- Email: [email protected]
Application of the luciferase cell culture bioassay for the detection of refined petroleum products
Marine Pollution Bulletin 44 (2002) 983–991 www.elsevier.com/locate/marpolbul
Reports
Application of the luciferase cell culture bioassay for the detection of refined petroleum products Michael H. Ziccardi
a,*
, Ian A. Gardner b, Jonna A.K. Mazet
a,b
, Michael S. Denison
c
a
b
Wildlife Health Center, School of Veterinary Medicine, University of California, Davis, CA 95616, USA Department of Medicine and Epidemiology, School of Veterinary Medicine, University of California, Davis, CA 95616, USA c Department of Environmental Toxicology, University of California, Davis, CA 95616, USA
Abstract A luciferase cell culture-based bioassay, developed to detect 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-like activity of halogenated and polycyclic aromatic hydrocarbons, was optimized to detect refined petroleum products and to determine their relative inducing potency. Quality control standards from 32 refined products (gasolines and diesels, jet fuels, lubricating oils, fuel oils and weathered products) and three commercial products were evaluated. Induction equivalents (I-EQs) were determined by direct comparison of the EC50 and EC20 values (based on the median and 20% TCDD maximal response, respectively) from dose-response curves for each product to those obtained with TCDD. Most petroleum products were active in the luciferase bioassay, with those products composed of fractions produced later in the distillation process (i.e. fuel oils) inducing higher levels. Additionally, weathering of products reduced their induction potency. Based on the high I-EQ estimates of many products, biological effects associated with exposure may have been previously underestimated using other diagnostic methods. Ó 2002 Elsevier Science Ltd. All rights reserved. Keywords: Petroleum; PAH; Oil spills; CYP1A1; Luciferase; Bioassay
1. Introduction Petroleum contamination of the environment has long been known to cause environmental damage, as well as acute and chronic effects to exposed organisms. Although acute effects of such exposures are well known (NRC, 1985; Jessup and Leighton, 1996), several intrinsic factors of petroleum products make their detection at low levels in the environment problematic, and the assessment of toxicity associated with chronic exposure difficult to establish. Crude and refined petroleum products are extremely complex mixtures of organic compounds (including alkane, alkene, naphthalene, and monocyclic and polycyclic aromatic hydrocarbon (PAH) components), the composition of which varies widely among products, fields, and refineries, and even within a single well (Hunt, 1979). Accordingly, the development of detection methods and identification schemes to determine product source, as *
Corresponding author. Tel.: +1-530-754-5701; fax: +1-530-7523318. E-mail address: [email protected] (M.H. Ziccardi).
well as the subsequent establishment of the toxicity associated with exposure to a specific product can vary widely among studies. An additional complicating factor is that the composition of these products, once introduced to the environment, is not static. Spilled or released oil undergoes a complex ‘‘weathering’’ process, during which the product undergoes evaporation, dissolution, dispersion, photo-oxidation, and additional breakdown/modification due to other biological, chemical and physical methods (NRC, 1985). Consequently, the chemical makeup of petroleum products in the environment and their associated toxicity can vary dramatically depending on the timing of exposure. Numerous analytical methods have been developed to determine the presence of, or exposure to, petroleum products in the marine environment. Direct methods, such as gas and liquid chromatography, have been extensively used to quantify levels of PAHs and other compounds within petroleum products and to ‘‘fingerprint’’ such products in either environmental samples (Short et al., 1996) or biological matrices (Varanasi, 1989). These methods, however, are often prohibitively expensive, require extensive time-consuming extraction
0025-326X/02/$ - see front matter Ó 2002 Elsevier Science Ltd. All rights reserved. PII: S 0 0 2 5 - 3 2 6 X ( 0 2 ) 0 0 1 2 8 - 5
984
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
methods and may require the collection of large quantities of often difficult to acquire tissue samples. Additionally, because they are designed to detect specific parent components or their breakdown products (often limited to those compounds known to cause morbidity in experimental exposure trials), direct methods can fail to detect entire classes of compounds that can produce similar biological/toxicological responses. Results from direct methods may therefore provide an inaccurate estimate of the organismal effects associated with the overall exposure to the product. Bioassay systems, another class of detection methods, are based on the ability of a petroleum product to produce a biological response in a living organism. One of the most widely used of these methods is the quantitation of cytochrome P4501A1 (CYP1A1) induction in tissues. This xenobiotic detoxification system is mediated by the Ah receptor (AhR), a soluble intra-cellular protein that binds PAHs and related chemicals with high affinity (Safe, 1990; Denison et al., 1998). In addition, this P4501A1 protein has been implicated as the ‘‘cellular switch’’ responsible for initiating the cascade of events leading to the specific toxic responses associated with exposure to these compounds (Safe, 1990; Fernandez-Salguero et al., 1996). Chemical-dependent activation of the AhR results in a sequence of events in the nucleus that stimulates CYP1A1 gene transcription and increases cellular concentration of P4501A1 protein and its associated monoxygenase activity, ethoxyresorufino-deethylase (EROD). The increase in CYP1A1-dependent catalytic activity most commonly results in enhanced detoxification of the inducing chemicals. Given this relationship, the quantitation of CYP1A1 and EROD has been used to evaluate the relative biological and toxicological potency and as a biomarker of exposure to PAHs and petroleum products (Addison et al., 1986; Stegeman et al., 1991). Although quantification of CYP1A1 activity may serve as an excellent ‘‘bioassay of effect’’ in potentially exposed organisms, intra- and inter-species differences in AhR ligand binding specificity, responsiveness and sensitivity to these chemicals, and the ability of the AhR to be activated by non-PAH/HAH-type chemicals complicate the interpretation of the results (Denison and Wilkinson, 1985; Garrison et al., 1996; Kikuchi et al., 1996; Denison et al., 1999). These analyses also provide no information as to which chemicals were responsible for the induction response, nor the level of exposure in vivo. Additionally, these methods, as with the analytical chemistry methods discussed above, often require sampling protocols that include the harvesting of tissues (i.e. liver) rich in CYP1A1 activity to achieve acceptable detection limits. Lastly, the documented ability of EROD activity to be competitively inhibited at high pollutant levels (as would be expected in samples from oil spill events) would also result in an underestimation
of the amount of inducing compounds (Kennedy et al., 1993). Recently, several recombinant cell bioassay systems using luciferase reporter genes under control of CYP1A1 upstream regulatory elements (including the promoter and/or dioxin responsive elements (DREs)) have been developed for the detection and relative quantitation of AhR-active chemicals and compounds through co-activation of luciferase and CYP genes (Postlind et al., 1993; Garrison et al., 1996; Murk et al., 1996). Induction of luciferase in these cell lines (measured by quantification of bioluminescence) occurs in a time-, dose- and AhR-dependent, and chemical-specific manner. Not only has this assay been shown to be more analytically sensitive than EROD-based bioassays, but also it lacks the majority of the limitations inherent in those systems. To date, most studies using these new luciferasebased bioassay systems have focused predominantly on the analysis of HAHs (Garrison et al., 1996; Murk et al., 1996; Sanderson et al., 1996) and/or PAHs (Jones and Anderson, 1999; Ziccardi et al., 2000). Also, few have examined their utility in the evaluation of complex mixtures, such as petroleum products. Here we describe studies in which this system has been optimized and validated for the detection and relative quantitation of AhR-active chemicals in complex petroleum products.
2. Materials and methods 2.1. Chemicals All petroleum products were considered hazardous and appropriate personal protective methods and materials were used throughout experiments. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) was obtained from Dr. S. Safe (Texas A&M University) and was dissolved in dimethyl sulfoxide (DMSO) from Aldrich Chemicals (Milwaukee, WI). All petroleum standards used in this study are listed in Tables 1–3 and were purchased from Accustandards Co. (St. Louis, MO). Whole petroleum products, listed in Table 3, were either purchased from local vendors or provided by industry representatives. Cell culture media and antibiotics were purchased from Life Technologies (Gaithersburg, MD), fetal bovine serum from Atlanta Biologicals (Atlanta, GA), luciferase substrate and assay buffers from Promega (Madison, WI) and fluorescamine from Molecular Probes (Eugene, OR). 2.2. Recombinant cell line Recombinant mouse hepatoma (H1L1.1c2) cells, containing the stably integrated PAH/HAH-inducible luciferase expression vector pGudLuc1.1, were described previously (Garrison et al., 1996). This vector contains
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
985
Table 1 Luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, gasoline and aviation fuel products calculated from dose-response curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Gas, regular leaded Gas, regular unleaded Gas, premium Gas, oxygenate free Diesel Diesel #1, low SO4 Diesel #2, low SO4 Diesel, Arctic Kerosene
8:28 10 4:81 106 6:13 106 5:85 105 4:56 105 8:15 105 1:25 107 4:42 106 NCa 4:25 105
1 10 0.172 0.135 1.414 1.816 1.015 0.066 0.187 NC 1.948
1:88 10 8:94 105 2:54 105 6:13 104 4:91 104 5:94 104 3:56 105 8:95 104 NC 2:73 105
1 106 0.210 0.739 3.064 3.828 3.162 0.528 2.100 NC 0.688
Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel,
NC NC NC NC NC NC NC NC NC
NC NC NC NC NC NC NC NC NC
NC NC 1:43 106 2:94 104 NC NC NC NC 1:50 105
NC NC 0.131 6.381 NC NC NC NC 1.255
a
JP 4 JP 5 JP 7 JP 8 JP 10 JP TS Jet Ref. aviation turbine
6
Not calculable due to induction less than 50% or 20% of TCDD.
