Impact assessment of dredging to remove coal fly ash at the Tennessee Valley Authority Kingston Fossil plant using fathead minnow elutriate exposures



On December 22, 2008, failure of an earthen containment structure resulted in the release of approximately 4.1 million m3 of coal fly ash into the Emory River and the surrounding area from the Tennessee Valley Authority Kingston Fossil Plant near Kingston, Tennessee, USA. The purpose of the present study was to assess the potential of dredging activities performed to remove the fly ash from the river to result in increased risk to pelagic fish, with special consideration of mobilization of metals. Elutriates were created using two sources of fly ash by bubbling with air over 10 d. This elutriate preparation method was designed to represent worst-case conditions for oxidation, metal release, and dissolution. Larval and juvenile Pimephales promelas underwent 10-d exposures to these elutriates. Larval end points included survival and biomass, and juvenile end points included survival, length, biomass, liver somatic index, and bioaccumulation. No significant toxicity was observed. Bioaccumulation of metals in juveniles was found to be primarily attributable to metals associated with particles in the gut. Results suggest little potential for toxicity to related fish species due to fly ash removal dredging activities given the extreme conditions represented by the elutriates in the present study. Environ. Toxicol. Chem. 2013;32:822–830. © 2013 SETAC


On December 22, 2008, approximately 4.1 million m3 of coal fly ash was released into the Emory River and the surrounding area from the Tennessee Valley Authority Kingston Fossil Plant near Kingston, Tennessee, USA, upon failure of an earthen containment structure 1. This fly ash has previously been shown to have elevated levels of metals 1. Recovery of the fly ash from the Emory River was accomplished primarily through the use of hydraulic cutterhead dredges. These dredges pumped the ash from the river bottom through a pipeline into an ash recovery ditch, where the heavier ash particles that settled rapidly were removed by a mechanical excavator for disposal. The remaining water and suspended ash flowed into a sluice channel and mixed with plant process water, then flowed into an ash settling pond that overflowed into a stilling pond for final effluent polishing, and finally was discharged back to the Emory River 2.

Because the potential exists for metal release and changes in metal speciation due to dredging activities 2, the purpose of the present study was to describe the potential effects of dredging fly ash on fish in the Emory River. To address this goal, larval and juvenile fathead minnows (Pimephales promelas) were exposed to two different suspensions (i.e., elutriates) created over a 10-d period using fly ash and waters taken directly from the Emory River and the ash recovery ditch. These elutriates were created using this extended 10-d period to represent worst-case conditions for oxidation, metal release, and dissolution, as well as metal speciation changes that may occur during dredging. Toxicity (survival and biomass) was assessed for the larval fish, and toxicity (survival, biomass, length, and liver somatic index) and metal bioaccumulation were assessed for the juveniles. While 24 metals were quantitated in fish tissue, particular focus in data interpretation was given to four metals of concern due to their known presence in fly ash and potential toxicological significance: arsenic (As), chromium (Cr), mercury (Hg), and selenium (Se).


Field sampling

All fly ash and site water samples were collected on June 12, 2009, and immediately packed on ice and transported overnight by truck to the U.S. Army Engineer Research and Development Center in Vicksburg, Mississippi. An initial attempt was made to use a Ponar grab sampler for fly ash collection; however, because the submerged fly ash had developed a crust due to tight consolidation on its surface, an acceptable grab sample using this method was not possible. Fly ash from the Emory River was collected by boat using a polypropylene shovel in areas where the fly ash was submerged by the river at depths of less than 1 m. Samples were collected from the submerged fly ash to a depth of at least 15 cm. Emory River fly ash samples were taken from five discrete locations approximately 2.2 miles upstream of the confluence of the Emory and Clinch Rivers near 35°54.877′N, 084°30.163′W, in an area where ash removal by hydraulic dredging was occurring (Fig. 1). These five samples were composited, homogenized by mixing with a polypropylene shovel until visually homogenous, and passed through a 1.25-cm sieve to remove any large debris from the samples. The homogenized samples were placed into two 18.9-L plastic buckets.

Figure 1.

Map of sampling locations near Kingston, Tennessee, USA.

Reference water for use in preparing the Emory River elutriate and in toxicity studies was collected from the Emory River approximately 12 miles upstream of the Kingston Fossil Plant site (35°55.995′N, 084°33.554′W). This water was collected using an electric pump at approximately mid-depth at a 1.5-m-deep site that was located approximately 7.6 m from the shoreline. The pump and associated hoses were flushed with site water for 15 min prior to sampling. Water was held in 208-L high-density polyethylene drums. Bags of ice were packed around the exterior of the drums for transportation to the U.S. Army Engineer Research and Development Center.

Fly ash and water samples were also collected from the ash recovery ditch (35°54.192′N, 084°31.020′W). The water was sampled using the same electric pump, which was flushed with water from the ash recovery ditch prior to sample collection. Ash recovery ditch water was collected in 16-L cubitainers and packed in coolers with ice for transportation. Due to safety concerns, fly ash was not sampled from the ash recovery ditch using a shovel. Instead, fly ash was sampled from the bottom of the ash recovery ditch at a depth of approximately 3 m using a mechanical excavator and then placed in an 18.9-L plastic bucket for transportation to the U.S. Army Engineer Research and Development Center.

