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Keywords:

  • Coal;
  • Natural gas;
  • Sediment;
  • Toxicity;
  • Mussel;
  • Amphipod;
  • Midge

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Sediment toxicity tests were conducted to assess potential effects of contaminants associated with coal mining or natural gas extraction activities in the upper Tennessee River basin and eastern Cumberland River basin in the United States. Test species included two unionid mussels (rainbow mussel, Villosa iris, and wavy-rayed lampmussel, Lampsilis fasciola, 28-d exposures), and the commonly tested amphipod, Hyalella azteca (28-d exposure) and midge, Chironomus dilutus (10-d exposure). Sediments were collected from seven test sites with mussel communities classified as impacted and in proximity to coal mining or gas extraction activities, and from five reference sites with mussel communities classified as not impacted and no or limited coal mining or gas extraction activities. Additional samples were collected from six test sites potentially with high concentrations of polycyclic aromatic hydrocarbons (PAHs) and from a test site contaminated by a coal ash spill. Mean survival, length, or biomass of one or more test species was reduced in 10 of 14 test samples (71%) from impacted areas relative to the response of organisms in the five reference samples. A higher proportion of samples was classified as toxic to mussels (63% for rainbow mussels, 50% for wavy-rayed lampmussels) compared with amphipods (38%) or midge (38%). Concentrations of total recoverable metals and total PAHs in sediments did not exceed effects-based probable effect concentrations (PECs). However, the survival, length, or biomasses of the mussels were reduced significantly with increasing PEC quotients for metals and for total PAHs, or with increasing sum equilibrium-partitioning sediment benchmark toxic units for PAHs. The growth of the rainbow mussel also significantly decreased with increasing concentrations of a major anion (chloride) and major cations (calcium and magnesium) in sediment pore water. Results of the present study indicated that (1) the findings from laboratory tests were generally consistent with the field observations of impacts on mussel populations; (2) total recoverable metals, PAHs, or major ions, or all three in sediments might have contributed to the sediment toxicity; (3) the mussels were more sensitive to the contaminants in sediments than the commonly tested amphipod and midge; and (4) a sediment toxicity benchmark of 1.0 based on PECs may not be protective of mussels. Environ. Toxicol. Chem. 2013;32:207–221. © 2012 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Native freshwater mussels, one of the most imperiled groups of animals, are in serious global decline and urgently need protection and conservation 1–3. The Tennessee and Cumberland River basins in the United States support the most diverse assemblages of freshwater mussels in the world, including 60 unionid mussel species 4. However, the endemic mussel populations of the basins have declined drastically since the 1980s, and 34 mussel species in the basin are currently listed as federally endangered or threatened 2, 4, 5. In addition, approximately 600 river-km of federally designated habitat have been established in the basins to help ensure the survival and recovery of five endangered mussel species 5. The basins have a long history of coal mining or natural gas extraction activities, and these activities are known to release a number of contaminants to surface waters via storm runoff and regulated point source discharges. Elevated concentrations of metals, polycyclic aromatic hydrocarbons (PAHs), and major ions, including calcium, magnesium, sodium, chloride, and sulfate, have been observed in sediment samples collected from historical coal mining or coalbed natural gas production areas 7–9. Previous field surveys have indicated that coal mining may be a source contributing toxic constituents to mussels (G. Johnson, U.S. Geological Survey, Knoxville, TN, USA, unpublished data; 2).

Juvenile mussels are in direct contact with sediment for the first several months after transformation of the gill of a fish host, making mussels at this life stage particularly susceptible to contaminants associated with sediments 10, 11. Although several earlier studies indicated the effects of coal mining–related contaminants in sediment on mussel behavior (feeding or respiration; 12) and survival 13, limited studies have been conducted to evaluate the toxicity of coal mining–associated contaminants in sediment to mussels, partially due to a lack of methods for conducting sediment toxicity tests with freshwater mussels. Since 2006, consensus-based methods have been available for conducting acute or chronic water-only toxicity tests with the early life stages of mussels 14, and several studies have demonstrated high sensitivity of mussels to some chemicals, such as copper, zinc, and ammonia, in water-only exposures 15–17. A series of studies, including the present study, were undertaken in our laboratory to develop the methods for conducting whole-sediment toxicity tests with juvenile mussels in basic accordance with the standard methods for conducting water-only toxicity test with freshwater mussels 14 and sediment toxicity tests with freshwater invertebrates 18, 19 using field-collected sediments or sediments spiked with contaminants 9, 20–22.

The objectives of the present study were to use the refined methods for conducting whole-sediment toxicity tests with juvenile mussels and to assess the potential effects of coal mining– or natural gas extraction–related contaminants (primarily metals and PAHs) in sediment collected from the upper Tennessee River basin and Cumberland River basin on two native mussel species (rainbow mussel, Villosa iris, and wavy-rayed lampmussel, Lampsilis fasciola) and two commonly tested benthic invertebrate species (amphipod, Hyalella azteca, and midge, Chironomus dilutus).

MATERIALS AND METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Sediment sampling sites were established to represent “worst-case” conditions and chosen primarily based on status of mussel communities, and secondarily based on historical or ongoing coal mining or natural gas extraction activities in the upper Tennessee River and South Fork Cumberland River basins in Tennessee and Virginia (see details about sediment collection sites in Supplemental Data, Table S1). Specifically, test sediment samples were collected from seven impacted sites where mussel communities were classified by personnel of the U.S. Fish and Wildlife Service as poor or extirpated and where associated coal mining or natural gas extraction activities existed. Reference sediment samples were collected from five nonimpacted sites where mussel communities were classified as good and where there were no or limited coal mining or natural gas extraction activities (Table 1 and Supplemental Data, Table S1). In addition, sediment samples from six sites suspected to have elevated concentrations of PAHs and chemicals of interest (such as metals) to the Virginia Department of Environmental Quality (VADEQ) were collected from the Tennessee River basin in Virginia. One extra sediment sample was also collected from a site affected by a coal ash spill from the Tennessee Valley Authority Kingston, USA, coal-fired power plant in December 2008 23. A control sediment sample collected from West Bearskin Lake in Minnesota, USA 24, was tested along with these field-collected sediments.

Table 1. Summary of sediment chemistry toxicity indicesa
Site IDSite descriptionAmmonia toxic unitb  Average metal PECQTotal PAH PECQΣESBTU for PAHs
RainbowWRLMΣSEM-AVSΣSEM-AVS/fOC<2-mm particles<0.25-mm particles
  • a

    Includes total ammonia toxic unit, molar difference between the sum of simultaneously extracted metals (SEM for Cd, Cu, Pb, Ni, Zn) and acid volatile sulfide (AVS) normalized to the fraction of organic carbon (ΣSEM-AVS/fOC), average probable effect concentration quotient (PECQ) for total recoverable metals (Cd, Cu, Ni, Pb, and Zn), PECQ for total polycyclic aromatic hydrocarbons (PAHs), and PAH equilibrium partitioning sediment benchmark toxic unit (ΣESBTU) in sediments (<2-mm particles) from the upper Tennessee and Cumberland River basins (two size fractions of sediments for metal analysis).

  • b

    Ammonia toxic unit was calculated based on total ammonia measured in this study and chronic value reported for amphipod, Hyalella azteca 39, midge, Chironomus dilutus 40, and rainbow mussel, Villosa iris, and wavy-rayed lampmussel (WRLM), Lampsilis fasciola 15 after adjustment to a common pH 8.0 and 25°C using Equations 5 and 12 in the U.S. Environmental Protection Agency (U.S. EPA) ambient water quality criteria for ammonia 41. Ammonia toxic units were all <0.1 mg/L for amphipods and midge exposed to sediments from all sites and not listed in the table.

  • c

    Indicates a value exceeding a toxicity threshold.

  • d

    Reference site. More details about reference sites and test sites are in the Supplemental Data, Table S1.

  • e

    U.S. EPA 53.

  • f

    MacDonald et al. 42.

  • g

    U.S. EPA 46 and Ingersoll et al. 39.

  • NM = not measured.