Table 2 Calculated luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, motor oil and fuel oil products calculated from doseresponse curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Oil, SAE30W Oil, SAE40W Oil, SAE50W Oil, SAE5W30 Oil, SAE10W30 Oil, SAE10W40 Oil, SAE20W50 Oil, Hydraulic Jack
8:28 10 8:16 105 1:08 106 8:11 105 1:67 107 4:65 107 1:63 106 1:52 107 1:28 106
1 10 1.014 0.767 1.020 0.050 0.018 0.508 0.054 0.649
1:88 10 1:11 105 2:37 105 1:71 105 2:68 106 7:29 106 1:38 105 1:74 106 1:94 105
1 106 1.686 0.793 1.096 0.070 0.026 1.365 0.108 0.966
Fuel Fuel Fuel Fuel Fuel Fuel
NCa NC 8:73 104 7:15 103 4:32 103 1:17 104
NC NC 9.483 115.8 191.5 70.91
NC 2:41 105 1:08 104 1:38 102 2:41 102 4:67 102
NC 0.778 17.47 1361 780.5 402.6
a
oil oil oil oil oil oil
#1 #2 #3 #4 #5 #6
6
Not calculable due to induction less than 50% or 20% of TCDD.
the firefly luciferase gene under PAH/HAH-inducible and AhR-dependent control of four DREs. 2.3. Sample analysis Dilutions of all products (10-fold dilutions (v/v) from original stock) were done using Optima-grade methanol (MeOH) obtained from Fisher Scientific (Pittsburgh, PA) or, in the case of the #6 fuel oil and 10W40 motor oil, first suspended in Optima-grade hexane (Fisher Scientific) to facilitate solubilization of the heavier
products, then diluted in MeOH and resuspended through pipetting and vigorous agitation. All petroleum standards were analyzed during a single period, with appropriate positive (1 nM TCDD) and negative (carrier solvent alone) controls run on each 96-well plate. A TCDD dose-response curve was also included in each experiment. The overall protocol for the analysis of samples using the 96-well luciferase bioassay format has been previously described (Ziccardi et al., 2000). Briefly, 100 mm plates of stable cell clones were trypsinized, resuspended in 10 ml of culture media, and cells diluted
986
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Table 3 Calculated luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, weathered unleaded gasoline and whole commercial petroleum products calculated from dose-response curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Gas, 0% weathered Gas, 25% weathered Gas, 50% weathered Gas, 75% weathered
8:28 10 NCa NC NC NC
1 10 NC NC NC NC
1:88 10 8:54 103 6:29 104 NC NC
1 106 21.99 2.984 NC NC
Gas, regular unleaded Oil, SAE10W40 Fuel oil #6
8:33 106 1:13 107 1:63 104
1.061 0.784 540.9
2:15 106 3:52 106 2:83 103
1.743 1.064 1324
a
6
Not calculable due to induction less than 50% or 20% of TCDD.
to 300,000 cells per ml. Aliquots (200 ll) of this suspension were added to wells in rows two through eight of sterile 96-well white CulturPlatesTM (Packard Instruments; Meriden, CT), and plates were incubated for 24 h prior to ligand addition, allowing cells to reach confluence. All wells were washed with 100 ll of sterile 1X phosphate-buffered saline (PBS), followed by addition of ligand in DMSO or methanol/hexane in MEM (final DMSO/methanol concentration was 1%), with 1 nM TCDD and either DMSO, methanol or hexane as positive and negative controls, respectively. Plates were then incubated for three hours at 37 °C, after which cells were washed twice with PBS, followed by the addition of Promega Lysis Buffer (25 ll) to each well, and plates were shaken at room temperature until cells were lysed (20 min). Luciferase activity was measured using an automated microplate luminometer (Dynatech ML3000; Chantilly, VA) in enhanced flash mode with 10 s delay and 10 s integration periods after the automatic injection of 50 ll of Promega stabilized luciferase reagent. Correction for intra-plate variation was accomplished through protein quantification using fluorescamine (at 500 lg/ml) in acetonitrile and detection with a multiplate fluorometer (Fluostar; SLT, Salzburg, Austria) at 400 nm excitation and 460 nm emission wavelengths. Inter-plate variation was corrected for by standardization of each plate to the luciferase activity induced by 1 nM TCDD. Final results were expressed as relative light units per mg protein. 2.4. Statistical methods Descriptive statistics and comparative analyses (Student’s t-test between samples and controls to determine significant induction above background) were calculated using Excel XP (Microsoft, Redmond, WA). Differences within petroleum classes and/or comparisons between standards and whole products were determined using multivariate analysis of variance (MANOVA) methods, with subsequent pairwise analysis (Tukey’s LSD and Scheffe’s tests) done between replicate wells, chemical
concentrations, petroleum products and/or product classifications using Statistica (Statsoft, Tulsa, OK) statistical software. Median (EC50 ) and 20% (EC20 ) effective concentrations of maximal TCDD response (for those chemicals showing induction equivalent to or greater than these levels) were calculated by fitting doseresponse data to a four-parameter Hill model by least squares using SigmaPlot (SPSS, Chicago, IL) as described previously (Ziccardi et al., 2000). Because the analyzed samples represent complex chemical mixtures, induction equivalency values (I-EQs) were expressed as micrograms of TCDD equivalents per gram of product and were calculated by dividing the product’s EC value by that calculated for TCDD.
3. Results To maximize the utility of this assay system for the rapid detection of PAHs and like compounds, prior optimization of the protocols has emphasized maximizing throughput in the system. Through the integration of a 96-well format, selection of a reporter system/ cell line that shows maximal response to PAHs at a three hour incubation period, minimization of lag and read times in the luminometer, and integration of several inter- and intra-plate correction methods (Ziccardi et al., 2000), we now have the ability to rapidly and accurately assess the ability of many different products and/ or samples to induce luciferase activity. Accordingly, the inducing activity of all petroleum controls in these experiments was assessed within a single sampling period, allowing for direct comparison using a single TCDD dose-response curve and maximizing our confidence in the subsequent I-EQ estimations. The luciferase bioassay detected activity in most petroleum products when added at elevated concentrations (above 1 105 picograms per well), with those compounds produced later in the distillation process generally resulting in higher levels of induction (Fig. 1A). The lighter grade products (which included gasolines and
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Fig. 1. Dose-response curves for induction of luciferase activity by (A) gasoline and diesel distillates, and (B) jet fuel standards. H1L1.1c2 cells in 96-well microplates (at 80–100% confluence) were incubated with TCDD or gasoline distillate standards at the indicated concentrations for three hours, after which luciferase activity (corrected for protein concentration) was determined as described in Section 2. Values represent the means S:D: of triplicate determinations.
diesel fuels) were moderate inducers with similar induction profiles, with the exception of arctic diesel. The median effective concentrations for these mixtures ranged from 1:25 107 to 4:25 105 pg/well, with I-EQ50 values ranging from 0.066 to 1.816 lg TEQ/g product (Table 1). In contrast, dose-response relationship studies using aviation fuels revealed increased induction responses at low concentrations, followed by declining luciferase levels at the highest concentrations tested (Fig. 1B). Several products reached 20% maximal TCDD luciferase expression levels, most notably that of JP-8, which produced significant induction levels (EC20 ¼ 2:94 104 pg/well) and a relatively large I-EQ20 (6.381 lg/g) (Table 1). Motor and lubricating oils caused levels of induction similar to those of gasoline and diesel compounds (Fig. 2A). All of these products showed approximately the same dose-response pattern, with EC50 values calculated from dose-response curves ranging from 4:65 107 to 8:11 105 pg/well and I-EQ50
987
Fig. 2. Dose-response curves for induction of luciferase activity by (A) lubricating oils, and (B) fuel oil distillate standards. H1L1.1c2 cells in 96-well microplates (at 80–100% confluence) were incubated with TCDD or gasoline distillate standards at the indicated concentrations for three hours, after which luciferase activity (corrected for protein concentration) was determined as described in Section 2. Values represent the means S:D: of triplicate determinations.
values from 0.018 to 1.020 lg/g product (Table 2). Lastly, fuel oils elicited a much greater overall response than the lighter control compounds (Fig. 2B). Additionally, there was a trend towards an increasing response as the products increased in viscosity, ranging from minimal response in fuel oil #1 to an EC50 value of 7:15 103 pg/well for fuel oil #4 (Table 2). However, as noted in the aviation fuels, induction levels decreased at the highest tested concentrations. In an attempt to determine the effect of progressive environmental degradation on the bioactivity of fuels, regular unleaded gasoline that had been ‘‘weathered’’ to 25%, 50% and 75% levels was analyzed. As shown in Table 3, the two less weathered products produced significant luciferase responses (with EC20 values of 8:54 103 and 6:29 104 , and I-EQ values of 21.99 and 2.984 lg/g), and the two more weathered samples produced little luciferase response. Unlike the activity seen in the other gasoline standards, however, luciferase
988
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
activity in the weathered products declined at the highest two concentrations. To determine whether the responses noted in the petroleum quality control standards were comparable to those evident in commercially available petroleum distillates, dose-response analyses were also carried out using regular unleaded gasoline, 10W40 motor oil and #6 fuel oil acquired either from local retailers or industry representatives. As seen in Table 3, appreciable luciferase response was noted for all three products, corresponding to EC50 values of 8:33 106 , 1:13 107 , and 1:63 104 , respectively. While the response patterns for each of these products were comparable to that observed with the appropriate petroleum standards, subsequent EC50 calculations for the whole products were several times higher for both the fuel oil and gasoline samples (Tables 1 and 2). Further statistical analyses showed significant overall differences between each product and its associated standard (MANOVA, P values for all effects <0.05) due to significant differences in the highest two concentrations for gasoline and #6 fuel oil, and the second and third highest concentrations for motor oil (LSD and Scheffe tests; all P values <0.05).