Elutriate preparation and characterization

Elutriates were created 1 d after the receipt of samples at the U.S. Army Engineer Research and Development Center. The Emory River elutriate (EMR-EL) and ash recovery ditch elutriate used in the P. promelas bioassays were created following minor modifications of the dredging elutriate test (DRET) procedure outlined by Palermo et al. 3. The DRET procedure is designed to evaluate contaminant releases by dredge-induced resuspension. The primary departure from the DRET procedure included the use of a 10-d mixing (via aeration) time instead of the 1-h mixing time given in the procedure. The purpose of the longer mixing time was to represent extreme conditions for metal release and dissolution so that a conservative assessment of toxicity and bioaccumulation due to dredging operations could be made. To minimize evaporative loss during the 10-d elutriate preparation, aeration with humidified air was used to suspend the fly ash material. Fly ash samples were homogenized using a rotary mixer in the laboratory prior to subsampling from sample containers. Specifically, 10 g (wet wt) of fly ash material was added per liter of Emory River or ash recovery ditch water to create the elutriate mixture. The Emory River fly ash sample was 53% solid material, and therefore, the elutriate sediment load was calculated to be approximately 5.3 g/L. The ARD-EL was prepared by mixing 10 g of ash recovery ditch sediment at 73% solids by weight for each 1 L of ash recovery ditch water with 0.7 g/L total suspended solids (TSS), yielding an elutriate TSS of approximately 8 g/L. All elutriates were created in triplicate to provide multiple samples for analytical determinations. Elutriates for the ash recovery ditch experiments were prepared in 19-L high-density polyethylene containers. Because of the volume of water needed for the juvenile fish bioassays, Emory River elutriates were prepared in 208-L high-density polyethylene drums. All elutriate preparations were performed in a 20°C environmental chamber.

During the 10-d elutriate preparation period, metal samples were collected at 1, 24, 48, 96, and 240 h. These samples were allowed to settle for 1 h, according to the standard DRET procedure. Then, the supernatant was decanted and submitted for metal analysis. Total and dissolved (fraction passing through a 0.45-µm filter) metal concentrations in the prepared elutriates, as well as the Emory River and ash recovery ditch waters, were thoroughly characterized to include As and Se speciation, as previously described in Bednar et al. 2. Briefly, metal concentrations for aqueous and solid samples were analyzed using inductively coupled plasma-atomic emission spectrometry (ICP-AES) or inductively coupled plasma-mass spectrometry (ICP-MS), as appropriate for the concentration ranges observed following U.S. Environmental Protection Agency (U.S. EPA) methods 6010C and 6020A 4. Speciation of Se and As was performed using high-performance liquid chromatography–ICP-MS as previously described 5, 6. Mercury concentrations were determined using atomic fluorescence spectrometry following U.S. EPA method 7474 4. Total organic carbon concentrations were determined following U.S. EPA method 9060A 4. Solid-phase metal concentrations were also quantitated from fly ash sampled from the Emory River, the ash recovery ditch, and the failed fly ash pile. For solid-phase analysis methods and results, see Bednar et al. 2.

Fish bioassays

The potential for effects due to elutriate exposure was assessed by 10-d exposures of two different life stages of P. promelas. All P. promelas were exposed to elutriate water collected on day 10 of the DRET elutriate preparation described above. Pimephales promelas is a recommended freshwater test species for elutriate toxicity testing in U.S. EPA and U.S. Army Corps of Engineers (U.S. ACE) guidance for evaluation of dredged material 7, 8 and is also a U.S. EPA–recommended species for the assessment of toxicity of effluents and receiving waters 9. Pimephales promelas were obtained from a commercial supplier (Aquatic Biosystems). Larval fish were selected to assess the potential toxicity of a sensitive life stage, while juvenile fish were tested to gain information on whole-body burdens. Larval fish were exposed to both elutriates (EMR-EL and ARD-EL), whereas juvenile fish were exposed only to EMR-EL. For all exposures, the 10-d extended elutriate preparation described above was considered the 100% elutriate treatment. In addition, this elutriate was diluted with Emory River reference water to prepare 50 and 10% elutriate treatments. Comparative treatments included the Emory River reference site water (0% treatment) and a performance control (dechlorinated tap water, Vicksburg, MS, USA, municipal source). Statistical comparisons of toxicity end points were made to the Emory River reference site water. Each treatment involved five experimental replicates, and exposures were conducted at 20 ± 1°C to represent conditions at the dredging site. The test acceptability criterion used for both larval and juvenile P. promelas was 80% survival in the performance control at the conclusion of the 10-d exposure as recommended by the U.S. EPA 9 7-d P. promelas survival and growth test (test method 1000.0).

To maintain water quality during the exposures, water exchanges (90% of total volume) using freshly diluted treatments (control, 0, 10, 50, 100%) were conducted on test days 3 and 7. Daily water exchanges were not practical due to both the amount of site water that would have been required for both bioassays (>800 L) and the settling of the solid-phase particles in the larval bioassay that necessitated transfer of larvae to new exposure chambers at each water renewal to avoid temporally increasing test concentrations. Prior to preparation of test concentrations, the stored 100% elutriates were thoroughly homogenized by mechanical mixers (for EMR-EL, Lightnin DuraMix, E78R2558N-RR; Mixing Equipment; for ARD-EL, model RW 20 DS1; IKA) for 5 min, and elutriate consistency at each renewal was confirmed by TSS measurements (conducted according to American Public Health Association 10 guidelines). During testing, larval and juvenile P. promelas received feeding rations of Artemia sp. nauplii and fish flakes (Zeigler® AquaTox Feed; Aquatic Eco-Systems), respectively. In compensation for forgoing daily water renewals, this ration was supplied every other day to maintain water quality. The experimental design used was in basic accordance with the 96-h P. promelas test method by the U.S. EPA and U.S. ACE 8 for assessing suspended dredge material effects. Survival was assessed daily, and deceased individuals were removed promptly.