WB (Control)Control sediment1.08c1.87c−0.58−280.220.110.0140.14
TN NR-2 (Refd)Mussel communities classified as not impacted; no or limited coal mining or gas extraction0.831.43c−0.64−710.100.080.0720.63
TN NR-6 (Ref)0.100.18−0.32−350.100.090.0020.04
VA CR-4 (Ref)0.540.94−2.35−980.140.190.0260.10
VA CR-8 (Ref)0.160.280.10340.070.090.0040.15
VA CR-9 (Ref)0.560.98−0.18−260.070.080.0220.26
TN coal ashFly ash spill3.27c5.65c0.69343c0.30NM0.0180.89
TN NR-4Mussel communities classified as not impacted; no or limited coal mining or gas extraction0.360.62−0.81−900.110.100.0540.50
TN NR-50.060.100.25490.110.090.1211.90c
VA CR-30.450.77−0.56−800.060.120.0780.90
VA CR-70.410.70−1.24−1240.170.350.1891.81c
VA PR-10.440.77−1.28−1830.070.090.0850.95
VA PR-21.25c2.16c0.0350.160.260.1712.28c
VA PR-30.270.470.05160.060.100.0260.79
VADEQ-2Suspected to have elevated concentration of PAHs and chemicals of interest to the Virginia Department of Environmental Quality0.340.580.37920.11NM0.1212.88c
VADEQ-40.300.520.31280.14NM0.1971.43c
VADEQ-80.270.47−10.26−7330.17NM0.2041.23c
VADEQ-90.230.40−2.30−1770.14NM0.1861.14c
VADEQ-110.651.13c−1.64−1820.14NM0.1571.37c
VADEQ-120.210.37−2.96−2110.20NM0.0270.22
Threshold1.01.01.7e>130e1.0f1.0f1.0f1.0g

Sediment collection

Composite sediment samples were collected by three sampling crews in June 2009 within approximately a 10-m radius of the designated coordinates at a site. Two precleaned 19-L polyethylene sample buckets were filled to approximately one-third full with site water and placed in a stable location adjacent to the sampling site. A wash bucket with a US no. 10 stainless steel screen (2.0-mm openings; Wildco part 190-J11-2000) was then placed inside one of the sample buckets 20, 21. A scoop sampler (8-cm diameter schedule 40 polyvinyl chloride) was slowly pushed through the surficial sediment to a depth of approximately 8 cm with an attempt to minimize loss of fine-grain particles; several scoops were placed in the wash bucket before sieving. The inner wash bucket was agitated to sieve the finer grain material into the surrounding sample bucket. This process was repeated until the sample bucket contained approximately 10 L of sediment with up to approximately 8 L of water. Within approximately 4 h of collection, sediment samples were transferred to a secure refrigeration truck at approximately 4°C in the dark. All samples were collected within 1 week and delivered by the refrigeration truck to the U.S. Geological Survey Columbia Environmental Research Center (CERC) for toxicity testing. The coal ash sample was collected from the Emory River at Roane County, Tennessee (Supplemental Data, Table S1) on June 16, 2009, kept at approximately 4°C in the dark, and shipped overnight to the CERC for testing.

Sediment processing

Sediment samples were held at 4°C for approximately one week before being processed for toxicity testing. Sample processing included decanting excess overlying water in the two buckets obtained at each sampling site, homogenizing the sediment in each bucket for approximately 5 min with a hand-held drill and stainless-steel auger, combining the homogenized sediment from the two buckets into one bucket and rehomogenizing the sediment in the bucket. The homogenized sediments were scooped into exposure beakers, and overlying water was added 1 d before test organisms were added. Subsamples of the sediments were also collected for physical and chemical characterization of the sediment, including grain size, moisture, total organic carbon (TOC), total recoverable metals, and PAHs.

Subsamples of sediment were also taken for water quality characterization and metal analysis of pore water. Porewater samples were isolated by centrifugation at 7,400 times standard gravity for 15 min at 4°C, collected with polypropylene syringes, and then filtered through a 0.45-µm-pore polyethersulfone membrane equipped with a 1-µm-pore glass fiber prefilter. The porewater samples for major cation and metal analyses were stabilized immediately by adding concentrated nitric acid (16 M) to each sample at a volume proportion of 1:100.

To facilitate retrieval of small juvenile mussels from sediment at the end of the exposures, sediments used for the mussel toxicity tests were press-sieved through a US no. 60 sieve (0.25-mm opening) before the start of the exposure (except for the extra TN coal-ash sample, which contained small particle sizes and did not need to be sieved for the mussel toxicity tests). Approximately 1 L of fine sediment (<0.25-mm particle sizes) was obtained by sieving the sediment with up to 1 L of water obtained from the overlying water in the bucket containing the original sediment sample. The samples of fine sediments were allowed to settle overnight and then homogenized in the bucket before placement in the exposure beakers for the mussel tests. To compare potential differences in metal concentrations between the bulk sediments (<2-mm particles) and the fine sediments (<0.25-mm particles), a subsample of the fine sediment from each sample was also collected for analysis of total recoverable metals.

Physical and chemical characterization of sediment and pore water

Moisture, grain size, and TOC were characterized by Laboratory and Environmental Testing in Columbia, Missouri, USA, following standard methods 25–27. The relative percentage differences (RPDs) between two sets of duplicate sample analyses were 1 and 12% for moisture, 0% for clay, sand, or silt, and 0 and 6.9% for TOC.

Porewater quality characteristics (dissolved oxygen, pH, conductivity, hardness, alkalinity, ammonia, hydrogen sulfide) were determined on centrifuged samples following standard methods 28. The concentration of total ammonia nitrogen (N) was determined with an Orion ammonia electrode and Orion EA940 meter (Thermo Electron). The concentrations of un-ionized ammonia nitrogen were calculated based on measured total ammonia, pH, and temperature using the formula in Emerson et al. 29. Hydrogen sulfide was estimated from total dissolved sulfide at the pH of each porewater sample using the conditional ionization constant for hydrogen sulfide in fresh water 28. The RPD of one duplicate sample analysis was 12.8% for ammonia and <1.8% for other water quality parameters.

Major cations (sodium, potassium, magnesium, calcium) and trace metals (aluminum, barium, beryllium, boron, cadmium, chromium, cobalt, copper, iron, lead, manganese, molybdenum, nickel, strontium, vanadium, and zinc) in pore-water samples were analyzed at Laboratory and Environmental Testing using inductively coupled plasma–atomic emission spectroscopy (ICP-AES; Model 7300 DV ICP-OES, PerkinElmer) according to the U.S. Environmental Protection Agency (U.S. EPA) method 200.7 30. Recoveries for two sample spikes for these analyses ranged from 87 to 105%. All measured results for a National Institute of Standards and Technology standard reference water were within 10% of the respective certified mean concentrations.

Major anions (bromide, chloride, fluorine, nitrate, nitrite, and sulfate) and dissolved organic carbon in porewater samples were analyzed by GPL Laboratories in Frederick, Maryland, USA. The RPD from one duplicate sample analysis was 7.4%. The recoveries for three sample spikes ranged from 70 to 100%, except for chloride, which had recoveries of 10, 11, and 98% in three samples, respectively. It is unclear why chloride recoveries were low in two of the spiked samples, and whether those recoveries were indicative of similarly low results in any of the samples.

Porewater samples were also collected on test day 7 from the amphipod sediment exposure beakers using “peepers” (dialysis chambers) prepared from nominal 2.9-ml polyethylene snap-cap vials in which a 5-mm hole was cut in the cap using a hole-punch tool 21. A peeper was inserted in each sediment in a chemistry beaker (containing amphipods and treated identically to other replicate toxicity beakers) on test day 0 (start of the exposures) and removed on test day 7. After removal of the peeper and decanting off the overlying water, the sediment of each chemistry beaker was transferred to a glass jar, which was sealed and stored at 4°C for later analysis for acid-volatile sulfide (AVS) and simultaneously extracted metals (SEMs) including cadmium, copper, lead, nickel, and zinc at the CERC. Concentrations of AVS in sediment samples were determined using purge and trap treatment with 1 N HCl, and measurement with a sulfide electrode 31. Concentrations of metals of peepers and SEMs in sediments were determined with inductively coupled plasma–mass spectrometry using the methods described in U.S. EPA 200.8 32 and May et al. 33. The RPDs for two duplicate porewater sample analyses for arsenic, cadmium, copper, lead, nickel, and zinc were <1.2% and the recoveries for spikes of these six metals on two porewater samples ranged from 94 to 102%. The RPD for a duplicated sediment analysis was 2.5% for AVS and ranged from 9 to 28% for the SEMs, except for copper, for which the RPD was 77% (it is unclear why the RPD was high in the duplicated copper samples).

Total recoverable metals (arsenic, cadmium, chromium, copper, lead, mercury, nickel, and zinc) in sediment were analyzed at the U.S. EPA Science and Ecosystem Support Division following U.S. EPA methods 200.8 32 and 1631E 34. The RPDs for the eight metals between two sets of duplicate sample analyses ranged from 0 to 27% (12 of 16 RPDs within 10%).

Concentrations of PAHs in sediment samples were analyzed at U.S. EPA Science and Ecosystem Support Division following U.S. EPA methods 3545A (extraction; 35) and 8270D (analysis; 36). Approximately 30 g of sediment were mixed with Hydromatrix (a drying agent). The mixture was then spiked with deuterated monitoring compounds and extracted with a 9:1 methylene chloride/acetone solution. The extract was concentrated and subjected to gel permeation chromatography cleanup. Internal standards were spiked into the concentrate, and the concentrate was then analyzed by full-scan gas chromatography–mass spectrometry in selective ion monitoring mode to achieve lower detection limits.