4. Discussion We used our recombinant luciferase cell culture bioassay to examine the ability of various refined petroleum products to activate the AhR signaling pathway and to estimate their relative biological/toxicological potency. Light and heavy gasoline fractions (comprised of gasoline, kerosene and diesel fuels) are largely made up of the first hydrocarbons distilled during the refining process and, as such, contain a very high proportion (25–40%) of smaller aromatic compounds (AFCEE, 1999). We observed that most of these different fuel mixtures produced similar induction profiles, with high induction levels noted at the higher concentrations (Fig. 1A; Table 1). At first glance, these data may seem to be contrary to our previous work that showed little luciferase induction caused by smaller PAH constituents (data not shown) and work of others that reported little CYP induction associated with smaller aliphatic hydrocarbons (Neff, 1979). However, these mixtures are often comprised of over 300 different hydrocarbon compounds (Barber et al., 1996), and there may be a significant proportion of larger-ringed structures in these mixtures that are responsible for the increased luciferase response. While the luciferase bioassay appears to work well for most of the gasoline and diesel fuels, arctic diesel, a product that includes additives necessary to remain fluid at extremely low temperatures, showed initial increased luciferase activity, followed by decreased induction at the highest concentrations (Fig. 1A). Similar declines in induction were noted for all aviation fuels, products
largely composed of gasoline, kerosene and additional additives necessary to satisfy specific requirements of the aircraft to be fueled (Fig. 1B). Indeed, only three of these products induced to 20% of the maximal TCDD response: JP-7 (a special non-distillate fuel designed for the SR-71 aircraft), JP-8 (a newer primary fuel used by the United States Air Force) and turbine fuel (also known as Jet A in commercial applications). At this 20% level, however, these products induced in a similar fashion to gasolines and produced fairly large I-EQ20 values (Table 1). The observed pattern in luciferase activity with these products may be attributable to specific additives in these products that may differentially affect the bioassay at the higher concentrations. Although it is possible that these compounds are toxic to the cells at these higher concentrations, no toxicity was observed by visual inspection of the cells prior to luciferase analysis. However, inhibition of cellular processes and/ or the AhR induction pathway by components in the fuels, without any apparent cell toxicity, is plausible. The ability of chemical components in these fuels to act as AhR antagonists and/or repress the AhR-dependent induction response via indirect pathways (Chen and Tukey, 1996; Long et al., 1998) could also account for the overall lower potencies of some products. Prior studies have identified a variety of PAHs, HAHs and organic compounds (such as organotins and nitroflavones) that reduce AhR activity either partially or completely (Santostefano et al., 1993; Lu et al., 1995; Bruschweiler et al., 1996a,b; Willett et al., 1998). The analysis carried out here only examined the agonist activity of these products; analysis of their abilities to act as effective antagonists would require testing of their ability to reduce TCDD activity. Chemicals in these fuels could also reduce luciferase activity directly via quenching of luciferin-generated light production, as has occurred in the analysis of a number of commercial products (Velazquez and Feirtag, 1997). Future work should focus on analysis of these additives in order to better understand this loss in bioactivity at elevated concentrations. Heavier residual petroleum products were also evaluated using this bioassay system. Motor oils, residual products containing very low levels of aromatics (1– 2%) (Potter and Simmons, 1998) showed induction at concentrations above 1 104 pg/well (Fig. 2A). Excluding 10W40, there was a statistically significant difference between single-weight and multi-weight products at concentrations above 1 104 pg/well (MANOVA; P values for all effects <0.05), with single-weight compounds producing higher luciferase induction. It is possible that this difference may be attributable to the presence of polymers used in multi-weight products causing decreases in luciferase production and/or detection in a similar fashion to that observed with jet fuels.
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Heating fuel oils, used in a variety of different applications where a clean burning, non-explosive fuel is required (such as in furnaces, heaters and ship boilers), were also assessed using this system. These products are graded from #1 to #6 based on viscosity, with fuel oil #1 (or a kerosene derivative) being the lightest and fuel oil #6 (or bunker C) being comprised largely of the residuum from the distillation process. Most of these compounds caused extremely large responses in the luciferase bioassay, with the response increasing with increasing viscosity (Fig. 2B). For example, for the three heaviest fuel oil products, relatively large I-EQ50 values of 70–191 and I-EQ20 values of 402–1361 lg/g of product were calculated (Table 2). The sizable differences between the I-EQ50 measurements are due to the slope of the TCDD curve being much steeper than that of all other petroleum compounds, therefore the true estimate of I-EQ most likely lies somewhere between the 50% and 20% values. All heating fuel oils produced a decline in bioassay response at the highest concentrations, which may either be attributable to the mechanisms mentioned earlier or it may result from increased interference/quenching of the bioluminescence detection as a result of the darker color at the highest concentrations. Weathering of petroleum products is known to volatilize off smaller aromatics, leaving most of the heavier products in the residual mixture (NRC, 1985). To determine the ability of the bioassay to detect such compounds, we analyzed samples of regular unleaded gasoline that had been artificially weathered to 25%, 50% and 75% levels. As shown in Table 3, decreased luciferase induction was observed with increased levels of weathering, therefore it appears that, in addition to these lighter PAHs, larger bioactive PAH compounds have also been eliminated and/or degraded. While there were noticeable differences in the dose-response relationship between the unweathered control evaluated here versus the unleaded gasoline standard described in Table 1, this disparity is most likely attributable to differences in the product itself. In order to ensure that these bioassay results, which used carefully formulated and produced standards, accurately reflected the response that would be seen in products likely to be involved in environmental contamination, three products from commercial suppliers were examined: 10W40 motor oil, regular unleaded gasoline and #6 (bunker C) fuel oil. Each of these products produced marked luciferase responses, with fuel oil > gasoline motor oil, however, each produced a dose-response that was statistically different from their respective quality-control standard (MANOVA, all P < 0:05). On further analysis, these overall differences were found to be due to significant differences in luciferase activity at the two highest concentrations for gas and fuel oil, and activity at the three highest for
989
motor oil (LSD and Scheffe; P < 0:05). These differences are likely attributable to intrinsic differences between the standards and whole products, such as decreased AhRactive ligands present in the whole products. However, because of the similarities in the shape of the doseresponse curves, we believe that these data support the general applicability of the standards data to crude products. Our results show that the H1L1.1c2 luciferase bioassay is a sensitive and valid technique for the detection of AhR-active chemicals present in petroleum products. These data also indicate that components of several of the heavier petroleum products can act as strong AhR ligands, thereby implying that problems associated with large-scale exposure to such compounds may extend beyond their physical presence in or on sensitive ecosystems or wildlife species. For perspective, if the I-EQ50 value calculated from the commercial fuel oil #6 is used, one gallon of this petroleum product would contain 1.9 g of I-EQs and, if an oil spill the size of the Exxon Valdez incident occurred with this product, 21 metric tons of I-EQs would be released into the environment. On a smaller scale, using these data, each gallon of unleaded gasoline would contain 2.5 lg of I-EQs. As the US Environmental Protection Agency has set an acceptable daily intake of 0.006 pg of TCDD equivalents per kg body weight per day (Webster and Commoner, 1995), for a 75 kg human, this would translate to the consumption of 0.8 ll of unleaded gasoline. While these figures clearly overestimate the toxicity associated with these products (as many AhR-active compounds can fail to produce TCDD-like toxicity), they do indicate the biological potency of petroleum as an inducer. These estimates emphasize the importance of additional research to separate and identify those compounds within these complex mixtures responsible for such elevated readings. Although our bioassay does not provide information about the toxic potency of chemicals and chemical mixtures, it can be used to assess indirectly the potential for these products to activate the AhR signaling pathway (i.e., CYP1A1 induction) in exposed organisms, and to provide an avenue in which to estimate in vivo exposure to AhR-active compounds or products. This CYP1A1 induction response has been shown to be an integral step in the detoxification and activation of PAHs and related compounds in higher vertebrates (Nebert et al., 1991). When an organism is exposed to a relatively high concentration of AhR ligands/agonists, however, the CYP1A1 system may remain elevated for an extended period of time. When this occurs, persistent metabolism may result in increased amounts of reactive metabolites that can cause toxicity and/or mutagenic effects if not resolved (Whitlock, 1999). The recent demonstration that increases in CYP1A1 levels and activity can result in increased oxidative stress in cells and
990
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
the production of oxidative DNA damage (Park et al., 1996; Shertzer et al., 1998) provides additional avenues in which AhR agonists can produce adverse effects. Although such massive exposures are rare, large-scale petroleum exposures to wildlife species during oil spill events could create such a situation. Hence, understanding the differing induction potencies of such complex mixtures may indirectly provide information needed to assess the potential deleterious effects of these mixtures in exposed vertebrates. Also, through the integration of previous modifications of this bioassay system directed towards the optimization of the bioassay for the detection of PAHs and HAHs in wildlife serum samples (Ziccardi et al., 2000), an application of this system for the semi-quantitation of internal exposure of marine wildlife to oil during spill events is possible. This analytic tool, while not specific for PAHs or petroleum compounds (as all AhR-active compounds would be detected), would minimize the limitations inherent in EROD-based bioassay techniques, use easily-accessed biological samples and matrices, and provide an approach for general assessment of the degree of internal exposure in the sampled animal. Therefore, the understanding of the luciferase response associated with these products and the specific AhR-active constituents in these mixtures will both lead to a better understanding of the potential deleterious effects to exposed organisms and provide a mechanism for detection of inapparent exposure of at-risk organisms.