Water-quality parameters (temperature, pH, conductivity, dissolved oxygen) were recorded for all replicates at test initiation and termination and for one replicate per treatment during test days 1 to 9. Ammonia was measured in one replicate per treatment over the duration of the exposure, and alkalinity and hardness were measured at test initiation. A model 315i meter (WTW) was used for pH and temperature, a model Oxi 330 meter (WTW) was used for dissolved oxygen, and an ECTestr low instrument (Oakton Instruments) was used for conductivity. Meters were calibrated before each use. LeMotte titration kits were used to measure ammonia (model PAN code 4795), alkalinity (model WAT-DR code 4491), and hardness (model PHT-DR-L1 code 4482).

Larval fish elutriate bioassay

Fish larvae (6 d old) were obtained and held for 72 h to allow acclimation to laboratory conditions. During each day of acclimation, a feeding ration of Artemia sp. nauplii was supplied and water quality was monitored and recorded. The age of the larval fish at test initiation was 9 d posthatching. The bioassay was conducted in general accordance with guidelines 8, 9, 11. Each of the five experimental replicates (300-ml glass beakers) per treatment was loaded with 10 larvae. Each replicate was gently aerated from an oil-free source (trickle flow, two to five bubbles per second) to provide some turbulent flow and to maintain dissolved oxygen levels. More rigorous aeration was not used to attempt to keep the solid-phase material in suspension because this may have caused energetic stress to the larval fish. At test termination, the end points assessed were survival and biomass. Additionally, a reference toxicity test using KCl was conducted to assess the larval fish health relative to historic control charts.

Juvenile fish elutriate bioassay

Juvenile fish were also obtained and held for 72 h to allow acclimation to laboratory conditions. The acclimating fish received a daily feeding ration of fish flakes, and water-quality parameters were recorded daily. Juvenile fish (4.2 ± 0.3 cm) were selected to supply adequate tissue mass for chemical analysis. While a specific testing protocol was not available for this juvenile fish exposure, it was conducted in general accordance with bioassay guidelines 8, 9, 11. Each of the five experimental replicates (3.75-L glass jars) per treatment was loaded with five juvenile fish and rigorously aerated from an oil-free source to represent turbulent lotic conditions, to provide adequate agitation to keep the solid-phase material suspended, and to maintain dissolved oxygen levels.

At test termination, the end points assessed were survival, whole-fish biomass, length, liver mass (to determine liver somatic index), and bioaccumulation. At termination, all fish were anesthetized using tricaine methanesulfonate and blotted dry. Three of the five fish per replicate were immediately frozen for analysis of chemical residues in whole-fish tissue. These fish were later composited and homogenized into a fine powder using a mortar and pestle over a liquid nitrogen bath and then submitted for chemical analysis. Dissections (i.e., liver removal) were conducted on the remaining two fish per replicate immediately following test termination.

Fish tissue metal analysis

Fish homogenates were digested by a modification of U.S. EPA 4 method 3050B, using additional hydrogen peroxide as needed to destroy lipids that may interfere with ICP-MS elemental analysis. The digestate was filtered through a #41 Whatman filter and diluted to 50 ml with 1% nitric acid. This digestate was further diluted with 1% nitric acid prior to ICP-AES and ICP-MS, as needed, such that the analyte concentrations were within the instrument calibration range. Mercury was determined by cold vapor atomic fluorescence following U.S. EPA 4 method 7474.

Analysis of gut metal content in juvenile fish

To assess the portion of metals in tissue measured during the juvenile P. promelas 10-d bioassay that were attributable to metal associated with ingested particles present in the gut, an additional 10-d juvenile P. promelas study was conducted. Determining what fraction of metals present in the gut versus the fraction present in the rest of the body is important because a significant portion of the metals in the gut may be passed out of the body and may not be bioavailable to the fish 12. Treatments in the present study included Emory River water (0% elutriate), EMR-EL (100% elutriate), and a dechlorinated tap water control. This bioassay was performed using the same methods as the original 10-d bioassay, except that in the second study each treatment consisted of one exposure chamber with five fish in each, and water was renewed once on day 5. At the end of the 10-d exposure, there was 100% survival in each treatment. Fish were killed using tricaine methanesulfonate, and their digestive tracts were removed (the liver was not removed). The remainder of the fish carcass was homogenized into a fine powder using a mortar and pestle over a liquid nitrogen bath. The digestive tracts were dried overnight in a 60°C oven, then ashed in a muffle furnace at 550°C for 3 h. The metal content in the ashed digestive tracts and the homogenized fish carcasses was then submitted for analytical determination using the same methods described for the original 10-d bioassay.

Statistical analysis

All statistical comparisons were conducted using SigmaStat software (version 3.5; Systat Software). Data normality and homogeneity were assessed by the Kolmogorov-Smirnov and Levene tests, respectively. All survival data were arc-sine–transformed. A one-way analysis of variance (ANOVA) was conducted to assess statistical significance between treatment levels (α = 0.05), and individual differences were elucidated using Dunnett's test. When necessary, data were transformed (square root, log10) to accommodate the assumptions of normality. When data distributions did not allow parametric statistics, the Kruskal-Wallis ANOVA on ranks was applied. Statistical significance of end points for the different treatment levels (10, 50, 100%) was compared relative to the site water (0% treatment). Tissue residues of juvenile fish were compared by one-way ANOVA and the Dunnett method as described above.