Sediment toxicity testing

Methods for conducting the sediment toxicity tests with the amphipod and the midge were based on procedures outlined by the U.S. EPA 18 and the American Society for Testing and Materials (ASTM; 19). The methods for conducting the sediment toxicity tests with the two mussel species were adapted from procedures outlined in the ASTM, U.S. EPA, and other publications 14, 15, 18–20. Test conditions are summarized in the Supplemental Data, Table S2.

Test beakers containing approximately 100 ml of sediment and 175 ml of overlying water were placed in water baths at 23°C under static conditions for 8 d before the start of the exposures to allow re-equilibration of sediment and the associated pore water 9, 20. Addition of overlying water to test beakers was started 24 h before the start of the exposures with an automated system that provided 120 ml of water to each replicate beaker every 8 h (approximately two water volume additions per day; 24). The overlying water was well water, which was diluted with deionized water to a water hardness of approximately 100 mg/L as CaCO3, alkalinity of 85 mg/L as CaCO3, conductivity of approximately 250 µS/cm at 25°C, pH of 8.2, and dissolved organic carbon of approximately 0.5 mg/L. Ambient laboratory light (wide-spectrum fluorescent lights) was approximately 200 lux at the water surface of the test beakers with a 16:8-h light/dark photoperiod.

The four organisms were used for toxicity testing with each sediment sample, except for the six additional VADEQ samples in which only the amphipod was tested. Approximately 7-d-old amphipods (mean initial length 1.72 ± 0.37 mm, n = 19) and 7-d-old midge (mean initial ash-free dry wt 0.063 mg/L) used to start the exposures were obtained from laboratory cultures at the CERC. Approximately 90-d-old rainbow mussel (mean length 1.60 ± 0.29 mm, n = 560) and approximately 80-d-old wavy-rayed lampmussel (mean length 1.60 ± 0.23 mm, n = 560) used to start the exposures were obtained from the laboratory cultures at the Virginia Department of Game and Inland Fisheries, Marion, Virginia (USA). Test organisms were acclimated to test water and temperature for 24 h (amphipods and midge) or 48 h (mussels) before the start of the exposures.

A total of 10 amphipods or midge were impartially transferred into each of four replicate beakers per sediment sample at the beginning of the exposures. Ten amphipods were also transferred into an additional beaker with a peeper (replicate chemistry beaker) for each sediment sample. These chemistry beakers were treated identically to other replicate beakers during the amphipod test until the chemistry beakers were removed on test day 7.

The size variation of mussels in the cultures was relatively high (0.9–2.4-mm shell length for rainbow mussel; 1–2-mm shell length for wavy-rayed lampmussel), which could potentially result in biased stocking in some replicates despite our efforts to impartially transfer mussels into test beakers. Thus, the initial sizes of mussels for each replicate were determined by photographing the mussels at the beginning of the toxicity test for subsequent length measurements. Specifically, a total of 10 mussels for each replicate were impartially transferred into a small dish containing test water, and the mussels as a group were photographed under a microscope equipped with a phototube and a camera for subsequent length measurement. The mussels were then transferred into each of four replicate beakers per species and per sediment sample. An additional replicate beaker was added in the rainbow mussel test for a histological study conducted at Virginia Polytechnic Institute and State University, in Blacksburg, Virginia, USA 37.

Amphipod and mussel tests were conducted for 28 d, and the midge test was conducted for 10 d. Amphipods were fed 1.0 ml of yeast–cerophyll–trout chow suspension (1.8 mg solids) once daily, and midge were fed 1.5 ml of a suspension of flake fish food (6.0 mg of dry solids) once daily 18, 19. Mussels were fed 2 ml of an algal mixture (Nannochloropsis concentrate and Shellfish Diet, Reed Mariculture; algal density ∼510 nl cell volume/ml) twice daily 14, 15.

Water quality in the overlying water in a replicate from each sediment sample for each test species was determined at the beginning and the end of the each exposure following standard methods 28. The mean water quality values were similar among sediment samples within a test species or among the four tested species. Mean pH ranged from 8.0 to 8.4, conductivity 248 to 290 µS/cm at 25°C, hardness 92 to 124 mg/L as CaCO3, and alkalinity 83 to 116 mg/L as CaCO3. However, mean concentrations of total ammonia in overlying water were generally lower in the samples from mussel tests (0.04–0.06 mg N/L) than in the samples from the amphipod test (0.05–0.17 mg N/L) and midge test (0.07–0.23 mg N/L). An exception was for the mussel test in the sample VA PR-2, where a high ammonia concentration of 1.08 mg N/L was observed at the beginning of the test, but a low concentration of 0.03 mg N/L was measured at the end of the test. Concentration of dissolved oxygen in overlying water was measured weekly in all treatments and mean concentrations ranged from 5.5 to 6.7 mg/L in the mussel test, 5.0 to 6.4 mg/L in the amphipod test, and 4.3 to 6.4 mg/L in the midge test, all above the acceptable concentrations of 4 mg/L for a water-only chronic toxicity test with mussels 14 and 2.5 mg/L for a sediment toxicity test with amphipods and midge 19.

At the end of the sediment toxicity tests, amphipods, midge, and mussels were isolated from each test beaker by washing the sediment through a US no. 50 sieve (0.3-mm opening), and rinsing the test organisms and remaining sediment into a glass tray. Live amphipods and midge in each replicate were counted. Surviving amphipods were preserved in 8% sugar formalin for subsequent growth measurement. Surviving midge in each replicate were placed in a preweighed aluminum pan for the measurement of ash-free dry weight (after ashing at approximately 550°C for 2 h). The mussels in each replicate beaker were recovered and transferred into a 50-ml glass beaker containing approximately 20 ml of water for the determination of survival (foot movement within a 5-min observation period) using a dissecting microscope. Surviving mussels in each replicate were photographed for subsequent length measurement as described above, and then preserved in 8% formalin for subsequent dry weight measurement. Surviving rainbow mussels in the additional replicate for histological study were fixed in 2.5% glutaraldehyde in 0.1 M sodium cacodylate buffer at a pH of 7.3 and then transferred in vials containing the fixative to the Virginia Polytechnic Institute and State University for histological processing and evaluation of tissues 37.

Lengths of surviving amphipods at the end of testing were measured from the base of the first antenna to the tip of the third uropod along the curve of the dorsal surface using a digitizing system with video micrometer software (Image Caliper, Resolution Technology) connected to a computer and a microscope. The biomass of surviving amphipods from each replicate was estimated as the sum of individual amphipod weights calculated from the empirical relationship: weight (mg) = (0.177 × length {mm}−0.0292)3 20.

Lengths of mussels photographed at the beginning and the end of the tests were measured as the maximum shell length using the digitizing system. The percentage of length increase of mussels over the 28-d exposures was calculated as the difference between mean final and initial lengths of mussels in a replicate divided by the mean initial length in the replicate. Dry weight of surviving mussels in each replicate was determined after the mussels were dried for 24 h at 60°C. The dry weight was measured to the nearest 0.001 mg with a microbalance (Model MX5, Mettler Toledo) for determination of biomass (total dry wt of surviving mussels in a replicate).

Reference toxicant testing

A subset of the amphipods, midge, and rainbow mussels used in sediment toxicity tests was evaluated by conducting acute static-renewal toxicity tests with a reference toxicant (reagent-grade NaCl) at the start of the sediment toxicity tests following standard test methods 14, 19, 38. The wavy-rayed lampmussel was not evaluated in the reference test due to the limited number of mussels available for testing. The reference toxicant tests were conducted in the ASTM reconstituted hard water (hardness 160–180 mg/L as CaCO3; 38) with six concentrations of NaCl (nominally 0, 0.5, 1.0, 2.0, 4.0, and 8.0 g/L). The concentrations of NaCl were confirmed by measuring salinity at the beginning of the tests. A water-only static 2-d test with the amphipod and a 4-d water-only static-renewal test with the midge or the mussels were conducted in a water batch at 20°C with ambient laboratory light (∼200 lux) and a 16:8-h light/dark photoperiod. Five organisms were tested in each of four replicate 50-ml glass beakers containing 30 ml of water (∼1 ml of fine sand was placed into each beaker for amphipod or midge tests). Water renewal was conducted after 48 h by replacing approximately 75% of the water volume. Test organisms were not fed during the exposures. Water quality was measured in the control, medium, and high NaCl concentrations at the beginning and the end of the tests. Dissolved oxygen ranged from 6.9 to 8.1 mg/L, pH 8.3 to 8.5, hardness 140 to 160 mg/L as CaCO3, and alkalinity 110 to 120 mg/L as CaCO3. At the end of the tests, survival was determined (amphipod and midge movement after gentle prodding; mussel foot movement within 5 min). Median effect concentrations (EC50s) for NaCl were calculated using TOXSTAT software (Version 3.5, Western EcoSystems Technology). The EC50 for NaCl (with a 95% confidence interval) was 6.2 (5.6–6.8) g/L for the midge, 5.7 g/L for the amphipod (no 95% confidence interval due to no partial kills), and 3.2 (2.8–3.5) g/L for the rainbow mussel. These values were within the range of EC50s for these test organisms in historic reference toxicant tests conducted at our laboratory in the ASTM reconstituted hard water (N. Wang, unpublished data).

