ASSESSMENT OF IN SITU STRESSORS AND SEDIMENT TOXICITY … · 2020-04-28 · Assessment of In Situ Stressors and Sediment Toxicity in the. Lower Housatonic River. Revised Draft Report

Assessment of In Situ Stressors and Sediment Toxicity in the Lower Housatonic River

Revised Draft Report

(Includes Annotated Review Comments)

by G. Allen Burton, Jr.

Institute for Environmental Quality 3640 Colonel Glenn Highway

Wright State University Dayton, Ohio 45435

November 2, 2001

Lower Housatonic River Project Pittsfield, MA

TABLE OF CONTENTS

TABLE OF CONTENTS 1

LIST OF FIGURES 2

LIST OF TABLES 3

LIST OF APPENDICES 4

ABSTRACT 5

1.0 INTRODUCTION 6

2.0 MATERIALS AND METHODS 9 2 1 SAMPLING METHODS 9 22CULTURING 10 2 3 TASK 1 SEDIMENT LABORATORY TOXICITY TESTING - SUMMER 1999 10

231 Hyalella azteca Life Cycle Assessment 10 232 Chironomus tentans Life Cycle Assessment 12

2 4 TASK 2 IN SITU TOXICITY AND BIOACCUMULATION 14 2 5 TASK 3 TOXICITY IDENTIFICATION EVALUATION (TIE) 17

251DAY1 18 252DAY2 18 253DAY3 19

2 6 STATISTICAL ANALYSIS 20 2 7 QUALITY ASSURANCE 20

3.0 RESULTS AND DISCUSSION: TASKS 1 AND 2 21 3 1 TASK 1 SEDIMENT MACROINVERTEBRATE LABORATORY TOXICITY TESTING - SUMMER 1999 21

311 Hyalella azteca Life Cycle Assessment 21 312 Chironomus tentans Life Cycle Assessment 22

3 2 TASK 2 IN SITU TOXICITY AND BIOACCUMULATION 23 321 48-hlow flow in situ 23 3 22 10-d low flow in situ 23 3 2 3 7-d bioaccumulation in situ 24

4.0 RESULTS OF TASK 3: TOXICITY IDENTIFICATION EVALUATION 31 4 1 DAY1 31

411 Initial Toxicity Test 31 4 2 DAY 2 31

421 Baseline Test 31 422 Oxidant Reduction Addition Test 31 423 EDTA Chelation Addition 32 424pH Adjusted Filtration 32

4 3 DAY3 32 431 Baseline Test 32 432pH Adjusted/C18 Solid Phase Extraction (SPE) 33 433pH Adjusted/Aeration 33

4 4 TASK 3 DISCUSSION AND CONCLUSIONS 33

5.0 PROJECT SUMMARY 36

6.0 LITERATURE CITED 39

FIGURES TABLES APPENDICES

Burton - Lower Housatonic River Project Page 1

LIST OF FIGURES

Figure 1: Locations of Wright State University toxicity sampling stations (011, 398, 019, 428, 389, 031; marked with asterisks) in relation to benthic invertebrate sampling locations

Figure 2: Survival of Hyalella azteca in chronic laboratory toxicity test, at three time periods (28-d, 35-d, 42-d)

Figure 3: Hyalella azteca 28-day dry weights in chronic laboratory toxicity test

Figure 4: Hyalella azteca 42-day dry weights in chronic laboratory toxicity test

Figure 5: Reproduction of Hyalella azteca in chronic laboratory toxicity test, based on mean number of neonates (totals and numbers per female)

Figure 6: Survival of Chironomus tentans in chronic laboratory toxicity test (20-day)

Figure 7: Dry weights for Chironomus tentans in chronic laboratory toxicity test (20-day)

Figure 8: Ash-free dry weights for Chironomus tentans in chronic laboratory toxicity test (20-day)

Figure 9: Emergence of Chironomus tentans in chronic laboratory toxicity test (20-day)

Figure 10: Survival of Hyalella azteca in 48-h low flow in situ toxicity tests conducted 14-16 June 1999.

Figure 11: Survival of Daphnia magna in 48-h low flow in situ toxicity tests conducted 14-16 June 1999.

Figure 12: Survival of Lumbnculus variegatus in 48-h low flow in situ toxicity tests conducted 14-16 June 1999

Figure 13: Survival of Chironomus tentans in 48-h low flow in situ toxicity tests conducted 14-16 June 1999.

Figure 14: Daphnia magna survival versus total sediment PCB concentrations in 48-hour low flow in situ (sediment) exposures

Figure 15: Survival of Hyalella azteca in 10-d low flow in situ toxicity tests conducted 17-27 June 1999.

Figure 16: Survival of Chironomus tentans in 10-d low flow in situ toxicity tests conducted 17-27 June 1999

Figure 17: Hyalella azteca survival versus total sediment PCB concentrations in 10-day low flow in situ (sediment) exposures

Figure 18: Chironomus tentans survival versus total sediment PCB concentrations in 10-day low flow in situ (sediment) exposures

Figure 19: Hyalella azteca survival versus total water column PCB concentrations in 10-day low flow in situ (water only) exposures

Figure 20: Lumbnculus variegatus PCB tissue burdens versus total sediment PCB concentrations in 7-day in situ bioaccumulation tests

Figure 21: Sediment-associated metals concentrations in 7-day in situ bioaccumulation exposures (sum of Cd, Cr, Cu, Hg, Ni, Pb, Ag, Zn)

Figure 22: Sediment-associatedexposures

PAH and SVOC concentrations in 7-day in situ bioaccumulation


LIST OF TABLES

Table 1: Test conditions for conducting a 42-d sediment toxicity test with Hyalella azteca

Table 2: Test conditions for conducting a 42-d sediment toxicity test with Hyalella azteca

Table 3: Test conditions for conducting a 42-d sediment toxicity test with Chironomus tentans

Table 4: General activity schedule for conducting a 42-d sediment toxicity test with Chironomus tetans

Table 5: TIE tests and controls

Table 6: Sediment chemical measurements for laboratory life-cycle assessments Total PCBs, % TOC and normalized PCB levels, for the test site

Table 7: H azteca 28-d survival following exposure to six Housatonic River sediments and one reference sediment (27 May - 24 June 1999)

Table 8: H azteca 35 (1 July 1999) and 42d (8 July 1999) survival and reproduction (total number of young from Day 35 - 42) following exposure to six Housatonic River sediments and one reference sediment

Table 9: H azteca dry weights following a 28-d exposure to seven sediment treatments 27 May 25 June 1999

Table 10: H azteca dry weights following a 42-d exposure to six Housatonic River sediments and one reference sediment, 27 May - 8 July 1999

Table 11: C tentans survival following a 20d exposure to six composite Housatonic River sediments and three reference sediments, 10-30 July 1999

Table 12: C tentans dry weights following a 20-d exposure to six Housatonic River sediments and three reference sediment, 10-30 July 1999

Table 13: C tentans ash free dry weight following a 20-d exposure to six Housatonic River sediments and three reference sediments 10-30 July 1999

Table 14: C tentans emergence following exposure to six Housatonic River sediments and one reference sediment, 10 July 16 August 1999

Table 15: Chemical measurement for in situ test site sedimentsPCB levels (where available) for the test site

total PCBs, % TOC and normalized

Table 16: Survival for four species following a 48-h in situ test at six sites on the Housatonic River 14- 16 June 1999

Table 17: Survival for two species following a 10-d in situ test at six sites on the Housatonic River 17-27 June 1999

Table 18: Values of octanol-water partitioning coefficient (Kow), organic carbon-water partitioning coefficent (KoC) and sediment-pore water partitioning coefficient (Kp) used to calculate pore water concentrations using equilibrium partitioning theory

Table 19: Estimated pore water, freely dissolved7-d in situ bioaccumulation test

PCB concentrations (Cpw) for sediments from the

Table 20: Pore water Bioconcentration Factors (BCF) for total PCBs in L vanegatus tissue following 7-d in situ bioaccumulation test

Table 21: Sediment Bioaccumulation and Biota/Sediment Accumulation Factors (BAF and BSAF) for total PCBs in L vanegatus tissue following 7-d in situ bioaccumulation test


Table 22: Measured and estimated PCB body residues in L variegatus tissue from a 7-d in situ bioaccumulation test

Table 23: Sediment Quality Guidelines (SQGs) or Water Quality Criteria (WQC) exceedances for samples from the 48-h, 7-d and 10-d in situ tests and the 48-h and 10-d in situ tests

Table 24: Acute PCB lethal body residues from published laboratory studies

Table 25: LC50 values (%) and toxic units resulting from TIE test manipulations

Table 26: Summary of results and conclusions for samples used for TIE test manipulations

Table 27: Total PAH and metals levels relative to Consensus-Based Sediment Quality Guidelines (SQG) and Water Quality Criteria (WQC) for sediments and extracted pore water used for TIE testing

LIST OF APPENDICES

Appendix 1: Weekly water chemistry measurements for 42d Hyallela azteca life-cycle assessment exposure to two control sediments and six composite sediments

Appendix 2: Daily temperature and dissolved oxygen measurements for the H azteca life-cycle assessment of six Housatonic River and two reference sediments, 27 May - 8 July 1999

Appendix 3: Weekly water chemistry measurements for 42d Chironomus tentans life-cycle assessment exposure to six Housatonic River Sediments and two reference sediments, 9 July-20 August 1999

Appendix 4: Daily temperature and dissolved oxygen measurements for a 42d C tentans life-cycle assessment of six Housatonic River sediments and two reference sediments 9 July -19 August 1999

Appendix 5: Water Chemistry measurements for a 48 h Hyalella azteca, Chironomus tentans Lumbnculus variegatus and Daphnia magna exposure at six Housatonic River sites and one laboratory control treatment

Appendix 6: Water Chemistry measurements for a 7 d (Lumbnculus variegatus) and 10d (Hyalella azteca and Chironomus tentans) exposure at six Housatonic River sites and one laboratory control treatment


ABSTRACT

As part of a comprehensive ecosystem assessment, the effects of sediment-associated polychlonnated biphenyl (PCBs) contamination in the Lower Housatonic River, near Pittsfield, Massachusetts, USA, were evaluated This assessment measured sediment toxicity in laboratory and field (in situ) exposures of surrogate test organisms and determined which class of chemicals contributed to toxicity using the Toxicity Identification Evaluation (TIE) approach Each approach provided unique information useful in assessments of ecosystem degradation Total PCB concentrations in reference site sediments ranged from less than the detection limit to 5 39 mg/Kg dry and test site sediments ranged from 0 7 to 521 7 mg/Kg dry depending on the exposure and test type Organisms for in situ studies were housed in flow-through chambers for 2 to 10 days and placed against the sediments, in the water column and in chambers filled partially with sediment and water In the laboratory, life cycle assessment tests were conducted with Hyalella azteca and Chironomus tentans for 4 to 6 weeks, following U S Environmental Protection Agency (USEPA) methods For laboratory exposures, organism response (e g , mortality) increased with PCB contamination For in situ exposures, organism responses also increased with PCB contamination for several species (/ e , significantly reduced survival at sites with low mg/Kg dry PCB concentrations), however, this response was observed primarily for sediment exposures and was not readily apparent in the water-only exposures Adverse effects were observed in laboratory exposures with sediments containing 8 7 mg/Kg dry PCB, or greater, at 20 to 42-d The USEPA TIE approach showed acute toxicity was significantly reduced in treatments that reduce organic compounds and metals in pore water of the two most contaminated sites As the EDTA chelation treatment (a metals toxicity indicator) failed to reduce toxicity, the TIE appears to implicate non-polar organics as the dominant causative agent Using both in situ and laboratory assays provided useful information on the source compartment (/ e sediments) and acute to chronic effect thresholds, thereby contributing to the weight-ofevidence assessment process


1.0 INTRODUCTION

The following report describes the approaches used for identifying major stressors and associated concentrations in the Lower Housatonic that cause adverse effects to benthic invertebrates. A weight of evidence approach was used that combined laboratory and field assessments of physical, chemical, and biological conditions with standard and non-standard test methods.

This study consisted of assessments of: 1) sediment toxicity using USEPA chronic test methods for Hyalella azteca and Chironomus tentans (USEPA 2000a, 2000b), 2) in situ exposures with Daphnia magna, C. tentans, H. azteca, and Lumbriculus variegatus (2 10 day) to determine effects and/or contaminant uptake from overlying water, suspended solids, and bedded sediment; and 3) laboratory Toxicity Identification Evaluation (Phase I) using Ceriodaphnia dubia to fractionate chemical stressor types. This study was coordinated with, and supported the outcomes of, other R.F. Weston project tasks, including: 1) benthic macroinvertebrate community surveys; 2) storm flow and transport modeling; 3) habitat and physical-chemical sampling; and 4) food-web modeling.

Hyalella azteca are routinely used to assess the toxicity of chemicals in sediments (e.g., Burton etal., 1989; Burton 1991; Burton era/. 1996a,b). Test duration and endpoints recommended in standardized methods for sediment testing with H. azteca include 10day (d) survival and 10- to 28-d survival and growth. Short-term exposures that only measure effects on survival can be used to identify high levels of contamination, but may not be able to identify moderately contaminated sediments. The method used in this study, however, can evaluate potential effects of contaminated sediment on survival, growth and reproduction occurring at lower stressor concentrations for H. azteca in a 42-d test.

The 42-d sediment exposure starts at Day 0 with 7- to 8-d-old amphipods. On Day 28, amphipods are isolated from the sediment and placed in water-only chambers where reproduction is measured on Day 35 and 42. Typically, amphipods are first in amplexus at about Day 21 to 28 with release of the first brood between Day 28 to 42. Endpoints measured include survival (Day 28, 35 and 42), growth (dry weight measured on Day 28 and 42), and reproduction (number of young/female produced from Day 28 to 42). Reproduction in amphipods is measured by exposing them in sediment until a few days before the release of the first brood. The amphipods are then sieved from the sediment and held in water to determine the number of young produced. This test design allows a quantitative measure of reproduction. One limitation to this design is that amphipods might recover from effects of sediment exposure during this holding period in clean water. However, amphipods are exposed to sediment during critical developmental stages before release of the first brood in clean water. The USEPA and ASTM state that a subset of endpoints may be measured with minor method modifications (USEPA 1998, ASTM 1999).


The midge Chironomus tentans has been used extensively in the short-term assessment of chemicals in sediments (e.g., Burton 1991, Burton et at. 1996a,b), and standard methods have been developed for testing with this species using 10-d exposures (USEPA 1994). Chironomus tentans is a good candidate for long-term toxicity testing because it normally completes its life-cycle in a relatively short period of time (25 to 30 d at 23°C), and a variety of developmental (growth, survivorship) and reproductive (fecundity) endpoints can be monitored. In addition, emergent adults can be readily collected so it is possible to transfer organisms from the sediment test system to clean, overlying water for direct quantification of reproductive success. In Europe and Canada, the chronic midge method ends before emergence and the USEPA gives the option for test termination (USEPA 1998). Survival is determined at 20 d and at test termination (about 50 to 65 d). Growth is determined at 20 d, which corresponds to the 10-d endpoint in the 10-d C. tentans growth test started with 10-d old larvae. From Day 23 to the end of the test, emergence is monitored daily. Each treatment of the life-cycle test is ended separately when no additional emergence has been recorded for 7 consecutive days (the 7-d criterion). When no emergence is recorded from a treatment, ending of that treatment should be based on the control sediment using this 7-d criterion. USEPA and ASTM state that minor modifications to the basic methods and a subset of endpoints may be used (USEPA 1998; ASTM 1999).

There are no standardized methods for toxicity and bioaccumulation testing of aquatic organisms in the field (in situ). However, there have been many investigators who have shown the usefulness of in situ testing of caged organisms for determinations of site toxicity and bioaccumulation (see citations in Burton et al. 1996c). In situ testing of indicator organisms provides some advantages over laboratory testing or surveys of indigenous community structure. Exposures are more realistic than laboratory testing, which reduces laboratory to field extrapolations uncertainties, and thereby provides simplified interpretations. As with laboratory tests, the in situ toxicity tests address varying species sensitivity to stressors (e.g., amphipods, midges, oligochaetes, mussels, and/or fish). They have the further advantages of adressing both the source of stressor exposures (water, suspended solids, and bedded sediment) and also the specific stressor categories (e.g., flow, suspended solids, photo-induced toxicity, ammonia, metals, and nonpolar organics) and relative importance of each. The exposure partitioning and resulting effects can be used to link effects in the indigenous communities. The information provided from in situ exposures can be used to validate and refine contaminant transport, fate, and bioaccumulation models. This information is coupled with laboratory toxicity, indigenous communities, tissue residues, habitat, physicochemical characterizations, and modeling predictions to provide a more comprehensive assessment of exposure and effects relationships.

The Toxicity Identification Evaluation (TIE) is a process by which water, effluent, or pore water samples are fractionated to isolate various classes of contaminants and then tested for toxicity. This process allows one to determine which groups of contaminants are primarily responsible for toxicity (USEPA 1991 a). These groups of contaminants include pH sensitive and volatile compounds (such as ammonia), metals, and nonpolar


organics. Toxicity is determined by exposing Ceriodaphnia dubia for 48 h and then measuring survival. The TIE Phase I approach used in this study followed modified draft USEPA guidelines for sediments evaluation (USEPA 19915). The pore water treatments include initial toxicity tests (within 24 h of sample receipt), baseline ambient pore water, pH adjusted with aeration, pH adjusted with filtration, pH adjusted with C18 filtration, thiosulfate addition, and ethylene-diamine-tetra-acetate (EDTA) addition fractions.


2.0 MATERIALS AND METHODS

The Housatonic River study sites were chosen on the basis of historic sediment contamination levels and sediment type, as well as proximity to known point source areas of concern. A total of six test sites were chosen for assessment and evaluated for potential toxicity in both laboratory and field (in situ) investigations. These stations are displayed on Figure 1 (sites with asterisks); the toxicity testing stations formed a subset of a larger number of stations sampled for benthic community structure and other sediment quality data. For the purposes of this report, stations are referenced by the last three digits of the Weston ID for that sampling location. For the laboratory sediment toxicity tests, field sites 398 and 011 (reference sites) and 019, 389, 023 and 031 (contaminated sites) were chosen for testing. For the in situ toxicity exposures, Site 023 was replaced with Site 428 (located upstream of Site 389) as Site 023 was too close to Site 031. This decision was made to evaluate how sediment toxicity varies on a spatial scale. Weston personnel recommended replacement of Site 428 with Site 023 as it was farther upstream from Site 031 and had similar historic levels of PCB contamination. The remaining sites used for laboratory testing were used for in situ testing as well.

2.1 Sampling Methods

Sediment samples for chemical analyses and laboratory toxicity testing were collected from the six test site locations by R.F. Weston personnel following project approved protocols. These methods included standard quality assurance and quality control measures to ensure that the sediment samples were not significantly altered and that cross contamination did not occur (ASTM 1999; Burton 1997; USEPA 1991a, 1991b, 1994, 1998, Weston 2000 [Need to add QAPP citation]). The sediment samples were collected using a standard core tube. The core tube was used to take five separate 4-5 cm deep sediment samples from each of the locations where in situ toxicity testing took place. A composite of five separate sediment portions was then homogenized in a sterilized stainless steel bowl and placed in 8-oz amber jars. Portions of each sediment sample were either used for chemical analysis of total PCBs or shipped to Wright State on ice for laboratory tests.

Sediment samples were collected for chemical analyses in concordance with the in situ portion of the study; however, these were not collected on the same day as sediments collected for laboratory toxicity testing. For in situ exposures, sediment samples were collected by Wright State and R.F. Weston personnel, using the same protocols stated above, but at the end of each exposure period (i.e., 3 times, when in situ chambers were retrieved at 48 h, 7 d and 10 d). In addition, unfiltered overlying site water was simultaneously collected from each test site in 1-L amber bottles and maintained at 4° C until shipment. Water and sediment samples were stored at 4°C. These samples were shipped to Dr. Tiernan's laboratory at Wright State within 24 hrs of collection. The 48-h, 7-d and 10-d sediment and water samples were analyzed for total PCBs. In addition, the 7-d water and sediment samples also were analyzed for all 209 PCB congeners by Dr. Tiernan's laboratory, and were then sent to Severn Trent Laboratories (STL) where they were analyzed for semi-volatile organic compounds (SVOCs),


including polycyclic aromatic hydrocarbons (PAHs), pesticides, metals and total organic carbon (TOC, only sediments). Throughout collection and shipment of the samples to WSU and STL, chain of custody procedure was followed for all samples.