Interpretation of elutriate and tissue metal concentrations

In the assessment of aqueous dissolved metal concentrations for the two elutriates, the highest dissolved concentration of each metal during the EMR-EL and ARD-EL 10-d elutriate preparation was compared with screening values to evaluate the most extreme scenario for exposure of fish to metals in the elutriate preparation. Most often, the highest measured concentration was the 10-d value. The criterion continuous concentration (CCC) from the U.S. EPA was selected as the preferred screening value protective of aquatic life. In addition to the screening values, literature values were also used for the comparison.

Metal concentrations measured in whole fish after exposure to the elutriates were evaluated through interpretation of literature-based metal critical body residues (CBRs). While metal tissue residues are generally poor estimators of an effect due to many organisms' ability to sequester or otherwise detoxify metals, they can be useful in a weight-of-evidence assessment to determine the role of a metal in potential toxicity. However, some metals, such as Se, are best evaluated through the use of tissue concentrations 13. Relevant CBRs (e.g., no observed effect residues [NOERs] and lowest observed effect residues [LOERs]) for fish were obtained through literature searches and using the U.S. ACE and U.S. EPA Environmental Residue-Effects Database ( This assessment was considered conservative because CBRs, which often do not include concentrations in the gut, were compared to the whole-body burdens of metals in fish exposed in the present study, which included metal-laden particles associated with the gut contents.

Screening of relevant literature studies used to compare water and tissue concentrations was performed following methods described by Steevens et al. 14. Where relevant screening criteria were not available, literature values were sought to relate metal concentrations in water and tissues to effects or the lack thereof observed in elutriate toxicity bioassays. Because there is a wide variety of laboratory and field-based studies reporting toxicity of metals, the available literature must be screened to obtain the most relevant comparable results to interpret the water and tissue concentrations. For the present study, we limited studies to those reporting laboratory-only experiments with fish. Preference was given to those examining the effects in P. promelas or another warm freshwater fish. For quality control, all studies must have reported measured concentrations and included a range of concentrations. Exposures of longer duration were preferred. End points considered in the screening of relevant literature studies included survival, growth, and reproduction.


Control survival in the bioassays with both life stages of P. promelas met acceptability criteria (10-d larval control survival = 86%, 10-d juvenile control survival = 100%). Additionally, all water parameters measured during the exposures were within the ranges specified by test guidance. The 48- and 96-h median lethal concentration values for larvae exposed to the reference toxicant KCl were both 0.69 (0.60–0.80) g KCl/L, which fell within two standard deviations (SD) from the mean lethal concentration derived from laboratory control charts (0.61–1.13 g/L). Total suspended solid levels were determined to be comparable for the resuspended 100% treatment and freshly diluted 50 and 10% treatments on each of the water renewal days (Table 1). Total and dissolved elutriate metal concentrations for the four metals of interest on day 10 of elutriate preparation are provided in Table 2. Dissolved metal concentrations from the 1-, 24-, 48-, 96-, and 240-h time points during the 10-d elutriate preparation period for EMR-EL and ARD-EL can be found in Supplemental Data, Tables S1 and S2, respectively. Total organic carbon concentrations in EMR-EL and ARD-EL increased over the 10-d elutriate mixing time: from 3.70 (± 0.26) to 11.50 (± 2.10) mg/L in EMR-EL and from 2.56 (± 0.16) to 10.83 (± 2.75) mg/L in ARD-EL. The observed increases in total organic carbon over time are thought to be a result of increases in microbial assemblages. A summary of measured water-quality parameters from the larval and juvenile P. promelas bioassays can be found in Table 3.

Table 1. Confirmation of elutriate homogenization consistency by total suspended solid concentrations (mean ± 1 standard deviation) prior to water renewals (days 0, 3, and 7)
ElutriateConcentrationMean total suspended solids (mg/L)
Day 0Day 3Day 7
  1. EMR-EL = Emory River elutriate; ARD-EL = ash recovery ditch elutriate; NA = data not available, sample not taken.

EMR-EL10%36.92 ± 10.6351.5 ± 15.250.00 ± 2.38
50%185.33 ± 16.63215.5 ± 33.81270.08 ± 18.00
100%500.42 ± 32.17453.08 ± 48.57527.58 ± 17.32
ARD-EL10%45.08 ± 14.6846.50 ± 2.8446.17 ± 39.81
50%NA157.58 ± 12.13213.83 ± 31.04
100%435.17 ± 12.79461.92 ± 425.07468.75 ± 14.34
Table 2. Total and dissolved metal concentrations (micrograms per liter, mean ± 1 standard deviation) for the four metals of interest in the Emory River and ash recovery ditch elutriates after 10 d of mixing
MetalEmory River elutriateAsh recovery ditch elutriate
As122.3 (±9.1)53.0 (±1.1)155.0 (±14.0)81.3 (±1.6)
Cr25.6 (±2.0)1.8 (±0.1)23.8 (±2.8)0.7 (±0.0)
Hg0.164 (±0.008)<0.0050.212 (±0.045)<0.005
Se6.9 (±0.4)5.5 (±0.5)12.0 (±1.0)12.7 (±0.1)
Table 3. Summary of water-quality parameters measured during the 10-d larval and juvenile Pimephales promelas exposuresa
Site IDConcentrationTemperature (°C)pHConductivity (µS/cm)Dissolved oxygen (mg/L)Ammonia (mg/L)Alkalinity (mg/L)Hardness (mg/L)
  • a

    Means ± one standard deviation from the mean are provided. Numbers in parentheses represent the minimum and maximum values.

    REF-WA = Emory River reference water; EMR-EL = Emory River elutriate; ARD-EL = ash recovery ditch elutriate.

