Arch. Environ. Contam. Toxicol. 37, 496–502 (1999)
A R C H I V E S O F
Environmental Contamination a n d Toxicology r 1999 Springer-Verlag New York Inc.
Comparative Toxicity of Fluoranthene to Freshwater and Saltwater Species Under Fluorescent and Ultraviolet Light R. L. Spehar,1 S. Poucher,2 L. T. Brooke,3 D. J. Hansen,4 D. Champlin,4 D. A. Cox3 1 2 3 4
U.S. Environmental Protection Agency, 6201 Congdon Blvd., Duluth, Minnesota 55804, USA Sciences Applications International Corporation, 165 Dean Knauss Dr., Narragansett, Rhode Island 02882, USA University of Wisconsin, 1800 Grand Ave., Superior, Wisconsin 54880, USA U.S. Environmental Protection Agency, 27 Tarzwell Dr., Narragansett, Rhode Island 02882, USA
Received: 5 January 1999/Accepted: 22 May 1999
Abstract. The acute and chronic toxicity of fluoranthene was determined for a diverse group of freshwater and saltwater species under both standard laboratory fluorescent light and ultraviolet (UV) light test conditions. Acute tests with 21 species demonstrated that fluoranthene was not lethal within its water solubility limit to most species tested under fluorescent light, but was lethal well below this limit to nearly all of the species tested under UV light. In general, the acute sensitivity of freshwater and saltwater species from the same class was similar, although UV light exposure changed the relative sensitivity of some species. Crustaceans were the most sensitive to fluoranthene, but in the presence of UV light, an oligochaete and a fish were the most sensitive. Overall, UV light increased acute fluoranthene toxicity approximately one to three orders of magnitude. In chronic tests, sublethal concentrations of fluoranthene were toxic under both fluorescent and UV light, but as in most acute tests, UV light increased chronic toxicity approximately an order of magnitude. Comparison of data from tests conducted in the laboratory and outdoors demonstrated that acute toxicity increased with increased UV light intensity.
Fluoranthene is a polycyclic aromatic hydrocarbon (PAH) and a principal constituent of PAH-contaminated aquatic systems (Sorrell et al. 1980; Hardy et al. 1987; Long 1992). As a component of fossil fuels and a product of incomplete combustion, its major sources to the aquatic environment are industrial and municipal effluents and atmospheric fallout (Neff 1979; Eisler 1987). Fluoranthene represents the greatest fraction of PAHs in some storm waters and combined sewer overflows (Wade 1995). In the Chesapeake Bay, ambient concentrations as great as 928 mg/L of fluoranthene, representing nearly 25% of the total PAH concentration, have been reported in the sea-surface microlayer (Hardy et al. 1987). In sediments of streams, lakes, and estuaries, fluoranthene concentrations as high as 2,370 µg/kg (dry sediment) have been demonstrated (Roberts et al. 1989).
Correspondence to: R. L. Spehar
Fluoranthene has been shown to be acutely toxic to some aquatic organisms (US EPA 1978, 1993). Median effect and lethal concentrations (EC and LC50s) for freshwater species range from 7.7 to 187 µg/L (Horne and Oblad 1983; Gendusa 1990; Suedel et al. 1993), and for saltwater species, from 40 to 500 µg/L (US EPA 1978; Rossi and Neff 1978). Studies reporting toxic concentrations far above fluoranthene’s water solubility limit of 206 µg/L for fresh water and 127 µg/L for salt water (May 1980) are not cited here. Little information exists on the toxicity of fluoranthene to plants (Ren et al. 1994; Huang et al. 1995) or its chronic toxicity to aquatic organisms (US EPA 1978; Oris et al. 1991). Most assessments of fluoranthene toxicity have been conducted under laboratory fluorescent light. However, it has been demonstrated that ultraviolet (UV) light increases the acute toxicity of fluoranthene to freshwater or saltwater organisms in both water and sediments (Kagan et al. 1985; Newsted and Giesy 1987; Ren et al. 1994; Ankley et al. 1994, 1995; Bell 1995; Huang et al. 1995; Boese et al. 1997; Swartz et al. 1997; Pelletier et al. 1997; Monson et al. 1999). For example, in laboratory studies, fluoranthene was 10 to 54 times more acutely toxic to invertebrates in the presence of UV light than when it was tested under fluorescent light (Ankley et al. 1995; Boese et al. 1997; Pelletier et al. 1997). Other studies conducted in the laboratory and the field have shown that aquatic sediments containing mixtures of PAHs were much more toxic to different species when UV light was introduced during the study (Davenport and Spacie 