2.2 Culturing

For all toxicity tests (i.e., laboratory and in situ tests), early life stages of test organisms (except Lumbriculus variegatus where adult worms were used) were implemented as prescribed. Culturing procedures followed USEPA methods for Hyalella azteca, Chironomus tentans, Daphnia magna and Lumbriculus variegatus (e g., USEPA 1993, 1994).

2.3 Task 1. Sediment laboratory toxicity testing - Summer 1999.

Laboratory life cycle testing began on May 27, 1999 for the H. azteca and July 9, 1999 for the C. tentans. At the time the laboratory tests were conducted, there were no formalized test methods for chronic toxicity testing of sediments. However, the U.S. Environmental Protection Agency (USEPA) and the American Society for Testing and Materials (ASTM) had identical draft methods that were in the process of final review These methods were finalized in early 2000 (USEPA 2000b). Therefore, the testing for this project followed the latest draft guidance, as of 1998, for chronic toxicity testing of Hyalella azteca and Chironomus tentans (USEPA 1998). Because these methods provide evidence of the effects of chronic exposures, they were deemed as a useful evaluation tool for this site.

2.3.1 Hyalella azteca Life Cycle Assessment Conditions for evaluating sublethal endpoints in a sediment toxicity test with H azteca are summarized in Table 1 and a general activity schedule is outlined in Table 2. The 42-d sediment toxicity test with H. azteca was conducted at 23°C with a 16:8 hr lightidark photoperiod at an illuminance of about 500 to 1000 lux (Table 1) Test chambers were 300-mL high-form lipless beakers containing 100 mL of sediment and 175 mL of overlying water. Amphipods in each test chamber were fed 1 0 mL of YCT (yeast-cerophyl-trout chow) daily (USEPA 1998) Each test chamber received 2 volume additions/d of overlying water.

Controls sediments included Ottawa sand and autoclaved Trout Farm (Fairborn, OH) In addition to laboratory controls, two field references were used for evaluating test performance and site differences ( Sites 011 and 398).

Review Comment: More documentation is needed regarding the source and composition of the two control sediments (Ottawa sand and Trout Farm), as well as the reason for testing two controls instead of one Water quality data for this control were included in the appendix, but only for the first 4 weeks because of a survival problem If survival was poor enough to warrant terminating this treatment then perhaps the Ottawa sand is not a suitable control, and text should document this In addition to the control sediments, the type of water used as overlying water in the test containers should be identified. Additional information is needed from WSU to clarify this issue


A total of 12 replicates, each containing ten, 7- to 14-d old amphipods were tested for each treatment. Twelve replicates were initiated on Day -1 to assess 28-d survival, of which 4 replicates were used for 28-d growth. The remaining 8 replicates were reused for measurement of survival and reproduction on Day 35, and survival, reproduction, and growth on Day 42.

Placement of Sediment into Test Chambers: The day before the sediment test was started (Day -1), each sediment was thoroughly homogenized and added to the test chambers. Each test chamber contained the same amount of sediment, determined by volume. Overlying water was gently added to each chamber on Day -1 in a manner that minimized suspension of sediment. The test began when the organisms were added to the test chambers (Day 0).

Acclimation: Test organisms were cultured and tested at the same temperature. Test organisms were cultured in the same water that was used in testing; therefore no acclimation was necessary.

Placing Organisms in Test Chambers: Amphipods were gently introduced into the overlying water below the air-water interface. Dry weight was measured on a subset of 30 amphipods prior to test initiation.

Feeding: For each beaker, 1.0 ml of YCT was added daily from Day 0 to Day 42 according to USEPA protocol. D.O. sag was briefly encountered during testing but was not due to food fouling. Detailed records of feeding rates and the appearance of the sediment surface were made daily.

Monitoring a Test: All chambers were checked daily and observations made to assess test organism behavior, such as sediment avoidance. However, monitoring effects on burrowing activity of test organisms is often difficult because the test organisms are often not visible during the exposure. The operation of the exposure system was monitored daily.

Overlying water quality was measured for: dissolved oxygen (mg/L), temperature (°C), conductivity (jimhos/cm) hardness (mg/L CaCO3), alkalinity (mg/L CaCO3), ammonia (mg/L total ammonia) and pH at the beginning of the sediment exposure portion of the test, weekly thereafter, then at the end of the test. Water quality measurements were also measured at the beginning and end of the reproductive phase (Day 29 to Day 42) (Appendices 1 and 2).

Dissolved oxygen was measured daily to ensure that chambers maintained a minimum reading of 2.5 mg/L. Aeration was required to maintain dissolved oxygen in the overlying water above 2.5 mg/L. Review Comment: Need to specify the day that aeration was started during the test.

Temperature was measured daily in at least one test chamber from each treatment. The daily mean test temperature was required to be 23 ± 1°C. The instantaneous


temperature was required to be 23 ± 3 °C. Aquarium heaters were used to maintain water bath temperatures within this range.

Ending a Test: Endpoints monitored included 28-d survival and growth of amphipods, 35-d survival and reproduction, and 42-d survival, growth, and reproduction (number of young/female from Days 28 to 42) of amphipods.

On Day 28, the sediment from 4 of the replicate beakers was sieved with an ASTM U.S. Standard #45 mesh sieve (355 j^m mesh) to remove surviving amphipods for growth determinations. Growth of amphipods was reported as mean dry weight/organism. The dry weight of amphipods in each replicate was determined on Days 28 and 42. Dry weight of amphipods was determined by transferring rinsed amphipods to aluminum pans, drying these samples for 24 h at 60°C, and weighing the dried amphipods on a balance to the nearest 0.01 mg. The mean dry weight of individual amphipods in each replicate was calculated from these data.

On Day 28, sediments from the remaining 8 beakers were sieved. The surviving amphipods from each sediment beaker were placed in 300-mL beakers containing approximately 250 ml_ of overlying water and a 5 cm x 5 cm piece of Nitex® screen or 3M® fiber mat. Each water-only beaker received 1.0 ml of YCT and two volume additions of water daily.

Due to the large number of Zumwalt dilutors required to accommodate all the sediment treatments, it was necessary to use several water baths to maintain the test temperature. Each bath was equipped with an aquarium heater and a power head to circulate water in the bath.

Reproduction of amphipods was measured on Day 35 and Day 42 in the water-only beakers by removing and counting the adults and young (neonates) in each beaker. On Day 35, the adults were then returned to the same water-only beakers. The number of adult females was determined by simply counting the adult males (mature male amphipods will have an enlarged second gnathopod) and assuming all other adults were females. The number of females was used to determine number of young/female from Day 28 to Day 42. Growth (dry weight) was also measured for these adult amphipods on Day 42.

2.3.2 Chironomus tentans Life Cycle Assessment Conditions for conducting a long-term sediment toxicity test with C. tentans are summarized in Table 3 and a general activity schedule is outlined in Table 4. The longterm sediment toxicity test with C. tentans was conducted at 23 °C with a 16L8D photoperiod at an illuminance of about 500 to 1000 lux (Table 3). Test chambers, sediment addition, water renewal, and water quality monitoring were as described above for H. azteca (See Section 2.3.1).

A total of 8 replicates, each containing 10, <24-h old larvae were tested for each sample. On Day -1, eight replicates were set up, of which four replicates were used for


20-d growth and survival endpomts and four replicates for determination of emergence Midges in each test chamber were fed 1 0 ml_ [confirm if this should be 1 5 mL] of a 4 g/L Tetrafm® suspension daily for the first week and then 15ml thereafter for the remainder of the test Endpomts monitored included 20-d survival, dry weight, ash free dry weight and percent emergence (adults) Control sediments included alpha-cellulose formulated sediment (USEPA 1998), Florissant soil, and Trout Farm sediment Review Comment: Confirm feeding ration for first week feeding 1 0 ml daily does not match text on Table 3

Review Comment: As with the Hyalella test more information is needed regarding the source and composition of the three control sediments (Trout Farm Florissant soil, alpha-cellulose formulated sediment) as well as the reason for testing three controls instead of one We also need to know which control sediment was used for statistical comparisons and the rationale for these decisions Finally the type of water used as overlying water in the test containers needs to be identified Additional information is needed from WSU to clarify this issue

Collection of Egg Cases: Egg cases were obtained from cultures of adult midges that were held in a male female sex ratio of 1 3 Adults were collected 4 days before starting a test The day after collection of adults, 6 to 8 of the larger egg cases were transferred to a petri dish with culture water and incubated (at 23°C) Hatching typically begins around 48 h and larvae typically leave the egg case 24 h after the first hatch

Hatching of Eggs: After the first 72-h period and upon the first visible migration of larvae out of egg cases, which indicates hatching, the cases were transferred from the incubation petri dish to another dish with clean test water The action of transferring the egg case stimulates the remaining larvae to leave the egg case within a few hours These larvae were used to start the test

Placing Organisms in Test Chambers: To start the test, larvae were carefully collected with a Pasteur pipette from the bottom of the incubation dish with the aid of a dissecting microscope Test organisms were then gently pipetted directly into overlying water beneath the air/water interface All larvae were transferred to exposure chambers within 4 h of emerging from the egg case

Feeding: Each beaker received a daily addition of 1 5 ml of Tetrafm® (4 mg/mL dry solids)

Dissolved Oxygen: Routine chemistries on Day 0 were taken before organisms were placed in the test beakers Test beakers were maintained at a DO concentration of greater than 2 5 mg/L to insure satisfactory performance If the DO level of the water fell below 2 5 mg/L for any one treatment, aeration was conducted in all replicates for the duration of the test (Appendix 3 or 4)

Monitoring Survival and Growth: At 20 d, 4 of the initial 8 replicates are selected for use in growth and survival measurements C tentans were collected using an ASTM

Burton - Lower Housatomc River Project Page 13

#45 sieve (355-|im mesh) to remove larvae from sediment. Surviving larvae were kept separated by replicate for dry weight measurements; if pupae were recovered, those organisms were included in survival data but not included in the growth data. The ash free dry weight (AFDW) of midges was determined for the growth endpoint. All living larvae per replicate were combined and dried to a constant weight (e.g., 60°C for 24 h). The sample was brought to room temperature in a dessicator and weighed to the nearest 0.01 mg to obtain mean weights per surviving organism per replicate. The dried larvae in the pan were then ashed at 550°C for 2 h. The pan with the ashed larvae was then re-weighed and the tissue mass of the larvae was determined as the difference between the weight of the dried larvae plus the pan and the weight of the ashed larvae plus the pan.

Monitoring Emergence: Emergence traps were placed on the reproductive replicates on Day 20. Emergence in control sediments typically begins on or about Day 23 and continues for about two weeks. For this portion of the test complete emergence and partial emergence was recorded. Complete emergence occurs when an organism has shed the pupal exuvia completely and escapes the surface tension of the water. If complete emergence has occurred but the adult has not escaped the surface tension of the water, the adult will die within 24 h. Therefore, 24 h elapses before this death is recorded. Partial emergence occurs when an adult has only partially shed the pupal exuvia. Between Day 23 and the end of the test, emergence of males and females, pupal and adult mortality was recorded daily for the reproductive replicates.

Ending a Test: The point at which the life cycle test is ended depends upon the sediments being evaluated. In clean sediments, the test typically requires 40 to 50 d from initial set up to completion if all possible measurement endpoints are evaluated. However, test duration will increase in the presence of environmental stressors that act to reduce growth and delay emergence. Where a strong gradient of sediment contamination exists, emergence patterns between treatments will likely become asynchronous, in which case each treatment needs to be ended separately. In this study, testing lasted for a total of 43 days. At this time, all beakers of the treatment were sieved through an ASTM #45 mesh screen (355 m) to recover remaining larvae, pupae, or pupal castes. According to draft chronic guidance criteria, a treatment may be terminated after one week elapses since the last adult has emerged from that test treatment (USEPA 1998). Performance control acceptability criteria requires that survival in the control treatment must be >60% of the initial stocked quantity (USEPA [2000] specifies >50% control emergence).

2.4 Task 2. In situ toxicity and bioaccumulation

The Task 2 activities were divided into 3 low flow testing periods, a 48-h, a 7-d and a 10-d exposure. The 7-d and 10-d exposures were initiated simultaneously. Testing periods were to include a high flow exposure period, in addition to the low flow period, but due to an uncharacteristically dry summer, the high flow sampling period was not conducted. The midge, Chironomus tentans ( 8 -12 days post hatch), the amphipod, Hyalella azteca (7 -14 days old), the oligochaete worm Lumbriculus variegatus (multiple ages) and the daphnid, Daphnia magna (48-hours old) were chosen for in situ


evaluation The age of the organisms, handling, and cultunng follows USEPA toxicity test methods (USEPA 1993, 1994) Due to organism health concerns related to shipping induced stress, organisms were purchased from a nearby test organism supplier (Aquatic Research Organisms, Hampton, NH) and transported to the on-site laboratory 24 hours prior to test initiation and monitored for health prior to deployment All test organisms were from USEPA reference toxicant tested stocks Quadruplicate chambers were deployed containing the early life stages of each organism and exposed for 2 -10 days Ten of each species were used for each replicate, however, C tentans and H azteca were grouped together and D magna and L vanegatus were exposed individually Ground up laboratory paper toweling was provided as a substrate for the amphipod/midge chambers (Chappie and Burton 1997)

Sample sites in Task 2 included a laboratory control for each organism [refer to Comment HI-7], 2 reference sites, and 4 test sites

Review Comment: Here and on page 16 for the in situ tests we need to explain what the "laboratory control" was Where were these conducted and how relevant are they to the field exposures? The concept of a lab control is not intuitive for in situ testing Additional information is needed from WSU to more fully describe the laboratory controls" for the in situ tests

Site selection criteria included the following sediments with grain sizes typical of the site, sediments from sites where PCB concentrations were likely to be at high levels, site of previous sampling, and site near proposed sampling for physicochemical characterizations, habitat, indigenous benthic communities and storm water sampling Site 428 was substituted for Site 023 for these in situ tests A 48-h low flow test was conducted with D magna, H azteca, C tentans and L vanegatus to evaluate potential toxic responses for each species A 10-d low flow test, also for survival, was conducted after completion of the 48-h screening test and included H azteca and C tentans only The third test was a 7-d L vanegatus bioaccumulation assay initiated on the same day as the 10-d midge/amphipod survival assay Past work on PCB-contammated sites by WSU indicated that uptake occurs rapidly in invertebrates and near steady-state within 4 days (Burton to provide reference)

The in situ chambers used for this study were constructed of clear core sampling tubes (cellulose acetate butyrate) cut to a length of approximately 15 cm Polyethylene closures capped each end Two rectangular windows (~85% of the core surface area) covered with 80-(am Nitex® mesh were placed on opposite sides of the core tube Organism exposures were limited to (1) the water column, via placement of the chamber in the top tray of the in situ basket, (2) interaction with both sediments and overlying water via placement of the chamber against the sediment surface by securing it to the lower in situ basket with one window facing the sediment and one mesh window facing the overlying water column (Figure 1a) or (3) in the sediment via filling chambers approximately one-third with sediment and the rest with overlying water (/ e the third method was only used for the in situ bioaccumulation exposures) In situ baskets were weighted down with bricks and anchored to the stream bottom with rebar Each set of


baskets was covered by a stainless steel flow deflector designed to divert strong currents of water and turbulence around the in situ chambers should a high flow event have occurred during exposure (Figure 1b). The functional design of the flow deflector would thus prevent the baskets from being swept away during short periods of high flow conditions. All treatments were conducted in replicates of four chambers per test species.

Photo 1: a) In situ chambers deployed in wire baskets and b) in situ chambers/baskets protected

with flow deflectors

Prior to chamber deployment, 10 of each organism (H. azteca, C. tentans and D magna) were gently added to 50-mL test tubes of culture water for ease of transport to field locations (one test tube contained one species only). Transportation of organisms to field sites by this method has proven to minimize handling and travel related stresses. For the 7-d L. variegatus bioaccumulation assay, either 1 or 2 g of tissue (equal to approximately 1:10 animal wet wtsediment organic carbon) was used in each chamber. In the field, site water temperatures were measured and additional acclimation took place in the field when necessary. Upon acclimation, in situ chambers capped on one end were immersed into the river allowing water to fill the chamber by infiltration through the mesh and test organisms were slowly delivered from the test tubes into the open end. The chambers were then capped. Before placement into in situ baskets, chambers were held below the water surface and purged of all internal air. After 2-10 days of exposure, in situ chambers were gently lifted out of the river and placed into coolers of site water for the return trip to the R.F. Weston laboratory for enumeration. Upon arrival at the lab, chambers were inspected for damage, rinsed on the outside and individually emptied into crystallizing dishes. The survivors of each species were enumerated and logged

In situ chambers were deployed at all field locations on a total of 2 occasions in which stream conditions were considered to be at base flow. The 48-h in situ assay began on the afternoon of June 14, 1999 and included H azteca, C. tentans, L. variegatus and D. magna. Chambers were retrieved in the afternoon of June 16, 1999 after 48 h of exposure. The second group of chambers was deployed on the morning of June 17, 1999 and included H. azteca, C. tentans and L. variegatus The L. variegatus chambers were retrieved on the morning of June 24, 1999 after 7 days of exposure for tissue


bioaccumulation L vanegatus tissue samples collected after the exposures were allowed to depurate in diluted mineral water (DMW) for 12 h, weighed (wet weights) to four decimal places [express this as a weight, eg, 0 01 mg], transferred to dry scintillation vials and stored at -15°C at R F Weston's facilities The tissue samples were then shipped on dry ice to Dr Tiernan's laboratory at Wright State for chemical analysis

H azteca and C tentans chambers were retrieved on the morning of June 27, 1999 after 10 d of exposure A laboratory water control was maintained for standard quality control purposes for each of the test organisms deployed in the field during all of the exposure periods

For each field site, water quality measurements were made at test initiation and again upon test termination Physicochemical measurements included temperature dissolved oxygen, pH, hardness, alkalinity, turbidity, conductivity and total ammonia (Appendices 5 and 6) All field water quality monitoring equipment was calibrated prior to each use according to USEPA , instrument specifications and/or the Weston QAPP

2.5 Task 3. Toxicity Identification Evaluation (TIE)

The Toxicity Identification Evaluation (TIE) is a process by which effluent or pore water samples are fractionated into various classes of contaminants and then tested for toxicity There were two primary objectives in the Phase I TIE (1) to detect the presence of toxicants and (2) broadly define the physical/chemical characteristics of the toxicant(s) If a significant change in toxicity is seen following these characterization procedures, additional tests are needed to further characterize the nature of toxicity If toxicity is not removed, other types of contaminants (not addressed by Phase I) may be present or a single toxicant may be present at such high levels that only a partial toxicity decrease is achieved If several Phase I tests succeed in partially removing toxicity, there may be several toxicants, each with different physicochemical characteristics, or a single toxicant that can be removed by more than one Phase I test (e g increased pH can cause a metal to precipitate out and EDTA can also remove the metal toxicity by chelation) The difference in Toxicity Units (TUs) for each fractionation helps determine which chemical class is the cause of toxicity (USEPA 1991 a) Separating contaminant classes allows one to characterize which class of contaminants is primarily responsible for toxicity (USEPA 1991a,b) Commonly detected contaminants are pH sensitive and volatile compounds (eg, ammonia), metals, oxidants/reductants, and nonpolar organics Toxicity is determined by exposing C dubia to a dilution series of the test water for 48 h and then measuring survival For this project, a TIE was conducted using draft USEPA guidelines for pore water (USEPA 1991b) Previous studies with nonionic organics, such as PCBs, have shown that bioavailabihty of contaminants to benthic organisms is strongly correlated with pore water concentrations in this freely dissolved fraction (Adams et al, 1985, Swartz et al , 1985, 1990, DiToro et al 1990, 1991 and Ankley et al , 1993, 1994) Pore water ahquots were used for initial toxicity tests (within 24 h of sample receipt), baseline ambient pore water, pH adjusted with aeration, pH adjusted with filtration, pH adjusted with Ci8 filtration, sodium thiosulfate addition, and


EDTA addition fractions. Diluted mineral water (DMW) was used for all pore water dilutions and tests.