ControlNA20.1 ± 0.2 (19.7–20.6)8.12 ± 0.14 (7.91–8.47)274 ± 7 (260–280)8.2 ± 0.8 (6.9–9.5)(<1–1)110100
REF-WA0%20.0 ± 0.4 (19.2–20.6)7.79 ± 0.20 (7.57–8.21)97 ± 6 (90–110)8.4 ± 0.9 (6.8–9.5)(<1–1)6420
EMR-EL10%20.0 ± 0.3 (19.4–20.5)7.62 ± 0.26 (7.12–8.02)96 ± 5 (90–100)8.3 ± 0.8 (6.6–9.5)(<1–1)  
50%20.1 ± 0.4 (19.6–20.8)7.63 ± 0.20 (7.20–7.94)114 ± 6 (100–120)8.1 ± 0.9 (6.4–9.3)(<1–1)  
100%20.0 ± 0.3 (19.6–20.7)7.74 ± 0.17 (7.30–7.95)129 ± 9 (120–140)8.1 ± 1.0 (6.6–9.4)(<1–1)10022
ARD-EL10%20.4 ± 0.3 (20.0–20.6)7.69 ± 0.14 (7.52–7.95)114 ± 5 (110–120)7.7 ± 1.0 (6.0–9.4)(<1–1)  
50%20.1 ± 0.4 (19.5–20.8)7.79 ± 0.12 (7.54–8.05)177 ± 19 (110–190)8.2 ± 1.1 (6.4–9.8)(<1–1)  
100%20.1 ± 0.5 (19.4–21.0)8.09 ± 0.12 (7.91–8.43)269 ± 12 (250–290)8.4 ± 0.9 (6.4–9.9)(<1–1)16076

Larval elutriate bioassay

In the larval fish exposure, daily survival counts were estimated in the 50 and 100% treatments due to the opaque nature of the media and settled solid-phase material. However, survival counts were fully assessed on water exchange days (days 3 and 7, data not shown) and at test termination (day 10). Feeding was observed in all elutriate treatments during the exposure. Relative to the site water treatment (0%), there were no significant decreases in 10-d larval survival (Fig. 2a). While slight decreases in survival were observed for the two highest ARD-EL treatments (100 and 50%) (Fig. 2a), they were not statistically significant. This trend was not observed in the EMR-EL treatments. Overall, larval growth did not occur relative to the initial weights of the fish, potentially due to the reduced feeding ration (once every other day vs twice daily), which was employed to maintain water quality with fewer water exchanges. Data were consequently reported as biomass. Only one statistically significant (p = 0.014) decrease in 10-d biomass relative to the site water was observed for EMR-EL 10%. However, significant reductions were not observed in the higher treatments (EMR-EL 50 and 100%) (Fig. 2b).

Figure 2.

Summary of larval Pimephales promelas survival percentage (A) and biomass (B) after 10-d exposure. Statistical comparisons of toxicity end points were made to the Emory River reference site water (0% elutriate dilution). “Control” refers to the laboratory performance control in dechlorinated tap water. All statistical comparisons were to the respective site water (0%). Asterisk (*) indicates a statistically significant decrease in the end point.

Juvenile elutriate bioassay

No mortality was observed during the 10-d exposure of juvenile fish. Fish length (p = 0.194), whole-fish biomass (p = 0.188), liver mass (p = 0.506), and liver somatic index (p = 0.789) were not significantly different in any of the treatments relative to the site water (0% treatment) (Table 4). The combination of the rigorous aeration and the current created by fish swimming kept the solid-phase material in suspension. The juvenile fish could not be enumerated in the 50 and 100% treatments on non–water exchange days due to the opaque nature of the suspension. Feeding was observed in all elutriate treatments during the exposure.

Table 4. Summary of mean juvenile Pimephales promelas survival and biomass (and one standard deviation from the mean) after 10-d exposure to Emory River elutriate
Elutriate treatmentPercent survivalLength (cm)Biomass (mg)Liver mass (mg)Liver somatic index
0%100 ± 04.2 ± 0.3888 ± 18816.83 ± 4.641.9 ± 0.6
10%100 ± 04.0 ± 0.4796 ± 17316.04 ± 4.632.0 ± 0.3
50%100 ± 04.3 ± 0.2959 ± 13818.46 ± 5.061.9 ± 0.5
100%100 ± 04.2 ± 0.3943 ± 16720.04 ± 8.092.2 ± 0.8

Whole juvenile fish body burdens

Analysis of the whole bodies of juvenile P. promelas revealed a dose-dependent response in the mean concentrations of the four metals of interest (Table 5). The increases in As and Se were significantly greater in all EMR-EL treatments (10, 50, 100%) relative to fish exposed to the Emory River reference water (0%). For Hg, the two highest treatments (50, 100%) showed significantly higher concentrations in the fish. While Cr did increase monotonically, no statistical significance was found in body burdens relative to the 0% treatment. Significant increases in whole-body burdens were observed for an additional 12 metals (Al, Ba, Cd, Co, Cu, Fe, K, Mg, Mo, Pb, Tl, and V). Although metals were detected in the flake fish food that was provided every 48 h (As 1.7, Ba 3.17, Cr 5.5, Hg 0.002, Se 1.3, V 0.36 mg/kg), the feeding ration was kept equal in all exposure chambers. Thus, the consistent and statistically significant dose-dependent relationship in the whole-body burdens of the fish suggests that the source of the metal burdens in fish was the test material. However, provided that this assessment was a whole-fish analysis in an exposure to metal-laden particles in suspension, the measured concentrations include the sum of metals accumulated in tissues and material that resided in the gut.