1991; Ankley et al. 1994; Monson et al. 1995; Ireland et al. 1996). Many PAHs, such as fluoranthene, can absorb UV light to become phototoxic to aquatic species (Newsted and Giesy 1987). It has been hypothesized that phototoxicity occurs once energy is transferred from the activated triplet state of the PAH molecule to molecular oxygen, thereby creating singlet oxygen, 1O2, which can react with biomolecules (e.g., amino acids) to cause cell damage (Newsted and Giesy 1987; Larson and Berenbaum 1988). The growing evidence that UV radiation enhances PAH toxicity warrants that phototoxic effects be considered when conducting hazard evaluations of aquatic systems contaminated with PAHs. Aquatic organisms can be at greater risk to these compounds than terrestrial organisms due to the potential for
Comparative Toxicity of Fluoranthene
their exposure to relatively higher concentrations found in aquatic systems (Arfsten et al. 1996). The observed effects of PAHs, such as fluoranthene, in sediments is of concern because regulation may be underprotective if phototoxicity is not accounted for (US EPA 1993). Although previous investigations provide the basis for such concern, comparative studies conducted under similar test conditions are needed to demonstrate the extent of these effects across a wider range of aquatic taxa and over longer exposure periods. The objective of this study was to determine the acute and chronic toxicity of fluoranthene to a variety of freshwater and saltwater species and to compare their sensitivity to this compound under both standard laboratory fluorescent light and UV light test conditions. The species tested were selected from taxonomic groups needed to fulfill requirements for deriving ambient aquatic life water quality criteria (Stephan et al. 1985). Subobjectives were to compare the effects of artificial and natural UV light on acute fluoranthene toxicity and to determine the influence of UV light intensity on this toxicity.
Materials and Methods Test Species The names and stage or age of the species tested are given in Table 1. Most freshwater species were obtained from cultures at the US EPA Mid-Continent Ecology Division, Duluth, MN, or at the University of Wisconsin–Superior, Superior, WI. Dragonflies were collected from the Eau Claire River near Gordon, WI. Certified disease-free rainbow trout and bluegills were purchased from commercial culture facilities. All organisms were acclimated to the test water conditions for at least 7 days prior to testing. Most saltwater species were obtained from cultures at the US EPA Atlantic Ecology Division, Narragansett, RI. American lobster larvae were obtained from gravid field-collected females. Adult lobsters were held under flow-through conditions in the laboratory until zoea were released. Winter flounder were either obtained from ova produced by spawning gravid field-collected adults or from a natural spawn of fish held for up to 1 year in the laboratory.
Water Characteristics Most freshwater species were tested in dechlorinated water from the municipal water supply of the City of Superior, WI, which has its origin from Lake Superior. Tests with the cladoceran and duckweed were conducted in reconstituted water (ASTM 1993). Mean water temperatures were 20–25°C for all freshwater species, except the rainbow trout which were 16.0–18°C. Mean dissolved oxygen (DO) values for acute and chronic tests were 6.5–9.0 mg/L for tests conducted at 20–25°C and 8.4–9.7 mg/L for tests conducted at 16–18°C. Total hardness, alkalinity, conductivity, and pH were measured once in acute tests and weekly in chronic tests. The range of mean values for all tests except those with the cladoceran and duckweed were: total hardness ⫽ 46.5–61.7 mg/L as CaCO3, total alkalinity ⫽ 48.0–61.8 mg/L as CaCO3, conductivity ⫽ 108–149 µmhos/cm, and, for all tests, pH ⫽ 7.10–8.42. For respective cladoceran and duckweed tests, means of total hardness ⫽ 169–219 and 83.9–85.8 mg/L as CaCO3, alkalinity ⫽ 88.2–127 and 57.0–59.0 mg/L as CaCO3, and conductivity ⫽ 481–694 and 224–226 µmhos/cm. Saltwater species were tested in filtered (10 micron) salt water from the West Passage of Narragansett Bay, RI. Salinity was 30–32 g/L and temperature was 20–25°C in all tests, except in the test with the winter flounder, which was conducted at 6°C. Mean DO values for individual
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acute and chronic tests were 4.5–5.1 mg/L for tests conducted at 20–25°C and 9.4–10.1 mg/L for tests conducted at 6°C.