Data analysis

Ammonia toxic unit (measured concentration divided by the estimated chronic effect concentration for a test organism) was calculated based on measured total ammonia in the present study and chronic value for amphipod 39, 40, and for rainbow mussel and wavy-rayed lampmussel 15 after adjustment to a common pH 8.0 and temperature of 25°C using equations 5 and 12 in the U.S. EPA chronic ambient water quality criterion (AWQC) for ammonia 41.

Sediment toxicity benchmarks, which were derived from empirically based probable effect concentrations (PECs) for metals or PAHs, were used to assess relationships between sediment chemistry and toxicity. The PECs are effect-based sediment quality guidelines established as concentrations of individual chemicals above which adverse effects in sediments are expected to frequently occur in field-collected sediments 42. Mean quotients based on PECs were calculated to provide an overall measure of chemical contamination and to support an evaluation of the combined effects of multiple contaminants in sediments 18, 42–45. Individual PEC quotients (PECQs) were calculated for each of five metals (SEM concentrations of Cd, Cu, Pb, Ni, and Zn) and 13 PAHs (acenaphthalene, acenaphthlene, anthracene, benzo[a]anthracene, benzo[a]pyrene, chrysene, dibenzo[a,h]anthracene, fluoranthene, fluorene, 2-methylnaphthalene, naphthalene, phenanthrene, and pyrene) in each sediment sample by dividing the dry-weight concentration of the chemical by the PEC for that chemical 42. To equally weight the contribution of metals or PAHs in the evaluation of sediment chemistry and toxicity, an average PECQ for metals or for total PAHs was calculated for each sample.

Mechanically based chronic sum equilibrium-partitioning sediment benchmark toxic units (ΣESBTUs) were also estimated for 34 parent and alkylated PAHs from the 16 measured parent PAHs 46. As recommended by the U.S. EPA 46, a 50% certainty factor was applied to the ΣESBTU that was calculated from the data on the 16 parent PAH concentrations. The ΣESBTU approach was developed to account for the biological availability of non-ionic organic compounds in different sediments and incorporates select biological effects concentrations in pore water (i.e., final chronic values; 46). The ΣESBTU approach is intended to support integration of benchmarks for 34 PAHs into a toxic unit model that accounts for the joint toxicity of various PAHs with the same mode of toxicity (i.e., nonpolar narcosis). A ΣESBTU benchmark of 1.0 was established, based on an estimated critical body burden of approximately 2 µmol total PAHs/g lipid for 28-d growth of amphipod (H. azteca) in water-only exposures and a final chronic value of approximately 2.2 µmol total PAHs/g of organic carbon in sediment (D. Mount, U.S. EPA, Duluth, MN, USA, personal communication).

Pearson product moment correlation was performed to evaluate the relationships between responses of test organisms and physical or chemical characteristics of sediment using SigmaPlot software (Ver 12, Systat). Spearman rank order correlation was used if basic assumptions for Pearson's correlation were not met, such as the variables were not normally distributed or an outlier appeared to deviate markedly from others. Statistically significant correlation was set as p <0.05.

Sediment samples were designated toxic or not toxic using a reference envelope approach 9, 39. Differences in mean survival, length, or biomass for each test species among all sediment samples were determined by analysis of variance or Kruskal–Wallis test if the assumptions for analysis of variance were not met) using SAS statistical software (SAS/STAT, Ver 9.2; SAS Institute). When the differences were significant among the samples (p <0.05), a sediment sample was classified as toxic if the mean response of one or more endpoints was less than the lowest mean response for the five reference samples.

RESULTS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Sediment characteristics

The results of the physical and chemical analyses of sediment and porewater samples are summarized in the Supplemental Data, Tables S3 to S8. The results for the extra sediment sample impacted by the coal ash are described separately from the results for other sediment samples.

Most of the sediment samples contained a high percentage of sand (>70%) and a low percentage of clay and slit (Supplemental Data, Table S3). The amounts of TOC (0.3–2.4%) or AVS (0.01–10.9 µmol/g), which can influence metal bioavailability and toxicity or PAH bioavailability and toxicity in sediments, were generally low in the sediment samples (Supplemental Data, Table S3). These physical characteristics were generally similar between the reference samples and the test samples.

In pore waters prepared by centrifugation, pH values ranged from 7.0 to 8.2 and were similar between reference samples and test samples (Supplemental Data, Table S4). The concentrations of hardness (72–428 mg/L as CaCO3) and dissolved organic carbon (3.0–26 mg/L), which can influence bioavailability and toxicity of metals or PAHs in pore water, were generally high in all reference and test samples (Supplemental Data, Table S4). The conductivities in the reference samples (244–578 µS/cm at 25°C) were generally lower than the conductivities in test samples (281–1,057 µS/cm at 25°C; Supplemental Data, Table S4). The concentrations of major cations and major anions in pore water were often higher in the test samples than in the reference samples, especially for the concentrations of Na and SO4, which were approximately 2- to 23-fold higher in the test samples (except for VADEQ-12) than the average concentrations in the reference samples (Supplemental Data, Table S4).

Concentrations of ammonia and dissolved metals in pore waters were compared with the U.S. EPA chronic AWQC to determine whether the concentrations of these elements were elevated. The concentrations of ammonia in pore water were generally low, except for the control sediment (WB) and one test sample (VA PR-2) with a concentration slightly above chronic AWQC for ammonia at the associated pH and a test temperature of 23°C 41 (Supplemental Data, Table S4). Ammonia toxic units of these two samples were above the toxicity benchmark of 1.0 for the two mussel species but not for the amphipod (Table 1). The ammonia toxic unit of >1.0 was also observed in a reference sample (TN NR-2 [Ref]) for the wavy-rayed lampmussel (Table 1).

Concentrations of five metals (Cd, Cu, Ni, Pb, and Zn) of interest in pore water prepared by centrifugation 8 d before the start of exposures and concentrations of the five metals plus As (a metal of concern of coal-associated contaminant) in pore water sampled using peepers samplers were low and less than analytical detection limits or below U.S. EPA chronic AWQC with two exceptions (Supplemental Data, Table S5). The hardness normalized Pb concentrations in five centrifuged samples (WB [control], TN NR-6 [Ref], VA CR-4 [Ref], VA PR-2, VA PR-3) were above the U.S. EPA chronic AWQC for Pb (Supplemental Data, Table S5). However, the Pb concentrations in peeper samples were low (all below the analytical detection limits; Supplemental Data, Table S5). Concentrations of the other four metals in peeper samples were also consistently lower than those measured in centrifuged samples (Supplemental Data, Table S5). One peeper sample (obtained from reference sediment VA CR-8) contained Cd at a concentration of 0.49 µg/L, which exceeded the chronic AWQC of 0.27 µg/L at hardness of 112 mg/L in the porewater sample. However, peeper sampler contamination may have contributed to the one unexpectedly high measured Cd value.

Concentrations of other chemicals of interest in pore water were generally low in all reference and test samples (Supplemental Data, Table S6). For example, the concentrations of hydrogen sulfide were less than 0.25 mg/L in all samples, and the concentrations of nitrate and nitrite were below the detection limit of 0.1 mg/L. Concentrations of aluminum and iron (>1 mg/L) were elevated in centrifuged pore waters of almost all reference samples and some test samples (Supplemental Data, Table S6). However, we believe that much of the aluminum in those centrifuged pore waters was probably present as a colloid, and that soluble iron probably was much lower during testing as a result of diffusional losses and due to oxidation and precipitation as Fe(OH)3 during pretest equilibration of the sediments 21.

Concentrations of eight total recoverable metals (As, Cd, Cr, Cu, Hg, Ni, Pb, and Zn) in the bulk sediment (sieved to <2-mm particles) and fine sediment (sieved to <0.25-mm particles) were low (all below the PECs; Supplemental Data, Table S7). The concentrations of a metal between the two size fractions of sediments were relatively similar (typically within a factor of 2; Supplemental Data, Fig. S1), although the concentrations of Pb in the bulk sediments tended to be lower than those in the fine sediments (Supplemental Data, Fig. S1). The average metal PECQs for Cd, Cu, Ni, Pb, and Zn ranged from 0.06 to 0.22 in the bulk sediments or from 0.08 to 0.35 in the fine sediments, all below the sediment toxicity benchmark of 1.0 (Table 1).