Table 5. Pimephales promelas whole-body burdens (milligrams per kilogram wet wt; mean ± 1 standard deviation) for arsenic (As), chromium (Cr), mercury (Hg), selenium (Se), and other metals after 10-d exposure to Emory River elutriate
  • a

    Significant increase relative to tissue concentrations measured in the 0% (Emory River reference) treatment. The detection limit was 0.10 mg/kg (detection limit for mercury was 0.005 mg/kg).

  • b

    Selenium is reported as a dry weight using an internally derived 0.238 conversion factor to compare with a U.S. Environmental Protection Agency 13 tissue benchmark value.

As0.12 ± 0.020.37 ± 0.041.21 ± 0.22a2.95 ± 0.76a2.84 ± 0.99a
Cr1.86 ± 1.123.53 ± 3.205.50 ± 3.005.45 ± 2.148.12 ± 5.13
Hg0.015 ± 0.0010.017 ± 0.0030.017 ± 0.0020.023 ± 0.004a0.023 ± 0.005a
Seb1.02 ± 0.061.12 ± 0.121.46 ± 0.11a2.14 ± 0.21a2.45 ± 0.42a
Ag<0.100.17 ± 0.27<0.10<0.100.13 ± 0.18
Al11.10 ± 3.5825.56 ± 9.06400 ± 112.17a955.80 ± 253.59a942.20 ± 352.93a
Ba1.83 ± 0.172.40 ± 0.528.20 ± 1.82a18.54 ± 4.99a20.24 ± 7.90a
Be<0.10<0.10<0.100.16 ± 0.040.16 ± 0.04
Ca5,810 ± 1415,544 ± 8505,766 ± 4355,566 ± 3805,106 ± 514
Cd<0.10<0.100.13 ± 0.01a0.37 ± 0.07a0.39 ± 0.09a
Co<0.10<0.100.37 ± 0.07a0.81 ± 0.19a0.83 ± 0.30a
Cu1.07 ± 0.141.30 ± 0.292.24 ± 0.37a4.09 ± 0.89a4.07 ± 1.03a
Fe35.42 ± 3.0971.10 ± 24.41a248.60 ± 51.26a517.60 ± 127.75a497.00 ± 175.74a
K2,202 ± 1412,304 ± 482,374 ± 1562,498 ± 1702,586 ± 159a
Mg260 ± 16266 ± 19293 ± 16338 ± 29a344 ± 36a
Mn1.20 ± 0.273.41 ± 1.253.41 ± 0.515.62 ± 1.33a5.22 ± 1.72
Mo<0.10<0.100.26 ± 0.19a0.28 ± 0.05a0.31 ± 0.12a
Na931 ± 56917 ± 48945 ± 47972 ± 24994 ± 57
Ni0.86 ± 0.571.71 ± 1.463.07 ± 1.153.83 ± 0.845.08 ± 2.78
Pb0.21 ± 0.10a0.36 ± 0.20a0.81 ± 0.30a1.47 ± 0.41a1.57 ± 0.56a
Sb0.20 ± 0.34<0.100.15 ± 0.22<0.100.45 ± 0.36
Tl<0.10<0.10<0.100.15 ± 0.030.21 ± 0.04a
V0.06 ± 0.030.10 ± 0.051.54 ± 0.40a4.06 ± 0.93a3.88 ± 1.28a
Zn24.94 ± 2.0427.98 ± 3.4227.12 ± 1.0827.82 ± 3.8726.96 ± 3.87

Gut metal content in juvenile fish

Mercury was not analyzed in the gut due to low mass of the recovered material. The total concentrations of As, Cr, and Se were much higher in the gut removed from the fish than from the remainder of the fish carcass. This effect was most dramatic when fish were exposed to EMR-EL where concentrations in the gut were 50 to 2,800 times higher than in the whole fish with the gut removed. This evidence would suggest that the exposure-dependent elevations in metal concentrations in fish were more likely from gut contents than actual metals integrated into the fish tissue.


The purpose of the present study was to determine the potential effects of fly ash removal by dredging on fish in the Emory River. Elutriate bioassay results may be interpreted through a weight-of-evidence evaluation that considers multiple lines of evidence (e.g., chemistry, toxicity, accumulation) to assess the potential for the suspended fly ash to adversely affect aquatic life. A weight-of-evidence approach provides a more comprehensive evaluation using all available data in contrast to reliance on a single measurement end point 15. In this assessment, elutriate bioassay results are complemented by additional lines of evidence including dissolved metal concentrations and metal body burdens in fish from the juvenile P. promelas elutriate bioassay.

No significant impact on survival was observed from either the larval or the juvenile P. promelas 10-d bioassay. The only statistically significant test end point reduction noted was in biomass at the EMR-EL 10% dilution in the larval bioassay; however, because significant reductions were not observed at the higher effluent concentrations (50 and 100%) and because overall larval growth did not occur over the 10-d test period due to a reduced feeding ration, it cannot be concluded that the EMR-EL elutriate would be expected to induce a harmful effect on P. promelas larvae.