Chemical Analytical grade fluoranthene (98% purity) was obtained from Aldrich Chemical Company (Milwaukee, WI) for freshwater tests and from Eastman Kodak Company (Rochester, NY) for saltwater tests.
Exposure Regime and Test Methods Freshwater acute and chronic tests were conducted at the University of Wisconsin–Superior, Superior, WI, and employed methods similar to those described by ASTM (1993). Acute tests with most species were flow-through and lasted for 96 h. Tests with the duckweed employed methods described by ASTM (1993) and Taraldsen and Norberg-King (1990). Organisms were not fed during the tests. A 48-h test with the cladoceran and a 96-h test with the amphipod (Hyalella azteca) were renewed daily. Chronic tests consisted of 21-day life-cycle tests with the cladoceran, which were renewed three times per week and 32-day early-life-stage, flow-through tests with the fathead minnow. Flowthrough tests were conducted with a continuous-flow diluter (Benoit et al. 1982). Fluoranthene stock was prepared by pumping uncontaminated Lake Superior water through a 22 ⫻ 500 mm glass column filled with fluoranthene-coated glass wool, which was used to provide five duplicated concentrations of fluoranthene with a dilution factor of 50%. Lake Superior water alone was used in all tests as a control. Acute tests consisted of 20 organisms per concentration. Each concentration in the life-cycle test with the cladoceran contained 10 individuals (one per replicate). Early-life-stage tests with the fathead minnow contained 60 individuals per treatment (15 per replicate). Saltwater tests were conducted at the US Environmental Protection Agency (EPA), Atlantic Ecology Division, Narragansett, RI, and employed methods similar to those described by ASTM (1993). Acute tests with most species lasted 96 h and were renewed daily. Static tests were conducted for 48 h with the coot clam and sea urchin and for 96 h with the winter flounder. Tests with the mysid were flow-through. Fluoranthene stocks were prepared in a mixture of acetone and triethylene glycol. Five concentrations with 50% dilution factors were prepared by diluting the fluoranthene stock with salt water. Both saltwater and solvent (4 µl acetone/L and 17 µl triethylene glycol/L) controls were employed. Thirty-one-day and 33-day life-cycle tests conducted with the mysid were flow-through and employed seven fluoranthene concentrations (in duplicate) and controls. Fifteen organisms were used per duplicate.
Light Regime Acute and chronic tests with freshwater species were conducted under both standard laboratory fluorescent light and UV light using a 12-h light and 12-h dark photoperiod. A combination of plant growth and cool-white bulbs provided a fluorescent light intensity of 581 ⫾ 140 lux. Fluorescent light intensity was measured with a Model 200 photometer (Photovolt, New York, NY). Ultraviolet light was provided by Q-Panel UVA-340 bulbs (The Q Panel Company, Cleveland, OH). The UV light intensity was measured with a model IL1700 radiometer (International Light Inc., Newburyport, MA) that was equipped with SUD038 (UVA) and SUD240 (UVB) detectors. In acute tests, UV light intensity was 359–587 µW/cm2 (UVA) and 63–80 µW/cm2 (UVB) for most tests and 783–850 µW/cm2 (UVA) and 104 µW/cm2 (UVB) in tests with the oligochaete and hydra. The UVA and UVB light intensities in chronic tests with the cladoceran and the fathead minnow were 283 and 47 µW/cm2 and 612 and 82 µW/cm2, respectively.
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Table 1. Toxicity of fluoranthene (in µg/L) to freshwater and saltwater species from acute tests conducted under fluorescent and UV light in the laboratory and outdoorsa Fluores./ Lab. UV
Lab. UV/ Outdoor UV
— 32
— —
—
⬎148
—
—
⬎2
—
—
78
—
Species
Stage/ Age (days)
Lab. Fluores.