Among 22 individual PAHs, only naphthalene had an elevated concentration above the PEC in five test samples (TN NR-5, VADEQ-4, VADEQ-8, VADEQ-9, and VADEQ-11; Supplemental Data, Table S8). The total PAH PECQs ranged from 0.002 to 0.20 and were below the sediment toxicity benchmark of 1.0 (Table 1). However, the ΣESBTUs for PAHs in 8 of 13 (62%) test samples exceeded a sediment toxicity benchmark of 1.0 (Table 1).

The TN coal ash sample had some unique characteristics. The sample contained high silt of 80% (Supplemental Data, Table S4). Porewater pH was high (9.1; Supplemental Data, Table S4). At this pH level and a test temperature of 23°C, the concentration of ammonia of 0.75 in the pore water (Supplemental Data, Table S4) would be above the chronic AWQC for ammonia 41. Ammonia toxic units of 3.27 for the rainbow mussel and 5.65 for the wavy-rayed lampmussel were well above a sediment toxicity benchmark of 1.0 (Table 1). An elevated concentration of As of 184 µg/L was observed in pore water collected from the peeper in the sample (Supplemental Data, Table S5) and the ΣSEM-AVS/foc value was 2.6-fold greater than the sediment toxicity benchmark (Table 1). In addition, the concentration of total recoverable As (71 mg/kg dry) was 2.2-fold greater than the PEC (Supplemental Data, Table S7).

In summary, the values from physical and chemical measurements of sediment samples were generally below toxicity benchmarks, except the ΣESBTU for PAHs was above the sediment toxicity benchmark in approximately 62% of sediments from the test samples. The TN coal ash sample had a high proportion of silt, elevated concentration of As, and high ΣESBTU for PAHs. The ammonia toxic unit of >1.0 for both tested mussels was observed in the control sediment (WB), a test sample (VA PR-2), and the TN coal ash sample.

Sediment toxicity

Mean survival of the amphipod and two mussel species in the control sediment ranged from 98 to 100% (Table 2), meeting the test acceptability criteria of ≥80% control survival for mussels in water-only testing 14 or for amphipods in sediment testing 18, 19. Mean control survival of midge was 88% and average ash-free dry weight was 0.80 mg/individual, meeting the test acceptability criteria of ≥70% survival and ≥0.48 mg ash-free dry weight per individual 18, 19. Live (foot movement) and dead (empty shell) mussels were easily recovered from the fine sediments after the 28-d exposures. Only one of a total of 700 rainbow mussels and two of a total of 560 wavy-rayed lampmussels were not found at the end of 28-d exposures.

Table 2. Mean responses (standard deviation in parenthesis; n = 4) of amphipod (Hyalella azteca), midge (Chironomus dilutus), rainbow mussel (Villosa iris), and wavy-rayed lampmussel (Lampsilis fasciola) in sediment toxicity tests
Site IDAmphipodMidgeRainbow musselWavy-rayed lampmussel
Survival (%)Length (mm)Biomass (mg)Survival (%)Biomass (mg)Survival (%)Length increase (%)Biomass (mg)Survival (%)Length increase (%)Biomass (mg)
  • a

    The value of an endpoint was less than the lowest value of the endpoint for the five reference sites (Ref) when the differences among all sites were significant (p <0.05; with or without the extra sample of TN coal ash).

WB (Control)98 (5.0)4.64 (0.19)5.03 (0.60)88 (15)7.97 (1.33)100 (0.0)56 (20)9.35 (2.63)100 (0.0)80 (14)13.83 (1.97)
TN NR-2 (Ref)98 (5.0)4.10 (0.03)3.35 (0.21)73 (22)8.17 (1.32)98 (5.0)65 (15)9.37 (1.03)100 (0.0)70 (14)10.73 (3.19)
TN NR-6 (Ref)98 (5.0)4.04 (0.10)3.18 (0.24)83 (13)7.42 (0.87)98 (5.0)78 (15)13.40 (1.54)95 (5.8)70 (11)10.02 (1.87)
VA CR-4 (Ref)95 (5.8)4.41 (0.28)4.19 (0.97)80 (18)8.12 (2.16)100 (0.0)48 (7.1)7.60 (1.31)100 (0.0)64 (10)9.19 (2.00)
VA CR-8 (Ref)100 (0.0)4.41 (0.05)4.30 (0.13)95 (5.8)8.55 (0.48)95 (5.8)52 (25)9.28 (5.48)90 (8.2)71 (7.2)9.89 (2.17)
VA CR-9 (Ref)93 (9.6)4.45 (0.17)4.12 (0.67)78 (21)7.19 (2.04)98 (5.0)46 (17)6.46 (2.49)100 (0.0)62 (18)10.35 (3.34)
TN coal ash68 (5.0)a4.75 (0.17)3.74 (0.34)95 (9.6)5.85 (0.17)a100 (0.0)44 (16)a8.84 (2.21)100 (0.0)37 (13)a8.49 (2.44)a
TN NR-495 (10)4.01 (0.06)a3.03 (0.31)a83 (17)8.35 (0.83)100 (0.0)79 (12)15.38 (1.00)100 (0.0)88 (5.9)14.05 (2.6)
TN NR-5100 (0.0)4.07 (0.11)3.35 (0.28)78 (22)6.96 (0.97)a100 (0.0)83 (8.2)14.78 (2.82)100 (0.0)54 (26)a12.84 (2.31)
VA CR-3100 (0.0)4.51 (0.12)4.60 (0.37)75 (37)5.76 (3.51)a98 (5.0)33 (20)a6.03 (1.16)a98 (5.0)63 (10)9.54 (1.52)
VA CR-7100 (0.0)4.37 (0.23)4.22 (0.72)88 (19)8.36 (0.97)98 (5.0)36 (24)a5.30 (2.60)a93 (9.6)45 (27)a5.92 (2.68)a
VA PR-175 (33)a4.54 (0.32)3.31 (1.04)83 (17)7.84 (1.20)98 (5.0)38 (9.4)a6.40 (1.00)a100 (0.0)72 (15)10.90 (2.75)
VA PR-298 (5.0)4.11 (0.12)3.39 (0.38)93 (9.6)11.10 (0.85)10 (14)a29 (0.6)a0.37 (0.47)a78 (22)a53 (11)a7.18 (1.90)a
VA PR-393 (9.6)4.37 (0.12)3.85 (0.26)95 (10)8.19 (0.93)100 (0.0)75 (14)11.53 (1.11)98 (5.0)87 (4.2)13.65 (0.97)
VADEQ-295 (5.8)4.00 (0.13)a3.01 (0.38)aNot testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
VADEQ-4100 (0.0)4.56 (0.25)4.80 (0.78)Not testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
VADEQ-885 (30)a4.01 (0.23)a2.87 (1.21)aNot testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
VADEQ-9100 (0.0)4.05 (0.29)3.33 (0.74)Not testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
VADEQ-1190 (8.2)a4.60 (0.13)4.46 (0.75)Not testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
VADEQ-1295 (5.8)4.45 (0.14)4.16 (0.19)Not testedNot testedNot testedNot testedNot testedNot testedNot testedNot tested
p value<0.05<0.001<0.001>0.05<0.01<0.05<0.001<0.001<0.05<0.001<0.001

Mean survival, length, or biomass of each test species differed significantly among sediments, except for midge survival (Table 2). Relative to the response of the organisms in the five reference samples, at least one response of one or more test organisms was reduced in 10 of 14 test samples (71%; Table 2). Specifically, reduced responses were observed in seven of eight test samples (88%) in which all four test species were tested, and in three of six samples (50%) in which only amphipods were tested (Table 2). A higher percentage of the eight test samples were classified as toxic to mussels (63% for rainbow mussels and 50% for wavy-rayed lampmussels) compared with amphipods (38%) and midge (38%) (Table 2). A higher percentage of the test samples were identified as toxic to mussels by endpoints of length increase or biomass (up to 63%) compared with the survival endpoint (13%; Table 2).

Comparisons of sediment characteristics to toxicity responses

Survival, length, or biomass of amphipods and midge were not significantly negatively correlated to any physical or chemical measurements in sediments (data not shown). In contrast, the responses of two mussel species were, in many cases, significantly negatively correlated to the sediment characteristics (Tables 3 and 4; note that the WB control sediment and the TN coal ash sample were not included in these correlation analyses). Therefore, only the correlations between responses of mussels (not amphipods or midge) and sediment characteristics are described below.

