Assessing Aromatic-Hydrocarbon Toxicity to Fish Early Life Stages

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Assessing Aromatic Hydrocarbon Toxicity to Fish Early Life Stages Using Passive Dosing Methods and Target Lipid / Chemical Activity Models Josh David Butler, Thomas F. Parkerton, Aaron Redman, Daniel J. Letinski, and Keith R. Cooper Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b01758 • Publication Date (Web): 11 Jul 2016 Downloaded from http://pubs.acs.org on July 12, 2016

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Environmental Science & Technology

ACS Paragon Plus Environment


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Assessing Aromatic Hydrocarbon Toxicity to Fish Early Life Stages

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Using Passive Dosing Methods and Target Lipid / Chemical Activity Models

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Josh D. Butlera*, Thomas F. Parkertonb, Aaron D. Redmana, Daniel J. Letinskia, Keith R. CooperC

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a

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1545 US Highway 22 East, Annandale, NJ, USA 08801-3059

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b

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800 Bell Street, Houston, Texas, USA 77002

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C

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14 College Farm Rd. New Brunswick, NJ, 08901

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*To whom correspondence may be addressed:

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ExxonMobil Biomedical Sciences, Inc.; 1545 US Highway 22 East; Annandale, NJ 08801;

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[email protected]; 980-730-1003

Toxicology & Environmental Sciences Division, ExxonMobil Biomedical Sciences, Inc.

Toxicology & Environmental Sciences Division, ExxonMobil Biomedical Sciences, Inc.,

Environmental Sciences Department, Rutgers University

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ABSTRACT

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Aromatic hydrocarbons (AH) are known to impair fish early life stages (ELS). However, poorly defined

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exposures often confound ELS test interpretation. Passive dosing (PD) overcomes these challenges by

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delivering consistent, controlled exposures. The objectives of this study were to: apply PD to obtain 5 d

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acute embryo lethality and developmental and 30 d chronic embryo-larval survival and growth effects

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data using zebrafish with different AHs; analyze study and literature toxicity data using target lipid (TLM)

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and chemical activity (CA) models; extend PD to a mixture and test the assumption of AH additivity. PD

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maintained targeted exposures over a six order of magnitude range in concentrations. AH toxicity 1 ACS Paragon Plus Environment

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increased with Log Kow up to pyrene (5.2). Pericardial edema was the most sensitive sub-lethal effect

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that often preceded embryo mortality although some AHs did not produce developmental effects at

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concentrations causing mortality. Cumulative embryo-larval mortality was more sensitive than larval

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growth with acute to chronic ratios < 10. More hydrophobic AHs did not exhibit toxicity at aqueous

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saturation. The relationship and utility of the TLM/CA models for characterizing fish ELS toxicity is

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discussed. Application of these models indicated concentration addition provided a conservative basis

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for predicting ELS effects for the mixture investigated.

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INTRODUCTION

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A key concern with aquatic exposures to aromatic hydrocarbons (AHs) is adverse effects on fish early life

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stages (ELS). In addition to reductions in embryo survival, AHs as well as crude oils or refined petroleum

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substances can produce developmental abnormalities including yolk sac and pericardial edemas and

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cardiovascular, spinal and craniofacial deformities via different mechanisms.1,2,3,4,5 6,7,8,9,10

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A limitation of past ELS studies is the lack of exposure measurements, i.e. tests based on nominal

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concentrations.1,3,5,7,11,12,13,14 Further, even in toxicity studies where analytical confirmation was

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performed, rapid declines in AH exposures due to loss processes,15,16 use of co-solvents17 or test

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concentrations in excess of aqueous solubility18 complicate interpretation and increase uncertainty in

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hazard assessment. Passive dosing (PD) techniques overcome these challenges by delivering well-

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controlled, solvent-free exposures in toxicity tests. 19,20,21,22,23,24,25,26 PD methods involve introducing AHs

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to an inert polymer, e.g. silicone, which is then added to the aqueous media. The loaded polymer then

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serves as a reservoir to buffer AH losses based on partitioning from the polymer to the test media.

