentally representative of toxicity dynamics occurring in situ, and a method by which the toxicity of industrial stormwater discharges can be assessed.

CHAPTER 6

CONCLUSION

Copper and zinc are common contaminants of concern present in stormwater effluent entering San Diego Bay. This study modified standard WET testing protocols to demonstrate the relative chronic and acute toxicity of copper and zinc under pulsed exposure regimes for both embryonic purple sea urchin (Strongylocentrotus purpuratus) and juvenile mysid shrimp (Americamysis bahia). The copper EC50 values for sea urchins under 3, 6, and 12-hr pulsed, and 96-hr continuous, exposures were found to be 296, 224, 114, and 14.8 µg/L, respectively. Under the same exposure regimes, the copper LC50 for mysid shrimp was found to be 1475, 795, 240, and 135 µg/L, respectively. The zinc EC50 for sea urchins under the same exposure scenarios was 27,120; 27,140; 16,330; and 146 µg/L. However, as described by Timourian (1968), the observed unresponsiveness may be due to a lag period of 8 to 10- hrs where purple sea urchin larvae do not uptake zinc. On the other hand, zinc LC50 values for mysid shrimp were 10,440; 3,914; 2,060; and 510 µg/L. Additionally, it was found that copper and zinc mixtures under both static and pulsed exposure regimes resulted in antagonistic toxicity. Using the TU binary metal mixture method, copper and zinc mixtures under 3, 6, 12-hr, and 96-hr exposures yielded TUs of 1.31, 1.53, 2.46, and 1.55, respectively. This leads to the conclusion that pulsed exposure bioassays applied to complex mixtures, representative of stormwater discharges, demonstrate the same toxicity dynamics. Consequently, this research found that shorter exposure regimes require higher toxicant concentrations in order to yield a toxic response in both mysid shrimp juveniles and purple sea urchin embryos equivalent to the static 96-hr exposures. It may be inferred that conventional 96-hr (or less) WET toxicity tests used for stormwater NPDES purposes may be substantially overestimating toxicity, especially in arid environments such as San Diego. However, research is needed to assess the contribution of copper and zinc to the receiving environment following stormwater events, the amount of bioavailable contaminants in the

45 environment following storms, and multiple pulsed exposures expressing the contaminant’s ability to cause latent toxicity in organisms. As this is the first known study assessing the toxicity of copper and zinc individually and in mixture under these specific exposure regimes for these specific species and endpoints, this study is limited in not having a historical basis by which to compare toxic endpoints. Lab controls, reference toxicant tests, and multiple replicates were used over all multi-concentration tests, providing a statistically significant sample size. However, there were not a statistically significant number of tests performed for each toxicant exposure regime to perform an ANOVA analysis, and thus determine significant differences between the toxicity of copper and zinc exposures at each time point. Additionally, the toxicity of mixed metal pulsed exposures is limited, as these tests should be replicated multiple times to provide a statistically significant data set needed to determine the relative antagonistic, synergistic, or additive nature of copper and zinc mixture under pulsed exposure regimes. In conclusion, the contaminants of concern copper and zinc enter San Diego via stormwater runoff. Regulatory measures are implemented as a means of curbing any deleterious effects of this effluent source. Through the enforcement of NPDES permits, utilization of WET testing protocols, and the implementation of BMPs, facilities bordering San Diego Bay are responsible for mitigating the impact of stormwater contamination in the Bay. Yet the WET testing criteria originally developed for continuous point-source industrial effluent are applied episodic stormwater discharges. These testing protocols exposed sensitive marine organisms to undiluted stormwater runoff for periods up to 96-hrs. However, a 55-year assessment of all storm events near San Diego Bay has shown that 95% of all storm events have been 12-hrs or less in length. As Katz et al. (2006), showed there is a strong correlation between storm length and the duration of EOP discharges. Accordingly, this study demonstrated that when modifying WET testing protocols to mirror realistic stormwater discharge scenarios, the toxic endpoints were orders of magnitude greater in the common test species. As a result, antiquated WET testing protocols maybe overestimating the toxicity of stormwater runoff released from the EOP.

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APPENDIX

DEFINITION OF TERMS Area under the curve AUC Best available technology BAT Best conventional pollutant control technology BCT Best management practices BMP California environmental laboratory accreditation program ELAP California State Water Resource Control Board SWRCB Clean Water Act CWA Complex effluent toxicity testing program CETTP Comprehensive environmental toxicity information system CETIS End-of-pipe EOP Enviromatrix Analytical, Inc. EMA High-density polyethylene HDPE Inductively coupled plasma atomic emission spectroscopy ICP-AES Inductively coupled plasma mass spectrometry ICP-MS Load allocations LA Margin of safety MOS

Median lethal concentration LC50 Median effective concentration EC50 Methods detection limit MDL National pollution discharge elimination system NPDES Reporting limits RL SPAWAR Systems Center Pacific SSCPac Time-averaged concentration TAC Total maximum daily load TMDL Toxicity identification evaluation TIE Toxic unit TU United States Environmental Protection Agency U.S. EPA Waste load allocations WLA Weck Laboratories Weck Whole effluent toxicity WET