Lab. UV
Outdoor UV
Freshwater tests Duckweed (Lemna minor) Hydra (Hydra americana)
2 Frond Nonbudding
⬎159b 2.2 (1.6–3.0) 1.2 (0.9–1.4) 82 (69–109) 1.6d (1.4–2.0) —
—c —
—
—
—
—
—
—
—
⬎110 7.7 (6.6–9.0) 12.2 (11.3–13.2) 12.3
— —
—
Oligochaete (Lumbriculus variegatus)
Adult
⬎166b 70 (57–87) ⬎178
Snail (Physella virgata)
Adult
⬎178
Cladoceran (Daphnia magna)
⬍1
Amphipod (Gammarus pseudolimnaeus)
Adult
Amphipod (Hyalella azteca)
7–14
Dragonfly (Ophiogomphus species) Rainbow trout (Oncorhynchus mykiss)
Nymph 30–50
108 (81–144) 44 (29–66) ⬎178 ⬎91
Fathead minnow (Pimephales promelas)
25–35
⬎212
Bluegill (Lepomis macrochirus)
Juvenile
⬎117
Adult Embryo-larval
⬎127e ⬎127d,e
Mysid (Mysidopsis bahia)
⬍1
Amphipod (Ampelisca abdita)
Juvenile
Grass shrimp (Palaemonetes species)
3
American lobster (Homarus americanus)
Larvae
Sea urchin (Arbacia punctulata)
Embryo-larval
31 (22–41) 67 (59–76) 142 (120–170) 317 (266–378) ⬎127d,e
Sheepshead minnow (Cyprinodon variegatus)
42
⬎127e
Inland silverside (Menidia beryllina)
21
Winter flounder (Pleuronectes americanus)
28
616 (541–702) ⬎188
Saltwater tests Polychaete worm (Neanthes arenaceodentata) Coot clam (Mulinia lateralis)
117d
— 2.8d (2.6–2.9) 1.4 (1.3–1.5) — 22 (20–23) 13 (11–16) 3.9d (3.8–4.0) 159 (132–190) 30 (28–33) 0.1 (0.09–0.1)
⬎12
— —
—
⬎17
—
—
⬎10
—
⬎45
— —
— — 1.7 (1.4–2.0) — 6.6 (6.2–7.1) 22 (18–26) 3.9d (3.8–4.1) 172 (147–201) 13 (11.3–15.0) —
—
22 —
0.8 —
6.5
3.3
24
0.6
⬎33
1.0
⬎0.8
0.9
21
2.3
⬎1,880
—
a
Values are 96-h LC50s (95% CL) except where specified Inhibition of growth (IC50) c Test not conducted or value not able to be calculated d 48-h LC50 e Solubility of fluoranthene in fresh water and salt water (May 1980) b
Acute and chronic tests with saltwater species were conducted under both laboratory fluorescent light and UV light using a 16-h light and 8-h dark photoperiod. Cool white and Q-Panel UVA-340 bulbs were used to provide fluorescent and UV light, respectively. Intensity of UV light in both acute and chronic saltwater tests was measured with a Macam Photometrics (Livingston, Scotland) model UV-103 radiometer. Because this meter had different detection wavelengths than the model IL1700 meter used in freshwater tests, side-by-side measurements were conducted using both meters to compare meter output. Four replicated measurements were taken with each meter at varying distances from the UV bulbs. For all distances, the ratio of measurements from the meters (IL1700/Macam UV-103) was 1.43 ⫾ 0.02 at UVA wavelengths, and 0.550 ⫾ 0.02 at UVB wavelengths and were used to correct light intensity readings for all saltwater exposures to
that of the model IL1700 radiometer. Both meters were periodically sent to the manufacturers for calibration. For saltwater laboratory tests, the intensity of UVA light was 465–724 µW/cm2 and UVB light was 68–109 µW/cm2. Acute toxicity tests using six saltwater species also were conducted outdoors at Narragansett, RI, between the months of June and September for comparison with tests conducted under laboratory UV light. Test methods for these tests were the same as those used in the laboratory. Light conditions outdoors, generally, were sunny during the experimental period. Outdoor UV light measurements were taken only in the tests with the mysid, grass shrimp, and inland silverside. Measured midday intensities in these tests were 1,273–2,660 µW/cm2 (UVA) and 76–182 µW/cm2 (UVB). Tests also were conducted under multiple UV light intensities using a combination of laboratory and
Comparative Toxicity of Fluoranthene
outdoor experiments to determine the effect of light intensity on the acute toxicity of fluoranthene to the mysid and the inland silverside. The lowest UVA level was achieved with cool white bulbs, which produced ⬃7 µW/cm2. A second intensity (64 µW/cm2) was obtained with Vita-Lite UV bulbs (Durotest Inc., Fairfield, NJ). Two other light intensities of 360 and 676 µW/cm2 were achieved by varying the height of Q-Panel UV bulbs from 26 to 45 cm from the water surface, respectively. Sunlight produced the highest UVA intensity level of 1,788 µW/cm2, which was obtained as the mean of measurements made outdoors at midday for the two tests.