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Since it is not possible to conduct toxicity studies on all possible AH structures, toxicity data on individual

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AHs have been used to develop predictive models.27 Once calibrated using reliable empirical data, such

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models can then be used for effects characterization of untested AHs or AH mixtures while reducing

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vertebrate testing.28 The target lipid model (TLM) has been used to derive environmental quality

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objectives in water, soils and sediments for AH substances29,30,31 and has also been applied to assess AH

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mixture toxicity at historically contaminated sites and oil spills using an additive toxic unit (TU) model.32

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,33,34

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risk assessments for petroleum substances.35,36 A second framework for AH effects assessment relies on

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the use of the chemical activity (CA) concept.37,38 Using CA as an alternative exposure metric, AH toxicity

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data are expressed in units that have shown to span a narrow range between 0.01 to 0.1.39 A further

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advantage of this approach is that the toxicity of AH mixtures can also be evaluated by simply summing

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CA of the components.40,41,42,43

This framework has also been used to predict oil toxicity in lab tests and to perform environmental

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The objectives of this study were to: apply PD to obtain acute and chronic effect data for zebrafish ELS

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with selected AHs; compare toxicity data on single AHs from this study and literature to TLM and CA

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frameworks; extend PD to a simple, defined mixture and test the assumption of AH additivity. Zebrafish

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were selected due to small size, low cost, high fecundity, transparent embryo, which facilitates visual

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identification of developmental abnormalities, and growing use in standardized testing and

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research.44,45,46

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MATERIALS AND METHODS

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Test Materials

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Silicone, i.e. polydimethylsiloxane (PDMS), O-rings (0.39g and 1.0g) were obtained from O-rings West

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(Seattle, WA). Naphthalene, 1-methylnaphthalene, biphenyl, phenanthrene, 1-methylphenanthrene,

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1,2,3,4,5,6,7,8,-octahydrophenanthrene, bicyclohexyl, pyrene, perhydropyrene, benzo(a)pyrene, pyrene

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were purchased from suppliers. Perhydrophenanthrene was synthesized internally by catalytic

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hydrogenation of phenanthrene. Stock AH solutions were prepared in methanol (HPLC grade) obtained


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from J.T. Baker (Phillipsburg, NJ). An overview of substances used in this study including suppliers,

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purities and relevant physicochemical properties estimated using SPARC47 are provided in Table S1.

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Test Organisms

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An outbred strain of zebrafish genotyped by Charles River Laboratories International, Inc. (Wilmington,

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MA) and obtained from Aquatic Research Organisms (Hampton, NH) was used in this study

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(documentation available upon request). Embryos were obtained from a breeding culture previously

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described.21 Spawning and fertilization occurred within 30 minutes after the onset of light. Eggs were

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collected and immediately rinsed with water, transferred to a crystallization dish and examined for

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viability before use in toxicity testing.

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Determination of Partition Coefficients

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Passive dosing was used to generate and maintain test substance aqueous concentrations. In order to

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accurately calculate predicted final water concentrations experimentally derived partition coefficients

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for ten compounds (Table S1 and S2) were determined by loading three separate 20 mL glass vials, each

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with four 0.39 g O-rings, with 10 mL of a known AH concentration in methanol. After a 72 h equilibration

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on a vial roller at 28℃, the methanol solution was poured off and an aliquot was diluted 1:20 with

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acetone and analyzed. Two O-rings from each vial were extracted four times each with 10 mL of acetone

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on a wrist-action shaker for one hour. An aliquot of each extract was then analyzed. The remaining two

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O-rings were rinsed with glass distilled water and wiped dry with lint-free tissue paper and then added

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to a 20 mL glass vial containing reconstituted water (see below) on an orbital shaker at 400 rpm for 24

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hours at 28℃. Partition coefficients between methanol and PDMS (KMeOH:PDMS) and between PDMS and

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water (KPDMS:Water) were calculated based on analysis of test substance concentrations in each phase.

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Toxicity Test Dosing



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O-rings served as the PD source for embryo/larval exposures. Methods detailed by Butler et al., 21 were

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employed to load test substances from the methanol/AH solution into O-rings and subsequently dose

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the aqueous phase. Methanol alone was equilibrated with O-rings used in the control replicates.

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Nominal concentrations of each compound were predicted by dividing methanol concentrations by the

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AH specific methanol-water partition coefficient ( see results section). Predicted concentrations were

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then compared to measured concentrations to assess accuracy and maintenance of targeted exposures.

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Test systems were designed to ensure that the mass of chemical absorbed into the O-rings was

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sufficient to prevent significant depletion following subsequent dosing of aqueous test media

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(Supplemental Information, Appendix A).