Table 3. Correlation coefficient between the response of mussels and sediment physical characteristics, porewater chemical characteristics, or polycyclic aromatic hydrocarbons (PAHs) and PAH toxicity indices for sediments from the upper Tennessee and Cumberland River basinsa
VariableRainbow musselWavy-rayed lampmussel
SurvivalLength increaseBiomassSurvivalLength increaseBiomass
  • a

    The WB control sediment and TN coal ash sample were excluded for the correlation analysis.

  • b

    Compound was used for the calculation of total PAH probable effect concentration quotient (PECQ; 42).

  • c

    Compound with a high water–octanol partition coefficient (log KOW >5.0) was used for the calculation of high KOW PAH PECQ 42.

  • d

    PAH equilibrium partitioning sediment benchmark toxic unit.

  • *

    Significant correlation (p < 0.05, n = 12).

Sand−0.23−0.26−0.26−0.30−0.140.03
Silt0.250.270.280.300.17−0.01
Clay0.140.180.150.260.00−0.11
Total organic carbon0.10−0.13−0.120.21−0.16−0.29
Hardness−0.16−0.64*−0.59*0.09−0.38−0.35
Alkalinity−0.31−0.61*−0.68*−0.24−0.07−0.34
Conductivity−0.06−0.51−0.55−0.21−0.26−0.28
Un-ionized ammonia−0.80*−0.66*−0.84*−0.60*−0.38−0.53
Dissolved organic carbon−0.500.01−0.19−0.45−0.04−0.33
Ca−0.07−0.59*−0.59*0.27−0.29−0.40
K−0.24−0.21−0.32−0.33−0.27−0.47
Mg−0.09−0.57*−0.58*0.09−0.22−0.25
Na0.13−0.41−0.41−0.19−0.19−0.04
Cl−0.04−0.63*−0.67*−0.38−0.38−0.47
SO4−0.07−0.34−0.33−0.17−0.210.05
Caprolactam0.11−0.06−0.060.20−0.07−0.14
Carbazole−0.53−0.53−0.55−0.56−0.70*−0.66*
Naphthaleneb−0.260.000.060.00−0.230.20
1,1-Biphenyl0.34−0.190.080.22−0.39−0.18
2-Chloronaphthalene0.080.070.050.20−0.09−0.11
2-Methylnaphthaleneb−0.16−0.080.020.08−0.300.14
Acenaphtheneb−0.72*−0.61*−0.68*−0.76*−0.65*−0.73*
Fluoreneb−0.56−0.64*−0.65*−0.58*−0.71*−0.71*
Acenaphthyleneb0.060.080.040.07−0.13−0.16
Dibenzofuran−0.18−0.21−0.10−0.01−0.400.00
Anthraceneb−0.79*−0.58*−0.69*−0.81*−0.61*−0.70*
Phenanthreneb−0.51−0.51−0.47−0.39−0.68*−0.45
Fluorantheneb,c−0.56−0.58*−0.64*−0.66*−0.64*−0.74*
Pyreneb,c−0.47−0.57−0.61*−0.59*−0.65*−0.74*
Chryseneb,c−0.34−0.54−0.53−0.47−0.71*−0.71*
Benzo[a]anthraceneb,c−0.47−0.56−0.60*−0.58*−0.68*−0.73*
Benzo[a]pyreneb,c−0.46−0.56−0.59*−0.57−0.68*−0.74*
Benzo[b]fluoranthene−0.30−0.51−0.52−0.44−0.65*−0.70*
Benzo[k]fluoranthene−0.48−0.53−0.58*−0.61*−0.66*−0.74*
Benzo[ghi]perylene−0.21−0.48−0.46−0.36−0.68*−0.68*
Dibenz[a,h]anthraceneb−0.38−0.54−0.55−0.51−0.69*−0.72*
Indeno [1,2,3-cd] pyrene−0.50−0.57−0.62*−0.60*−0.67*−0.75*
Total PAH PECQ−0.50−0.49−0.48−0.45−0.68*−0.51
Mean high KOW PAH PECQc−0.47−0.57−0.60*−0.58*−0.67*−0.74*
ΣESBTU for PAHsd−0.58*−0.37−0.38−0.50−0.60*−0.32
Table 4. Pearson correlation coefficient between the response of mussels (Villosa iris, Lampsilis fasciola) and concentrations of simultaneously extracted metals (SEM), the SEM and acid volatile sulfide (AVS) normalized to the fraction of organic carbon (ΣSEM-AVS/fOC), average metal probable effect concentration quotient (PECQ) fraction of organic carbon (ΣSEM-AVS/fOC), total recoverable (TR) metals in sediments from the upper Tennessee and Cumberland River basins
VariableRainbow musselWavy-rayed lampmussel
SurvivalLength increaseBiomassSurvivalLength increaseBiomass
  • a

    SEMs in fine sediment samples were not measured.

  • *

    Significant correlation (p <0.05, n = 12).

Bulk sediment (<2-mm particles)
SEM-As0.10−0.27−0.21−0.15−0.52−0.68*
SEM-Cd−0.34−0.26−0.32−0.23−0.47−0.49
SEM-Cu−0.16−0.44−0.43−0.29−0.66*−0.73*
SEM-Ni−0.42−0.36−0.42−0.39−0.48−0.57
SEM-Pb−0.02−0.45−0.42−0.11−0.57−0.70*
SEM-Zn−0.56−0.63*−0.70*−0.60*−0.63*−0.78*
ΣSEM-AVS−0.270.260.18−0.330.050.28
ΣSEM-AVS/fOC−0.250.350.24−0.33−0.030.23
TR-As0.19−0.23−0.09−0.15−0.54−0.60*
TR-Cd−0.23−0.08−0.11−0.14−0.33−0.30
TR-Cr0.060.110.110.09−0.42−0.29
TR-Cu−0.48−0.36−0.40−0.61*−0.62*−0.74*
TR-Hg0.09−0.03−0.010.21−0.33−0.28
TR-Ni−0.430.150.04−0.39−0.31−0.25
TR-Pb−0.09−0.37−0.36−0.18−0.62*−0.70*
TR-Zn−0.69*−0.45−0.55−0.68*−0.62*−0.66*
Average metal PECQ (all eight metals)−0.38−0.21−0.26−0.43−0.59*−0.61*
Average metal PECQ (Cd, Cu, Ni, Pb, Zn)−0.49−0.23−0.30−0.50−0.56−0.59*
Fine sediment (<0.25-mm particles)a
TR-As0.02−0.41−0.31−0.25−0.61*−0.69*
TR-Cd−0.27−0.44−0.46−0.40−0.57−0.65*
TR-Cr−0.34−0.59*−0.58*−0.42−0.68*−0.82*
TR-Cu−0.25−0.51−0.50−0.42−0.64*−0.78*
TR-Hg−0.23−0.19−0.24−0.08−0.19−0.28
TR-Ni−0.72*−0.44−0.55−0.75*−0.49−0.63*
TR-Pb−0.13−0.55−0.52−0.26−0.66*−0.77*
TR-Zn−0.57−0.60*−0.66*0.66*−0.62*−0.76*
Average metal PECQ (all eight metals)−0.41−0.56−0.58*−0.53−0.65*−0.77*
Average metal PECQ (Cd, Cu, Ni, Pb, Zn)−0.45−0.56−0.59*−0.55−0.64*−0.77*

In general, no significant correlations were observed between the responses of mussels and the sediment particle size, TOC, and chemical characteristics in pore waters of centrifuged samples (Table 3). A few exceptions were that the survival, length increase, or biomass (primarily length increase and biomass of rainbow mussels) were significantly correlated to the hardness, alkalinity, un-ionized ammonia, Ca, Mg, or Cl in the pore waters (Table 3; other variables with no significant correlations were not included in Table 3). However, concentrations of 15 of 22 measured PAHs and the PAH toxicity indices (total PAH PECQ, high KOW PAH PECQ, and ΣESBTU) were significantly negatively correlated with at least one response of mussels (Table 3). Among these 15 individual PAHs, the concentrations of nine PAHs were significantly correlated with one or more responses of both mussel species (Table 3).

Significant negative correlations were also observed between the responses of mussels and concentrations of up to four simultaneously extracted or total recoverable metals (i.e., As, Cu, Pb, and Zn) in bulk sediment (<2-mm particles; Table 4). The significant correlations were observed more frequently when the metals in the fine sediment (<0.25-mm particles; Table 4) were compared. In addition, the mussel responses were significantly correlated to seven of eight metals (i.e., As, Cd, Cr, Cu, Ni, Pb, and Zn) in fine sediments (Table 4). The average metal PECQs were also significantly correlated to the mussel responses, whereas ΣSEM-AVS and ΣSEM-AVS/fOC were not significantly correlated to any responses of mussels (Table 4).