Interpretation of aqueous metal concentrations and fish bioaccumulation data


The highest mean (± SD) concentrations of As measured in EMR-EL and ARD-EL were 53.0 (± 1.1) and 81.9 (± 1.8) µg/L, respectively. Thus, all measured aqueous dissolved concentrations of As were below the CCC of 150 µg/L recommended by the U.S. EPA 16. The highest mean concentration of dissolved As measured in either elutriate preparation, 81.9 (± 1.8) µg/L, is 1.8-fold less than the CCC. Also, literature effect values for aqueous exposure of P. promelas to As were much higher than the conservative CCC value. Arsenic measured in the elutriates was present as As(V) and not the more toxic As(III). However, because toxicity information was more readily available for As(III) in the literature, concentrations of As(V) in the elutriates are compared with effect values for the more toxic As(III). This provided a conservative screening level comparison. The 30-d P. promelas growth no observable effect concentration (NOEC) and lowest observable effect concentration (LOEC) were 2,130 and 4,300 µg/L, respectively, for As(III) in a P. promelas early life stage exposure in which the test organisms were exposed as eggs and then for 30 d postfertilization (about 24 d posthatch) 17. Based on this information, the concentrations of As in water were below all screening thresholds and, thus, not expected to result in adverse effects to P. promelas.

The whole-body burden of As ranged from 0.37 (± 0.04) mg/kg wet weight in fish exposed to the Emory River reference water to 2.95 (± 0.76) mg/kg wet weight in fish exposed to the 50% EMR-EL dilution. The 100% dilution yielded a mean As body burden of 2.84 (± 0.99) mg/kg wet weight. The body burdens of As measured in fish exposed to the Emory River reference water were well below CBR values for As in fish; however, the highest body burden of As measured in the EMR-EL exposure, 2.95 (± 0.76) mg/kg wet weight, is above or approximately equal to CBR values for As in Lepomis macrochirus (bluegill) and Oncorhynchus mykiss (rainbow trout) 18–20. However, the elevated level of As in tissues is likely due to contributions of As from gut contents during analysis and unlikely to indicate toxicity to fish due to As in our test system.


The only Cr species found in the present study was Cr(III), which is widely reported to be less toxic than the more heavily studied Cr(VI) 21, 22. Dissolved Cr (<0.2 µg/L) was not detected above detection limits in the Emory River reference water or the ash recovery ditch water.

In EMR-EL, mean (± SD) dissolved Cr concentrations increased over the 240-h elutriate preparation period from a concentration that was less than detection limits (<0.2 µg/L) at the 1-h time point to 1.8 (± 0.1) µg/L after 240 h of extended elutriate mixing under oxic conditions. Dissolved Cr levels were slightly lower in ARD-EL and increased over time, ranging from less than detection limit (<0.2 µg/L) to 0.7 (± 0.0) µg/L at the 1- and 240-h time points, respectively.

These values are less than the U.S. EPA's CCC 16 for dissolved Cr of 74 and 21 µg/L at water hardness of 100 (standard water hardness used in U.S. EPA 16) and 22 mg/L as CaCO3 (the lowest water hardness measured by titration in a site water), respectively. All measured dissolved concentrations are well below a 48-h NOEC 21 and a 30-d maximum allowable toxicant concentration (MATC) 23 reported for P. promelas. Overall, Cr would not be expected to be of toxicological concern because the measured concentrations in the 100% elutriates were more than 11 times lower than the CCC and 1,000 times lower than a MATC reported in the literature.

In the EMR-EL bioaccumulation exposure of juvenile P. promelas, mean measured Cr whole-body burdens, including gut contents, ranged from 3.53 (± 3.20) to 8.12 (± 5.13) mg/kg wet weight in the lowest (0% elutriate) and highest (100% elutriate) treatments, respectively. The median Cr body burden in the 100% elutriate treatment was comparable to a 108-d NOER reported for Oncorhynchus tshawytscha 22. Additionally, Roling et al. 24 generated a tissue residue benchmark value at which a 15% reduction in growth was modeled (ER15) of 44 mg/kg for Fundulus heteroclitus exposed to hexavalent (VI) Cr. Although this freshwater fish was exposed to the more toxic Cr(VI) species, this ER15 value is still five times higher than the measured body burdens of the less toxic Cr(III) in the present study. It is not expected that the whole-body burdens of Cr are of toxicological significance in the present study. In addition, most of the Cr detected was associated with the fish gut contents.


Only inorganic Hg was detected in the present study, which is reported to be less toxic than methylmercury 25. Dissolved Hg was below reporting limits (<0.005 µg/L) in the Emory River reference water and ash recovery ditch water. In the EMR-EL, mean (± SD) dissolved Hg concentrations did not increase over time during elutriate preparation. In fact, the only detectable concentrations in EMR-EL were for the 1-h (0.014 ± 0.001 µg/L) and 24-h (0.011 ± 0.007 µg/L) time points. Similarly, Hg was only detected in the ARD-EL at the 1-h (0.008 ± 0.005 µg/L) and 24-h (0.010 ± 0.007 µg/L) time points. These values are near reporting limits and should be interpreted cautiously.

The highest measured values of Hg are much less than the U.S. EPA's CCC for dissolved Hg of 0.77 µg/L 16. The measured dissolved concentrations are well below the median, long-term NOECs and LOECs generated for P. promelas 17, 26. Spehar and Fiandt 23 reported a 32-d MATC for P. promelas growth of 0.89 µg/L, a value 64 times higher than the Hg concentrations in the EMR-EL and ARD-EL. However, MATCs as low as 0.26 µg/L, a value that is still 19 times higher than the dissolved Hg concentrations in the elutriates, have been reported by others 26. Overall, the dissolved Hg concentrations detected in the test elutriates are well below the water-quality criterion (55 times lower), and other fish effect levels reported in the literature and are not likely to be toxicologically significant.