R.F. Weston personnel collected sediment samples from each of the study sites (i.e., where in situ studies were based) for Phase I Toxicity Identification Evaluation (TIE) tests. The top 5 cm of sediment at test Sites 398, 019, 428, 389 and 031 were collected the week of August 9, 1999 and shipped to Soil Technologies, Inc. (STI). Pore water was extracted from these sediments using either the centrifuge method or nitrogen pressure. Samples 031 and 389 (fine grained sediments) were centrifuged at 7,000 rpm for 30 minutes. Pore waters were extracted from samples 428, 019 and 398 anaerobically in a nitrogen chamber. In addition, samples 428, 019 and 398 (sandy sediments) were also centrifuged for 30 minutes at 5,500 rpm to further remove particulates. All pore water samples were stored in 1-L glass amber bottles and shipped to Wright State University on August 13, 1999 for TIE test manipulations. Samples were received August 15, 1999. Review Comment: White the pore water samples were used for TIE testing within 24 h of receipt, that wasn't until at least 48 h after their extraction. Given the changes in toxicity observed between Day 1 and Day 3 of the Phase I TIE, it would be useful to evaluate and discuss potential impact on results, if any.

Also, need to explain why two methods used for pore water extraction, and discuss potential for any impact on the outcome of the TIE results? Why is there no mention here of the fact that the pore water container for Site 428 was broken (and presumably not tested)?

2.5.1 DAY 1 Day 1, initial toxicity tests for all samples were started on August 15, 1999. Prior to initial toxicity tests, the temperature was raised for all pore water samples in a 24 °C water bath. Water quality measurements were recorded on all samples. Ceriodaphnia dubia (5 individuals/replicate) were introduced into test dilutions of a control treatment (i.e., culture water and no pore water) and 6.25, 12.5, 25, 50 and 100% pore water treatments. C. dubia were fed a mixture of Selenastrum capricornutum and cereal leaves (Sel-Cero).

2.5.2 DAY 2 Day 2 test manipulations were performed on August 16, 1999 and included; oxidant reduction (sodium thiosulfate), EDTA addition, and pH adjusted filtration. Up to three types of "controls" were used to check for artifact toxicity; "toxicity blanks" (no manipulation), the baseline toxicity test and dilution water (DMW) controls (Table 5). The same test manipulations were performed on the dilution water controls (except baseline and initial tests) to determine if test manipulations contributed to toxicity. If toxicity was affected by the sample matrix, a toxicity control or blank would not help in identifying artifact toxicity. In all cases, baseline tests (i.e., unaltered pore water aliquots) served as controls for all tests while providing information on temporally changing toxicity of the initial samples. The baseline test serves as a "toxicity blank" and control for the oxidant reduction test. If the baseline test is less toxic than the manipulated sample, the test procedure has artificially increased toxicity. The EDTA


addition test and oxidant reduction (sodium thiosulfate) addition test dilution water blanks are not considered relevant as "controls", instead the baseline test is used as a control (USEPA 1991 a). However, EDTA and sodium thiosulfate were added to DMW as a treatment control (Table 5).

The same sample dilution series that was used for the initial toxicity test (Day 1) was also used for the Day 2 test manipulations because LCso values were >25%. Corresponding Day 2 baseline test results were compared to the Day 1 initial toxicity test results to determine if toxicity had changed over time.

The oxidant reduction (sodium thiosulfate) addition test (48 h) was performed for identification of toxicity due to oxidants and cationic metals. Oxidants generally reduced or neutralized in this test are chlorines, ozone, bromine, iodine, manganous ions and some electrophilic organics. In addition, some cationic metals can be chelated by sodium thiosulfate, (e.g., Cd, Cu, Ag and Hg) but complexation can be slow (up to 96 h). Two methods of sodium thiosulfate additions are recommended and the dilution/3 X 3 matrix (pore water and thiosulfate concentrations) approach was used in this test (USEPA 1991a,b).

The 48-h EDTA chelation test was conducted to identify toxicity due to certain cationic metals. The pH adjusted filtration test was conducted to identify the effects of filterable compounds. Although positive pressure filtration is recommended, the vacuum method was required to sample turbidity and to achieve holding time requirements. This method modification may have reduced concentrations of volatile compounds from the subsequent toxicity assays.

2.5.3 DAY3 USEPA methods suggest all characterization tests be performed at approximately the same time to avoid confounding results due to degradation of sample toxicity over time It was necessary to perform pH adjusted/Ci8 solid phase extraction and aeration tests on Day 3 (August 17, 1999). These tests were delayed due to the late arrival of the samples and the need for initial toxicity test 24-h LC50 results to be obtained to calculate thiosulfate and EDTA test concentrations. Therefore, the baseline pore water toxicity test was repeated on Day 3 to determine storage effects. Due to the 24 hour wait for the initial toxicity test results, the pH adjusted aliquots were allowed an additional equilibration time (as suggested in the USEPA 1991a manual) and pH adjustments were made as necessary when pH drift was detected. Corresponding Day 3 baseline test results were compared to the Day 1 initial toxicity test results and Day 2 baseline test results to determine toxicity changes.

The pH adjusted/Cis SPE characterization test was conducted for non-polar organics, pesticides and some metal chelates that are relatively non-polar. The Ci8 column (an octadecyl resin) can also extract organic acids and bases. By adjusting sample pH to low (pH = 3) and high (pH = 9), the un-ionized form in some of these compounds predominates and sorbs in the column.


The pH adjusted/aeration characterization test for volatile, sublatable or oxidizable compounds was also conducted. Aeration was performed at three pH levels because some compounds can be oxidized or removed more easily at a particular pH.

2.6 Statistical Analysis

For the laboratory and in situ toxicity tests (Sections 2.3 and 2.4), data meeting the assumptions of normality and homogeneity of variance by analysis of variance (ANOVA) were followed by Dunnett's test (Toxstat version 3.0). Data not meeting assumptions of normality were analyzed using Steel's many-one rank test. Statistical significance level, p = 0.05. The Pearson correlation coefficient, R, was used as a quantitative measure of the relationship between sediment PCB concentration and animal lethality. All reported mean values were calculated as arithmetic means.

Review Comment: This section needs to clarify whether statistical comparisons were made to the laboratory control or to one or both of the reference sediments.

For the TIE toxicity tests (Section 2.5), the LC50 values were estimated via the Spearman-Karber model with 95% confidence levels. Toxic Units (TUs) were calculated asTU = 100/LC50.

2.7 Quality Assurance

Protocols for the chronic toxicity test methods were followed as outlined (ASTM 1999, USEPA 1998). Other quality assurance issues are addressed in the Quality Assurance Project Plan for the U.S. Environmental Protection Agency's Freshwater Sediment Toxicity Methods Evaluation (Burton 1997) and as described in the study SOP in the SWIP [define].


3.0 RESULTS AND DISCUSSION: TASKS 1 AND 2

During sediment homogenization of in situ sediment samples and when replicate beakers were sieved at test termination (see Lab test methods) the sediments from field Sites 031 and 389 produced an oil sheen on the water surface and an organic odor This was not observed in any of the other collected samples As stated previously, analysis for other contaminants were performed and a comparison to current Sediment Quality Guidelines (SQGs)

3.1 Task 1. Sediment macroinvertebrate laboratory toxicity testing - Summer, 1999

From the upstream reference sites (398, 011) to the most downstream site (031), PCB concentrations for laboratory test sediments ranged from less than the detection limit to 213 mg/Kg dry for laboratory tests (Table 6) A concentration gradient of total PCBs made it possible to compare exposure gradients with effect levels

3.1.1 Hyalella azteca Life Cycle Assessment Both the laboratory control (Trout Farm sediment) and the reference site (398) met the 28-d H azteca acceptable control survival criteria of 80% (USEPA 1998, 2000b) Site 398 was located on the West Branch and consistently had high survival in all Task 1 and Task 2 assessments Site 011 was also chosen as an upstream reference but was located on the East Branch in the town of Dalton The sediment at Site 011 was larger grained than downstream sediments and contained no detectable total organic carbon It had consistently lower survival than 398, but higher survival than downstream sites The 28-d survival was 81 8% in the laboratory control sediment and 83 3% in the upstream reference Site 398 Review Comment: Up to this point Sites 398 and 011 have been referred to as upstream reference sites However at this point Site 011 is described as having different physicochemical characteristics from the downstream sediments and from this point onward it is grouped with the four 'exposure" sites rather than the control(s) and Site 398 when results are presented (and Site 398 is now the only one referred to as a reference sediment) This report must be consistent in terms of how Site 011 is treated especially if statistical comparisons are made and how the reference sediment(s) are referred to This issue requires clarification from WSU about how they handled the data from both reference sediments it also ties in with the previous comment regarding statistical comparisons

Survivals at 28 d in the remaining sediment exposures were 65 8% at Site 011 48 3% at Site 019, 0% at Site 389, 25 8% at Site 023, and 22 5% at Site 031 (Table 7, Figure 2) At 28 d, survivals at all sites were statistically different from reference site 398 Survival in reference site control sediment 398 was 83%, 80% and 75% at Days 28 35 and 42, respectively (Table 8, Figure 2) Survival within a given site did not significantly differ between Days 35 and 42 of the exposure period Exposure to sediment from Site 389 resulted in 100% Hyalella mortality in less than 28 d However, survival between Days 35 and 42 at remaining sites varied less than 4% Site 019 (8 7 mg/Kg dry PCB) had the lowest PCB contamination with detectable mortality at Days 35 and 42


Growth, represented as dry weight, was measured on a subset of organisms from each location at Days 0, 28 and 42 (Tables 9 and 10, Figures 3 and 4). The 28-d mean dry weight of the control (Trout Farm) replicates was 0.56 mg/organism. The control performance acceptability criterion required that control organism dry weight average 0.15 mg/organism after 28 d of sediment exposure. As sediment PCB levels increased, organism 28-d mean dry weights decreased. There were no surviving animals in Site 389 after 28 d of exposure. At all other sites, organism dry weights slightly increased between Day 28 and Day 42. Growth of the control organisms seemed to cease around 0.55 mg/organism after Day 28. The control dry weight at Day 42 was 0.56 mg/organism and 0.53 mg/organism at the reference site. All site treatments were above the established control acceptability criteria after 42 d except for Site 389, which had no survival.

Following 28 d of sediment exposure, surviving amphipods were transferred to clean culture water and allowed to recuperate for two additional weeks while reproductive success (mean number of neonates/female) was evaluated. The control performance acceptability criterion for neonate production is 2 neonates/female between 28 and 42 d of exposure. The control treatments averaged 13.4 young/female between Days 28 to 42, and the reference site (398) averaged 9 neonates/female. Sites 011, 019, 023 and 031 averaged 7.7, 6.3, 2.7 and 0.1 neonates/female, respectively (Table 8, Figure 5). Site 023 was the site with lowest PCB contamination impairing total reproduction from Days 28-42.

Dry weight and mean neonate/female control acceptability criteria endpoints were met for both the laboratory control treatment and the field reference sample (Site 398). Site 389 (i.e., site with highest PCB concentrations) yielded 100% mortality within the first 28 days of sediment exposure, so could not be continued in the water-only exposure.

3.1.2 Chironomus tentans Life Cycle Assessment Acceptability criteria for both the laboratory control (Trout Farm sediment) and the reference site (398) were met for C. tentans 20-d survival. The 20-d survival was 85% in the control sediment and 77.5% in the reference sediment. The 20-d survival was 52.5 % at Site 011, 5.0% at Site 019, 0% at Site 389, 0% at Site 023, and 7.5% at Site 031 (Table 11, Figure 6). Site 019 (8.7 mg/Kg dry PCB) was the lowest level of PCB contamination with adverse effects on 20-d survival. As observed in the amphipod exposures, C. tentans survival seemed to be inversely proportional to sediment PCB concentration.

The 20-d mean dry weights of organisms in the control and reference sediment (Site 398) were 2.20 and 1.97 mg/organism, respectively, and were well above the minimum performance acceptability control criteria (0.6 mg). Organisms from Sites 011, 019 and 031 yielded dry weights of 1.82, 0.035 and 0.04 mg/organism, respectively. Because complete mortality resulted following exposure to sediments from Sites 389 and 023, there were no dry weight measurements for these sites (Table 12, Figure 7). As with survival, the growth was also affected at Sites 019 and 031. Control organisms


averaged 0.56 mg/organism, ash free dry weight (AFDW) which is above the minimum performance acceptability criteria (0.48 mg/organism). The organisms exposed to reference site sediments grew larger than the organisms exposed to the laboratory control sediments with an average AFDW of 0.81 mg/organism (Table 13, Figure 8). Exposure to sediments from Sites 011 and 019 resulted in AFDWs of 0.28 mg/organism and 0.02 mg/organism, respectively. Only one organism was retrieved from the Site 031 treatment and the balance was not sensitive enough to detect the AFDW of this sample. There was complete midge mortality in sediments collected from Sites 389 and 023. Growth (AFDW) for midges exposed to all test site sediments for 20 d was significantly different from growth at the reference site (398).

C. tentans emergence in the control was 67.5% and 35% in the reference sediment. The control performance acceptability criterion for emergence of >50% was met. Adult emergence in sediments 011 and 023 was 30% and 2.5%, respectively. Sites 019, 031 and 398 had no adult emergence and no pupae were recovered (Table 14, Figure 9). The site with lowest PCB contamination affecting emergence was 019 (8.7 mg/Kg dry PCB).

3.2 Task 2. In situ toxicity and bioaccumulation

From the upstream reference sites (398, 011) to the most downstream site (031), total PCB concentrations at in situ test site sediments ranged from less than the detection limit to 522 mg/Kg dry between sites, however, from sediment chemical analyses, there also appeared to be a high degree of intra-site sediment PCB variability (e.g., Table 6 vs. 15).

3.2.1 48-h low flow in situ The 48-h laboratory control survival results were as follows: 75% for H. azteca (Figure 10), 92.5% for the D. magna (Figure 11) and 97.5% for both L. variegatus (Figure 12) and C. tentans (Figure 13, Table 16). In the 48-h in situ exposures, daphnids and amphipods were the most sensitive species. D. magna and H. azteca survival was >80% for water column exposures at all sites and >82.5% at sites 398, 011 and 019 in the sediment exposures. However, daphnids exposed to sediments at sites 428 (7.3 mg/Kg dry PCBs), 389 (139.2 mg/Kg dry PCBs) and 031 (521.7 mg/Kg dry PCBs) exhibited markedly reduced survivals. Average daphnid survival was 47.5% at Site 428 and 0% at Sites 031 and 389. Decreased D. magna survival following against sediment exposure correlated with increasing sediment PCB concentration (Figure 14). Amphipod survival was similar to daphnids at Sites 428 (40%) and 389 (22.5%) for sediment exposure. C. tentans and L. variegatus did not appear to be severely affected following 48 h of exposure at any of the sites and survival was 85% or higher for both species in both water and sediment exposures. During this exposure, lower survivals occurred primarily in sediment chambers rather than water column chambers.

3.2.2 10-d low flow in situ The 10-d in situ exposure with H. azteca and the C. tentans was initiated on June 17, 1999. As stated in the methods, test species were exposed to both the overlying water column and the surficial sediments. The 10-d laboratory control survival was 100% for


H. azteca (Figure 15) and 95% for C. tentans (Figure 16) (Table 17). Exposure at Sites 398, 011 and 019 yielded greater than 80% survival for both species regardless of exposure compartment (e.g., against sediment or in water column). However, at Site 428, H. azteca and C. tentans exposed to sediments began to exhibit signs of stress showing a decline in survival to 52.5% (Figure 15) and 77.5%, respectively (Figure 16) (Table 17). Following exposures in the against sediment chambers, 0% of the H. azteca and 7.5% of the C. tentans survived at Site 389 (52.2 mg/Kg dry PCBs), and only 2.5% of the H. azteca and 20% of the C. tentans survived at Site 031 (111.9 mg/Kg dry PCBs). Decreasing survival and increasing total PCB concentration correlated for both H. azteca and C. tentans exposed against sediments (Figures 17 and 18).

In general, amphipod and midge survival following water column exposure was higher when compared to survival resulting from sediment exposures. Exposures at upstream sites (Site 398 at 5.8 ng/L PCB, Site 011 at 5.7 ng/L PCB and Site 019 at 129.1 ng/L PCB) resulted in survival of 80% or greater for each species. At Site 428 (143.8 ng/L PCB) H. azteca showed 70% survival and C. tentans had 60% survival. H. azteca survival dropped to 62.5% at Site 389 (299.2 ng/L PCB), but was 85% at Site 031 (98.7 ng/L PCB). Midge survival, however, was 92.5% at Site 031 and 97.5% at Site 389 in the water. There was a relationship between decreased H. azteca survival following water column exposures and total PCB concentrations (Figure 19). Review Comment: According to previous text in this report, overlying water was sampled along with the sediments at the end of the 48-h, 7-d and 10-d in situ tests. If so, then we should discuss more than just the 10-d water data presented here and include all water data in Table 15. Also, raw data for all the sediment and water chemistry analyses should be included in the report to facilitate independent data evaluation; these data need to be provided by WSU.

3.2.3 7-d bioaccumulation in situ The relationship between total PCB tissue concentrations in L. variegatus and total PCB sediment concentrations is shown in Figure 20. Tissue PCB residues were normalized to lipid content and total sediment PCB concentrations were normalized to organic carbon content. However, a large degree of both spatial and temporal heterogeneity was detected for sites and sediment samples. Total PCB levels at Site 031, for example, were 71 mg/Kg dry for the lab studies, and 521.7 mg/Kg dry for the 48-h, 16.9 mg/Kg dry for the 7-d and 111.9 mg/Kg dry for the 10-d in situ studies (Tables 6 and 15). A general trend of increasing PCBs contamination, however, was observed from upstream to downstream sites for all studies. Results of PCB congener analysis showed PCB levels in both sediments and tissue residues contained at least eight PCB congeners considered either highest priority or most "dioxin-like" of the PCB compounds by the US EPA.

Review Comment: This last sentence about PCB congener data does not fit with the rest of the paragraph; how does it relate to the L. variegatus results? More importantly, the complete results of these congener-specific analyses should be reported in tables or an appendix. Figures (not shown) developed by WSU show that some quantity of


PCS congeners in the "highest priority or most dioxm-like categories were detected but do not give any indication of the number of congeners that were detected

There is also some uncertainty regarding total organic carbon (TOC) levels For Site 389, TOC was measured at 4 0% for the lab studies and 6 0% for the TIE tests, but was initially reported as non-detected for the 7-d in situ study This meant that calculations requiring OC-normahzation could not be performed for this sample and were therefore not included in earlier drafts of this report A TOC concentration of 4% has now been identified for the 7-d in situ test

Review Comment: The source and validity of the 7-d TOC value reported for Site 389 must be confirmed, as well as all calculations and conclusions based on that value According to previous drafts TOC was not detected for the 7-d in situ study this would be surprising since TOC values were quite high for the laboratory and TIE studies We have adopted a value of 4% for this parameter until the uncertainty is resolved

3.2.3.1 PCBs in Pore Water The equilibrium partitioning approach (Di Toro et al 1990) was used to estimate the pore water concentrations of PCBs based on site-specific measurements of PCBs and TOC in sediment samples from Sites 011, 019, 428 and 031 (Tables 18 and 19) According to equilibrium partitioning theory, the freely dissolved pore water concentrations of nonpolar organic compounds can be estimated using the following relationships

cpw = CS/KP (1), where Cpw = concentration in the pore water (mg/L), Cs = concentration in the sediment (mg/Kg dry), and Kp = the sediment-pore water partition coefficient (L/kg) Kp is determined by the equation

Kp = foc-Koc (2), where foc = fraction organic carbon in the sediments and Koc = the organic carbon-water partition coefficient Koc is determined from the octanol-water partition coefficient (Kow) by the following relationship

log Koc = 0 00028 + (0 983-log Kow) (3) The values of Kow were taken from the published literature (Veith et al 1979, Hawker and Connell 1988, Gabric et al 1990, Boese et al 1997, Fisher et al 1999, USEPA 2000a) Pore water concentrations were not calculated for Sites 398 and 389 because TOC measurements concurrent with the sediment contaminant analyses were not available

Review Comment: This statement about TOC data not being available for Site 389 does not match Table 15, which shows 4% TOC for this station Need to resolve the TOC for this station and make appropriate changes throughout report and tables if necessary


3.2.3.2 Bioconcentration ofPCBs from Pore Water The estimated pore water concentrations (Table 19) were used to calculate the site-specific bioconcentration factors (BCFs) for L. variegatus tissues (Table 20). The BCF is determined from the following relationship:

BCF = Ca,Ss/CpW (4),

where Ca,Ss = steady-state tissue concentration of contaminant in the organism (mg/Kg wet) and Cpw = concentration of contaminant in the pore water (mg/L). The BCFs reported in Table 20 were calculated using the mean total PCB tissue concentrations and are reported on both a whole-body, wet weight basis (BCFww) and a lipid weight basis (BCF|). The assumptions of the BCF are that the tissue concentrations are at steady-state with environmental concentrations and, for PCBs, that the contaminants are not metabolized.