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Acute toxicity tests

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Embryos were exposed individually to naphthalene, 1-methylnaphthalene, biphenyl, phenanthrene,

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pyrene, 1,2,3,4,5,6,7,8-octahydrophenanthrene, benzo(a)pyrene, and chrysene. Each test consisted of 5

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concentrations and a control, except for benzo(a)pyrene and chrysene which were investigated only at

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the respective solubility limits. Embryos were exposed in 20 mL headspace vials with a 0.39 g O-ring

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resting on the bottom. Vials were then capped with Teflon coated screw caps and placed into an

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environmental chamber at 28±1°C. Twenty replicates with one embryo per vial comprised each

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treatment. In vial microscopic observations were performed at 24-h intervals. Lethal (coagulation, lack

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of heartbeat) and sub-lethal (pericardial and yolk sac edemas and tail curvature) endpoints included

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those defined in the OECD (2006) FET draft guideline up to 5 days post fertilization (dpf).48 The

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photoperiod was set to a 12:12 light/dark cycle with a light intensity of 250 lux. Test media for the acute

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tests was prepared using Instant Ocean® sea salt, 60 µg/mL stock salts. 49 Water quality parameters

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(dissolved oxygen, temperature, conductivity, pH) were monitored and maintained within test

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guidelines requirements. For all acute toxicity tests the dissolved oxygen was ≥80%, pH ranged from 7.2-


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7.6, Conductivity was 130-150 µS/cm and temperature ranged from 25.8-26.4℃. Duplicate samples for

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analytical confirmation of AH exposures were taken for each treatment at test start and termination.

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Chronic Individual Compound and Mixture Toxicity Tests

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Embryo/larvae were exposed for 30 d to five AHs individually (1-methylnaphthalene, phenanthrene, 1,

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2, 3, 4, 5, 6, 7, 8-octahydrophenanthrene, chrysene, benzo(a)pyrene). Tests with the first three

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substances consisted of 3 concentrations and a control while the remaining two substances were

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performed as limit tests at test media saturation. A separate test was performed using a mixture of 7

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AHs and 3 hydrogenated analogs (Table S1) with three treatment concentrations and a control. Half of

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the compounds included in the mixture test were included in single compound experiments while the

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remainder of the test compounds were not tested individually. These exposures, therefore, represent a

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test of the predictive toxicity model derived from single compound tests for estimating the contribution

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of untested components to the resulting mixture toxicity (see Data Analysis section). Moderately hard

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reconstituted water was used as the test media from day 5 through 30 of each chronic test.50 Water

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quality parameters (dissolved oxygen, temperature, conductivity, pH) were monitored and maintained

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within test guidelines requirements. For all chronic toxicity tests the dissolved oxygen was ≥80%, pH

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ranged from 7.4-7.8, Conductivity was 320-375 µS/cm and temperature ranged from 25.1-26.1℃.

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Embryos were initially exposed in vials for 5 d as described for acute experiments. Upon 5 days post

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fertilization (dpf), embryos that hatched into larval fish as well as any unhatched embryos exposed in

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vials were transferred into four replicate 300 mL flow-through chambers, described previously by Butler

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et al.21 and depicted in Figure S1 each with up to 5 fish per chamber (i.e. fewer larvae were used in

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treatments with earlier embryo mortality). Four-liter mixing vessels containing 30 loaded O-rings (1.3

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each), were prepared for passively dosing each treatment. Each individual flow-through chamber was



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also supplemented with three, 1.3g loaded O-rings to buffer potential losses in this recirculating system.

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The test media flow (5 mL/min) through each chamber was supplied by a peristaltic pump. Mixing

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vessels were supersaturated with oxygen to > 20 mg/L for ensuring adequate oxygen concentrations

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during exposures and exchanged every 5-7 d with a new set of loaded O-rings to avoid biofouling and O-

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ring depletion. Upon transfer to flow through chambers, fish were fed Paramecium

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multimicronucleatumn.51 Starting at 12 dpf the diet was supplemented with Artemia nauplii. Larvae

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were fed ad libitum, twice daily throughout the test. The photoperiod was set to a 12:12 light/dark cycle

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with a light intensity of 250 lux.