The relationships between mussel growth (shell length increase) and selected metal or PAH toxicity indices are illustrated in Figure 1. The two samples with a PECQ for metals >0.2 were those samples that were classified as toxic to mussels (Fig. 1A and B, and Table 3). Additionally, two of three samples tested with rainbow mussels and all three samples tested with wavy-rayed lampmussels were toxic, with a PECQ for total PAHs >0.1 (Fig. 1C and D) or with a ΣESBTU for PAHs >1.0 (Fig. 1E and F).

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Figure 1. (A–F) Relationship between shell length increase of rainbow mussel (Villosa iris) or wavy-rayed lampmussel (Lampsilis fasciola) and average probable effect concentration quotient (PECQ) for total recoverable metals (Cd, Cu, Ni, Pb, and Zn), PECQ for total polycyclic aromatic hydrocarbons (PAHs), or PAH equilibrium-partitioning sediment benchmark toxic unit (ΣESBTU) in sediment toxicity tests. No regression line is shown in E because no significant correlation was found (p >0.05).

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Correlation analysis among concentrations of major cations (Ca, K, Mg, Na) or major ions (Cl, SO4), conductivity, PECQs for metals, PECQs for high KOW PAHs, PECQ for total PAHs, and ΣESBTU for PAHs showed that the concentrations of the major cations were not significantly correlated with the metal or PAH toxicity indices, whereas the major anions were significantly correlated with some of the toxicity indices (Supplemental Data, Table S9). However, all four metal and PAH toxicity indices were significantly correlated each other (r2 ranging from 0.57 to 0.94, p ranging from <0.001 to <0.05; Supplemental Data, Table S9).

DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

The results of the present study indicate that the concentrations of dissolved metals in pore water generally were below U.S. EPA chronic AWQC, the concentrations of total recoverable metals or PAHs were below the PECs, and the metal or PAH toxicity indices did not exceed sediment toxicity benchmarks that have been established primarily with the amphipod (H. azteca) and with the midge (C. dilutus; 43, 45). An exception was that the ΣESBTU for PAHs exceeded the sediment toxicity benchmark of 1.0 in 57% of test samples (Table 1; n = 14, not counting the TN coal ash sample). Neither the control sediment nor the reference samples had a ΣESBTU for PAHs >1.0. The relatively high ESBTU values might be mainly due to the low percentage of TOC (0.3–2.4%) in the test sediments.

Although the concentrations of metals and PAHs in the whole-sediment samples were relatively low, a total of 77% of test samples (10 of 13 samples, Table 2; not including the coal ash sample) were identified as toxic to at least one of the test species relative to five reference samples, or a total of 86% of test samples tested with all four species (six of seven samples, Table 2) were identified as toxic to at least one of the test species relative to five reference samples. Because a higher proportion of samples was classified as toxic to mussels than to the commonly tested amphipods and midge (Table 2), the mussels were more sensitive under the conditions in the test sediments compared with the other two organisms. The overall agreement with the classification of sediment toxicity based on the responses of the amphipods or midge and the two mussel species ranged from 62 to 77% when all reference samples and test samples were used for the comparison (Table 5A), and ranged from 38 to 63% when only test samples were used for the comparison (Table 5B). Importantly, there was an up to 38% incidence of toxicity to mussels when the samples were not identified as toxic to amphipods or midge (Table 5). These results were consistent with those reported for exposures of sediment samples from metal-contaminated sites in southeast Missouri, USA 9 and for exposures of sediment samples containing elevated concentrations of major anions and major cations in the Holston River at Saltville, Virginia, USA 47. In contrast, the amphipod (H. azteca) was more sensitive than a unionid mussel (Lampsilis siliquoidea) in sediment toxicity tests conducted with exposure to metal-contaminated sediments from the Tri-State Mining District in Missouri, Kansas, and Oklahoma, USA 20 or with exposure to nickel-spiked sediments 21.

Table 5. Percentage of agreement with classification of sediment toxicity between responses of mussels (rainbow mussel, Villosa iris, and wavy-rayed lampmussel, Lampsilis fasciola) and commonly tested amphipod (Hyalella azteca) or midge (Chironomus dilutus) across all reference and test sites (A) or only within test sites (B), and percent agreement between mussel responses based on laboratory toxicity tests and histological evaluations (C; 37)a
 AmphipodMidge
ToxicNot toxicToxicNot toxic
  • a

    Value indicates agreement between a comparison.

A. Mussels vs amphipod or midge compared across all reference and test sites (n = 13)
Rainbow musselToxic15a2315a23
Not toxic854a854a
  Overall agreement: 69Overall agreement: 69
Wavy-rayed lampmusselToxic8a2315a15
Not toxic1554a862a
  Overall agreement: 62Overall agreement: 77
B. Mussels vs amphipod or midge compared within test sites (n = 8)
Rainbow musselToxic25a3825a38
Not toxic1325a1325
  Overall agreement: 50Overall agreement: 50
Wavy-rayed lampmusselToxic13a3825a25
Not toxic2525a1338a
  Overall agreement: 38Overall agreement: 63
C. Toxicity test vs histological evaluation compared across all control, reference, and test sites (n = 14)
 Gill filament 
AbnormalNormal
Rainbow musselToxic21a14
Not toxic2143aOverall agreement: 64

The high proportion of 77% of test samples classified as toxic by laboratory testing with mussels, amphipods, or midge in the present study indicated a good agreement between the field observations and the results of laboratory sediment toxicity tests. Previous studies also showed high agreement between a survey of mussel populations (taxa richness) in a river contaminated by metals and the results of laboratory sediment toxicity testing conducted with mussels 9, 48. In the present study, mussel survival or growth decreased with increasing concentrations of some of the individual metals or PAHs (Tables 3 and 4) and with increasing PECQs for metals and PECQs for total PAHs (Fig. 1). These results indicate that the metals and PAHs in the sediments may have contributed to the toxicity to the mussels. Determining the relative contribution of metals or PAHs to the observed toxicity was not possible given the significant correlation between the metal and PAH toxicity indices (Supplemental Data, Table S9).

Freshwater unionid mussels are sensitive to ammonia 15, 17. However, it was unlikely that the ammonia in pore water or overlying water contributed to the toxicity to mussels in any test samples in the present study. The concentrations of ammonia in pore waters sampled before the start of exposures were generally low (Supplemental Data, Table S4) although a few samples, including the control sediment and one reference sample (TN NR-2 [Ref]), had high concentrations of ammonia (ammonia toxic unit >1.0; Table 1). However, the concentrations of ammonia in overlying waters were low in all sediment samples. Importantly, the mussels exposed to the control sediment and the reference sediment (TN NR-2 [Ref]) with high concentrations of ammonia in pore water survived and grew well over the 28-d exposures (Table 2). It was likely that the ammonia concentrations in pore water and overlying water were reduced by water renewal (two additional water volumes per day) during the 28-d exposures. Furthermore, there is considerable uncertainty in the estimated concentrations of ammonia or pH in pore water isolated from sediment using centrifugation 17, 20.

Significant correlations between the mussel responses and the concentrations of two major cations (Ca, Mg) or a major anion (Cl) were only observed in the rainbow mussel test (Table 3). However, a visual evaluation of the relationships between mussel growth (length increase) and the major cations (Ca, K, Mg, Na) or major anions (Cl, SO4) in scatterplots indicated the trends of decreasing growth of mussels with increasing concentrations of the major ions (Supplemental Data, Figs. S2 and S3). Importantly, a higher proportion of toxic samples was often observed at the high end of the concentration ranges of the major ions or conductivity (Supplemental Data, Figs. S2 and S3). Integrative parameters, such as conductivity and salinity, are often used to measure concentrations of major ions in field studies, and a few studies have demonstrated that the decline of mayflies and other aquatic macroinvertebrates correlates with water conductivity or salinity concentrations in residential or mining areas 49–51. The concentrations of the major cations (Ca, K, Mg, and Na) were not correlated to the concentrations of metal or PAHs (Supplemental Data, Table S9), and thus, these individual cations or a combined effect of these cations may also have contributed to the toxicity of the sediment samples. A previous study showed reduced mussel survival in reconstituted waters with elevated concentrations of major cations and major anions in acute water-only toxicity tests 47. In contrast to the major cations, the concentrations of the major anions (Cl and SO4) were correlated to the concentrations of metals or PAHs in the present study (Supplemental Data, Table S9), and thus, the effects of the major anions on mussels were uncertain. However, the concentrations of sulfate in the eight sediments of the present study (66–220 mg SO4/L at hardness of >140 mg/L as CaCO3; Supplemental Data, Table S4) were approximately three- to ninefold lower than the 20% effective concentration of 600 mg SO4/L at hardness of 100 mg/L as CaCO3 for mussels (Lampsilis abrupta) in a 28-d water-only sulfate toxicity test (N. Wang, unpublished data). It appeared unlikely that sulfate alone contributed the observed toxicity of the sediment samples to mussels. In laboratory toxicity tests conducted with reconstituted waters representative of ambient water associated with runoff from mountain top mining of coal in Kentucky, Virginia, West Virginia, and Pennsylvania, USA, the mussel (L. siliquoidea) was more sensitive in acute or chronic exposures to mixtures of elevated major ions compared with the amphipod (H. azteca) or the cladoceran (Ceriodaphnia dubia; C. Ingersoll, unpublished data). Additional field samples should be evaluated to characterize major cations and major anions in sediment pore waters and surface water, and then, effects of major cations and major anions on mussels could be determined by testing mussels in reconstituted waters to represent key porewater quality characteristics without the addition of other potential contaminants (e.g., metals and PAHs) as performed in the previous study with surface water 47.