The values of BCF for total PCBs in the oligochaetes in the present study (log BCF^ = 4.38 - 5.80; log BCF, = 6.40 - 7.57; Table 20) are similar to published BCFs for total PCBs in field-collected macroinvertebrates (e.g., zooplankton, amphipods) (log BCF^ = 3.55 - 5.42, log BCF, = 4.23 - 6.14) (Kucklick et a/. 1996). For oligochaetes exposed in the laboratory to anthropogenically contaminated sediments containing di-, tri-, tetra-, penta-, hexa- and octachlorinated PCBs, individual congener log BCF^ values ranged from 3.17 (dichlorobiphenyl) to 5.45 (octachlorobiphenyl) (Connell era/. 1988).

Although BCF values provide an index of contaminant accumulation from aqueous exposures, they do not account for uptake from other exposure routes (e.g., ingestion). Therefore, for sediment-exposed species that interact directly with sediments, or selectively ingest sediment particles, the BCF value may overestimate the contribution of water-borne contaminant in the total body burden. More general bioaccumulation factors that consider aqueous and all other routes of exposure have been developed and are described for L. variegatus below.

3.2.3.3 Bioaccumulation Factors The site-specific bioaccumulation factors (BAF; Table 21) for L. variegatus were determined from measured sediment and mean tissue levels of total PCBs. The BAF is calculated from the following equation:

BAF = Ca,Ss/Cs (5),

where Ca,Ss = steady-state tissue concentration of contaminant in the organism (mg/Kg wet) and Cs = concentration of contaminant in the sediments (mg/Kg dry). The BAFs reported in Table 21 were calculated on both a whole-body, wet weight basis (BAFww) and a lipid weight basis (BAFi). The assumptions of the BAF are the same as those listed previously for the BCF. Note that the BAF is not calculated based on organic carbon-normalized sediment concentrations, which is useful in that any available sediment concentrations can be used. Therefore, the BAF as calculated above represents the simplest model of bioaccumulation and will vary with sediment type. The


accuracy of any predictions for a given organism will be limited if the organic matter or other sediment characteristics differ between sites (Lee 1992).

BAFi values reported in Table 21 range from 3.84 - 743. These values for L, variegatus tissues in the present study are within or above the ranges measured in oligochaetes (BAP,, 0.882 - 14.9) and chironomid larvae (BAF,, 2.08 - 17.1) collected from a PCB contaminated lake (Bremle and Ewald 1995). However, the highest BAFs occurred at the least contaminated reference site (Site 011; 7.12 ng total PCBs/g sediment), which illustrates, in general, the limited utility of the BAF.

Review Comment: It is not apparent where the value of 7.12 ng/g PCB comes from for Site 011, since according to Table 15, PCBs were not detected in sediments at this site. If this PCB value comes from a different analysis (e.g., the congener-specific one) then all the PCB data from that analysis should also be reported and compared to the data in Table 15 to see if there is any intra-site consistency. If the 7.12 ng/g PCB value is not appropriate for use, then it should not be used for BAF or BSAF calculations and the ranges and results will need to be adjusted accordingly. We have eliminated the BAF and BSAF values for this station in the Tables until this is resolved.

A more useful indicator of bioaccumulation is the biota/sediment accumulation factor (BSAF), which is the quotient of the lipid-normalized, steady-state tissue concentration in an organism and the organic-carbon normalized sediment concentration of a contaminant. L. variegatus BSAFs (Table 21) were calculated using lipid-normalized mean total PCBs in tissues, and only for sites at which sediment TOC levels were determined concurrently with sediment PCB analyses (i.e., Sites 011, 019, 428 and 031). TOCs at Sites 389 and 398 were reported as not detects. The assumptions of the BSAF model, as applied to total PCBs, are that the concentrations of PCBs are at steady-state between organism lipids and sediment organic carbon and that PCBs are not metabolized. A theoretical maximum BSAF (goc/giipid) of 1.7, based on the partitioning of neutral organic compounds between organic carbon and lipids, has been proposed (McFarland and Clarke 1986). A BASF < 1.7 indicates less partitioning into lipids than theoretically predicted whereas a BSAF > 1.7 indicates greater uptake than can be explained by simple partitioning.

BSAFs ranged from 0.80 - 6.05 (Table 21) with the lowest value for sediment-exposed L. variegatus at Site 031 (Cs,oc = 392 mg/Kg OC) and the highest BSAF determined at Site 019 (Cs,oc = 38.4 mg/Kg OC). These levels are within the range of mean BSAFs for total PCBs in nine species of bivalves (0.4 - 10.9), chironomid larvae (0.11 - 2.41), and oligochaete worms (0.07 - 1.01) sampled elsewhere (Lee 1992; Bremle and Ewald 1995).

The BAFs and BSAFs (Table 21) were consistently highest at the reference site (011) in the 7-d bioaccumulation study using L. variegatus as a surrogate infaunal invertebrate. Similar observations were made in the Bremle and Ewald (1986) study on indigenous midge larvae and oligochaetes collected from PCB-contaminated lakes. Such results are not surprising, because it has been frequently observed in field studies and in


laboratory-spiked sediment tests, that an inverse relationship exists between BSAF and sediment contaminant concentration and/or TOC (Rubenstein et al. 1987; McElroy and Means 1988; Ferraro era/. 1990a,b). Thus, "cleaner" sites often result in higher BSAFs than more contaminated sites. This may be due to a sublethal physiological response or a change in organism behavior (e.g., decreased feeding rate, decreased sediment reworking, contaminant avoidance) with increasing levels of sediment contamination (Keiltyefa/. 1988a,b).

3.2.3.4 Body Residues ofPCBs in L. variegatus Body residues (BR) of total PCBs measured in L. variegatus (Table 22) were calculated by converting the mass-based concentrations of each congener group (i.e., di- through deca-chlorinated biphenyls) to molar-based concentrations and then summing the congeners to obtain a total PCB body residue. The mean BRs (Table 22) ranged from 0.07 umol/Kg wet weight at Site 011 to 12.35 umol/kg wet weight at Site 428. The mean BR values were 0.35, 0.07, 11.17, 12.35, 4.43 and 8.85 umol/Kg wet weight at Sites 398, 011, 019, 428, 389 and 031, respectively. On a lipid-normalized basis, the mean BR values were 51.3, 7.47, 646, 1060, 379 and 912 umol/Kg lipid at Sites 398, 011, 019, 428, 389 and 031, respectively.

The critical body residue (CBR) is the tissue concentration at which mortality will occur in 50% of an exposed population. McCarty et al. (1992a,b) described the CBR (mmol/kg wet weight) for aqueous exposures to neutral, lipophilic chemicals with the following acute toxicity-based relationship:

CBR = BCF.LC50 (6),

where BCF = the bioconcentration factor (eq. 4) and LC50 = the aqueous concentration (mmol/L) of a contaminant at which 50% mortality occurs within an exposed population.

The body residue (BR) in a sediment-dwelling organism can be predicted by:

BR = BCF.Cpw (7).

The BCFs reported in Table 20 were calculated with 7-d tissue data and estimated site-specific pore water concentrations (Cpw) that were derived from sediment data in this study (eq. 1; Table 19). However, prediction of L. variegatus BRs with equation (7) using our interdependent BCF and Cpw values would not be biologically meaningful. Therefore, this calculation was carried out for sediment-exposed L. variegatus by using published, congener-specific BCFs for oligochaetes (Connell et al. 1988) and the estimated site-specific pore water concentrations (Table 19). The proportion that each congener represented in the total pore water PCB concentration at each site was accounted for in the calculations and the results are summarized in Table 22. Because BCFs were reported for tri-, tetra-, penta-, hexa- and octachorobiphenyl in Connell et al. (1988), only the estimated pore water concentrations for these congeners were used in the calculations. However, the estimated BRs from these calculations are useful for comparisons to the actual measured BRs because this set of congeners (tri-hexa and


octa-chlorinated biphenyls) represented between 84 - 93% of the total PCB body residues across the test sites (Table 22) The estimated di-hexa- and octa-chlorinated biphenyl BR was within a factor of 1 21, 1 12 and 1 69 of measured BRs for Sites 011, 428 and 031, respectively The model (eq 7) underestimated BR at Site 019 by a factor of 12 With the exception of Site 019, it was concluded from the measured BRs that L vanegatus exposed to contaminated sediments in the study area bioaccumulated PCBs to expected levels

To further evaluate the L vanegatus BR data, site-specific CBRs for whole-sediment exposures at Sites 011, 019, 428 and 031 were estimated using recently published consensus-based sediment effects concentrations (SECs, mg/Kg dry) (MacDonald et al 2000) (Table 23) These estimated CBRs were then compared to the BRs measured for L vanegatus in the present study The approach described by McCarty et al (1992a,b) (eqs (6) and (7)) was adapted for use with the BSAF (analogous to BCF) and the SEC for PCBs (analogous to LC50) by the following relationship

CBR = BSAF,.SEC (8)

Site-specific CBRs were calculated on a mol/Kg wet weight basis for Sites 011, 019, 428 and 031 by accounting for the hpid contents of the exposed L vanegatus and the TOC of the sediments Therefore, equation (8) was modified to

1 (Q}r"DD — V I Ctr^-DOAC ••"•~ — •!•"" ___L^_,- V^ 7 / ' UbK - > bbU'BbAhtV ^-r^.^*. ....'P,

where / = the /th congener, 7 = total number of congeners (mono-decachlormated biphenyl, 10), BSAFt = 1 7 (goc/gipd, the thermodynamically-based theoretical maximum), SEC = mg total PCBs/Kg dry, frac Lipid = hpid fraction in tissues, frac TOC = fraction of total organic carbon in the sediments, MW = molecular weight of the /th

congener (g/mol) and p, = the proportion of the /th congener represented in the total measured PCB BR BSAF = 1 7 was chosen because this site-specific CBR calculation was intended to serve as an estimate of lethal tissue dose at equilibrium Three SECs (mean ± 1 SD) have been derived for total PCBs, including the threshold effects concentration (TEC, 0 040 ± 0 054 mg/Kg dry), the midrange effects concentration (MEC, 0 40 ± 0 33 mg/Kg dry) and the extreme effects concentration (EEC, 1 7 ± 2 0 mg/Kg dry) (MacDonald et al 2000) The mean SECs were used in the calculations

Review Comment: There is a lot of material presented in Table 23 but only a small portion of it is discussed here It is not clear how the Tables are supposed to be used to support the statements in the following paragraph

The TEC-, MEC-, and EEC-based estimates of CBR (Table 23) were not exceeded at the reference site (011) However, all three site-specific CBR estimates were exceeded when compared to the measured BRs for L vanegatus at Sites 019, 428 and 031


(Table 22). Therefore, it appears that potentially toxic levels of PCBs were accumulated in L. variegatus. However, these values are lower than published acute CBRs from laboratory studies (Fisher et al. 1999; Borgman etal. 1990; Wezel et al. 1995; Hwang era/. 2001; Nebekerand Puglisi 1974) (Table 24).

All CBR literature values to date have been derived from water-only laboratory studies, usually with only one spiked chemical contaminant. This does not account for the same uptake exposure type unique to sediments, nor do they account for any additivity of mixtures inherent to real life environmental conditions. It is hypothesized that body burdens are additive and the mode of action for narcotic compounds is known to vary (polar vs. nonpolar) with respect to the type of hydrogen bonding with in the membrane as well as distribution within different types of lipids in an animal. No systematic studies have been published relating narcosis to body burdens of organisms exposed to mixtures of chemicals (van Wezel et al. 1995).


4.0 RESULTS OF TASK 3: TOXICITY IDENTIFICATION EVALUATION

Results for all of the Phase I TIE tests are summarized in Tables 25 and 26.

Review Comment: In addition to the Tables, the complete results of the Phase I TIE (i.e., responses in each treatment and control) need to be included as an appendix to allow independent evaluation of the data. These summary tables do not provide results for all the treatments performed under the Phase I TIE. The raw data and complete results (i.e., LC50 and TU values for each treatment) need to be provided by WSU.

4.1 Day 1

Pore water samples from Sites 031, 389, 019 and 398 were received in good condition. The sample container for Site 428 was broken during shipment and was reported to R.F. Weston and Soil Technologies, Inc. (STI). All custody seals were intact. Sample temperatures ranged from 8 to 11 °C. Samples from Sites 031 and 389 were turbid, colored and exhibited an oily sheen and odor. Samples from Sites 019 and 398 were relatively clear and without color, oil or odor.

4.1.1 Initial Toxicity Test Initial 24 h toxicity test results indicated that only the pore water samples from Sites 031 and 389 were acutely toxic to C. dubia. The LCso values calculated following exposure to a pore water dilution series (dilution series = control, or no pore water, 6.25, 12.5, 25, 50 and 100% pore water) were 29.9% and 30.7%, respectively (Table 25). Review Comment: Need to confirm that these are the 24-h LC50 values for the initial toxicity tests. The 48-h LC50s should also be included here and in the Table for comparison to the Day 2 and Day 3 baseline test results. Need data from WSU to address this.

4.2 Day 2

Day 2 test manipulations were conducted on August 16, 1999.

4.2.1 Baseline Test The Day 2 baseline LC50 values were 26.8 and 25.6% and TUs were 3.7 and 3.9 for Site 031 and 389, respectively (Table 25). These results indicated that there was very little change in toxicity of the original samples by Day 2 of the TIE.

4.2.2 Oxidant Reduction Addition Test The LC50 values for samples 031 and 389 were lower (TUs higher by +1 - 2) than the corresponding Day 2 baseline test results. Survival increased in Sample 031 over decreasing sodium thiosulfate concentrations and Sample 389 survivals were all 0% as sodium thiosulfate concentrations decreased (Table 25). The DMW control yielded 100% survival. Although they are not ordinarily considered as controls in the USEPA TIE protocol, treatment controls were conducted to determine if the thiosulfate additions caused toxicity. No toxicity was observed in the thiosulfate treatment control.


Review Comment: Need to explain what the "treatment controls'' were for the oxidant reduction test - was this sodium thiosulphate added to DMW? In the last sentence, the baseline test showed increased toxicity on Day 3 but not on Day 2

4.2.3 EDTA Chelation Addition In the sample from Site 031, an unknown, golden yellow precipitate was observed in the 100% and 60% dilutions. The 30% dilution was yellow colored with a yellow film on the cup bottom, and the pore water in the 15% dilution was yellow. In Sample 389, the pore water in the 60% dilution was yellow, the 30% dilution was slightly colored, and the 15% dilution was slightly cloudy.

Review Comment: Need to explain why a different dilution series used for the EDTA chelation test (34,1,3-4) The statement about LC50 values for Sample 389 being lower than the baseline cannot be evaluated because there are no LC50 values for these treatments in the Table.

The LC50 values were lower for Sample 031 showing increased toxicity as EDTA addition concentrations increased. The lowest EDTA concentration yielded the same LC50 value and TU as the Day 2 baseline test (Table 25). All Sample 389 LC50 values were lower than the Day 2 baseline test. These results indicate that toxicity increased overall in this test. No toxicity was observed in the EDTA controls.

4.2.4 pH Adjusted Filtration Sample 031 turned golden yellow and a brown, cloudy precipitate was formed in pH 3 during manipulation while adding HCI. Organisms in this Sample also showed a significant increase in survival in pH 3 and pH 11 adjusted samples. The pHi (6.75) Sample LC50 value was one-half that of the initial and Day 2 baseline test and percent survival in the 100% sample was 0%, indicating increased toxicity with filtration. Sample 389 test survival increased to 100% in all (pH 3, pH, and pH 11). In this sample, the pH, (6.80) filtered survival was higher than both the initial toxicity and baseline tests indicating that toxicity was decreased by the manipulation treatment (Table 25). Overall, toxicity decreased considerably in the pH 3 and pH 11 filtration in both 031 and 389 compared to the initial and Day 2 baseline tests. In pH, samples, the 031 sample toxicity increased after filtration but in the 389 sample, toxicity decreased. The graduated pH treatment was not conducted due to lack of sample.

4.3 Day 3

Day 3 test manipulations were conducted on August 17, 1999.

4.3.1 Baseline Test The Day 3 baseline LC50 values for Samples 031 and 389 were both lower than the initial toxicity test and the Day 2 baseline LC50 values. These results indicate that the original pore water samples became more toxic over the 3-d testing period. All dilutions


had golden colored water with a gradation in intensity from highest (100%) to lowest (12 5%) sample concentration Review Comment: The change in toxicity between the Day 2 and Day 3 baseline tests could be significant, this warrants discussion (i e discuss whether there is a chance that the pore water for these two samples (being extracted using a different procedure from the other samples) could have any effect on toxicity)

4.3.2 pH Adjusted/da Solid Phase Extraction (SPE) A golden yellow precipitate was observed in the pH 3 adjusted sample in 100%, 50% and 25% dilutions The 48-h test results showed significant decreases in toxicity in both Samples 031 and 389 The pH 3 adjusted samples from 031 and 389 had survivals of 80% and 100%, respectively For pH,, samples from Sites 031 and 389 survivals were 60% and 100%, respectively, and pH 11 adjusted samples for 031 and 389 were 0% and 80%, respectively (Table 25) This indicates a high overall decrease in toxicity in the initial and pH adjusted samples (031 and 389) in all but the Sample 031 pH 11 treatment Review Comment: Here and elsewhere confirm whether the pH was increased to pH 9orpH11 for the C78 SPE treatments this is not reported consistently

The results of this manipulation indicate that the Ci8 filtration reduced the toxicity of the original samples The pH adjustment/C18 test toxicity controls were filtered dilution waters and one of each of the three pH adjustments (pH 3, pHh pH 9) These blanks indicated artifact toxicity likely as a result of toxics leached from the C18 column in the pH 3 treatment Toxicity was observed in only the pH 3 dilution blank for both 031 and 389, whereas a low level of toxicity occurred in the pH 3 treatment for 031 and none occurred in the pH 3 treatment for 389

4.3.3 pH Adjusted/Aeration In Sample 031, survival decreased as pH increased Survivals for pH 3, pH,, and pH 11 were all higher than the Day 3 baseline toxicity tests Sample 389 survival results were variable Toxicity was slightly reduced overall in this test This could indicate volatile, surfactant or oxidizable toxic materials were present in lower concentrations The DMW controls (no pH manipulation) were performed instead of the manipulated pH blanks recommended in the USEPA methods No toxicity was observed in the blanks

Review Comment: We require LC50 values reported here or in Tables to support the statements about Sample 031 Also toxicity was more than slightly reduced in one treatment Also, need to provide an explanation for why the recommended pH blanks were not included for these tests

4.4 Task 3: Discussion and Conclusions

When samples are manipulated during the TIE procedures the pore water aliquots should not become more toxic as was observed, however, in some treatments Artifacts that may occur and contribute to toxicity are excessive ion strength from acid/base addition during pH adjustment, formation of toxic products by addition of acids/bases,


inadequate mixing of the solutions, contaminants leached from filters, pH probes, SPE columns, and reagents added and their contaminants (USEPA 1991a,b).