Acknowledgements The authors thank Michael Sowby, Dr. David Jessup and the California Department of Fish and Game, Office of Spill Prevention and Response (CDFG-OSPR) for their continued support of the development of this system, and Dr. George Clark and David Brown of Xenobiotic Detection Systems for technical assistance and advice. This research was supported by CDFGOSPR (FG-6404OS and FG-393OS), CDFG’s Oil Spill Response Trust Fund through the Oiled Wildlife Care Network at the Wildlife Health Center, School of Veterinary Medicine, University of California at Davis; the University of California Toxic Substances Research and Training Program, the National Institutes of Environmental Health Sciences (ES07685) and Superfund Basic Research program grant (ES04699 and ES05707).
References Addison, R.F., Brodie, P.F., Edwards, A., Sadler, M.C., 1986. Mixed function oxidase activity in the harbour seal (Phoca vitulina) from Sable Is. N.S.. Comparative Biochemistry and Physiology C 85, 121–124.
AFCEE, 1999. Petroleum fuels: basic composition and properties. Air Force Center for Environmental Excellence, TI#19206, Brooks AFB. Barber, R. et al., 1996. Motor Gasolines Technical Review. Chevron Products Company, FTR-1, San Francisco. Bruschweiler, B.J., Wurgler, F.E., Fent, K., 1996a. Inhibition of cytochrome P4501A by organotins in fish hepatoma cells PLHC-1. Environmental Toxicology and Chemistry 15, 728–735. Bruschweiler, B.J., Wurgler, F.E., Fent, K., 1996b. Inhibitory, effects of heavy metals on cytochrome P4501A induction in permanent fish hepatoma cells. Archives of Environmental Contamination and Toxicology 31, 475–482. Chen, Y.-H., Tukey, R.H., 1996. Protein kinase C modulates regulation of the CYP1A1 gene by the aryl hydrocarbon receptor. Journal of Biological Chemistry 271, 26261–26266. Denison, M.S., Phelan, D., Elferink, C.J., 1998. The Ah receptor signal transduction pathway. In: Denison, M.S., Helferich, W.G. (Eds.), Toxicant–Receptor Interactions. Taylor and Francis, Philadelphia, pp. 3–33. Denison, M.S., Seidel, S.D., Rogers, W.J., Ziccardi, M., Winter, G.M., Heath-Pagliuso, S., 1999. Natural and synthetic ligands for the Ah receptor. In: Puga, A., Wallace, K.B. (Eds.), Molecular Biology of the Toxic Response. Taylor and Francis, Philadelphia, pp. 393– 410. Denison, M.S., Wilkinson, C.F., 1985. Identification of the Ah receptor in selected mammalian species and induction of aryl hydrocarbon hydroxylase. European Journal of Biochemistry 147, 429–435. Fernandez-Salguero, P.M., Hilbert, D.M., Rudikoff, S., Ward, J.M., Gonzalez, F.J., 1996. Aryl-hydrocarbon receptor-deficient mice are resistant to 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced toxicity. Toxicology and Applied Pharmacology 140, 173–179. Garrison, P.M., Tullis, K., Aarts, J.M., Brouwer, A., Giesy, J.P., Denison, M.S., 1996. Species-specific recombinant cell lines as bioassay systems for the detection of 2,3,7,8-tetrachlorodibenzop-dioxin-like chemicals. Fundamental and Applied Toxicology 30, 194–203. Hunt, J.M., 1979. Petroleum Geochemistry and Geology. W.H. Freeman, San Francisco. Jessup, D.A., Leighton, F.A., 1996. Oil pollution and petroleum toxicity to wildlife. In: Fairbrother, A., Locke, L.N., Hoff, G.L. (Eds.), Noninfectious Diseases of Wildlife. Iowa State University Press, Ames, pp. 141–156. Jones, J.M., Anderson, J.W., 1999. Relative potencies of PAHs and PCBs based on the response of human cells. Environmental Toxicology and Pharmacology 7, 19–26. Kennedy, S.W., Lorenzen, A., James, C.A., Collins, B.T., 1993. Ethoxyresorufin-o-deethylase and porphyrin analysis in chicken embryo hepatocyte cultures with a fluorescence multiwell plate reader. Analytical Biochemistry 211, 102–112. Kikuchi, H. et al., 1996. Differences in inducibility of CYP1A1-mRNA by benzimidazole compounds between human and mouse cells: evidences of a human-specific signal transduction pathway for CYP1A1 induction. Archives of Biochemistry and Biophysics 334, 235–240. Long, W.P., Pray-Grant, M., Tsai, J.C., Perdew, G.H., 1998. Protein kinase C activity is required for aryl hydrocarbon receptor pathway-mediated signal transduction. Molecular Pharmacology 53, 691–700. Lu, Y.F., Santostefano, M., Cunninhgam, B.D.M., Threadgill, M.D., Safe, S., 1995. Identification of 30 -methoxy-40 -nitroflavone as a pure aryl hydrocarbon (Ah) receptor antagonist and evidence for more than one form of the nuclear Ah receptor in MCF-7 human breast cancer cells. Archives of Biochemistry and Biophysics 316, 470–477. Murk, A.J., Jonas, A., Brouwer, A., Leonards, P.E.G., Denison, M.S., 1996. Application of the CALUX (chemical activated luciferase
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991 gene expression) assay for measuring TCDD-equivalents in sediment, pore water and blood plasma samples. Organohalogen Compounds 27, 291–296. Nebert, D.W. et al., 1991. The P450 superfamily update on new sequences: gene mapping and recommended nomenclature. DNA and Cell Biology 10, 397–398. Neff, J.M., 1979. Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates, and Biological Effects. Applied Science Publishers, London. NRC, 1985. Oil in the Sea: Inputs, Fates and Effects. National Academy Press, National Research Council, 0309034795, Washington. Park, J.-Y.K., Shigenaga, M.K., Ames, B.N., 1996. Induction of cytochrome P4501A1 by 2,3,7,8-tetrachlorodibenzo-p-dioxin or indolo(3,2-b)carbazole is associated with oxidative DNA damage. Proceedings of the National Academy of Sciences of the United States of America. vol. 93, pp. 2322–2327. Postlind, H., Vu, T.P., Tukey, R.H., Quattrochi, L.C., 1993. Response of human CYP1-luciferase plasmids to 2,3,7,8-tetrachlorodibenzop-dioxin and polycyclic aromatic hydrocarbons. Applied Pharmacology 118, 255–262. Potter, T.L., Simmons, K.E., 1998. Composition of Petroleum Mixtures. The Association for the Environmental Health of Soils, Amherst. Safe, S., 1990. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and related compounds: environmental and mechanistic considerations which support the development of toxic equivalency factors (TEFs). Critical Reviews in Toxicology 21, 51–88. Sanderson, J.T., Aarts, J.M., Brouwer, A., Froese, K.L., Denison, M.S., Giesy, J.P., 1996. Comparison of Ah receptor-mediated luciferase and ethoxyresorufin-o-deethylase induction in H4IIE cells: implications for their use as bioanalytical tools for the detection of polyhalogenated aromatic hydrocarbons. Toxicology and Applied Pharmacology 137, 316–325.
991
Santostefano, M., Merchant, M., Arellano, L., Morrison, V., Denison, M.S., Safe, S., 1993. Alpha-naphthoflavone-induced CYP1A1 gene expression and cytosolic aryl hydrocarbon receptor transformation. Molecular Pharmacology 43, 200–206. Shertzer, H.G. et al., 1998. Dioxin causes a sustained oxidative stress response in the mouse. Biochemical and Biophysical Research Communications 253, 44–48. Short, J.W., Jackson, T.J., Larsen, M.L., Wade, T.L., 1996. Analytical methods used for the analysis of hydrocarbons in crude oil, tissues, sediments, and seawater collected for the natural resources damage assessment of the Exxon Valdez oil spill. American Fisheries Society Symposium 18, 125–132. Stegeman, J.J., Smolowitz, R.M., Hahn, M.E., 1991. Immunohistochemical localization of environmentally induced cytochrome P450IA1 in multiple organs of the marine teleost Stenotomus chrysops (Scup). Toxicology and Applied Pharmacology 110, 486–504. Varanasi, U., 1989. Metabolism of Polycyclic Aromatic Hydrocarbons in the Aquatic Environment. CRC Press, Boca Raton. Velazquez, M., Feirtag, J.M., 1997. Quenching and enhancement effects of ATP extractants, cleansers, and sanitizers on the detection of the ATP bioluminescence signal. Journal of Food Protection 60, 799–803. Webster, T., Commoner, B., 1995. Overview: the dioxin debate. In: Schecter, A. (Ed.), Dioxins and Health. Plenum Publishing, New York, pp. 1–50. Whitlock, J.P., 1999. Induction of cytochrome P4501A1. Annual Review of Pharmacology and Toxicology 39, 103–125. Willett, K.L., Randerath, K., Zhou, G.-D., Safe, S.H., 1998. Inhibition of CYP1A1-dependent activity by the polynuclear aromatic hydrocarbon (PAH) fluoranthene. Biochemical Pharmacology 55, 831–839. Ziccardi, M.H., Gardner, I.A., Denison, M.S., 2000. Development and modification of a recombinant cell bioassay to directly detect halogenated and polycyclic aromatic hydrocarbons in serum. Toxicological Sciences 54, 183–193.