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For each chronic test, aqueous samples were collected for AH analysis from each treatment and control

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at test start and on day 5 prior to the addition of larval fish to the flow through chamber. While larval

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fish were held in flow-through chambers, subsequent water samples were taken for AH analysis from

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each new and corresponding old mixing vessel at least once every 5-7 d. Chronic test endpoints for the

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first 5 days included those described for the acute test. Beginning on day 6, hatched larvae were

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observed daily for immobility, lack of respiratory movement and reaction to mechanical stimulus per the

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OECD guideline.52 To evaluate potential larval growth effects, total lengths of surviving larvae were

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measured at test termination.

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Test Substance Analysis for Exposure Confirmation

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Two, 40 mL glass vials were collected at each sampling point for quantifying AH exposures. Analysis for

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naphthalene, biphenyl, phenanthrene and 1-methylnaphthalene in water was performed using static

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headspace gas chromatography with flame ionization detection (HS-GC-FID) using a Perkin Elmer

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Autosystem XL gas chromatograph. Analysis of octahydrophenanthrene was performed by HS-GC-FID

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modified with a trap accessory to provide additional sensitivity. Pyrene, benzo(a)pyrene, chrysene and


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all mixture components were analyzed by automated direct immersion (DI) solid phase microextraction

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(SPME) coupled with gas chromatography with mass selective detection (GC-MSD).

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Water samples were collected in duplicate in glass volatile organic analysis (VOA) vials at each sampling

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point to quantify AH exposure concentrations. Aliquots were transferred to ca 20 mL glass, septum

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sealed vials for either headspace or direct immersion SPME along with the internal standard.

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The HS-GC-FID was equipped with a 30 mm x 0.53 mm id, 1.5 µm film DB-5 analytical column (J&W

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Scientific) which was connected directly to the transfer line of a Perkin-Elmer TurboMatrix 40 Trap

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Headspace Sampler. For octahydrophenanthrene analysis, a sorbent trap containing fused silica beads

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and Carbopack C was placed in-line of the headspace sampler. All water samples for headspace analysis

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and standards spiked into water were equilibrated for 45 m at 90°C. The needle and transfer line

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temperatures were both 175°C, the pressurization time was 3 m. The injector temperature was 50°C and

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column pressure was 28 psi. The FID was 275°C and the oven temperature was held at 40°C for 1 m and

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then ramped up to 275°C at 20°C/minute. Data were acquired and processed using Perkin Elmer

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TotalChrom Workstation software.

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A DI-SPME autosampler was coupled with an Agilent 6890N gas chromatograph and Agilent 5975 mass

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selective detector operated in the selective ion monitoring mode (SIM). The analytical column was a 25

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m x 0.2 mm id, 0.33 µm DB-5MS (J&W Scientific) capillary column. The GC-MSD system was equipped

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with a Gerstel CIS 4 inlet and interfaced with a Gerstel MPS 2 autosampler operated in the DI-SPME

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mode using a 7-µm PDMS fiber (Supelco) with a 90-minute extraction time at 30oC. The SPME fiber was

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automatically desorbed for 5 minutes at 300oC. The GC oven temperature was programmed from 50oC

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for 5 mto 340oC at 12oC/minute. After an initial 25 psi pressure pulse for one minute, the helium carrier

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gas pressure was set at 13.8 psi corresponding to a constant flow of approximately 0.8 mL/m for the



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duration of the run. Standards analysis for both methods resulted in a linear response over

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concentration ranges bracketing the respective water concentrations.

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TLM and CA Model Frameworks

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The TLM describes the LC50 of a nonionic chemical using the relationship:

LC50 =

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(1)

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where:

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LC50= lethal (or effective) concentration in water causing 50% response (mmol/ Lwater)

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CTLBB = critical target lipid body burden corresponding to the acute effect (mmol/kglipid)

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KTLW = target lipid water partition coefficient (kg lipid/ Lwater)

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A linear free energy relationship is used to estimate the target lipid water partition coefficient from the

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octanol-water partition coefficient. Following rearrangement of equation (1) and substitution of the

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correlation between KTLW and Kow,, LC50 is expressed as:

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Log LC50 = log (CTLBB) - 0.94 Log Kow + Δc

(2)

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The TLM indicates that toxicity depends on both the sensitivity of the species, as defined by the CTLBB,

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and the partitioning properties of the test substance. The Δc term is a correction that accounts for

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partitioning differences not accounted for by Kow and has been empirically derived for various functional

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groups by fitting equation (2) to toxicity data for a large array of chemicals. In the case of mono and

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poly AHs, Δc = -0.109 and -0.352, respectively. 29