Chemical Analysis Fluoranthene concentrations were measured in exposure water in all freshwater tests and are reported as mean concentrations. Fluoranthene was measured at 0, 24, 48, 72, and 96 h in flow-through acute tests and twice weekly in each replicate of the early-life-stage test with the fathead minnow. Both new and old solutions were measured (expressed as means of new solutions) during acute and chronic renewal tests. Fluoranthene measurements were made by reverse-phase high pressure liquid chromatography (HPLC). Samples were injected onto a LichroCart 125-4 cartridge column with LiChrospher100 RP-18 (5 µm) packing (E. Merck, Darmstadt, Germany). An isocratic elution was performed with acetonitrile:water (85:15, v/v) at a rate of 1.5 ml/min. A Waters M 490 UV detector set at a wavelength of 236 nm and a Spectra-Physics SP4270 Integrator (San Jose, CA) were used to quantify peak areas. Analytical stock solutions were obtained by dissolving fluoranthene in acetone (Mallinckrodt Inc., Chesterfield, MO). Spiked recoveries from all tests ranged between 86 and 104% and were used to correct measured concentrations. The mean detection limits were 0.49–5.0 µg/L for tests using fluorescent light and 0.13–2.1 µg/L in tests using UV light. In flow-through acute and chronic tests, the concentration of fluoranthene was maintained within 20% of the mean measured concentration. Fluoranthene loss between renewals in acute tests with the cladoceran was 40–50% in tests using fluorescent light and 54–80% in UV light tests. Fluoranthene concentrations decreased between renewals in chronic tests by approximately 70 to 85% in fluorescent light tests and up to 100% in UV light tests. Fluoranthene concentrations in saltwater tests with fluorescent light were measured in each treatment at the beginning and end of each acute test with the mysid and winter flounder and weekly during the life-cycle test with the mysid. Measurements were made using a Hewlett Packard 5890A GC with a flame ionization detector and a 30 m DB-5 capillary column and are reported as mean concentrations. The injection temperature was 270°C, and detection temperature was 325°C. The initial oven temperature, held for 1 min, was 120°C and was increased by 20°C per min to a final temperature of 315°C and held for ⬃4 min. Fluoranthene standards and internal standards of pyrene were used to correct measured concentrations for recoveries, which averaged between 60 and 70%. Samples from the life-cycle test with the mysid were flash-evaporated (up to 1,000:1) prior to injection on the GC. The detection limit for the method was 0.1 µg/L. Fluorimetry was used to measure fluoranthene in acute UV light tests with the mysid and winter flounder and in a UV life-cycle test with the mysid. Fluoranthene concentrations in other saltwater UV light tests were not measured. Samples to which internal standards were added were extracted and reduced 500:1 (v/v). Freon solvent was exchanged with hexane. A Farrand Spectrofluorometer, MK1 (Farrand Optical Co., Valhalla, NY) was used at excitation and emission wavelengths of 285 and 450 nm, respectively. Recoveries of matrix spikes in samples averaged 78%. The limit of detection for this method was 0.06 µg/L. Fluoranthene concentrations in the winter flounder static acute test decreased up to a factor of 4 from the beginning of the test to the end.