Reports
Application of the luciferase cell culture bioassay for the detection of refined petroleum products Michael H. Ziccardi
a,*
, Ian A. Gardner b, Jonna A.K. Mazet
a,b
, Michael S. Denison
c
a
b
Wildlife Health Center, School of Veterinary Medicine, University of California, Davis, CA 95616, USA Department of Medicine and Epidemiology, School of Veterinary Medicine, University of California, Davis, CA 95616, USA c Department of Environmental Toxicology, University of California, Davis, CA 95616, USA
Abstract A luciferase cell culture-based bioassay, developed to detect 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-like activity of halogenated and polycyclic aromatic hydrocarbons, was optimized to detect refined petroleum products and to determine their relative inducing potency. Quality control standards from 32 refined products (gasolines and diesels, jet fuels, lubricating oils, fuel oils and weathered products) and three commercial products were evaluated. Induction equivalents (I-EQs) were determined by direct comparison of the EC50 and EC20 values (based on the median and 20% TCDD maximal response, respectively) from dose-response curves for each product to those obtained with TCDD. Most petroleum products were active in the luciferase bioassay, with those products composed of fractions produced later in the distillation process (i.e. fuel oils) inducing higher levels. Additionally, weathering of products reduced their induction potency. Based on the high I-EQ estimates of many products, biological effects associated with exposure may have been previously underestimated using other diagnostic methods. Ó 2002 Elsevier Science Ltd. All rights reserved. Keywords: Petroleum; PAH; Oil spills; CYP1A1; Luciferase; Bioassay
1. Introduction Petroleum contamination of the environment has long been known to cause environmental damage, as well as acute and chronic effects to exposed organisms. Although acute effects of such exposures are well known (NRC, 1985; Jessup and Leighton, 1996), several intrinsic factors of petroleum products make their detection at low levels in the environment problematic, and the assessment of toxicity associated with chronic exposure difficult to establish. Crude and refined petroleum products are extremely complex mixtures of organic compounds (including alkane, alkene, naphthalene, and monocyclic and polycyclic aromatic hydrocarbon (PAH) components), the composition of which varies widely among products, fields, and refineries, and even within a single well (Hunt, 1979). Accordingly, the development of detection methods and identification schemes to determine product source, as *
Corresponding author. Tel.: +1-530-754-5701; fax: +1-530-7523318. E-mail address: [email protected] (M.H. Ziccardi).
well as the subsequent establishment of the toxicity associated with exposure to a specific product can vary widely among studies. An additional complicating factor is that the composition of these products, once introduced to the environment, is not static. Spilled or released oil undergoes a complex ‘‘weathering’’ process, during which the product undergoes evaporation, dissolution, dispersion, photo-oxidation, and additional breakdown/modification due to other biological, chemical and physical methods (NRC, 1985). Consequently, the chemical makeup of petroleum products in the environment and their associated toxicity can vary dramatically depending on the timing of exposure. Numerous analytical methods have been developed to determine the presence of, or exposure to, petroleum products in the marine environment. Direct methods, such as gas and liquid chromatography, have been extensively used to quantify levels of PAHs and other compounds within petroleum products and to ‘‘fingerprint’’ such products in either environmental samples (Short et al., 1996) or biological matrices (Varanasi, 1989). These methods, however, are often prohibitively expensive, require extensive time-consuming extraction
0025-326X/02/$ - see front matter Ó 2002 Elsevier Science Ltd. All rights reserved. PII: S 0 0 2 5 - 3 2 6 X ( 0 2 ) 0 0 1 2 8 - 5
984
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
methods and may require the collection of large quantities of often difficult to acquire tissue samples. Additionally, because they are designed to detect specific parent components or their breakdown products (often limited to those compounds known to cause morbidity in experimental exposure trials), direct methods can fail to detect entire classes of compounds that can produce similar biological/toxicological responses. Results from direct methods may therefore provide an inaccurate estimate of the organismal effects associated with the overall exposure to the product. Bioassay systems, another class of detection methods, are based on the ability of a petroleum product to produce a biological response in a living organism. One of the most widely used of these methods is the quantitation of cytochrome P4501A1 (CYP1A1) induction in tissues. This xenobiotic detoxification system is mediated by the Ah receptor (AhR), a soluble intra-cellular protein that binds PAHs and related chemicals with high affinity (Safe, 1990; Denison et al., 1998). In addition, this P4501A1 protein has been implicated as the ‘‘cellular switch’’ responsible for initiating the cascade of events leading to the specific toxic responses associated with exposure to these compounds (Safe, 1990; Fernandez-Salguero et al., 1996). Chemical-dependent activation of the AhR results in a sequence of events in the nucleus that stimulates CYP1A1 gene transcription and increases cellular concentration of P4501A1 protein and its associated monoxygenase activity, ethoxyresorufino-deethylase (EROD). The increase in CYP1A1-dependent catalytic activity most commonly results in enhanced detoxification of the inducing chemicals. Given this relationship, the quantitation of CYP1A1 and EROD has been used to evaluate the relative biological and toxicological potency and as a biomarker of exposure to PAHs and petroleum products (Addison et al., 1986; Stegeman et al., 1991). Although quantification of CYP1A1 activity may serve as an excellent ‘‘bioassay of effect’’ in potentially exposed organisms, intra- and inter-species differences in AhR ligand binding specificity, responsiveness and sensitivity to these chemicals, and the ability of the AhR to be activated by non-PAH/HAH-type chemicals complicate the interpretation of the results (Denison and Wilkinson, 1985; Garrison et al., 1996; Kikuchi et al., 1996; Denison et al., 1999). These analyses also provide no information as to which chemicals were responsible for the induction response, nor the level of exposure in vivo. Additionally, these methods, as with the analytical chemistry methods discussed above, often require sampling protocols that include the harvesting of tissues (i.e. liver) rich in CYP1A1 activity to achieve acceptable detection limits. Lastly, the documented ability of EROD activity to be competitively inhibited at high pollutant levels (as would be expected in samples from oil spill events) would also result in an underestimation
of the amount of inducing compounds (Kennedy et al., 1993). Recently, several recombinant cell bioassay systems using luciferase reporter genes under control of CYP1A1 upstream regulatory elements (including the promoter and/or dioxin responsive elements (DREs)) have been developed for the detection and relative quantitation of AhR-active chemicals and compounds through co-activation of luciferase and CYP genes (Postlind et al., 1993; Garrison et al., 1996; Murk et al., 1996). Induction of luciferase in these cell lines (measured by quantification of bioluminescence) occurs in a time-, dose- and AhR-dependent, and chemical-specific manner. Not only has this assay been shown to be more analytically sensitive than EROD-based bioassays, but also it lacks the majority of the limitations inherent in those systems. To date, most studies using these new luciferasebased bioassay systems have focused predominantly on the analysis of HAHs (Garrison et al., 1996; Murk et al., 1996; Sanderson et al., 1996) and/or PAHs (Jones and Anderson, 1999; Ziccardi et al., 2000). Also, few have examined their utility in the evaluation of complex mixtures, such as petroleum products. Here we describe studies in which this system has been optimized and validated for the detection and relative quantitation of AhR-active chemicals in complex petroleum products.
2. Materials and methods 2.1. Chemicals All petroleum products were considered hazardous and appropriate personal protective methods and materials were used throughout experiments. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) was obtained from Dr. S. Safe (Texas A&M University) and was dissolved in dimethyl sulfoxide (DMSO) from Aldrich Chemicals (Milwaukee, WI). All petroleum standards used in this study are listed in Tables 1–3 and were purchased from Accustandards Co. (St. Louis, MO). Whole petroleum products, listed in Table 3, were either purchased from local vendors or provided by industry representatives. Cell culture media and antibiotics were purchased from Life Technologies (Gaithersburg, MD), fetal bovine serum from Atlanta Biologicals (Atlanta, GA), luciferase substrate and assay buffers from Promega (Madison, WI) and fluorescamine from Molecular Probes (Eugene, OR). 2.2. Recombinant cell line Recombinant mouse hepatoma (H1L1.1c2) cells, containing the stably integrated PAH/HAH-inducible luciferase expression vector pGudLuc1.1, were described previously (Garrison et al., 1996). This vector contains
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
985
Table 1 Luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, gasoline and aviation fuel products calculated from dose-response curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Gas, regular leaded Gas, regular unleaded Gas, premium Gas, oxygenate free Diesel Diesel #1, low SO4 Diesel #2, low SO4 Diesel, Arctic Kerosene
8:28 10 4:81 106 6:13 106 5:85 105 4:56 105 8:15 105 1:25 107 4:42 106 NCa 4:25 105
1 10 0.172 0.135 1.414 1.816 1.015 0.066 0.187 NC 1.948
1:88 10 8:94 105 2:54 105 6:13 104 4:91 104 5:94 104 3:56 105 8:95 104 NC 2:73 105
1 106 0.210 0.739 3.064 3.828 3.162 0.528 2.100 NC 0.688
Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel, Fuel,
NC NC NC NC NC NC NC NC NC
NC NC NC NC NC NC NC NC NC
NC NC 1:43 106 2:94 104 NC NC NC NC 1:50 105
NC NC 0.131 6.381 NC NC NC NC 1.255
a
JP 4 JP 5 JP 7 JP 8 JP 10 JP TS Jet Ref. aviation turbine
6
Not calculable due to induction less than 50% or 20% of TCDD.