By themselves, none of the manipulation treatments provide conclusive results. However, acute toxicity was clearly removed by the two pH adjusted/filtration tests. The test results of these previously described samples indicate that the pH adjusted/filtration and pH adjusted/Ci8 SPE filtration tests significantly decreased toxicity. The TUs from the initial toxicity test and the Day 2 and Day 3 baseline tests were decreased substantially. Filterable compounds can include non-polar organics, such as PCBs and some metals. Colloids in these samples could transport non-polar organics and would be filterable. Higher survival in the filtration test was most likely due to organic colloids in the samples being filtered out and/or pH toxicity alteration of organics (the hydrolysis rate of organic compounds can be pH dependent and organics may degrade) (Table 26). Review Comment: It would help for interpretation if all the LCSOs and TUs for these treatments were reported in the Tables.

The pH adjusted filtration test (filterable compounds) decreased TUs by +3 points in the pH manipulated 031 sample, indicating a significant decrease in toxicity. Only the 031 sample result was not improved by the filtration test. This result could be attributed to artifact toxicity and/or as this sample was not pH manipulated, the organic compounds were not degraded. Review Comment: The statements in the paragraph require review. Only one TU was reported for this treatment, and it increased. The fact that the pH, treatments (i.e., no pH adjustment) had increased toxicity is also a concern for possible artifacts unrelated to the contaminants of concern.

The pH adjusted/Cie SPE filtration test also implicates non-polar organics, pesticides and some metals, if present. The pH adjusted C18 SPE filtration test TUs in samples 031 and 389 were decreased from +6 to 7 points when compared to the Day 3 baseline test results. As is evidenced by the comparison of Day 2 and Day 3 baseline tests, by Day 3 toxicity had increased by almost 5 TUs in Sample 031 and by +3 TUs in Sample 389. The C-|8 SPE test conducted on Day 3 removed up to 100% of the toxicity in both samples.

Review Comment: Only one TU was reported to support these statements; they should all be provided in Table 26.

Metal ions form bridging compounds that will flocculate at higher pH and can be filtered out. At higher pH, negatively charged particles (e.g., colloids) will bind with metals and form larger molecules that are removed by filtration (Manahan 1994). Increasing pH may remove toxicity by causing metal(s) to precipitate and become filterable or unavailable. Speciation of compounds may or may not affect toxicity but changes in solution pH (test manipulations) may affect toxicity of any of the compounds tested.


Returning the manipulated sample pH to its initial pH may still alter its toxicity These pH adjustments can cause a reduction, loss, or increase in toxicity (USEPA 1991a,b) (Table 26)

The confounding results of the oxidant reduction and EDTA chelation tests do not strongly support metals as a cause of acute toxicity. However, several metals of concern were detected in the sediment samples and extracted pore waters used for the TIE tests. The levels of some of these metals exceed recommended SQGs, in sediments, and WQC, in pore waters (Table 27). In addition, some levels of total PAHs were detected in Site 389 and Site 031 for sediment (7.65 and 4 11 mg/Kg dry, respectively) and pore water (15.1 and 19.0 ug/L, respectively) These sediment levels were below recommended SQGs for total PAHs (Table 27) Detection of total PCBs in sediments (Site 031 =210 mg/Kg dry and Site 389 = 93 mg/Kg dry) and pore water (Site 031 = 100 ug/L and Site 389 = 180 ug/L) far exceeded both SQGs and WQC (Table 27). Therefore, the general conclusion of these test results indicate that toxicity was decreased when non-polar organics were removed. A more comprehensive Phase II or Phase III TIE would be required to definitively identify compounds that were associated with altered biological responses Review Comment: It is not clear where the above-mentioned sediment and pore water PCB data come from. The sediment data do not match summary data from Tables 6 or 15 for laboratory and in situ tests If these data originate from a separate set of analyses, then there should be a table containing all those data Unless the concentration units are wrong, the measured pore water PCB concentration for Site 031 is 1000 times higher than what was estimated in Tables 20 and 21 If this is correct how does this impact the estimated CBR for L vanegatus which was derived from those estimates?


5.0 PROJECT SUMMARY

The H azteca and C tentans laboratory life-cycle assessment tests showed lethal to sublethal toxicity existed in the study area to both test species The strong and consistent relationships between PCB sediment contamination and species responses suggest PCBs were a major stressor Based on whole sediment PCB concentrations, adverse biological effects to H azteca and C tentans were in the low parts per million ranges However, there was significant sediment heterogeneity of PCBs and TOC at the test sites, which results in some uncertainty in the establishment of concentration-response relationships It is likely that the dynamics of the river and sediment environments result in varying degrees of contaminant exposure over time Downstream sites are more likely to have a wider range of contaminant levels The TIE and in situ studies supported the laboratory toxicity test conclusions, indicating that PCBs are a dominant stressor in the Housatonic River

The results of the 7-d bioaccumulation study provide evidence that PCBs in the study area can be accumulated in the tissues of aquatic organisms to toxic levels capable of causing mortality The tissue concentrations of total PCBs measured in Lumbnculus vanegatus were above SQG-based predictions of CBRs at three of the four contaminated sites The 48-h and 10-d in situ exposures did not lead to toxicity in L vanegatus (not used for 10-d test), however, toxicity was observed for the more sensitive Hyalella azteca, Daphnia magna and Chironomus tentans (only in the 10-d test) It is likely that H azteca and C tentans had accumulated BR levels of total PCBs that were similar to L vanegatus or to the estimated SQG-based CBRs This accumulation would be possible because the lipid levels of daphnids (zooplankton, 1 3% by wet weight, Kucklick et al 1996, D magna, 7 2% by dry wt, Cauchie et at 1999), chironomids (Chironomus tentans 3rd and 4th mstar larvae, 0 82 - 1 08% by wet weight, West et al 1997) and amphipods (Acanthogammarus sp , 1 6% by wet weight Kucklick et al 1996, H azteca, 1 8% by wet weight, Lotufo et al, 1999) are within the same range as lipids in the L vanegatus used in the present study (0 97 - 1 72%) CBRs are expected to be similar across species (McCarty and Mackay 1993 Fisher et al 1999), however, L vanegatus are tolerant to organic contaminants when compared to other aquatic species (Schuytema ef a/ 1990, Phipps et al 1993) Therefore, it is possible that lethal levels of total PCBs, similar to the estimated CBRs and measured BRs, were accumulated by H azteca, C tentans and D magna, whereas these same levels elicited only sublethal effects (e g, behavioral, physiological) in L vanegatus during the short (7-d) period of exposure Review Comment: Note that there are no data on measurement of these sublethal effects in the 7-d test nor survival over the full test duration so we do not have empirical assessment of these effects endpomts

Chemical contaminants other than PCBs, such as PAHs, metals and chlorinated benzenes, were detected in samples of sediment and surface water from the study area (Figures 21 and 22) Even though the results of this project did not establish a relationship between sediment or surface water concentrations of other contaminants (PAHs, metals or chlorinated benzenes) and tissue residues or organism responses,


other contaminants were observed are at potential levels of concern based on screening to conservative criteria. Several of the PAHs, metals and chlorinated benzene detections were above SQGs or WQC. Therefore, it is possible that these chemicals may also have contributed to sediment toxicity. However, the frequency and magnitude of exceedances for these other chemicals was much lower than for PCBs in sampled media.

Review Comment: The data on other contaminants measured in the sediments should be presented in a table. Figures 21 and 22 are not as detailed as would be preferable to assess potential for confounding effects of other COPECs.

Total PAH exceedances in sediment samples from the 7-d bioaccumulation test were observed for Sites 011 and 019. However, survival of the test organisms was high at Sites 011 and 019 even though the total PAH threshold effects concentration (TEC, 290 mg/kg OC; Swartz 1999) was exceeded. In samples of sediments and pore water from Sites 389 and 031 used for TIE testing, both the sediment extreme effects concentration (EEC, 10,0000 mg/Kg OC; Swartz 1999) and acute WQC (2.0 jag/L) were exceeded. Thus, it is possible that PAH levels at Sites 389 and 031 contributed to the toxicity observed in the 48-h and 10-d in situ exposures. Review Comment: This paragraph compares data from different sampling events for different sites (i.e., data from the 7-d test for Sites 011 and 019, and data from the TIE for Sites 389 and 031). Given the variability of the PCB data, this type of comparison carries considerable uncertainty that should be acknowledged.

Metals exceedances in sediment and pore water samples were noted at some downstream sites. No SQGs for metals in the 7-d bioaccumulation sediment samples were exceeded for Sites 011 and 389, and only mercury and lead exceeded an SQG at reference Site 398 and Site 019, respectively. However, sediment levels at Sites 428 and 031 exceeded consensus TECs for 6 and 5 metals, respectively. Chemical analysis of the TIE samples resulted in exceedances of 7 SQGs and 6 WQC for site 389, and 8 SQGs and 4 WQC for Site 031. Therefore, it is possible that metals also contributed to the toxicity observed in the 48-h and 10-d in situ exposures. However, there were no statistically significant correlations between toxicity responses and metal concentrations.

The chlorinated benzenes, 1,2,4-trichlorobenzene (Site 389 only) and 1,4dichlorobenzene ( Sites 398, 428, 389 and 031), were detected in sediment samples from the 7-d bioaccumulation study. The level of 1,2,4-trichlorobenzene at Site 389 (0.16 mg/Kg dry) was above the reported threshold (0.092 mg/Kg dry), acute (0.091 mg/Kg dry) and chronic (0.0091 mg/Kg dry) effects SQGs for this contaminant (NYDEC 1994; USEPA 1996). 1,4-dichlorobenzene at Sites 398 (0.041 mg/Kg dry) and 389 (0.11 mg/Kg dry) was above reported threshold (0.0035 mg/Kg dry), acute (0.012 mg/Kg dry) and chronic (0.0012 mg/Kg dry) effects levels (NYDEC 1994; USEPA 1996). TOC-normalized chlorinated benzene concentrations were calculated for 1,4dichlorobenzene at Sites 428 (2.13 mg/Kg OC) and 031 (1.41 mg/Kg OC). These concentrations were above the site-specific threshold, acute and chronic effects levels for Sites 428 (0.068, 0.23 and 0.023 mg/Kg OC, respectively) and 031 (0.081, 0.28 and


0.028 mg/Kg OC, respectively) (NYDEC 1994; USEPA 1996). A probabilistic model of chlorobenzene sediment effects concentrations predicts that 1,4-dichlorobenzene at Sites 428 and 031 would have been toxic at the 20th - 30th and the 10th - 20th percentiles of effects distributions, respectively (Fuchsman et al. 1999). Thus, it is possible that chlorinated benzenes contributed to the observed toxicity if it is assumed that their levels in the sediments during the 48-h and 10-d in situ toxicity exposures were similar to the concentrations measured in the 7-d bioaccumulation test samples. However, the variability associated with PCB concentrations suggests that such assumptions should be made with caution.

It has been established in the laboratory that PCBs are acutely toxic to aquatic invertebrate species through narcosis and that the lethal CBR concentrations range from 0.1 - 1.9 mmol/kg wet weight (Table 24). However, in the field tests conducted in the study area, H. azteca, C. tentans and D. magna mortality was observed and estimated CBRs derived from recent sediment quality guidelines were in the (imol/kg range. Thus, it appears that within the complex mixture of sediment-associated chemicals detected in the present in situ study, PCBs may not have been acting by only narcosis to cause the observed mortality endpoint. Landrum et al. (1989) observed similar results with amphipods in a study of mixtures of narcotic chlorinated hydrocarbons, and non-narcotic chemicals. It is unknown whether other suspect contaminants in the study sediments (e.g., PAHs, metals, chlorinated benzenes) acted additively, antagonistically, or synergistically with the PCBs. When xenobiotics exist in contaminated environments, they are often in mixtures with several other chemical classes, so the effective concentrations of individual compounds are difficult to determine (Burton 1991). The BRs of PCBs accumulated in situ were below laboratory-derived values for acute lethality of single PCB congeners or Aroclor® mixtures. Because the PCBs existed in a complex mixture of other chemicals in the study area, it was not unexpected that deleterious effects in aquatic organisms occurred despite lower than acute PCB BR levels.

The weight of evidence supports the conclusion that PCBs are causing toxicity in the study area. In addition to the results of the laboratory and in situ toxicity and bioaccumulation studies, that used a number of different species, the sediment and water chemistry data supported this conclusion. Measured sediment and water concentrations of total PCBs at the study sites exceeded recently published consensus-based sediment effects concentrations. Because these are empirically based, conservative single chemical guidelines, they merely suggest that effects will not be observed at lower levels. However, when predictions from SQGs are combined with the biological data from in situ and laboratory toxicity, bioaccumulation testing and TIE Phase I exposures, a strong case is presented that implicates PCBs as a dominant causative agent in the observed mortality. This study has demonstrated that the PCBs in the study area are present at levels that are sufficient to cause toxicity to aquatic organisms.


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Chappie, D J and G A Burton, Jr 1997 Optimization of in situ bioassays with Hyalella azteca and Chironomus tentans Environ Toxicol Chem 16559-564

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Connell, D W , M Bowman and D W Hawker 1988 Bioconcentration of chlorinated hydrocarbons from sediments by ohgochaetes Ecotoxicol Environ Saf 16293302

Di Toro, D M , J D Mahoney, D J Hansen, K J Scott, M B Hicks, S M Mayr and M S Redmond 1990 Toxicity of cadmium in sediments The role of acid volatile sulfide Environ Toxicol Chem 9 1487-1502

Di Toro, D M , C S Zarba, D J Hansen, W J Berry, R C Swartz, C E Cowan, S P Pavlou, H E Allen, N A Thomas and P R Paqum 1991 Technical basis for establishing sediment quality criteria for non-ionic organic chemicals using equilibrium partitioning Environ Toxicol Chem 10(12)1541-1583

Ferraro, S , H Lee II, R Ozretich, D T Specht 1990a Predicting bioaccumulation potential A test of a fugacity-based model approach Arch of Environ Contam Toxicol 19386-394

Ferraro, S , H Lee II, L Smith, R Ozretich, D T Specht 1990b Accumulation factors for eleven polychlorinated biphenyl congeners Bull Environ Contam Toxicol 46 276-283

Fisher, S W , S W Chordas III and P F Landrum 1999 Lethal and sublethal body residues for PCB intoxication in the oligochaete, Lumbnculus vanegatus Aquatic Toxicol 45 115-126

Fuchsman, P C , D J Duda and T R Barber 1999 A model to predict threshold concentrations for toxic effects of chlorinated benzenes in sediment Environ Toxicol Chem 19(9)2100-2109

Gabric, A J , D W Connell and P R F Bell 1990 A kinetic model for bioconcentration of lipophilic compounds by ohgochaetes Wat Res 24(10) 1225-1231

Hawker, D W and D W Connell 1988 Influence of partition coefficient of lipophilic compounds on bioconcentration kinetics with fish Wat Res 22(6) 701-707

Hwang, H , S W Fisher and P F Landrum 2001 Identifying body residues of HCBP associated with 10-d mortality and partial life cycle effects in the midge, Chironomus npanus Aquatic Toxicology 52(3-4)251-267


Keilty, T. J., White, D. S., Landrum, P. F. 1988a. Sublethal Responses to Endrin in Sediment By Limnodrilus- Hoffmeisteri (Tubificidae), and in Mixed-Culture With Stylodrilus-Heringianus (Lumbriculidae). Aquatic Toxicology. 13(3):227-250.

Keilty, T. J., White, D. S., Landrum, P. F. 1988b. Sublethal Responses to Endrin in Sediment By Stylodrilus- Heringianus (Lumbriculidae) As Measured By a 137 Cesium Marker Layer Technique. Aquatic Toxicol. 13(3):251-270.

Kucklick, J. R., Harvey, H. R., Ostrom, P. H., Ostrom, N. E. and Baker, J. E. 1996. Organochlorine dynamics in the pelagic food web of Lake Baikal. Environ. Toxicol. Chem. 15(8): 1388-1400.

Landrum, P. F., et at. 1989. Bioavailability and toxicity of a mixture of sediment-associated chlorinated hydrocarbons to the amphipod, Pontoporeia hoyi. Pp. 315-329 in Aquatic Toxicology and Hazard Assessment: 12th Volume. ASTM STP 1027. U. M. Cowgill and L. R. Williams (eds.). American Society for Testing and Materials, Philadelphia, PA, USA.

Lee, H., II. 1992. Models, muddles, and mud: Predicting bioaccumulation of sediment-associated pollutants. Pp. 267-293 in Sediment Toxicity Assessment, G. A. Burton, Jr. (ed.). Lewis Publishers, Inc. Chelsea, Ml.

Lotufo, G. R., P. F. Landrum, M. L. Gedeon, E. A. Tigue and L. R. Herche. 1999. Comparative toxicity and toxicokinetics of DDT and its major metabolites in freshwater amphipods. Environ. Toxicol. Chem. 19(2):368-379.

Manahan, S.E. Environmental Chemistry, 6th Edition. CRC Press, Boca Raton, FL. 1994.

MacDonald, D. D., L. M. DiPinto, J. Field, C. G. Ingersoll, E. R. Long and R. C. Swartz. 2000. Development and evaluation of consensus-based sediment effect concentrations for polychlorinated biphenyls. Environ. Toxicol. Chem. 19(5):4031413.

MacDonald, D. D., C. G. Ingersoll and T. A. Berger. 2000. Development and evaluation of consensus-based sediment quality guidelines for freshwater ecosystems. Arch. Environ. Contam. Toxicol. 39(1):20-31.

McCarty, L. S., Mackay, D., Smith, A. D., Ozburn, G. W. and Dixon, D. G. 1992a. Residue-Based Interpretation of Toxicity and Bioconcentration QSARs From Aquatic Bioassays - Neutral Narcotic Organics. Environ. Toxicol. Chem. 11(7):917-930.

McCarty, L. S., Ozburn, G. W., Smith, A. D., and Dixon, D. G. 1992b. Toxicokinetic modeling of mixtures of organic chemicals. Environ. Toxicol. Chem. 11 (7): 10371047.

McCarty, L. S. and Mackay, D. 1993. Enhancing Ecotoxicological Modeling and Assessment. Environ. Sci. Technol. 27:1719-1728.

McElroy, A. E. and J. C. Means. 1988. Factors effecting the bioavailability of hexachlorobiphenyls to benthic organisms. Pp. 149-158 in Aquatic Toxicology and Hazard Assessment: 10th Vol., ASTM STP 971. W. J. Adams, G. A.


Chapman and W. G. Landis (eds.). American Society for Testing and Materials. Philadelphia, PA, USA.

McFarland, V. A. and J. U. Clarke. 1986. Testing bioavailability of polychlorinated biphenyls using a two-level approach. Pp. 220-229 in Water Quality R & D: Successful Bridging between Theory and Application. R. G. Wiley (ed.). Hydrologic Engineering Research Center. Davis, CA, USA.

Nebeker, A.V. and F.A. Puglisi. 1974. Effect of polychlorinated biphenyls (PCBs) on survival and reproduction of Daphnia, Gammarus, and Tanytarsus. Trans. Amer. Fish. Soc. 103:722-728. NY State Department of Environmental Conservation (NYDEC). 1994. Technical guidance for screening contaminated sediments. Division of Fish and Wildlife and Division of Marine Resources, Albany, NY, USA.

Phipps, G. L, G. T. Ankley, D. A. Benoit and V. R. Mattson. 1993. Use of the aquatic oligochaete Lumbriculus variegatus for assessing the toxicity and bioaccumulation of sediment-associated contaminants. Environ. Toxicol. Chem. 12:269-279.

Rubenstein, N. I., J. L. Lake, R. J. Pruell, H. Lee II, B. Taplin, J. Heltshe, R. Bowen, S. Pavignano. 1987. Predicting bioaccumulation of sediment-associated organic contaminants: Development of a regulatory tool for dredged material evaluation. U.S. EPA 600/X-87/368.

Schuytema, G. S., D. F. Krawczyk, W. L. Griffis, A. V. Nebeker, M. L. Robideaux. 1990. Hexachlorobenzene uptake by fathead minnows and macroinvertebrates in recirculating sediment/water systems. Arch. Envrion. Contam. Toxicol. 19:1-9.

Stumm, W. and J.J. Morgan. 1981. Aquatic Chemistry - An Introduction Emphasizing Chemical Equilibria in Natural Waters. John Wiley & Sons, Inc., New York, NY.