Toxicity was observed to both mussel species at low PECQs for metals (<0.4) and low PECQs for total PAHs (<0.2). All of the samples were toxic to the mussels when the PECQs for metals were >0.2 (Fig. 1A and B), and 67 to 100% of samples were toxic to the mussels when the PECQs for total PAHs were >0.1 (Fig. 1E and D), well below a sediment toxicity benchmark of 1.0 established primarily with tests conducted with the amphipod and midge 43, 45. These results indicate that mussels may be more sensitive to metals and PAHs than other organisms commonly tested in sediment toxicity tests, and therefore, these sediment toxicity benchmarks based on PECs may not be protective of mussels.

In contrast to PECQ for total PAHs, a sediment toxicity benchmark of 1.0 for PAHs estimated using ΣESBTU appeared to be more protective of the mussels. When the ΣESBTUs for PAHs exceeded the benchmark of 1.0, 67 to 100% of samples were toxic to mussels (Fig. 1E and F). However, when ΣESBTU for PAHs was >1.0, only 38% of samples were toxic to amphipods and 33% of samples were toxic to midge (Tables 1 and 2); these percentages were lower than the frequencies of toxicity (50–63%) at a ΣESBTU benchmark of >1.0 for the two commonly tested species reported in other studies 39, 52.

More test samples were classified as toxic to mussels based on growth (percentage of length increase) or biomass (a combined response of survival and dry wt) compared with survival (Table 2), indicating that sublethal endpoints were more sensitive than a lethality endpoint for mussels, which is consistent with previous studies on mussel sensitivity to contaminants in water exposures 17, 22). The ability to identify toxic samples was similar between the biomass and percentage of length increase (Table 2), and the correlations of the two endpoints to metals or PAHs were also similar (Tables 3 and 4). Due to the likely differences in initial weights of mussels among individual replicate beakers, the potential influence of initial size differences on biomass (calculated based on final wt) is unknown, and the mussel biomass data in the present study should be used with caution. Ideally, similar sizes of mussels should be used to start a toxicity test. A study evaluating the relationships of shell lengths and dry weights of juvenile mussels of different ages and species is needed. The length–weight relationships are particularly important when limited numbers of mussels with large size variation are available to start a toxicity test.

Adverse effects were observed in all four test species in the sediment sample impacted by the coal ash spill in Kingston, Tennessee (Table 2). Elevated concentration of metals (e.g., As was 2.2-fold higher than the PEC for As; Supplemental Data, Table S7) in the sediment might have contributed to the observed toxicity of this sediment sample to mussels, amphipods, and midge. The concentrations of metals measured in the present study (Supplemental Data, Table S7) were similar to those reported in a study on the coal ash spill, in which the reported levels of metals in coal ash samples were up to 30-fold higher than the levels in local soil 24.

Results of the histological evaluation of rainbow mussels sampled at the end of the 28-d exposure in the present study were summarized by Henley 37. Although microscopical inspections of diverticula of digestive glands revealed no abnormalities and kidneys, ganglia, posterior adductor muscles, heart–rectum complexes, and vesicular connective tissues showed no evidence of pathologies, abnormalities of gill filaments, including absence or reduction in abundance of cilia, deterioration of epithelial layers, or hemocoel spaces, and fusion of filaments were observed in mussels from one of the reference samples (VA CR-4 [Ref]) and from five of 14 test samples (TN coal ash, TN NR-5, VA CR-7, VA PR-2, and VA PR-3). The overall agreement based on the histological evaluations and the sediment toxicity tests was 64% (Table 5C). The results indicate that the histological response of the mussels was a useful endpoint to help evaluate potential effects of sediment-associated contaminants on mussels.

The starting age (two- to three-month-old) and size (∼1.6-mm length) of juvenile mussels used in the present study were similar to those recommended by the ASTM to start chronic water-only toxicity tests 14. The decision to start sediment exposures with this life stage of mussels was based on high control survival and high sensitivity of this life stage to contaminants in 28-d chronic water-only exposures 14–17, 22). Another advantage to starting exposures with mussels that are approximately 1 mm in length is that the small juveniles, which have not fully developed to filter feeders, use foot (pedal) feeding and burrow into the sediment 10, 11. A concern with tests using small mussels in bulk sediments (<2-mm particles) is that retrieval of mussels at the end of the 28-d exposure can be difficult. The fine sediments (<0.25-mm particles) used in the present study resulted in recovery of >99% of the mussels introduced at the beginning of the exposures.

One concern in using the fine sediment in mussel tests was the potential differences of chemical characteristics between the fine and bulk sediment samples. Concentrations of each of eight metals in the two size fractions of sediments in the present study were relatively similar (Supplemental Data, Fig. S1), indicating that these metals, and perhaps other chemicals, in the fine sediments likely were not substantially different from those in the bulk sediments. Subsequent 28-d toxicity testing of mussels in bulk sediments with <2-mm particles has been successfully conducted at our laboratory (21; C. Ingersoll, unpublished data), indicating that if sediment exposures are started with mussels with a shell length between 1.5 and 2.0 mm, presieving of sediments to <0.25-mm particles is not necessary. However, observations we have made with the older juveniles of rainbow mussel (∼four months old and ∼3-mm shell length) indicate that animals at this life stage begin to burrow less in sediment and tend to begin filtering overlying water at the sediment–water interface, like adult mussels (e.g., are less dependent on foot feeding behavior within sediment). Thus, the larger (older) juveniles in a sediment toxicity test may behave differently from younger juveniles that tend to burrow more into the sediment. We are currently conducting a series of studies to further evaluate (1) the influence of sediment physical and chemical characteristics (e.g., particle size, organic content, porewater quality) on the response of juvenile mussels in sediment toxicity tests; and (2) the influence of mussel species, size, age, burrowing behavior, feeding, or duration of exposure on the responses of juvenile mussels in sediment toxicity tests.

Only a limited number of sediments were evaluated in the present study, which targeted sites where mussel populations had been classified as impacted and where coal mining or natural gas extraction activities had occurred. A more comprehensive understanding of the effects of energy resource extraction on the upper Tennessee River and Cumberland River basins' mussels could be gained by additional studies that (1) evaluate a gradient of sites potentially impacted by coal mining or natural gas extraction activities and are coupled with quantitative measures of mussel populations at these sites; (2) evaluate mussel sensitivity in sediments spiked with contaminant(s) of concern associated with coal mining or natural gas extraction activities (e.g., metals, PAHs); and (3) evaluate the toxicity of field-collected sediment porewater samples and determine the effect of major ions on test organisms in reconstituted water representative of key water quality characteristics of field-collected samples.

CONCLUSIONS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Whole-sediment toxicity tests were successfully conducted with juvenile mussels following methods adapted from standard ASTM methods for conducting water-only toxicity tests with freshwater mussels and for conducting sediment toxicity tests with freshwater invertebrates. The results based on responses of test organisms in laboratory testing were generally consistent with the field observations of impacted mussel populations inhabiting sites located in proximity to coal mining or natural gas extraction activities. Total recoverable metals, PAHs, or major ions, or all three in sediments might have contributed to the sediment toxicity. The two tested mussels were more sensitive to the contaminants in sediments than the commonly tested amphipods and midge. Therefore, sediment toxicity benchmarks based on PECs may not be protective of mussels. Quantitative surveys of potential impacts of coal mining or natural gas extraction activities on local mussel communities, together with laboratory toxicity tests with additional field sediment samples or with sediments spiked with chemicals of interest, are needed to further evaluate the concentration–response relationships and provide a basis for developing site-specific sediment toxicity benchmarks for mussels inhabiting the upper Tennessee and Cumberland River basins.

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

We thank the Toxicology Branch and Environmental Chemistry Branch of the Columbia Environmental Research Center for technical assistance and R.C. Lott at Virginia Department of Environmental Quality for suggestions on study design, sediment sample collections, and chemical analysis. We also thank the Virginia Department of Game and Inland Fisheries, Aquatic Wildlife Conservation Center for providing juvenile mussels for testing. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government.

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  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

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