Table 2 Calculated luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, motor oil and fuel oil products calculated from doseresponse curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Oil, SAE30W Oil, SAE40W Oil, SAE50W Oil, SAE5W30 Oil, SAE10W30 Oil, SAE10W40 Oil, SAE20W50 Oil, Hydraulic Jack
8:28 10 8:16 105 1:08 106 8:11 105 1:67 107 4:65 107 1:63 106 1:52 107 1:28 106
1 10 1.014 0.767 1.020 0.050 0.018 0.508 0.054 0.649
1:88 10 1:11 105 2:37 105 1:71 105 2:68 106 7:29 106 1:38 105 1:74 106 1:94 105
1 106 1.686 0.793 1.096 0.070 0.026 1.365 0.108 0.966
Fuel Fuel Fuel Fuel Fuel Fuel
NCa NC 8:73 104 7:15 103 4:32 103 1:17 104
NC NC 9.483 115.8 191.5 70.91
NC 2:41 105 1:08 104 1:38 102 2:41 102 4:67 102
NC 0.778 17.47 1361 780.5 402.6
a
oil oil oil oil oil oil
#1 #2 #3 #4 #5 #6
6
Not calculable due to induction less than 50% or 20% of TCDD.
the firefly luciferase gene under PAH/HAH-inducible and AhR-dependent control of four DREs. 2.3. Sample analysis Dilutions of all products (10-fold dilutions (v/v) from original stock) were done using Optima-grade methanol (MeOH) obtained from Fisher Scientific (Pittsburgh, PA) or, in the case of the #6 fuel oil and 10W40 motor oil, first suspended in Optima-grade hexane (Fisher Scientific) to facilitate solubilization of the heavier
products, then diluted in MeOH and resuspended through pipetting and vigorous agitation. All petroleum standards were analyzed during a single period, with appropriate positive (1 nM TCDD) and negative (carrier solvent alone) controls run on each 96-well plate. A TCDD dose-response curve was also included in each experiment. The overall protocol for the analysis of samples using the 96-well luciferase bioassay format has been previously described (Ziccardi et al., 2000). Briefly, 100 mm plates of stable cell clones were trypsinized, resuspended in 10 ml of culture media, and cells diluted
986
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Table 3 Calculated luciferase induction (median and 20% of maximal TCDD response, or EC50 and EC20 , respectively) and TCDD induction equivalency (I-EQ50 and I-EQ20 , in micrograms of TCDD equivalence per gram product) for TCDD, weathered unleaded gasoline and whole commercial petroleum products calculated from dose-response curves generated using the luciferase bioassay Product
EC50 (pg/well) 1
I-EQ50 (lg/g)
EC20 (pg/well) 1
I-EQ20 (lg/g)
TCDD Gas, 0% weathered Gas, 25% weathered Gas, 50% weathered Gas, 75% weathered
8:28 10 NCa NC NC NC
1 10 NC NC NC NC
1:88 10 8:54 103 6:29 104 NC NC
1 106 21.99 2.984 NC NC
Gas, regular unleaded Oil, SAE10W40 Fuel oil #6
8:33 106 1:13 107 1:63 104
1.061 0.784 540.9
2:15 106 3:52 106 2:83 103
1.743 1.064 1324
a
6
Not calculable due to induction less than 50% or 20% of TCDD.
to 300,000 cells per ml. Aliquots (200 ll) of this suspension were added to wells in rows two through eight of sterile 96-well white CulturPlatesTM (Packard Instruments; Meriden, CT), and plates were incubated for 24 h prior to ligand addition, allowing cells to reach confluence. All wells were washed with 100 ll of sterile 1X phosphate-buffered saline (PBS), followed by addition of ligand in DMSO or methanol/hexane in MEM (final DMSO/methanol concentration was 1%), with 1 nM TCDD and either DMSO, methanol or hexane as positive and negative controls, respectively. Plates were then incubated for three hours at 37 °C, after which cells were washed twice with PBS, followed by the addition of Promega Lysis Buffer (25 ll) to each well, and plates were shaken at room temperature until cells were lysed (20 min). Luciferase activity was measured using an automated microplate luminometer (Dynatech ML3000; Chantilly, VA) in enhanced flash mode with 10 s delay and 10 s integration periods after the automatic injection of 50 ll of Promega stabilized luciferase reagent. Correction for intra-plate variation was accomplished through protein quantification using fluorescamine (at 500 lg/ml) in acetonitrile and detection with a multiplate fluorometer (Fluostar; SLT, Salzburg, Austria) at 400 nm excitation and 460 nm emission wavelengths. Inter-plate variation was corrected for by standardization of each plate to the luciferase activity induced by 1 nM TCDD. Final results were expressed as relative light units per mg protein. 2.4. Statistical methods Descriptive statistics and comparative analyses (Student’s t-test between samples and controls to determine significant induction above background) were calculated using Excel XP (Microsoft, Redmond, WA). Differences within petroleum classes and/or comparisons between standards and whole products were determined using multivariate analysis of variance (MANOVA) methods, with subsequent pairwise analysis (Tukey’s LSD and Scheffe’s tests) done between replicate wells, chemical
concentrations, petroleum products and/or product classifications using Statistica (Statsoft, Tulsa, OK) statistical software. Median (EC50 ) and 20% (EC20 ) effective concentrations of maximal TCDD response (for those chemicals showing induction equivalent to or greater than these levels) were calculated by fitting doseresponse data to a four-parameter Hill model by least squares using SigmaPlot (SPSS, Chicago, IL) as described previously (Ziccardi et al., 2000). Because the analyzed samples represent complex chemical mixtures, induction equivalency values (I-EQs) were expressed as micrograms of TCDD equivalents per gram of product and were calculated by dividing the product’s EC value by that calculated for TCDD.
3. Results To maximize the utility of this assay system for the rapid detection of PAHs and like compounds, prior optimization of the protocols has emphasized maximizing throughput in the system. Through the integration of a 96-well format, selection of a reporter system/ cell line that shows maximal response to PAHs at a three hour incubation period, minimization of lag and read times in the luminometer, and integration of several inter- and intra-plate correction methods (Ziccardi et al., 2000), we now have the ability to rapidly and accurately assess the ability of many different products and/ or samples to induce luciferase activity. Accordingly, the inducing activity of all petroleum controls in these experiments was assessed within a single sampling period, allowing for direct comparison using a single TCDD dose-response curve and maximizing our confidence in the subsequent I-EQ estimations. The luciferase bioassay detected activity in most petroleum products when added at elevated concentrations (above 1 105 picograms per well), with those compounds produced later in the distillation process generally resulting in higher levels of induction (Fig. 1A). The lighter grade products (which included gasolines and
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Fig. 1. Dose-response curves for induction of luciferase activity by (A) gasoline and diesel distillates, and (B) jet fuel standards. H1L1.1c2 cells in 96-well microplates (at 80–100% confluence) were incubated with TCDD or gasoline distillate standards at the indicated concentrations for three hours, after which luciferase activity (corrected for protein concentration) was determined as described in Section 2. Values represent the means S:D: of triplicate determinations.
diesel fuels) were moderate inducers with similar induction profiles, with the exception of arctic diesel. The median effective concentrations for these mixtures ranged from 1:25 107 to 4:25 105 pg/well, with I-EQ50 values ranging from 0.066 to 1.816 lg TEQ/g product (Table 1). In contrast, dose-response relationship studies using aviation fuels revealed increased induction responses at low concentrations, followed by declining luciferase levels at the highest concentrations tested (Fig. 1B). Several products reached 20% maximal TCDD luciferase expression levels, most notably that of JP-8, which produced significant induction levels (EC20 ¼ 2:94 104 pg/well) and a relatively large I-EQ20 (6.381 lg/g) (Table 1). Motor and lubricating oils caused levels of induction similar to those of gasoline and diesel compounds (Fig. 2A). All of these products showed approximately the same dose-response pattern, with EC50 values calculated from dose-response curves ranging from 4:65 107 to 8:11 105 pg/well and I-EQ50
987
Fig. 2. Dose-response curves for induction of luciferase activity by (A) lubricating oils, and (B) fuel oil distillate standards. H1L1.1c2 cells in 96-well microplates (at 80–100% confluence) were incubated with TCDD or gasoline distillate standards at the indicated concentrations for three hours, after which luciferase activity (corrected for protein concentration) was determined as described in Section 2. Values represent the means S:D: of triplicate determinations.
values from 0.018 to 1.020 lg/g product (Table 2). Lastly, fuel oils elicited a much greater overall response than the lighter control compounds (Fig. 2B). Additionally, there was a trend towards an increasing response as the products increased in viscosity, ranging from minimal response in fuel oil #1 to an EC50 value of 7:15 103 pg/well for fuel oil #4 (Table 2). However, as noted in the aviation fuels, induction levels decreased at the highest tested concentrations. In an attempt to determine the effect of progressive environmental degradation on the bioactivity of fuels, regular unleaded gasoline that had been ‘‘weathered’’ to 25%, 50% and 75% levels was analyzed. As shown in Table 3, the two less weathered products produced significant luciferase responses (with EC20 values of 8:54 103 and 6:29 104 , and I-EQ values of 21.99 and 2.984 lg/g), and the two more weathered samples produced little luciferase response. Unlike the activity seen in the other gasoline standards, however, luciferase
988
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
activity in the weathered products declined at the highest two concentrations. To determine whether the responses noted in the petroleum quality control standards were comparable to those evident in commercially available petroleum distillates, dose-response analyses were also carried out using regular unleaded gasoline, 10W40 motor oil and #6 fuel oil acquired either from local retailers or industry representatives. As seen in Table 3, appreciable luciferase response was noted for all three products, corresponding to EC50 values of 8:33 106 , 1:13 107 , and 1:63 104 , respectively. While the response patterns for each of these products were comparable to that observed with the appropriate petroleum standards, subsequent EC50 calculations for the whole products were several times higher for both the fuel oil and gasoline samples (Tables 1 and 2). Further statistical analyses showed significant overall differences between each product and its associated standard (MANOVA, P values for all effects <0.05) due to significant differences in the highest two concentrations for gasoline and #6 fuel oil, and the second and third highest concentrations for motor oil (LSD and Scheffe tests; all P values <0.05).