Swartz, R. C., G.R. Ditsworth, D.W. Schults and J.O. Lamberson. 1985. Sediment toxicity to a marine infaunal amphipod: Cadmium and its interaction with sewage sludge. Mar. Environ. f?es.18:133-153.

Swartz, R. C., D.W. Schults, T.H. DeWitt, G.R. Ditsworht and J.O. Lamberson. 1990. Toxicity of fluoranthene in sediment to marine amphipods: A test of the equilibrium partitioning approach to sediment quality criteria. Environ. Toxicol. Chem. 9:1071-1080.

Swartz, R. C. 1999. Consensus sediment quality guidelines for polycyclic aromatic hydrocarbon mixtures. Environ. Toxicol. Chem. 18(4):780-787.

USEPA. 1986. Quality criteria for water 1986. EPA 440/5-86-001. Office of Water. Washington, DC. May 1, 1986.

USEPA. 1987. Update #2 to Quality criteria for water 1986. Office of Water. Washington, DC. May 1, 1987.

USEPA. 1991 a. Methods for aquatic toxicity identification evaluations. Phase 1, Toxicity characterization procedures 2nd edition. EPA/600/6-91/003. Office of Research and Development, Washington, DC.


USEPA. 1991b. Sediment toxicity identification evaluations: Phase 1 (characterization), phase 2 (identification), and phase 3 (confirmation). Modification of effluent procedures. EPA 600/6-91-007. Environmental Research Laboratory, Duluth, MN. Tech Report No. 08-91.

USEPA. 1993. Methods for measuring the acute toxicity of effluents and receiving waters to freshwater and marine organisms, 4th edition. EPA/600/4-90/027F. Environmental Monitoring and Support Laboratory. US Environmental Protection Agency, Cincinnati, OH. August 1993. USEPA. 1994. Procedures for Assessing the Toxicity and Bioaccumulation of Sediment-Associated Contaminants with Freshwater Invertebrates. EPA 600/R-94/024. U.S. Environmental Protection Agency, Duluth, MN,

USEPA. 1996. Eco Update: Ecotox thresholds. EPA 540/F-95/083. Office of Solid Waste and Emergency Response. Washington, DC.

USEPA. 1998. Methods for measuring the toxicity and bioaccumulation of sediment-associated contaminants with freshwater invertebrates. Duluth, MN, Draft.

USEPA. 2000a. Appendix to Bioaccumulation Testing and Interpretation for the Purpose of Sediment Quality Assessment. Status and Needs. Chemical Specific Summary Tables. EPA-823-R-00-002. Office of Water (4305) and Office of Solid Waste (5307W). Washington, DC. Feb 2000. 816 pp.

USEPA. 2000b. Methods for measuring the toxicity and bioaccumulation of sediment-associated contaminants with freshwater invertebrates. Second edition. EPA/600/R-99/064. Office of Water, Washington, DC. March 2000. 192 PP-

Van Wezel, A. P., D. A. M. de Vries, S. Kostense, D. T. H. M. Sijm and A. Opporhuizen. 1995. Intraspecies variation in lethal body burdens of narcotic compounds. Aquatic Toxicology. 33:325-342.

Veith, G. D., D. L. DeFoe and B. V. Bergstedt. 1979. Measuring and estimating the bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36:10401048.

West, C. W., Ankley, G. T., Nichols, J. W., Elonen, G. E. and Nessa, D. E. 1997. Toxicity and bioaccumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin in long-term tests with the freshwater benthic invertebrates Chironomus tentans and Lumbriculus variegatus. Environ. Toxicol. Chem. 16(6): 1287-1294.


Figure 1: Locations of Wright State University toxicity sampling stations (011, 398, 019, 428, 389, 031; marked with asterisks) in relation to benthic invertebrate sampling locations.

River Mile: N/ H9-SE000615

NOTE: WSU Stations use last three digits of Weston IDs shown above (e.g.. H3-SEEC0031 is the

Weston ID for WSU Station 031.

NOTE: Station 023 (not shown) was located just upstream of 031, as was used for the laboratory

portion of the testing only.

* = Wright State University sediment toxicity sampling station.


Review Comments for All Tables/Appendices:

(Table 13) There is quite a difference between the dry weight and AFDW values for the three control treatments (i.e.. relative changes in dry weight were not similar). Some comment on this is warranted.

(Table 15) As stated elsewhere, need to confirm whether the 7-d TOO value for Site 389 (4%) is a valid result. Table 15 should also include the PCS water data for the 48-h, 7-d and 10-d sampling events, even if the 10-d Hyalella test was the only one sensitive to the water exposure.

(Table 18) Need to clearly identify the source of these data According to Table 15, sediment PCBs were undetected at Site 011, so these data must come from the congener-specific analyses

(Table 19) The equation in the footnote specifies pore water concentration as "mg/L" but this table heading uses "ng/L", confirm that this is correct Is it appropriate to report the estimated concentrations to so many decimal places'?

(Table 21) To match Table 23, the tissue residue data should be listed for all 6 sites, even if the calculations cannot be completed for all 6 sites

(Table 22) The source of the value of 0.0071 mg/kg needs to be identified, since elsewhere it states that total PCS was non-detected for this station Again, confirm whether or not the TOO value of 4% is correct for Site 389

(Table 25) All of the LC50 values and TUs from each test should be reported here If survival was 100% in undiluted sample, then the LC50 will be >100% and the TU will be 1. Confirm whether the LCSOs for the initial toxicity tests are for 24 h or 48 h, both should be provided so that the 48-h results can be compared for all tests

(Table 26) Filtration also increased toxicity for pH, treatment for Site 031 and decreased toxicity for pH 3 treatment for Site 389. Where do the 'percent removal" values come from for the C18 SPE treatment, and what do they mean?

(Appendix 6) Temperature ranged from 16 4 - 28 3'C in the 7-d and 10-d in situ tests What are the implications for such a large temperature variation?


Figure 2: Survival of Hyalella azteca in chronic laboratory toxicity test, at three time periods (28-d, 35-d, 42-d). Value in italics represents total PCB concentration in mg/kg.

Test acceptability Control Survival >80%

Trout 011 398 019 389 023 031 Farm

Station Locations (Arranged North to South)

• 28-day (24 June 99) D35-day (1 July 99) D42-day (8 July 99)

Figure 3: Hyalella azteca 28-day dry weights in chronic laboratory toxicity test. Value in italics represents total PCB concentration in mg/kg.

- 8.7 NA

ND ND

°'6 i

#: ;s O)o.4

£ ? 0 . 2 re 0) jst Acceptability: 5 ontrol>0 15mg

0.0

Initial Trout 011 398 019 031 Farm


71

Figure 4: Hyalella azteca 42-day dry weights in chronic laboratory toxicity test. Value in italics represents total PCB concentration in mg/kg.

213 (no data _| due to zero

survival)

0.0 ± Trout 011 398 019 389 023 031 Farm


Figure 5: Reproduction of Hyalella azteca in chronic laboratory toxicity test, based on mean number of neonates (totals and numbers per female). Value in italics represents total PCB concentration in mg/kg.

o CO +

re o 0)

0)JO E 213 (no data

due to zero 31

C survival)re

Trout 011 398 019 389 023 031 Farm


128-42 day mean neonates per female D 28-42 day mean neonates (total)

Figure 6: Survival of Chironomus tentans in chronic laboratory toxicity test (20-day). Value shown in italics represents total PCB concentration in mg/kg.

Survival >70%

Trout a-cellu- Floris- 011 398 019 389 023 031 Farm lose sant


Figure 7: Dry weights for Chironomus tentans in chronic laboratory toxicity test (20-day). Value shown in italics represents total PCB concentration in mg/kg.

_ 5.0 Test Acceptability

NAV) Control >0 6 mgi 4.0

"55 NA NA

f3.0

0) T I 2.0

•o c 1.0 ra 213 (no 31 (noo>

8.7 survivors) survivors) ^^^^M0.0

Trout a-cellu- Floris- 011 398 010199 389 023 Farm lose sant

Station Locations(ArrangedNorth to South)

Figure 8: Ash-free dry weights for Chironomus tentans in chronic laboratory toxicity test (20-day). Value shown in italics represents total PCB concentration in mg/kg.

3 n Test acceptability Control 5"

>0 48 mg w

S NA

-§- ND '

O) ' 0> NA

T£. 1.0 is ,•D f ND m

C I re 0) ,-l'tf .1 _L 1• 2t3fno 31 (no 5 IS 71;« / ^ 1I 8.7 survivors) survivors)

O n - ^™ ••



Figure 9: Emergence of Chironomus tentans in chronic laboratory toxicity test (20-day). Value shown in italics represents total PCB concentration in mg/kg.

NA

100 Q Test acceptability Control

NA Emerg >50% ±, 80 T NA ND

'—' Rn • •&o> ou ."w*"^ 1 c

^ ., if ND u 0)

0) E ¥*" iu 20 'AS,

".'I- t• I1• 31 • ? m 87 213 T 71 .

1 • I1 ^ ^0 * •



Figure 10: Survival of Hyatella azteca in 48-h low flow in situ toxicity tests conducted 14-16 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

120

V) 100

5 80

5 60+ 'E co 40

20 -I

0 Lab 011 398 019 428 389 031

Control


D Water Column I Against Sediment

Figure 11: Survival of Daphnia magna in 48-h low flow in situ toxicity tests conducted 14-16 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

120 NA ND 0.4

139.2 527 7o CO 100

± TT

80

re 60 I D 40 CO

ra o> 20 -| 5

0 Lab 011 398 019 428 389 031

Control



Figure 12: Survival of Lumbriculus variegatus in 48-h low flow in situ toxicity tests conducted 14-16 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

Q (O

5 'E 3 W c n

Lab 011 398 019 428 389 031 Control


D Water Column • Against Sediment

Figure 13: Survival of Chironomus tentans in 48-h low flow in situ toxicity tests conducted 14-16 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

a V)

«J

I 3

CO C re 0)

Lab 011 398 019 428 389 Control


D Water Column • Against Sediment

Figure 14: Daphnia magna survival versus total sediment PCB concentrations in 48-hour low flow in situ (sediment) exposures.

•i nn j1 UU <

on • .oU J

Note: Non-detected

60 0.1 mg/kg

/in4U

on£.(}

n _ ^ ^

0.1 1 10 100 1000

Total PCBs (mg/kg)

Figure 15: Survival of Hyaletla azteca in 10-d low flow in situ toxicity tests conducted 17-27 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

120

Mea

n S

urv

ival (

%)

Lab 011 398 019 428 389 031 Control


D Water Column lAgainst Sediment

Figure 16: Survival of Chironomus tentans in 10-d low flow in situ toxicity tests conducted 17-27 June 1999. Value shown in italics represents total PCB concentration in mg/kg.

o CO

TO

I CO C TO 0)

Lab 011 398 019 428 389 031 Control



Figure 17: Hyalella azteca survival versus total sediment PCB concentrations in 10-day low flow in situ (sediment) exposures.

•i nn 1 UU

> .s* ou

1 on

Note: Non-detected fc finou PCB replaced by 3 CO • 0.1 mg/kg C /in4U V^

on£\J

m •

C) 20 40 60 80 100 1Z 10

Total PCBs (mg/kg)

• •

Figure 18: Chironomus tentans survival versus total sediment PCB concentrations in 10-day low flow in situ (sediment) exposures.

^ nn1UU ( . •

—». on !£ •

1

"re Note: Non-detected > or* « bU •

0.1 mg/kg 3

CO r- /in .c 'tu re o

on2U

•0 .1

C) 20 40 60 80 100 12 JO

Total PCBs (mg/kg)

Figure 19: Hyalella azteca survival versus total water column PCB concentrations in 10-day low flow in situ (water only) exposures.

•\ r\r\1UU

• on

^ 80 0^

~m • 5> /?n • 1 CO >- /in C »HJ re o> 5

ZUon .

n (3 50 100 150 200 250 300 3! 50

Total PCBs (mg/kg)

Figure 20: Lumbriculus variegatus PCB tissue burdens versus total sediment PCB concentrations in 7-day in situ bioaccumulation tests.

[Note: Stations 398 and 011 not shown due to either non-detected OC or PCB in sediment sample]

cnnOUU

</) CO(j Ann 4UU

428

Q. 0, £-s 3 TJ CA "«" CL inn . oUU

• 031

H1 o> •o ^ • 019 ., g*~ E

2UU

OB ~~" • 389

i—o -i nn iuu z

_0 n

C) 100 200 300 400 5C )0

Normalized Sediment PCBs (mg/kg OC)

Figure 21: Sediment-associated metals concentrations in 7-day in situ bioaccumulation exposures (sum of Cd, Cr, Cu, Hg, Ni, Pb, Ag, Zn). Concentrations in mg/kg dry weight.

700

600

500

400

300

200

100

0

398 011 019 428 389 031

Site

Figure 22: Sediment-associated PAH and SVOC concentrations in 7-day in situ bioaccumulation exposures. Concentrations in mg/kg dry weight.

300 -]

250

200

150

100

50 , ,

_,...._,.„.... 0

398 011 019 428 389 031

Site

Table 1: Test conditions for conducting

Parameter

Test type

Temperature

Light quality

Illuminance

Photopenod

Test chamber

Sediment volume

Overlying water volume

Renewal of overlying water

Age of organism

Number of organisms

Number of replicate chambers/treatment

Feeding

Aeration

Overlying water

Test chamber cleaning

Overlying water quality

Test duration

Endpomts

Test acceptability

a 42-d sediment toxicity test with Hyalella azteca

Conditions

Whole-sediment toxicity test with renewal of overlying water

23 ± 1°C

Wide-spectrum fluorescent lights

500-1000 lux

16 h light 8 h dark

300 ml high-form hpless beaker

100mL

175 mL in sediment exposure from Day 0-28 (175-275 ml in the water-only exposure from Day 28-42)

2 volume additions per 24 hours continuous or intermittent (e g , one volume addition every 12 hours)

7-8 days old at the start of the test

10 per replicate

12 (4 for 28-d survival and growth and 8 for 35-d and 42-d survival growth and reproduction) Reproduction is more variable than growth or survival, hence more replicates might be needed to establish statistical differences among

1 0 mL YCT (1800 mg/L stock) daily to each test beaker

None, unless dissolved oxygen in overlying water drops below

Culture water, well water, surface water or site water (use of reconstituted water is not recommended)

If screens clog during test gently brush from outside beaker

Initial and final hardness, alkalinity conductivity and total ammonia ( at Day 0 and 28) Daily temperature and weekly conductivity DO and pH 3 times/week (DO should be measured more often if it drops 1 mg/L below last

42 days

28-d survival and growth, 35-d survival and reproduction, and 42-d survival, growth and reproduction

Minimum mean control survival of 80% on Day 28 At Day 28 the minimum control mean dry weight must be at least 0 15 mg

Table 2: General activity schedule for conducting a 42-d sediment toxicity test with Hyalella azteca

Day Activity

Pre-Test -8 Separate known-age amphipods from the cultures and place in holding chambers

Begin preparing food for the test The <24-h amphipods are fed 10 mL of YCT (1800 mg/L stock solution) and 10 ml of Selanastrum capncornutum (about 3 0 x 10E7 cells/mL) on the first day of isolation and 5 mL of both YCT and S capncornutum on the

-7 Remove adults and isolate <24-h-old amphipods (if procedures outlined in Section 10 3 4 of EPA [2000] are followed)

-6 to -2 Feed and observe isolated amphipods monitor water quality (e g temperature and dissolved oxygen)

-1 Feed and observe isolated amphipods monitor water quality Add sediment into each test chamber place chambers into exposure system and start renewing overlying water

Sediment Test 0 Measure total water quality (pH temperature dissolved oxygen hardness alkalinity

conductivity ammonia) Transfer ten 7- to 8-d-old amphipods into each test chamber Release organisms under the surface of the water Add 1 0 ml of YCT (1800 mg/L stock) into each test chamber Archive 20 organisms for length determination or archive 80 test organisms for dry weight determination Observe behavior of test organisms

1 to 27 Add 1 0 mL of YCT to each test beaker Measure temperature daily conductivity weekly and dissolved oxygen (DO) and pH three times per week Observe behavior of test organisms

28 Measure temperature dissolved oxygen pH hardness alkalinity conductivity and ammonia End the sediment-exposure portion of the test by collecting the amphipods with a #40 mesh sieve (425-micron mesh U S standard sieve size) Use four replicates for growth measurements count survivors and preserve organisms in sugar formalin for growth measurements Use eight replicates for reproduction measurements place survivors in individual replicate water-only beakers and add 1 0 mL of YCT to each test beaker/d and 2 volume additions/d of overlying water

Reproduction Phase 29 to 35 Feed daily (1 0 mL of YCT) Measure temperature daily conductivity weekly and DO

and pH three times a week Measure hardness and alkalinity weekly Observe behavior of test organisms

35 Record the number of surviving adults and remove offspring Return adults to their original individual beakers and add food

36 to 41 Feed daily (1 0 mL of YCT) Measure temperature daily conductivity weekly and DO and pH three times a week Measure hardness and alkalinity weekly Observe behavior of test organisms

41 Measure total water quality (pH temperature dissolved oxygen hardness alkalinity conductivity ammonia)

42 Record the number of surviving adults and offspring Surviving adult amphipods on Day 42 are preserved in sugar formalin solution The number of adult males in each beaker is determined from this archived sample This information is used to calculate the number of young produced per female per replicate from Day 28 to Day 42

Table 3: Test conditions for conducting a long-term sediment toxicity test with Chironomus tentans.