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Statistics Median effect and lethal concentrations were calculated using the trimmed Spearman-Karber method (Hamilton et al. 1977), and regression analysis was used to determine the relationship between acute lethality and UV light intensity. Chronic test results were checked for normality and homogeneity of variance and were analyzed with one-way analysis of variance using standard statistical procedures (US EPA 1994). Percentage data were first transformed using arc sine square root. Significant statistical differences relative to the controls were subsequently identified at p ⱕ 0.05 (US EPA 1994). Chronic effects on saltwater organisms were compared against combined controls (with and without solvent) because statistical differences in responses between the two control groups were not observed. Chronic values, used to estimate threshold effect concentrations, were calculated as the geometric mean of the no observed effect concentration (NOEC) and the lowest observed effect concentration (LOEC).
Results Acute Tests Results of acute lethality tests conducted with 21 species are shown in Table 1. Control survival in all tests were acceptable according to ASTM (1993) for the species and methods stated. In fluorescent light tests, LC50s were ⬍100 µg/L for the hydra, two amphipods (Hyalella azteca and Ampelisca abdita) and the mysid and between 108 and 142 µg/L for the amphipod (Gammarus pseudolimnaeus), cladoceran, and the grass shrimp. The LC50s calculated for the American lobster and inland silverside were ⬃ two to five times higher than the saltwater solubility limit for fluoranthene. Fluoranthene was not lethal to several freshwater species, including the duckweed, oligochaete, snail, dragonfly, rainbow trout, fathead minnow, and bluegill, or to several saltwater species, including the polychaete worm, coot clam, sea urchin, sheepshead minnow, and winter flounder, within the limit of fluoranthene’s solubility. However, behavioral effects including reduced response to probing or loss of equilibrium were observed at lower concentrations than those causing lethality for some freshwater species, including the amphipod (G. pseudolimnaeus), rainbow trout, fathead minnow, and bluegills. Median effect concentrations for these species, based on these observations, were 43, 26, 69, and 44 µg/L, respectively. Fluoranthene was much more acutely lethal to most freshwater and saltwater species in the presence of UV light. The LC50s were ⱕ159 µg/L for 15 of the 21 species tested and ranged from ⬎0.8 to ⬎1,880 times lower than values observed in fluorescent light tests (Table 1). Median lethal concentrations could not be calculated for freshwater amphipods, a saltwater amphipod, or a saltwater polychaete. Comparisons of acute values illustrated that UV light, generally, affected sensitive freshwater and saltwater species to a similar degree. Ratios of LC50s obtained for most fluorescent and UV light tests ranged from ⬎2 to 78 for freshwater species and ⬎0.8 to ⬎45 for saltwater species. Greater ratios (⬎148 to ⬎1,880) were observed for the oligochaete and the winter flounder, which were the most sensitive species to fluoranthene under UV light. In contrast, fluoranthene was not lethal to the duckweed and dragonfly under UV light, although many dragonflies appeared paralyzed with curved spines at all concentrations tested. An EC50 for dragonflies based on these observations was ⬍19 µg/L.
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Chronic Tests Chronic toxicity tests using fluorescent light were conducted with fluoranthene and the cladoceran, fathead minnow, and mysid. Control survival in all tests were acceptable according to ASTM (1993) for the species tested. Survival effects on the cladoceran were not observed at 6.9, 17, and 73.2 µg/L, but all animals were dead at the highest concentration of 148 µg/L after 21 days. In contrast, lower concentrations of 35.3 and 73.2 µg/L significantly reduced growth (total length) relative to the controls after this time. Reproduction effects were not observed at the concentrations tested. An early-life-stage test conducted with the fathead minnow