4. Discussion We used our recombinant luciferase cell culture bioassay to examine the ability of various refined petroleum products to activate the AhR signaling pathway and to estimate their relative biological/toxicological potency. Light and heavy gasoline fractions (comprised of gasoline, kerosene and diesel fuels) are largely made up of the first hydrocarbons distilled during the refining process and, as such, contain a very high proportion (25–40%) of smaller aromatic compounds (AFCEE, 1999). We observed that most of these different fuel mixtures produced similar induction profiles, with high induction levels noted at the higher concentrations (Fig. 1A; Table 1). At first glance, these data may seem to be contrary to our previous work that showed little luciferase induction caused by smaller PAH constituents (data not shown) and work of others that reported little CYP induction associated with smaller aliphatic hydrocarbons (Neff, 1979). However, these mixtures are often comprised of over 300 different hydrocarbon compounds (Barber et al., 1996), and there may be a significant proportion of larger-ringed structures in these mixtures that are responsible for the increased luciferase response. While the luciferase bioassay appears to work well for most of the gasoline and diesel fuels, arctic diesel, a product that includes additives necessary to remain fluid at extremely low temperatures, showed initial increased luciferase activity, followed by decreased induction at the highest concentrations (Fig. 1A). Similar declines in induction were noted for all aviation fuels, products
largely composed of gasoline, kerosene and additional additives necessary to satisfy specific requirements of the aircraft to be fueled (Fig. 1B). Indeed, only three of these products induced to 20% of the maximal TCDD response: JP-7 (a special non-distillate fuel designed for the SR-71 aircraft), JP-8 (a newer primary fuel used by the United States Air Force) and turbine fuel (also known as Jet A in commercial applications). At this 20% level, however, these products induced in a similar fashion to gasolines and produced fairly large I-EQ20 values (Table 1). The observed pattern in luciferase activity with these products may be attributable to specific additives in these products that may differentially affect the bioassay at the higher concentrations. Although it is possible that these compounds are toxic to the cells at these higher concentrations, no toxicity was observed by visual inspection of the cells prior to luciferase analysis. However, inhibition of cellular processes and/ or the AhR induction pathway by components in the fuels, without any apparent cell toxicity, is plausible. The ability of chemical components in these fuels to act as AhR antagonists and/or repress the AhR-dependent induction response via indirect pathways (Chen and Tukey, 1996; Long et al., 1998) could also account for the overall lower potencies of some products. Prior studies have identified a variety of PAHs, HAHs and organic compounds (such as organotins and nitroflavones) that reduce AhR activity either partially or completely (Santostefano et al., 1993; Lu et al., 1995; Bruschweiler et al., 1996a,b; Willett et al., 1998). The analysis carried out here only examined the agonist activity of these products; analysis of their abilities to act as effective antagonists would require testing of their ability to reduce TCDD activity. Chemicals in these fuels could also reduce luciferase activity directly via quenching of luciferin-generated light production, as has occurred in the analysis of a number of commercial products (Velazquez and Feirtag, 1997). Future work should focus on analysis of these additives in order to better understand this loss in bioactivity at elevated concentrations. Heavier residual petroleum products were also evaluated using this bioassay system. Motor oils, residual products containing very low levels of aromatics (1– 2%) (Potter and Simmons, 1998) showed induction at concentrations above 1 104 pg/well (Fig. 2A). Excluding 10W40, there was a statistically significant difference between single-weight and multi-weight products at concentrations above 1 104 pg/well (MANOVA; P values for all effects <0.05), with single-weight compounds producing higher luciferase induction. It is possible that this difference may be attributable to the presence of polymers used in multi-weight products causing decreases in luciferase production and/or detection in a similar fashion to that observed with jet fuels.
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
Heating fuel oils, used in a variety of different applications where a clean burning, non-explosive fuel is required (such as in furnaces, heaters and ship boilers), were also assessed using this system. These products are graded from #1 to #6 based on viscosity, with fuel oil #1 (or a kerosene derivative) being the lightest and fuel oil #6 (or bunker C) being comprised largely of the residuum from the distillation process. Most of these compounds caused extremely large responses in the luciferase bioassay, with the response increasing with increasing viscosity (Fig. 2B). For example, for the three heaviest fuel oil products, relatively large I-EQ50 values of 70–191 and I-EQ20 values of 402–1361 lg/g of product were calculated (Table 2). The sizable differences between the I-EQ50 measurements are due to the slope of the TCDD curve being much steeper than that of all other petroleum compounds, therefore the true estimate of I-EQ most likely lies somewhere between the 50% and 20% values. All heating fuel oils produced a decline in bioassay response at the highest concentrations, which may either be attributable to the mechanisms mentioned earlier or it may result from increased interference/quenching of the bioluminescence detection as a result of the darker color at the highest concentrations. Weathering of petroleum products is known to volatilize off smaller aromatics, leaving most of the heavier products in the residual mixture (NRC, 1985). To determine the ability of the bioassay to detect such compounds, we analyzed samples of regular unleaded gasoline that had been artificially weathered to 25%, 50% and 75% levels. As shown in Table 3, decreased luciferase induction was observed with increased levels of weathering, therefore it appears that, in addition to these lighter PAHs, larger bioactive PAH compounds have also been eliminated and/or degraded. While there were noticeable differences in the dose-response relationship between the unweathered control evaluated here versus the unleaded gasoline standard described in Table 1, this disparity is most likely attributable to differences in the product itself. In order to ensure that these bioassay results, which used carefully formulated and produced standards, accurately reflected the response that would be seen in products likely to be involved in environmental contamination, three products from commercial suppliers were examined: 10W40 motor oil, regular unleaded gasoline and #6 (bunker C) fuel oil. Each of these products produced marked luciferase responses, with fuel oil > gasoline motor oil, however, each produced a dose-response that was statistically different from their respective quality-control standard (MANOVA, all P < 0:05). On further analysis, these overall differences were found to be due to significant differences in luciferase activity at the two highest concentrations for gas and fuel oil, and activity at the three highest for
989
motor oil (LSD and Scheffe; P < 0:05). These differences are likely attributable to intrinsic differences between the standards and whole products, such as decreased AhRactive ligands present in the whole products. However, because of the similarities in the shape of the doseresponse curves, we believe that these data support the general applicability of the standards data to crude products. Our results show that the H1L1.1c2 luciferase bioassay is a sensitive and valid technique for the detection of AhR-active chemicals present in petroleum products. These data also indicate that components of several of the heavier petroleum products can act as strong AhR ligands, thereby implying that problems associated with large-scale exposure to such compounds may extend beyond their physical presence in or on sensitive ecosystems or wildlife species. For perspective, if the I-EQ50 value calculated from the commercial fuel oil #6 is used, one gallon of this petroleum product would contain 1.9 g of I-EQs and, if an oil spill the size of the Exxon Valdez incident occurred with this product, 21 metric tons of I-EQs would be released into the environment. On a smaller scale, using these data, each gallon of unleaded gasoline would contain 2.5 lg of I-EQs. As the US Environmental Protection Agency has set an acceptable daily intake of 0.006 pg of TCDD equivalents per kg body weight per day (Webster and Commoner, 1995), for a 75 kg human, this would translate to the consumption of 0.8 ll of unleaded gasoline. While these figures clearly overestimate the toxicity associated with these products (as many AhR-active compounds can fail to produce TCDD-like toxicity), they do indicate the biological potency of petroleum as an inducer. These estimates emphasize the importance of additional research to separate and identify those compounds within these complex mixtures responsible for such elevated readings. Although our bioassay does not provide information about the toxic potency of chemicals and chemical mixtures, it can be used to assess indirectly the potential for these products to activate the AhR signaling pathway (i.e., CYP1A1 induction) in exposed organisms, and to provide an avenue in which to estimate in vivo exposure to AhR-active compounds or products. This CYP1A1 induction response has been shown to be an integral step in the detoxification and activation of PAHs and related compounds in higher vertebrates (Nebert et al., 1991). When an organism is exposed to a relatively high concentration of AhR ligands/agonists, however, the CYP1A1 system may remain elevated for an extended period of time. When this occurs, persistent metabolism may result in increased amounts of reactive metabolites that can cause toxicity and/or mutagenic effects if not resolved (Whitlock, 1999). The recent demonstration that increases in CYP1A1 levels and activity can result in increased oxidative stress in cells and
990
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991
the production of oxidative DNA damage (Park et al., 1996; Shertzer et al., 1998) provides additional avenues in which AhR agonists can produce adverse effects. Although such massive exposures are rare, large-scale petroleum exposures to wildlife species during oil spill events could create such a situation. Hence, understanding the differing induction potencies of such complex mixtures may indirectly provide information needed to assess the potential deleterious effects of these mixtures in exposed vertebrates. Also, through the integration of previous modifications of this bioassay system directed towards the optimization of the bioassay for the detection of PAHs and HAHs in wildlife serum samples (Ziccardi et al., 2000), an application of this system for the semi-quantitation of internal exposure of marine wildlife to oil during spill events is possible. This analytic tool, while not specific for PAHs or petroleum compounds (as all AhR-active compounds would be detected), would minimize the limitations inherent in EROD-based bioassay techniques, use easily-accessed biological samples and matrices, and provide an approach for general assessment of the degree of internal exposure in the sampled animal. Therefore, the understanding of the luciferase response associated with these products and the specific AhR-active constituents in these mixtures will both lead to a better understanding of the potential deleterious effects to exposed organisms and provide a mechanism for detection of inapparent exposure of at-risk organisms.