In the EMR-EL bioaccumulation exposure of juvenile P. promelas, mean measured Hg body burdens ranged from 0.017 (± 0.003) to 0.023 (± 0.005) mg/kg wet weight in the lowest (0% elutriate) and highest (100% elutriate) treatments, respectively. These body burdens are 87- and 130-fold lower than median NOERs and LOERs reported in the literature 25, 26 and more than eight times lower than the tissue threshold effect level modeled by Beckvar et al. 27. These data suggest that body burdens of Hg in juvenile P. promelas are not high enough to be of toxicological significance. Furthermore, metal burdens in the present study were primarily associated with the gut contents.


Dissolved Se (<0.2 µg/L) was not detected above detection limits in the Emory River reference water. The measured dissolved Se concentration in the ash recovery ditch water was 5.9 µg/L. In the EMR-EL, mean dissolved Se concentrations increased over the 240-h elutriate preparation period from 0.7 (± 0.1) µg/L at the 1-h time point to 5.5 (± 0.5) µg/L at 240 h. Dissolved Se levels were higher in the ARD-EL and increased over time, ranging from 5.6 (± 0.3) to 12.7 (± 0.1) µg/L at the 1- and 240-h time points, respectively. The only species of Se observed in both elutriate preparations was Se(IV). The toxicity of Se species in aquatic exposures is generally organic selenium (selenomethionine) > selenite (Se[IV]) > selenate (Se[VI]) 28, 29.

At least at some time point(s) over the 240-h elutriate preparation period, the concentrations of Se in the ash recovery ditch water as well as in both the EMR-EL and the ARD-EL exceeded the U.S. EPA's and the state of Tennessee's CCC of 5 µg/L for Se in water. Thus, there is some concern for the potential of Se to contribute to chronic toxicity. However, as outlined by the U.S. EPA 13, it is technically more valid to use a whole-body tissue concentration of Se as the chronic criterion because diet is the primary route of Se exposure that controls chronic toxicity to fish. Thus, the Se body burdens measured at the conclusion of the 10-d juvenile P. promelas toxicity and bioaccumulation study (discussed below) should be used as the primary assessment of potential Se impacts to fish in the present study.

While exceedances of water CCCs were observed, all measured dissolved Se concentrations were well below the NOECs reported by Norberg-King 30 in 7-d studies with larval P. promelas that ranged from 377 to 1,450 µg/L. The LOECs for the same studies for the growth end point ranged from 836 to 2,920 µg/L 30. Thus, the highest concentration of Se measured in the ARD-EL, 12.73 (±0.1) µg/L, is at least 65-fold less than the LOECs observed in the present study.

In the EMR-EL bioaccumulation study, the highest mean measured Se body burden was 2.91 (± 0.49) mg/kg dry weight. Wet weights of P. promelas measured in the study were converted to dry weights using an experimentally determined factor of 23.8% to compare with the U.S. EPA 13 chronic exposure criterion for Se in tissue that is given in dry weight (7.91 mg/kg dry wt). The highest measured Se body burden in the EMR-EL of 2.91 mg/kg dry weight is 2.7-fold less than the U.S. EPA chronic exposure criterion for Se in tissue 13. The highest mean measured Se body burden in fish exposed to the Emory River water was 1.33 (± 0.14) mg/kg dry weight.

All measured tissue Se values in the present study are below literature-based CBRs for P. promelas. The CBRs used in this comparison are for larval and adult P. promelas obtained via water 31, 32 and dietary exposure 33, 34. This suggests that the whole-body burden of Se, which included gut contents, is unlikely to be of toxicological concern as measured in the current investigation. While some reported NOER tissue-based CBRs are below tissue Se concentrations measured in fish exposed to the Emory River reference water and EMR-EL, no effect levels such as NOERs can be an artifact of spacing of treatments in experimental design, so exceedances of these values do not necessarily imply that a significant toxic response will occur.

Sources of uncertainty

While the weight-of-evidence approach for the evaluation of toxicity, chemistry, and bioaccumulation is one of the most rigorous methods for integration of data, there is significant uncertainty in its application to field conditions as well as other aquatic species. Uncertainty in the evaluation of water chemistry comes from the variability in water quality (i.e., hardness, pH) for published studies, as well as the present study, compared with field conditions. Furthermore, the conditions represented in the elutriate bioassay were extreme; dilution and settling at the site are expected to result in metal concentrations lower than those in the present study. Uncertainty is also encountered in the estimation of potential effects to fish using measured aqueous and tissue metal concentrations from elutriate bioassays that are exposures to complex mixtures of contaminants when these measured metal concentrations are compared to literature-based effect values that are often derived from single chemical exposures.

Uncertainty in the evaluation of metal in tissues comes from the short duration of exposure in the elutriate bioassay and assumption regarding the species of metal. The concentration of metals in tissue of fish in relationship to toxicity is poorly understood 35. In an organism, metals have the capacity to partition in freely dissolved, protein-bound, or metal-rich granule fractions. Except for organometals (e.g., methylmercury) or Se, it is difficult to relate tissue concentrations to biological effects 36, 37.


In this weight-of-evidence approach—where water chemistry, metal speciation, toxicity, and bioaccumulation data are integrated—the results suggest little potential for toxicity to related fish species due to fly ash removal dredging activities, given the extreme conditions represented by elutriates in the present study.


Tables S1 and S2. (229 KB DOC).


This work was performed for the Tennessee Department of Environment & Conservation and was part of an interagency agreement with the Tennessee Valley Authority. Permission was granted by the chief of engineers to publish this material. The authors thank J. Burr (Tennessee Department of Environment & Conservation) for assistance in field sampling and A. Davis (U.S. Army Engineer Research and Development Center, Environmental Laboratory) for geographic information systems support.