Parameter

Test type

Temperature

Light quality

Illuminance

Photopenod

Test chamber

Sediment volume

Overlying water volume

Renewal of overlying water

Age of organism

Number of organisms

Number of replicate chambers/treatment

Feeding

Aeration

Overlying water

Test chamber cleaning

Overlying water quality

Test duration

Endpomts

Test acceptability

Conditions

Whole-sediment toxicity test with renewal of overlymq water

23 ± 1°C

Wide-spectrum fluorescent lights

500-1000 lux

16 h light 8 h dark

300-mL high-form hpless beaker

100 mL

175 ml

2 volume additions/d, continuous or intermittent (e g one volume addition every/12h)

<24-h-old larvae

10 per replicate chamber - Note that USEPA (2000) advocates 12

8 used in this study - Note that USEPA (2000) advocates 16 (12 at Day -1 and 4 for auxiliary males on Day 10)

1 5 ml Tetrafm* goldfish food (6 0 mg dry solid) daily to each test beaker

None, unless dissolved oxygen in overlying water drops below 2 5 mq/L

Culture water, well water, surface water or site water (use of reconstituted water is not recommended)

If screens cloq durmq test qently brush from outside beaker

Initial and final hardness, alkalinity, conductivity and total ammonia (at Day 0 and Day 20) Daily temperature and weekly conductivity DO and pH 3 times/week (DO should be measured more often if it drops 1 mg/L below last measurement)

About 40 - 50 days Each treatment is ended separately when no additional emergence has been recorded for seven consecutive days When no emergence is recorded from a treatment, termination of that treatment should be based on the control sediment using this 7-d criterion

20-d survival and growth, female and male emergence and adult mortality

Minimum average size of C tentans in the control sediment at 20 d must be at least 0 6 mg/surviving organism as dry weight or 0 48 mg/survivmg organism as ash-free dry weight Emergence should be >50% Time to death after emergence is < 6 5 d for males and < 5 1 d for females

Table 4: General activity schedule for conducting a long-term sediment toxicity test with Chironomus tentans

Day Activity

-4 Start reproduction flask with cultured adults (1 3 male female ratio) For example for 15 to 25 egg cases 10 males and 30 females are typically collected Egg cases typically range from 600 to 1500

-3 Collect egg cases (a minimum of 6 to 8) and incubate at 23 °C

-2 Check egg cases for viability and development

-1 1 Check egg cases for hatch and development 2 Add 100 mL of homogenized test sediment to each replicate beaker and place in corresponding treatment holding tank After sediment has settled for at least 1 h add 1 5 mL Tetrafin slurry (4g/L solution) to each beaker Overlying water renewal begins at this time

0 1 Transfer all egg cases to a crystallizing dish containing control water Discard larvae that have already left the egg cases in the incubation dishes Add 1 5 ml food to each test beaker with sediment before the larvae are added Add 10 larvae to each replicate beaker (beakers are chosen by random block assignment) Let beakers sit (outside the test system) for 1 h following addition of the larvae After this period gently immerse all beakers into their respective treatment holding tanks

2 Measure temperature pH hardness alkalinity dissolved oxygen conductivity and ammonia at start of test

1-End On a daily basis add 1 5 mL food to each beaker Measure temperature daily Measure the pH and dissolved oxygen three times a week during the test If the DO has declined more than 1 mg/L since previous reading increase frequency of DO measurements and aerate if DO continues to be less than 2 5 mg/L Measure hardness alkalinity conductivity ammonia weekly

6 For auxiliary male production start reproduction flask with culture adults (e g 10 males and 30 females 1 3 male to female ratio)

7-10 Set up schedule for auxiliary male beakers (4 replicates/treatment) same as that described above for Day - 3 to Day 0

19 In preparation for weight determinations ash weigh-pans at 550 °C for 2 h Note that the weigh boats should be ashed before use to eliminate weighing errors due to the pan oxidizing during ashing

20 Randomly select four replicates from each treatment and sieve the sediment to recover larvae for growth and survival determinations Pool all living larvae per replicate and dry the sample to a constant weight (e g 60 °C for 24 h) Install emergence traps on each of the remianmg reproductive replicate beakers Measure temperature pH hardness alkalinity dissolved oxygen conductivity

21 The sample with dried larvae is brought to room temperature in a dessicator and weighed to the nearest 0 01 mg The dried larvae in the pan are then ashed at 550°C for 2 h The pan with the ashed larvae is then reweighed and the tissue mass of the larvae determined as the difference between the weight of the dried larvae plus pan and the weight of the ashed larvae plus pan

23-End Record daily emergence of males and females pupal and adult mortality and time to death for previously collected adults

33-End Transfer males emerging from the auxiliary male replicates to individual inverted petri dishes The auxiliary males are used for mating with females from corresponding treatments from which most of the males had already emerged or in which no males emerged

40-End After 7 d of no recorded emergence in a given treatment end the treatment by sieving the sediment to recover larvae pupae or pupal exuviae When no emergence occurs in a test treatment that treatment can be ended once emergence in the control sediment has ended using the 7-d criterion

Table 5: TIE tests and controls

Day TIE Test Control Blank

August 15 (Day 1) Initial Toxicity DMW DMW

August 16 (Day 2) Baseline DMW DMW

Sodium Thiosulfate Baseline & DMW + Sodium Thiosulfate DMW

EDTA Baseline & DMW + EDTA DMW

pH/Filtration Baseline DMW

August 17 (Day 3) Baseline DMW DMW

pH/Aeration Baseline DMW

pH/C18SPE DMW + pH adjust DMW

Note DMW - deminerahzed water (USEPA 2000b), EDTA-ethylene-diamme-tetra-acetate SPE-Solid Phase Extraction

Table 6: Sediment chemical measurements for laboratory life-cycle assessments: Total PCBs, % TOC and normalized PCB levels, for the test site.

Total PCB Levels (n=1) WSU Site No. River Mile mg/Kg TOC (%) mg/Kg OC

011 N/A ND ND ND 398 N/A ND 4 04 ND 019 132.34 8 7 0 3 2900 0 389 126.38 2130 40 53250

023a Upstream of 12565 310 5 5 5636 031 12565 710 52 13654

aSite 023 was used for the laboratory exposures and site 428 was used for in situ exposures

Table 7: H. azteca 28-d survival following exposure to six Housatonic River sediments and one control sediment (27 May - 24 June 1999).

Treatment Replicate

Trout Farm 4 (Control) 6

11 9 8 7 12 10 2 3 5

011 9 10 11 1 2 5 12 6 8 7 4 3

398 8 3 10 1 5 7 9 2 11 4 12 6

019 9 8 11 7 6 2 10 12 5 1 3

28d Survival (%)

Per Replicate Mean St Dev

100 81 8 1471 80 100 70 70 80 90 80 90 50 90

90 658 1379 80 60 70 60 60 70 50 70 40 80 60

80 833 1073 90 80 70 90 90 90 80 60 90 100 80

1697 20 30 50 50 70 40 60 30 70 50

70 483

4 40

Table 7 (cont'd): H. azteca 28-d survival following exposure to six Housatonic River sediments and one control sediment (27 May - 24 June 1999).

Treatment Replicate

389 11 12 9 10 2 3 5 1 4 8 7 6

023 12 9 11 10 1 5 2 6 8 3 7 4

031 2 8 4 3 6 11 12 9 1 7 10 5

Per Replicate

00 0 0 0 0 0 0 0 0 0 0

2020 40 20 20 10 40 60 30 20 10 20

2030 0 10 50 20 50 10 10 40 10 20

28d Survival (%)

Mean St Dev

00 000

258 1443

225 1658

- - -

Table 8: H. azteca 35-d (1 July 1999) and 42-d (8 July 1999) survival and reproduction (total number of young from Day 35-42) following exposure to six Housatonic River sediments and one control sediment (sediment 389 also tested, but yielded 0% survival following 28 days)

35-d Survival (%) 42-d Survival (%) # Young Females/ Young/ Mean # Young

Treatment Replicated Replicate Mean St Dev Replicate Mean St Oev 35d 42d Total Replicate Female Treatment St Dev

Trout 2 80 813 991 60 775 139 30 26 56 3 187 134 45 Farm 11 100 100 23 27 50 5 100 (Control) 4 80 80 45 20 65 4 163

12 90 90 22 29 51 4 128 6 70 60 15 21 36 3 120 5 80 80 30 42 72 5 144 7 80 80 33 38 71 4 178 10 70 70 5 5 10 2 50

011 12 70 688 146 70 675 167 21 2 23 5 46 77 75 1 70 70 2 4 6 3 20 9 90 90 5 4 9 2 45 10 80 80 0 2 2 4 05 3 50 50 18 7 25 2 125 8 50 40 21 51 72 3 240 4 80 80 16 11 27 5 54 5 60 60 13 11 24 3 80

398 10 70 800 120 70 76 3 141 34 9 43 4 108 90 24 12 80 70 19 16 35 3 11 7 6 60 50 9 25 34 3 11 3 7 100 100 29 4 33 4 8 3 5 80 80 22 20 42 4 105 9 80 80 11 16 27 5 54 4 90 80 20 3 23 3 77 3 80 80 18 7 25 4 63

019 1 70 500 185 70 500 185 11 6 17 4 4 3 63 30 2 60 60 32 17 49 4 123 3 20 20 0 0 0 0 6 50 50 0 10 10 2 50 7 40 40 8 0 8 1 80 9 70 70 10 9 19 3 63 11 30 30 8 0 8 2 40 12 60 60 4 0 4 1 40

023 9 10 238 130 10 243 140 0 0 0 1 00 27 23 1 10 10 0 0 0 0 4 20 20 1 0 1 1 1 0 6 50 50 8 4 12 4 30 2 30 30 0 - 0 - 11 30 30 2 10 12 2 60 3 20 20 1 6 7 2 35 10 20 - 0 0 0 0

031 8 30 286 168 30 27 1 160 0 0 0 3 00 01 03 6 50 50 0 0 0 3 00 7 40 40 0 2 2 3 07 11 20 20 0 0 0 2 00 5 0 0 0 0 2 20 20 0 0 0 1 00 12 40 30 0 0 0 0

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Table 10: H. azteca dry weights following a 42-d exposure (27 May - 8 July 1999) to five Housatonic River sediments and one control sediment (sediment 389 was also tested but yielded 0% survival, and no dry weight data exist for this treatment).

Treatment

Initial

Trout Farm

011

398

019

023

Replicate

1 2 3

2 4 5 6 7 10 11 12

1 3 4 5 8 9 10 12

3 4 5 6 7

9 10 12

1 2 3 6 7 9 11 12

1 2 3 4 6 9 11

2 6 7 8 11 12

Animals Per Pan

8 9 10

4 8 8 6 7 6 10 9

6 5 7 6 4 9 7 8

8 8 7 4 10 8 7 6

6 7 2 5 4 7 3 6

1 3 2 2 6 1 3

2 4 4 3 2 3

Mean H. azteca Dry Wt. (mg)Per Individual Per Treatment

00100 0010800133 00090

07600 0561705362 05150 06666 05528 04833 04360 05433

04350 0 454205020 04371 04650 05425 04444 04328 03750

04975 0531004500 04928 06150 05130 05537 05228 06033

03766 0 530404885 07950 04480 06050 04085 06600 04616

09100 0 524503866 03750 05550 03350 05100 06000

03900 0 396403425 03225 02800 05600 04833

Standard Deviation

000

010

0 05

006

0 14

0 20

0 1 1 031

Table 11: C. tentans survival following a 20d exposure to six composite Housatonic River sediments and three control sediments (10-30 July 1999).

Treatment

Trout Farm (Control)

a-Cellulose (Control)

Florissant (Control)

011

398

019

389

023

031

Survival (%) Replicate Mean

90 85 100 90 60

50 725 80 60 100

60 725 80 70 80

60 525 40 40 70

80 77 5 70 60 100

0 5 0

20 0

0 0 0 0 0

0 0 0 0 0

20 7 5 0 10 0

Standard Deviation

173

222

96

150

17 1

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Table 14: C. tentans emergence following exposure to six Housatonic River sediments and one reference sediment, 10 July -16 August 1999 (all emergence took place between 1-13 August 1999)

Number Treatment/Site Replicate Emerged

a-Cellulose 1 8 (Control) 2 4

3 7 4 6

Florissant 1 3 (Control) 2 6

3 6 4 1

Trout Farm 1 3 (Control) 2 10

3 4 4 10

398 1 7 2 4 3 0 4 3

011 1 3 2 2 3 3 4 4

019 1 0 2 0 3 0 4 0

023 1 0 2 0 3 1 4 0

031 1 0 2 0 3 0 4 0

389 1 0 2 0 3 0 4 0

Date of Last Emergence

9-Aug 7-Aug 6-Aug 7-Aug

1 3-Aug 9-Aug 11-Aug 1 3-Aug


11-Aug 12-Aug

-8-Aug


_ ---

_ -

4-Aug -

_ ---

. ---

MeanEmergence (%)

62 5

Standard Deviation

1 71

40 0 2 45

67 5 37 7

350 289

300 8 2

0 0

2 5 5 0

0 0

00

Table 15: Chemical measurement for in situ test site sediments total PCBs, % TOG and normalized PCB levels (where available) for the test site

Total PCB Levels (n=1)

48 hb 7 d 10 db

Wright State ID River Mile mg/Kg mg/Kg TOC (%) mg/Kg OC mg/Kg

011 N/A ND ND 032 ND ND 398 N/A 0 4 5 4 ND N/A 01 019 132.34 0 9 0 7 1 75 385 139 428a 13032 7 3 170 5 16 330 1 4 389 126.38 1392 71 4 0 175 522 031 125.65 521 7 169 431 392 111 9

aSite 428 was used for in situ exposures in place of site 023 that was used for laboratory tests bNo TOC values associated with 48-h and 10-d sediment chemical analysis

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Table 18: Values of octanol-water partitioning coefficient (Kow), organic carbon-water partitioning coefficent

(KoC) and sediment-pore water partitioning coefficient (Kp) used to calculate pore water concentrations using

equilibrium partitioning theory3

Kp atstations11

PCB# Common Name log Kowb log Kocc 011 019 428 031

5 2,3-DiCB 500 492 259 1440 4246 3546 9 2,5-DiCB 510 501 325 1806 5324 4447 14 3,5-DiCB 530 521 511 2839 8372 6993 15 4,4-DiCB 533 524 547 3039 8960 7484

Mean DiCB 5 18 509 411 2287 6725 5678 StDev 0 16 016 140 779 2296 7978

18 2,2',5-TnCB 524 5 15 446 2479 7309 6105 28 2,4,4'-TnCB 560 551 1008 5599 16509 13790 31 2,5,4'-TnCB 560 551 1008 5599 16509 13790

Mean TnCB 548 539 827 4559 73443 77228 StDev 021 020 324 7802 5372 4437

40 2,2',3,3'-TetraCB 590 580 1987 11041 32556 27193 47 2,2',4,4'-TetraCB 6 10 600 3125 17363 51195 42762 52 2,2',5,5'-TetraCB 584 574 1735 9639 28422 23740 53 2,2,5,6'-TetraCB 590 580 1987 11041 32556 27193 66 2,3',4,4'-TetraCB 590 580 1987 11041 32556 27193 77 3,3',4 4'-TetraCB 6 10 600 3125 17363 51195 42762

Mean TetraCB 596 586 2325 72975 38080 37807 StDev 0 11 0 11 628 3488 70284 8590

101 2,2',4,5,5'-PentaCB 638 627 5890 32723 96487 80593 105 2,21,3',4,41-PentaCB 665 654 10853 60293 177779 148493 118 2,3',4,4',5-PentaCB 674 663 13305 73916 217946 182044 126 3,3',4,4',5-PentaCB 653 641 8178 45435 133969 111901

Mean PentaCB 657 646 9557 53092 756545 730758 StDev 0 16 0 15 3278 77879 52778 44034

128 2,21I3,3',4,4'-HexaCB 674 663 13305 73916 217946 182044 138 2,2',3,44',5'-HexaCB 683 671 16311 90617 267190 223176 151 2,2',3,5,5',6-HexaCB 664 653 10610 58944 173800 145170 153 2,2',414',5,5'-HexaCB 692 680 19996 111091 327559 273601 155 2,2',4,41,6,6'-HexaCB 640 629 6163 34239 100955 84325 156 2,3,3',4,4',5-HexaCB 700 688 23966 133143 392581 327912 169 3,3',4,4',5,5-HexaCB 740 727 59264 329244 970800 810881

Mean HexaCB 685 673 27374 778742 350779 292444 StDev 031 037 77772 98399 290136 242343

170 2,2',3,3',4,4',5-HeptaCB 727 7 15 44157 245317 723335 604181 180 2,2',3,4,4',5,5'-HeptaCB 736 724 54134 300745 886767 740691

Mean HeptaCB 732 7 19 49746 273037 805051 672436 StDev 006 006 7055 39793 115564 96527

194 2,2',3,3'4,41,5,51-OctaCB 780 767 146552 814176 2400655 2005198

NanoCB6 799 785 246457 1369205 4037199 3372157

209 2,2',3,31,4,4',5,5'6,6'-DecaCB 8 18 804 346362 1924235 5673744 4739115

'Equilibrium Partition equation for sediments Cpw = C^KP where Cpw is cone in pore water (mg/L) Csis cone in sediments (mg/Kg) and Kpis frac

OC'Koc (L/Kg)

"«<» values from Boese et al , 1997 ET&C 16(&) 1545-1553 Gabnc et al , 1990 Wat Res 24(10)1225-1231 Hawker and Connell 1988 Wat Res 22(6)701-707, Verth et al , 1979 J Fish Res Board Can 36 1040-1048, Fisher et al , 1999 Aquatic Toxicol 45 115-126, USEPA, 2000

clog K* = 0 00028 + 0 983*(k>g «„,) dKp not calculable for sites 398 or 389 due to no available sediment TOC

'Estimated as mean of OctaCB and DecaCB values

Table 19: Estimated pore water, freely dissolved, PCB concentrations (Cpw) for sediments from the 7-d in situ bioaccumulation test (Cpw calculated using equilibrium

a b \ partitioning equations )

Cpw(ng/L)

PCB Congener 011 019 428 031

MonoCBc

DiCB TnCB TetraCB PentaCB HexaCB HeptaCB OctaCB NanoCB DecaCB

0 0

00433 00308 00761 0 151

00584 000135

0 0

0 0

0296 1 84 1 38 1 47 1 21

00700 0 00789 000125

0 1 80 550 232 178 189 7 12

0 352 00212

0 000640

00420 1 92 143 644 159 220 747

0524 00245

0 000534

Total 03606 6275 7465 1266

'Equilibrium Partition equation for sediments Cpw = CS/KP where Cpw is cone in pore water (mg/L) Cs is cone in sediments (mg/Kg) and Kp is frac OC*KOC (L/Kg)

"Sediment concentration used in calculations were from samples for the 7 d in situ bioaccumulation test cMonoCB Cpw were calculated using Kp values for DiCB (see Table 21)

Table 20: Pore water Bioconcentration Factors (BCF) for total PCBs in L vanegatus tissue following 7-d in situ bioaccumulation test

Pore Water Concentration"'6

Tissue Residues

BCFC

Log BCF

ug/L: Cpwd

ug/Kg wet: Cae

ug/Kg Lipid: C,f

wet wt. basis lipid basis

wet wt. basis lipid basis

011

00004

57

5295

158111 14685210

520 717

Site (Upstream to019

0 0063

3997

232634

637076 37075220

5 80 757

Downstream) 428 031

0 0747 0 1266

4409 3043

380626 314517

59060 24044 5098662 2485215

4 7 7 4.38 6 7 1 640

aPore water concentrations were estimated based on equilibrium partitioning theory using sediment data from this study and published Kow values bPore water concentrations not estimated for sites 398 or 389 due to no available sediment TOC CBCF = Ca/Cpw (wet wt basis) or C|/Cpw (lipid wt basis) dCpvv = concentration in the pore water eCa = concentration in the animal (wet weight basis) fC, = concentration in the lipid

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Table 23 Sediment Quality Guidelines (SQGs) or Water Quality Criteria (WQC) exceedances for samples from the 48-h 7-d and 13-d in situ tests and the

48-h and 10-d in situ tests

Guidelines/

Criteria Test Group Compound 398 011 019

Site

428 38S 031

SedimentQuality Guidelines

(SQGs)'

7-d PAHs

Metals

Total PAHsJrQa/lip QCJ.