Acknowledgements The authors thank Michael Sowby, Dr. David Jessup and the California Department of Fish and Game, Office of Spill Prevention and Response (CDFG-OSPR) for their continued support of the development of this system, and Dr. George Clark and David Brown of Xenobiotic Detection Systems for technical assistance and advice. This research was supported by CDFGOSPR (FG-6404OS and FG-393OS), CDFG’s Oil Spill Response Trust Fund through the Oiled Wildlife Care Network at the Wildlife Health Center, School of Veterinary Medicine, University of California at Davis; the University of California Toxic Substances Research and Training Program, the National Institutes of Environmental Health Sciences (ES07685) and Superfund Basic Research program grant (ES04699 and ES05707).
References Addison, R.F., Brodie, P.F., Edwards, A., Sadler, M.C., 1986. Mixed function oxidase activity in the harbour seal (Phoca vitulina) from Sable Is. N.S.. Comparative Biochemistry and Physiology C 85, 121–124.
AFCEE, 1999. Petroleum fuels: basic composition and properties. Air Force Center for Environmental Excellence, TI#19206, Brooks AFB. Barber, R. et al., 1996. Motor Gasolines Technical Review. Chevron Products Company, FTR-1, San Francisco. Bruschweiler, B.J., Wurgler, F.E., Fent, K., 1996a. Inhibition of cytochrome P4501A by organotins in fish hepatoma cells PLHC-1. Environmental Toxicology and Chemistry 15, 728–735. Bruschweiler, B.J., Wurgler, F.E., Fent, K., 1996b. Inhibitory, effects of heavy metals on cytochrome P4501A induction in permanent fish hepatoma cells. Archives of Environmental Contamination and Toxicology 31, 475–482. Chen, Y.-H., Tukey, R.H., 1996. Protein kinase C modulates regulation of the CYP1A1 gene by the aryl hydrocarbon receptor. Journal of Biological Chemistry 271, 26261–26266. Denison, M.S., Phelan, D., Elferink, C.J., 1998. The Ah receptor signal transduction pathway. In: Denison, M.S., Helferich, W.G. (Eds.), Toxicant–Receptor Interactions. Taylor and Francis, Philadelphia, pp. 3–33. Denison, M.S., Seidel, S.D., Rogers, W.J., Ziccardi, M., Winter, G.M., Heath-Pagliuso, S., 1999. Natural and synthetic ligands for the Ah receptor. In: Puga, A., Wallace, K.B. (Eds.), Molecular Biology of the Toxic Response. Taylor and Francis, Philadelphia, pp. 393– 410. Denison, M.S., Wilkinson, C.F., 1985. Identification of the Ah receptor in selected mammalian species and induction of aryl hydrocarbon hydroxylase. European Journal of Biochemistry 147, 429–435. Fernandez-Salguero, P.M., Hilbert, D.M., Rudikoff, S., Ward, J.M., Gonzalez, F.J., 1996. Aryl-hydrocarbon receptor-deficient mice are resistant to 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced toxicity. Toxicology and Applied Pharmacology 140, 173–179. Garrison, P.M., Tullis, K., Aarts, J.M., Brouwer, A., Giesy, J.P., Denison, M.S., 1996. Species-specific recombinant cell lines as bioassay systems for the detection of 2,3,7,8-tetrachlorodibenzop-dioxin-like chemicals. Fundamental and Applied Toxicology 30, 194–203. Hunt, J.M., 1979. Petroleum Geochemistry and Geology. W.H. Freeman, San Francisco. Jessup, D.A., Leighton, F.A., 1996. Oil pollution and petroleum toxicity to wildlife. In: Fairbrother, A., Locke, L.N., Hoff, G.L. (Eds.), Noninfectious Diseases of Wildlife. Iowa State University Press, Ames, pp. 141–156. Jones, J.M., Anderson, J.W., 1999. Relative potencies of PAHs and PCBs based on the response of human cells. Environmental Toxicology and Pharmacology 7, 19–26. Kennedy, S.W., Lorenzen, A., James, C.A., Collins, B.T., 1993. Ethoxyresorufin-o-deethylase and porphyrin analysis in chicken embryo hepatocyte cultures with a fluorescence multiwell plate reader. Analytical Biochemistry 211, 102–112. Kikuchi, H. et al., 1996. Differences in inducibility of CYP1A1-mRNA by benzimidazole compounds between human and mouse cells: evidences of a human-specific signal transduction pathway for CYP1A1 induction. Archives of Biochemistry and Biophysics 334, 235–240. Long, W.P., Pray-Grant, M., Tsai, J.C., Perdew, G.H., 1998. Protein kinase C activity is required for aryl hydrocarbon receptor pathway-mediated signal transduction. Molecular Pharmacology 53, 691–700. Lu, Y.F., Santostefano, M., Cunninhgam, B.D.M., Threadgill, M.D., Safe, S., 1995. Identification of 30 -methoxy-40 -nitroflavone as a pure aryl hydrocarbon (Ah) receptor antagonist and evidence for more than one form of the nuclear Ah receptor in MCF-7 human breast cancer cells. Archives of Biochemistry and Biophysics 316, 470–477. Murk, A.J., Jonas, A., Brouwer, A., Leonards, P.E.G., Denison, M.S., 1996. Application of the CALUX (chemical activated luciferase
M.H. Ziccardi et al. / Marine Pollution Bulletin 44 (2002) 983–991 gene expression) assay for measuring TCDD-equivalents in sediment, pore water and blood plasma samples. Organohalogen Compounds 27, 291–296. Nebert, D.W. et al., 1991. The P450 superfamily update on new sequences: gene mapping and recommended nomenclature. DNA and Cell Biology 10, 397–398. Neff, J.M., 1979. Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates, and Biological Effects. Applied Science Publishers, London. NRC, 1985. Oil in the Sea: Inputs, Fates and Effects. National Academy Press, National Research Council, 0309034795, Washington. Park, J.-Y.K., Shigenaga, M.K., Ames, B.N., 1996. Induction of cytochrome P4501A1 by 2,3,7,8-tetrachlorodibenzo-p-dioxin or indolo(3,2-b)carbazole is associated with oxidative DNA damage. Proceedings of the National Academy of Sciences of the United States of America. vol. 93, pp. 2322–2327. Postlind, H., Vu, T.P., Tukey, R.H., Quattrochi, L.C., 1993. Response of human CYP1-luciferase plasmids to 2,3,7,8-tetrachlorodibenzop-dioxin and polycyclic aromatic hydrocarbons. Applied Pharmacology 118, 255–262. Potter, T.L., Simmons, K.E., 1998. Composition of Petroleum Mixtures. The Association for the Environmental Health of Soils, Amherst. Safe, S., 1990. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and related compounds: environmental and mechanistic considerations which support the development of toxic equivalency factors (TEFs). Critical Reviews in Toxicology 21, 51–88. Sanderson, J.T., Aarts, J.M., Brouwer, A., Froese, K.L., Denison, M.S., Giesy, J.P., 1996. Comparison of Ah receptor-mediated luciferase and ethoxyresorufin-o-deethylase induction in H4IIE cells: implications for their use as bioanalytical tools for the detection of polyhalogenated aromatic hydrocarbons. Toxicology and Applied Pharmacology 137, 316–325.
991
Santostefano, M., Merchant, M., Arellano, L., Morrison, V., Denison, M.S., Safe, S., 1993. Alpha-naphthoflavone-induced CYP1A1 gene expression and cytosolic aryl hydrocarbon receptor transformation. Molecular Pharmacology 43, 200–206. Shertzer, H.G. et al., 1998. Dioxin causes a sustained oxidative stress response in the mouse. Biochemical and Biophysical Research Communications 253, 44–48. Short, J.W., Jackson, T.J., Larsen, M.L., Wade, T.L., 1996. Analytical methods used for the analysis of hydrocarbons in crude oil, tissues, sediments, and seawater collected for the natural resources damage assessment of the Exxon Valdez oil spill. American Fisheries Society Symposium 18, 125–132. Stegeman, J.J., Smolowitz, R.M., Hahn, M.E., 1991. Immunohistochemical localization of environmentally induced cytochrome P450IA1 in multiple organs of the marine teleost Stenotomus chrysops (Scup). Toxicology and Applied Pharmacology 110, 486–504. Varanasi, U., 1989. Metabolism of Polycyclic Aromatic Hydrocarbons in the Aquatic Environment. CRC Press, Boca Raton. Velazquez, M., Feirtag, J.M., 1997. Quenching and enhancement effects of ATP extractants, cleansers, and sanitizers on the detection of the ATP bioluminescence signal. Journal of Food Protection 60, 799–803. Webster, T., Commoner, B., 1995. Overview: the dioxin debate. In: Schecter, A. (Ed.), Dioxins and Health. Plenum Publishing, New York, pp. 1–50. Whitlock, J.P., 1999. Induction of cytochrome P4501A1. Annual Review of Pharmacology and Toxicology 39, 103–125. Willett, K.L., Randerath, K., Zhou, G.-D., Safe, S.H., 1998. Inhibition of CYP1A1-dependent activity by the polynuclear aromatic hydrocarbon (PAH) fluoranthene. Biochemical Pharmacology 55, 831–839. Ziccardi, M.H., Gardner, I.A., Denison, M.S., 2000. Development and modification of a recombinant cell bioassay to directly detect halogenated and polycyclic aromatic hydrocarbons in serum. Toxicological Sciences 54, 183–193.