showed that 21.7 µg/L significantly reduced survival 67%, length 27%, and dry weight 50% compared to the controls after 32 days. Statistical effects were not observed at ⱕ10.4 µg/L. In the 31-day life-cycle test with the mysid, the highest concentration of 18.8 µg/L completely inhibited reproduction, but did not reduce survival or growth (dry weight). Significant adverse effects were not observed at concentrations ⱕ11.1 µg/L. Chronic toxicity tests for the same species were conducted in the presence of UV light. Significant adverse effects were not observed in the 21-day test with the cladoceran, at test concentrations of ⬍0.3, 0.6, and 1.4 µg/L, although the mean number of young per surviving adult was reduced 21% to 25% in the highest two concentrations, respectively, compared to the controls. An early-life-stage test with the fathead minnow showed that the two highest concentrations of 4.8 and 7.3 µg/L significantly reduced survival 97 to 100%. Adverse effects were not observed at concentrations of ⱕ1.4 µg/L. In the life-cycle test with the mysid, statistical differences from the control were not observed at test concentrations ⱕ0.6 µg/L. The NOECs, LOECs, and chronic values for tests with the cladoceran, fathead minnow and mysid are shown in Table 2. Chronic values for fluorescent light tests with these species ranged from 14.4 to 24.5 µg/L. Definitive chronic values could not be calculated for the cladoceran and mysid under UV light. However, test results indicated that significant mortality would likely have occurred if concentrations were only slightly greater. In acute UV light tests, mortality was 80% for the cladoceran at 2.4 µg/L and 70% for the mysid at 1.5 µg/L; concentrations that were only 1.7 and 2.5 times higher than the NOECs of 1.4 and 0.6 µg/L measured in chronic tests for these species, respectively. The acute LC50 was 1.6 µg/L for the cladoceran and 1.4 µg/L for the mysid. Both acute and chronic tests with each species were conducted by the same investigators in the same water with the same aged organisms, and test concentrations between experiments overlapped to provide a continuum of measured water concentrations. Since the LC50s for these species were the lowest effect concentrations in this treatment series, they were used along with the NOECs to estimate, for comparison purposes, chronic values of 1.4 µg/L for the cladoceran and 0.9 µg/L for the mysid. These values are close to the chronic value of 2.6 µg/L calculated for the fathead minnow when exposed to fluoranthene and UV light.
UV Light Intensity Comparison of results from tests conducted with laboratory and outdoor UV light are shown are shown in Table 1. The LC50s
R. L. Spehar et al.
Table 2. Effect endpoints and concentrations (in µg/L) from chronic fluoranthene tests conducted under fluorescent and UV light conditions Fluorescent Species
UV
Chronic Chronic NOECa LOECb Valuec NOEC LOEC Value
Cladoceran (Daphnia magna) 17.0 Fathead minnow (Pimephales promelas) 10.4 Mysid (Mysidopsis bahia) 11.1
35.3
24.5
1.4d
1.5d
1.4d
21.7
15.0
1.4
4.8
2.6
18.8
14.4
0.6d
1.4d
0.9d
a
No observed effect concentration Lowest observed effect concentration c Geometric mean of NOEC and LOEC d Estimated (see text) b
for these test conditions were not significantly different for the mysid, American lobster, sea urchin, and sheepshead minnow (95% CL overlapped; laboratory UV/outdoor UV ratios ranged from 0.8–1.0), but were two to three times lower in outdoor tests than with the inland silverside and grass shrimp. Results of acute tests conducted under multiple UV light intensities with the mysid and the inland silverside illustrated that fluoranthene toxicity was directly related to light intensity (Figure 1). The LC50s for the mysid were 58, 12, 12, 2.8, and 1.7 µg/L and for the inland silverside were 620, 103, 49, 30, and 13 µg/L at UV light intensities of 7, 64, 360, 676, and 1,788 µW/cm2, respectively. Regression analysis indicated that this relationship was linear and the slopes were the same (not statistically different) for both species (Figure 1). The mysid was ⬃10 times more sensitive to fluoranthene than the inland silverside within this range of UV light intensity.