Arsenic

TEC-MEC MEC EEC <TEC

Cadmium TEC-PEC

Mercury TEC-PEC1 <TEC' TEC-PEC TEC-PEC

Chromium TEC-PEC TEC-PEC

Copper TEC-PEC TEC-PEC

Lead >PEC TEC PEC TEC-PEC

Nickel TEC-PEC <TEC'

Zinc TEC PEC TEC-PEC

CBnz 1 4-DiCBnz >ECT >Nr-A >ECT >Nr-A »ECT >Nr--A >ECT >NY-A

1 24-TnCBnz >ECT >Nf-A

PCBs Total (mg/Kg) >EEC TEC-MEC 'EEC >EEC >EEC

48-h PCBs Total (mg/Kgi MEC-EEC MEC-EEC 'EEC >EEC >EEC

10-d PCBs Total Img/Kg) 'EEC MEC-EEC 'EEC >EEC

Water

Quality Criteria(WQC)"

48-h

7-d Surface Water 7-d Pore Water '

PCBs

PCBsPCBs

Total (ug/L)

Total (ug/L) Total (ug/L)

C-A

C-A

C-A

C-A C-A

C-A

C-A

C-A

C-A C-A

10-d PCBs Total (ug/L) C-A C A c-A C-A

*SQG definitKjm (Mac Donald*./ 2000 MacDonald**/ in review Swartz 1S99 USEPA 1996 NYDEC 199«l

TEC Threshold Effect Conrncentritton

MEC « Median EHect Concentraflon

EEC « Extreme Effect Concentration

PEC Probable Effect Concentration

ECT • Ecotox Thre*r>oldi

NY-A « New Y ortt Department at Envron Coniervation Acute tnd Chronic Sediment value* exceeded

Âpproaches or is very near noted SQG 'CBnr « chlorooated benzene*

"WQC d*fin*t»n« (USEPA 1986 1987)

C « Chrome CcHena

'pore water concentrations estimated using equihbrum partitioning (see Table 20)

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Table 26 Summary of results and conclusions for samples used for TIE test manipulations

Test

Initial Toxicity

Filtration

EDTA Addition

Oxidant Reduction

Aeration

C,8SPE

Site 031

Result

LCSO = 29.87%

decreased loxicity in pH 3 411

(pH 3 formed precipitate)

increased toxicity (precipitate w/addition)

increased loxicity

decreased toxicity at pH 3 only (pH 3 & 1 1 precipitated)

greatly reduced toxicity for all pH: pH = 90% removal

pH 3 = 75% removal pH 1 1 = 70% removal

(TIE Sample 771)

Conclusion

toxic sample

some toxicity related to filterable solids (neutral toxicity solubilization, pH )

possible artifact toxicity, metals toxicity unlikely

possible artifact toxicity, oxidizing toxicity unlikely

possible artifact toxicity, oxidizing, volatile, surfactant toxicity possible

very likely, organic toxicants present and causing toxicity

Site 389

Result

LC6C = 30 73%

decreased toxicity for pHand pH 11

increased toxicity

increased toxicity

decreased toxicity for pH 11 only

greatly reduced toxicity for all pH. pH = 100% removal

pH3= 100% removal pH 1 1 = 85% removal

(TIE Sample 772)

Conclusion

toxic sample

some toxicity related to filterable solids (acidic toxicity solubilization, pH 3)

possible artifact toxicity, metals toxicity unlikely

possible artifact toxicity, oxidizing toxicity unlikely

possible artifact toxicity, oxidizing, volatile, surfactant toxicity possible

very likely, organic toxicants present and causing toxicity

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Appendix 1: Weekly water chemistry measurements for 42-d Hyalella azteca life-cycle assessment exposure to two control sediments and six composite sediments

Temperature pH Conductivity DO Alkalinity Total NH3 Hardness Date Treatment (°C) (umhos/cm) (mg/L) (mg/L) (mg/L) (mg/L)

27-May-99 Trout Farm 224 8 15 400 76 164 2 180 210 Ottawa Sand 227 836 345 76 128 0436 175 398 230 826 320 74 120 0398 169 11 220 831 330 74 136 0252 173 19 21 7 821 325 7 5 108 0472 167 23 21 7 821 290 77 128 0838 156 31 224 8 15 285 7 5 104 1 230 139 389 225 813 325 7 2 128 2270 156

8-Jun-99 Trout Farm 229 799 360 73 160 0306 213 Ottawa Sand 226 8 11 310 7 5 140 0441 189 398 227 8 17 250 72 152 0308 193 11 234 829 350 7 3 140 0 184 191 19 237 8 15 360 72 136 0 161 185 23 23 1 824 350 7 4 136 0 130 177 31 232 8 12 310 7 0 132 0 128 177 389 232 809 300 73 136 0 113 173

15-Jun-99 Trout Farm 222 840 390 7 4 176 0 184 228 Ottawa Sand 224 830 310 74 132 0538 187 398 222 827 290 74 148 0224 195 11 228 839 310 76 132 0 184 187 19 229 834 310 7 5 128 1 020 175 23 228 834 300 7 8 128 0 111 175 31 228 8 14 300 7 1 116 0 158 167 389 227 823 300 7 0 124 0234 179

22-Jun-99 Trout Farm 21 6 839 390 61 172 0 169 227 Ottawa Sand 21 7 828 350 6 3 138 0360 187 398 21 5 828 350 6 2 148 0258 197 11 21 7 826 350 64 138 0308 187

19 22 1 824 330 62 134 0234 185 23 21 9 830 370 6 3 122 0313 182 31 222 824 300 60 124 0 127 170 389 22 1 829 330 6 1 136 0 180 183

24-Jun-99 Trout Farm 230 834 395 7 5 173 0 173 211 Ottawa Sand 230 8 10 325 7 4 120 0476 163 398 230 826 340 7 5 140 0449 175 11 240 829 315 7 5 128 0215 182

19 260 830 310 7 3 120 0202 187

23 240 826 330 7 3 92 0 123 175

31 230 821 305 7 2 124 0 110 175 389 230 809 300 7 0 124 0 134 163

Page 31 of 42

Weekly Chemistry continued

Temperature pH Conductivity DO Date Treatment (°C) (umhos/cm) (mg/L)

1-Jul-99 Trout Farm 225 821 330 7 5 Ottawa Sand* - - - -398 225 836 310 7 5 11 230 837 300 7 4 19 230 836 300 7 5 23 235 840 310 7 4 31 235 835 310 7 5 389* - - - -

8-Jul-99 Trout Farm 230 807 310 68 Ottawa Sand* - - - -

398 230 816 300 7 0 11 230 8 12 300 6 8 19 230 821 300 7 0 23 230 821 290 7 0 31 230 823 300 7 0 389* - - - -

'Ottawa Sand and sediment 389 not continued following June 25 due to low survival

Alkalinity (mg/L)

160 -

152 136 132 136 132

-

128 -

120 120 124 116 100

-

Total NH3

(mg/L)

1 370 -

0793 0706 0689 0925 0627

-

0012 -

0009 0008 0006 0006 0005

-

Hardness (mg/L)

172 -

184 168 172 172 168

-

177 -

165 165 161 165 157

-

Page 32 of 42

Appendix 2: Daily temperature and dissolved oxygen measurements for the H azteca life-cycle assessment of six Housatonic River and two control sediments (27 May - 8 July 1999)

May-99 Jun-99 Parameter Site 27 28 29 30 31 1 2

Temperature Trout Farm 224 232 222 245 235 21 5 21 9

(°C) Ottawa Sand 227 232 21 8 250 235 21 5 21 1 398 230 227 21 8 245 23 3 21 5 21 9 11 220 234 222 250 228 21 5 21 4 19 21 7 229 21 8 245 238 21 5 224 23 21 7 226 222 243 233 21 0 225 31 224 229 21 8 243 23 1 21 0 22 1 389 225 226 222 240 232 21 0 223

Dissolved Trout Farm 76 7 5 7 5 4 9 7 2 81 8 5 Oxygen (mg/L) Ottawa Sand 76 7 8 76 4 9 7 0 7 9 8 5

398 7 4 7 5 76 4 7 6 9 81 8 5 11 7 4 7 5 76 4 9 7 1 7 9 8 3 19 7 5 7 5 7 4 4 9 6 8 8 0 84 23 7 7 7 6 7 5 5 0 6 9 8 1 8 3 31 7 5 7 5 7 4 50 6 9 8 1 8 2 389 72 7 5 7 5 5 0 6 9 81 81

Jun-99 continued 3 4 5 6 7 8 9 10

Temperature Trout Farm 228 242 220 240 226 229 228 230

(°C) Ottawa Sand 229 242 223 242 226 226 229 230 398 226 242 22 1 24 1 225 227 226 235 11 23 1 246 222 247 235 234 222 240 19 234 246 222 242 232 237 23 1 240 23 227 240 223 238 233 23 1 229 235 31 225 240 21 7 242 23 1 232 230 240 389 229 240 21 7 242 233 232 229 235

Dissolved Trout Farm 8 3 7 9 70 8 0 76 7 3 7 4 -

Oxygen (mg/L) Ottawa Sand 8 3 7 7 7 2 8 0 7 7 7 5 7 5 -398 8 2 76 68 7 8 7 7 7 2 66 -11 8 3 7 7 69 8 0 7 4 7 3 7 4 -

19 8 0 7 5 66 8 0 7 5 7 2 7 0 -

23 8 3 72 66 8 0 74 7 4 7 4 -31 7 9 7 0 6 5 7 9 7 3 7 0 66 -

389 76 7 5 6 7 7 9 74 7 3 6 9 -

Page 33 of 42

DO/Temperature continued

Jun-99 continued 11 12 13 15 16 17 18 19

Temperature Trout Farm 240 222 232 222 235 22 1 205 21 0

(°C) Ottawa Sand 240 225 23 1 224 240 22 1 209 21 3 398 240 222 23 1 222 240 220 203 21 0 11 225 224 235 228 250 228 21 1 21 3 19 225 223 236 229 240 227 209 21 0 23 225 224 236 228 245 225 21 0 21 0 31 250 224 235 228 235 228 209 21 0 389 250 224 234 227 240 226 21 0 21 0

Dissolved Trout Farm 76 76 77 74 70 67 92 82 Oxygen (mg/L) Ottawa Sand 81 80 8 0 74 7 1 7 2 9 1 86

398 74 78 7 8 7 4 6 2 7 0 9 2 82 11 7 7 7 9 7 9 76 7 2 7 1 8 8 82 19 76 79 78 75 7 1 66 90 80 23 76 7 8 7 9 7 8 7 3 7 1 9 0 7 8 31 7 2 7 8 7 8 7 1 7 1 6 9 9 1 8 1 389 76 7 8 7 8 70 7 0 6 9 8 9 8 2

Jun-99 continued 20 21 22 23 24 25* 26 27

Temperature Trout Farm 240 220 21 6 220 230 23 240 245

(°C) Ottawa Sand 240 21 9 21 7 220 230 - - -398 240 22 1 21 5 220 230 23 238 245 11 240 220 21 7 225 240 24 245 250 19 243 227 22 1 225 260 24 250 250 23 243 225 21 9 230 240 24 245 250 31 240 22 1 222 225 230 24 245 250 389 240 222 22 1 225 230 - - -

Dissolved Trout Farm 80 77 6 1 7 0 7 5 77 77 78 Oxygen (mg/L) Ottawa Sand 80 77 6 3 76 7 4 - - -

398 80 7 8 6 2 6 9 7 5 7 7 7 8 76 11 80 76 64 72 75 74 76 75 19 80 7 1 6 2 7 0 7 3 7 5 77 7 4 23 7 5 76 6 3 66 7 3 7 5 7 7 7 5 31 8 0 7 7 60 6 9 7 2 7 5 7 7 7 5

389 78 7 5 61 - 70 - - -

Page 34 of 42

DO/Temperature continued

Jun-99 continued 28 29 30 1

Jul-99 2 3 4

Temperature (°C)

Trout Farm Ottawa Sand 398 11 19 23 31 389

230 -

230 230 240 240 240

-

230 -

225 240 240 240 240

-

220 -

21 0 220 220 220 220

-

22 5-

22 523023 0235235

-

22 0

22 0 225 22 0 230 230

23 0 230 23 0 240 230

24 0

24 0 240 24 0 245 245

Dissolved Oxygen (mg/L)


7 5 -

75 74 73 7 3 7 3 -

8 0 -

79 8 0 8 0 7 9 80 -

6 7 -

60 56 5 5 5 4 5 4

-

7 5.

75747 57 47 5-

7 7

78 75 7 7 7 7 7 6

7 3

75 72 7 4 7 4 7 3

7 4

76 74 7 5 7 5 7 4

5 Jul-99 6

cont. 7 8

Temperature (°C)


245 -

245 245 245 248 248

-

240 -

240 240 235 240 240

-

23 0-

23 023 023023 023 0

-

23 0

23 0 23 0 230 23 0 23 0

Dissolved Oxygen (mg/L)


8 5 -

8 4 8 3 104 89 93 -

7 7 -

7 8 7 8 7 8 7 8 7 8 -

7 5-

7 774767 57 5-

6 8

7 0 68 70 7 0 7 0

"Ottawa Sand and sediment 389 not continued following June 25 due to low survival

Page 35 of 42

Appendix 3: Weekly water chemistry measurements for 42d Chironomus tentans life-cycle assessment exposure to six Housatonic River Sediments and two control sediments (9 July - 20 August 1999).

Temperature pH Conductivity D.O. Alkalinity Date Sample (°C) (umhos/cm) (mg/L) (mg/L)

9-Jul-99 Trout Farm 21 8 778 340 390 136 Flornsant 21 2 770 330 760 112 a-Cellulose 21 2 823 350 8 10 166 398 21 5 789 350 670 154 11 21 5 825 300 805 128 19 21 5 778 300 665 116 23 21 8 765 310 660 116 31 21 6 745 290 630 102 389 220 768 340 630 132

16-Jul-99 Trout Farm 220 8 18 325 760 164 Flornsant 230 822 325 8 10 120 a-Cellulose 225 826 300 8 10 140 398 225 826 340 780 140 11 230 825 400 780 136 19 230 825 310 770 120 23 235 825 300 750 116 31 230 823 280 770 100 389 230 828 300 770 116

23-Jul-99 Trout Farm 230 823 350 675 160 Flornsant 230 819 330 670 136 a-Cellulose 230 823 340 700 148 398 230 826 360 722 148 11 240 826 360 750 140 19 240 823 345 741 136 23 240 828 320 735 128 31 240 826 310 733 120 389 240 831 330 726 136

30-Jul-99 Trout Farm 230 810 350 752 122 Flornsant 225 808 320 750 106 a-Cellulose 226 8 16 330 764 128 398 227 827 340 776 120 11 227 820 310 743 124 19 229 8 18 330 756 104 23 228 8 14 320 754 104 31 229 8 16 310 764 101 389 230 821 320 760 108

Total NH3

(mg/L)

0006 0000 0000 0001 0000 0000 0000 0000 0004

0350 0054 0056 0010 0039 0007 0009 0008 0008

3470 1 320 0794 0231 0 118 0251 0 117 0 105 0072

5 160 0628 0741 0427 0315 0 120 0078 0078 0084

Hardness (mg/L)

193 189 177 195 181 179 175 151 169

177 169 177 181 173 169 169 157 161

180 168 184 200 180 184 176 160 172

184 160 168 188 164 168 164 156 164

Page 36 of 42

Weekly chemistry continued

Temperature pH Conductivity DO Alkalinity Total NH3 Hardness Date Treatment (°C) (umhos/cm) (mg/L) (mg/L) (mg/L) (mg/L)

6-Aug-99 Trout Farm 230 791 310 682 108 0386 160 Flornsant 230 801 300 721 112 1 240 160 a-Cellulose 230 789 310 638 124 0959 168 398 230 801 310 7 19 120 0386 168 11 230 8 13 305 728 108 0290 160 19 230 806 295 720 104 0260 160 23 240 806 290 737 100 0 184 152 31 230 808 295 741 104 0 140 148 389 240 821 300 739 108 0 117 160

13-Aug-99 Trout Farm 230 827 325 746 120 0337 170 Flornsant 230 8 13 325 7 12 128 0382 166 a-Cellulose 230 807 350 684 144 0469 182 398 230 8 18 325 746 140 0670 174 11 230 8 18 310 767 108 1 020 162 19 230 823 325 736 116 0518 182 23 230 828 300 746 108 0347 150 31 230 828 300 725 108 0390 162 389 230 833 325 704 168 0378 162

20-Aug-99 Trout Farm 230 833 380 739 144 0 107 194 Flornsant 230 8 18 400 752 132 0489 194 a-Cellulose 230 819 400 690 176 0 140 234 398 225 833 390 740 140 0225 198 11 230 833 375 744 136 0 189 190 19 230 833 380 738 132 0 106 194 23 230 831 350 739 132 0075 182 31 230 827 310 722 124 0062 182 389 230 834 380 730 140 0048 198

Page 37 of 42

_ _

Appendix 4: Daily temperature and dissolved oxygen measurements for a 42d C tentans life-cycle assessment of six Housatonic River sediments and two control sediments (9 July - 19 August 1999)

Jul-99 Parameter Site 9 10 11 12 13 14 15 16

Temperature Trout Farm 21 8 220 21 0 21 0 220 220 220 220 (°C) Flornsant 21 2 220 208 208 220 220 220 230

a-Cellulose 21 2 220 21 0 21 0 220 220 21 5 225 398 21 5 225 220 220 220 23 0 220 225 11 21 5 225 223 223 220 230 220 230 19 21 5 230 225 225 230 230 230 230 23 21 8 230 225 225 230 230 230 235 31 21 6 230 230 230 230 230 230 230 389 220 230 230 230 230 23 0 230 230

Dissolved Trout Farm 390 770 750 750 750 7 10 740 760 Oxygen (mg/L) Flornsant 760 770 790 790 770 740 790 8 10

a-Cellulose 810 790 790 790 780 750 780 8 10 398 670 785 770 770 770 750 760 780 11 805 800 825 825 740 760 770 780 19 665 790 785 785 770 700 760 770 23 660 790 770 770 770 720 760 750 31 630 780 760 760 770 720 760 770 389 630 780 750 750 780 7 10 760 770

Jul-99 continued 17 18 19 20 21 22 23 24

Temperature Trout Farm 21 0 220 230 21 0 220 220 230 235

(°C) Flornsant 21 0 220 229 21 0 220 230 230 235 a-Cellulose 21 0 220 220 21 5 220 230 230 236 398 220 230 230 220 230 230 230 24 1 11 220 22 0 240 22 5 230 240 240 242 19 220 230 232 225 240 230 240 24 1 23 220 230 239 225 240 240 240 241 31 220 230 21 9 225 240 240 240 241 389 220 230 230 225 240 240 240 24 1

Dissolved Trout Farm 677 672 7 10 689 675 683 Oxygen (mg/L) Flornsant - - 750 736 793 7 15 670 694

a-Cellulose - - 776 723 762 710 700 674 398 - - 741 719 760 703 722 682 11 - - 730 748 769 734 750 693 19 - - 778 739 803 726 741 680 23 - - 759 776 763 733 735 690 31 - - 835 750 759 725 733 688 389 - - 783 741 755 732 726 701

Page 38 of 42

DO/Temperature continued Jul-99

25 26 27 28 29 30 31

Temperature Trout Farm 246 23 1 222 230 240 230 246 Flornsant 246 230 223 230 230 225 243 a-Cellulose 249 230 223 230 230 226 245 398 25 1 234 225 230 - 227 246 11 252 239 229 240 240 227 248 19 254 239 230 230 250 229 249 23 254 24 1 23 1 240 250 228 247 31 254 243 232 240 250 229 248 389 256 242 233 240 250 230 248

Dissolved Trout Farm 741 685 727 7 12 681 752 693 Oxygen (mg/L) Flornsant 701 757 750 758 724 750 646

a-Cellulose 7 11 743 723 735 7 13 764 7 16 398 718 736 736 752 - 776 688 11 731 756 747 783 7 15 743 702 19 707 739 737 754 704 756 7 17 23 735 746 736 749 732 754 721 31 727 720 736 751 724 764 729 389 739 741 735 768 720 760 732

Aug-99 1 2 3 4 5 6 7

Temperature Trout Farm 240 220 200 230 230 230 240 Flornsant 240 220 200 230 230 230 240 a-Cellulose 240 21 0 200 230 230 230 240 398 230 220 200 230 230 230 240 11 240 220 200 230 230 230 240 19 250 240 240 230 230 230 240 23 250 240 240 240 240 23 0 24 0 31 250 240 240 240 230 230 250 389 250 240 240 240 240 230 250

Dissolved Trout Farm 799 780 953 734 682 769 757 Oxygen (mg/L) Flornsant 828 430 836 209 721 111 749

a-Cellulose 785 859 929 657 638 1 13 725 398 829 890 948 735 7 19 790 748 11 838 880 953 745 728 8 10 760 19 792 7 11 843 733 720 800 762 23 800 705 840 737 737 792 760 31 794 285 845 628 741 800 756 389 787 722 848 743 739 796 756

8

240 240 240 240 240 240 240 240 240

702 693 635 689 724 701 693 691 693

Page 39 of 42

DO/Temperature continued Aug-99

9 10 11 12 13 14 15 16

Temperature Trout Farm 230 230 230 230 220 247 240 220 (°C) Flornsant 230 230 230 230 220 240 240 220

a-Cellulose 230 230 230 230 220 240 240 220 398 230 230 230 230 230 240 240 220 11 230 230 230 230 230 240 240 220 19 230 230 230 230 240 240 240 220 23 230 230 230 230 230 240 240 220 31 230 230 230 230 230 250 240 220 389 230 230 230 230 230 240 240 220

Dissolved Trout Farm 750 6 19 664 746 750 685 749 747 Oxygen (mg/L) Flornsant 727 523 633 7 12 750 730 751 750

a-Cellulose 685 546 623 684 7 17 650 7 11 705 398 743 604 673 746 720 7 13 750 750 11 797 615 670 767 640 7 15 758 748 19 811 649 685 736 700 7 14 748 754 23 799 633 687 746 640 7 12 750 753 31 794 647 678 725 730 703 747 761 389 8 13 644 686 704 730 695 744 760

Aug-99 continued 17 18 19

Temperature Trout Farm 220 230 230 (°C) Flornsant 220 230 230

a-Cellulose 220 230 230 398 220 225 225 11 220 230 230 19 220 230 230 23 220 230 230 31 230 230 230 389 220 230 230

Dissolved Trout Farm 748 760 739 Oxygen (mg/L) Flornsant 747 788 752

a-Cellulose 728 730 690 398 736 111 740 11 738 749 744 19 730 111 738 23 729 780 739 31 730 770 722 389 731 778 730

Page 40 of 42

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