Discussion The results of this study show that fluoranthene was, generally, much more toxic in the presence of UV light than fluorescent light. For most species, acute lethality increased by an order of magnitude, which is similar to the degree of effect observed in UV light tests conducted with fluoranthene by other investigators (Ankley et al. 1995; Boese et al. 1997; Pelletier et al. 1997), although greater increases of two to three orders of magnitude were observed for the oligochaete and winter flounder (the most sensitive freshwater and saltwater species tested under UV light). UV light also increased the chronic toxicity of fluoranthene by approximately an order of magnitude, but the degree of this effect is uncertain due to the limited database. No other information was found in the literature on the effects of UV light on chronic fluoranthene toxicity to aquatic organisms for comparison. UV light changed the relative acute sensitivity of some species to fluoranthene. Under fluorescent light, fluoranthene was not lethal to lower invertebrate forms and fish, but in the presence of UV light, their sensitivity approached or was below that of some sensitive crustaceans. Reasons for such shifts in sensitivity are not clear, but, in part, may be related to the species inability to detoxify phototoxic reactions through
Comparative Toxicity of Fluoranthene
Fig. 1. Median lethal concentrations (LC50s) of fluoranthene versus UVA light intensity for the inland silverside (open circles) and mysid (solid circles). Error bars represent 95% CL
biochemical mechanisms including enzyme repair (Larson and Berenbaum 1988; Greenberg et al. 1997). Organisms such as the oligochaete and winter flounder became extremely sensitive to fluoranthene under UV light because they probably lack well-developed defense mechanisms due to their natural association with bottom sediments, where they are usually shielded from UV exposure. In contrast, morphological attributes such as the thick integument of the duckweed and snail likely contributed to their relative tolerance in UV light tests. Previous test results regarding the phototoxic effects of fluoranthene to aquatic organisms have often involved acute exposures at UV light intensities that are much less than that of natural sunlight at its peak intensity (Newsted and Giesy 1987; Diamond et al. 1995; Ankley et al. 1995, 1997; Boese et al. 1997; Swartz et al. 1997; Monson et al. 1999). In the current study, the acute toxicity of several saltwater invertebrate and fish species was compared under higher laboratory UV light intensities (up to 723 µW/cm2 UVA and 109 µW/cm2 UVB) and UV sunlight intensities of up to 2,660 µW/cm2 UVA and 182 µW/cm2 UVB. Fluoranthene was similarly toxic under both light conditions to four species, but was two to three times more toxic to the inland silverside and grass shrimp tested outdoors. The similarity in organism sensitivity observed from these experiments suggest that laboratory tests may be useful in predicting field effects that are attributable to solar radiation. However, the greater effects observed for two species tested under natural sunlight suggests that higher light intensities can increase phototoxicity. Present experiments with the mysid and inland silverside and studies with other organisms (Ankley et al. 1995; Monson et al. 1999) support this observation. Toxicity to the mysid and inland silverside increased ⬃30 to 50 times with increased UVA light intensity (Figure 1). Regression analysis showed that this relationship was linear. Studies with oligochaetes (Ankley et al. 1995) and frogs (Monson et al. 1999) illustrated that time-dependent mortality due to fluoranthene was linear with respect to the product of initial tissue concentration and UV light intensity. Slopes of the regressions in these tests and those in the present study were similar, indicating that phylogenetically different species responded similarly to fluoranthene and increased UV light intensity.
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The present results and those from other studies indicate that hazard assessments of phototoxic compounds like fluoranthene should take into account effects from UV light. At this time, however, the degree of this effect in aquatic ecosystems is uncertain due to environmental factors that might enhance or mitigate UV light penetration to the site of toxic action in the organism. For example, dissolved organic carbon concentrations ⬎2 mg/L significantly attenuate UV light in fresh and salt water, but at concentrations of ⬍1–2 mg/L, UV light can penetrate a substantial portion of the depth of lakes, and in coastal marine areas, may approach 20 m (Williamson et al. 1996). Organisms in aquatic environments containing low levels of dissolved organic carbon and high levels of phototoxic compounds may be at higher risk because of the potential for greater exposure to highly toxic conditions. Although many organisms have developed protective mechanisms against UV light, behavioral, morphological, or biochemical responses may not afford enough protection when organisms contaminated with fluoranthene and other phototoxic PAHs cannot avoid UV light, such as during low stream flows and tides or when they must migrate up the water column to feed. The present outdoor experiments and previous studies by Ankley et al. (1995, 1997) and Monson et al. (1999) suggest that organisms may be more at risk when exposed to phototoxic compounds at higher UV light intensities. These authors conclude that hazard assessments such as water quality criteria for such compounds should define potential toxicity as a function of the product of accumulated dose and UV light intensity and could be based on models that are applied to data from tests using single UV light intensities. We agree with this conclusion and concur that further study is needed to determine the effects of UV light intensity in long-term exposures with a variety of species and phototoxic compounds. At this time, it is not clear if the determined linear relationship between toxicological effects versus light intensity can be extrapolated for long-term exposures of such compounds.
Acknowledgments. This manuscript has been reviewed in accordance with EPA policy. Mention of trade names does not imply endorsement by the US government.
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