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In the chemical activity (CA) framework, toxicity is defined as a fractional solubility of the test substance

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relative to the reference liquid state:37

LA50 =

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(3)

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where:

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LA50= lethal (or effective) activity in water causing 50% response (unitless)

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SL = aqueous solubility of the sub-cooled liquid (mmol/Lwater)

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For hydrocarbons, SL can also be estimated from Kow.53

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Log SL =3.54 - 1.10 Log Kow

(4)

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Substituting this relationship into (3) and solving for LA50 yields:

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Log LC50 = Log (LA50) + 3.54 - 1.10 Log Kow

(5)

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Equating (2) and (5) shows the predicted relationship between the TLM and CA frameworks:

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Log LA50 = Log (CTLBB) + 0.16 Log Kow - 3.54 + Δc

(6)

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Thus, LA50 depends on the species and endpoint-specific CTLBB and is predicted to be positively

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correlated to substance Log Kow as modulated by chemical class-specific correction factors that account

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for substance partitioning to target lipid.

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Data Analysis

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Acute (5-d LC50) and chronic (30 d EC10) effect concentrations were calculated by probit analysis using

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SAS

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determine significant differences from the control. The logistic equation was used was used to fit dose-

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. The T-test with Bonferroni adjustment56or Fisher’s Exact Test 57 58using TOXSTAT 59software to



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response zebrafish mortality data using a least squares regression method included in the GraphPad

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Prism 5.0 software.60

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Empirical LC50 data were adjusted based on the TLM prescribed class corrections (e.g., logLC50 - Δc, by

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re-arranging Eqn. 2) and then fit by linear regression using equation (2) to derive the CTLBB that

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characterizes the acute sensitivity of zebrafish embryos29. This relationship was further extended to

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predict chronic toxicity by subtracting the empirically derived mean log of the Acute to Chronic Ration

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(ACR). For the mixture test, exposure concentrations for each substance were converted into Toxic

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Units (TUs) by dividing the targeted exposure concentration of each component by the TLM-derived

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acute or chronic effect concentration. Summed TUs (∑TUs) were then compared with observed effects

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to assess concordance with concentration addition. Test concentrations were selected so that each

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compound contributed about equally to TUs and three treatment levels plus a control were selected to

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provide ∑TUs bracketing unity which corresponds to treatment levels above and below the appropriate

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effect level (LC50 for acute and LC10 for chronic). The ∑CA was used as an alternate exposure metric for

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evaluating mixture toxicity by dividing exposures for individual mixture components by the

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corresponding sub-cooled liquid aqueous solubility (Table S1) and then summing the CA of all mixture

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components.

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RESULTS & DISCUSSION

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Partition Coefficients

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Measured AH partition coefficients are summarized in Table S2. Log KMeOH:PDMS ranged between -0.070 to

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1.41 while Log KPDMS:water varied from 2.98 to 5.16. Experimental KMeOH:Water values were found to

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correlate to the octanol-water partition coefficient:

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K : = 0.825 × K + 0.390 r2 = 0.96, n=10

(7)

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This relationship is consistent with KMeOH:Water values reported by Smith et al. 61 (Figure S2). Equation (7)

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was used to estimate KMeOH:Water for the remaining AHs for which measured partition coefficients were

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not determined.

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Analytical confirmation of test concentrations

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To evaluate the ability of PD to deliver and maintain exposure concentrations, predicted nominal

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concentrations were compared to analytically determined concentrations in test media. Results

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indicated excellent agreement between targeted and measured concentrations across test compounds

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that span a wide range of physicochemical properties and exposure treatments spanning over six orders

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of magnitude (0.1 to 19,436 µg/L) (Figure S3; Table S3). These results indicate that the equilibration

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periods selected for dosing o-rings and water were sufficient to achieve equilibrium. The mean ratio of

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predicted to observed concentrations across all tests was 1.29±0.66; n=70 (Table S3). PD maintained

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exposures with a mean percent difference in measured concentrations of 9±6% between the start and

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end of 5 d embryo tests (Table S4). Figure 1 further illustrates the fidelity of exposures over 30 d chronic

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tests (Figure 1, A-D).

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Acute Exposures

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No control mortality was observed in any test. Most test substances showed statistically significant sub-

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lethal developmental effects on embryos. Pericardial edema was the most sensitive developmental

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effect noted and often preceded mortality, but this was not always the case as several AHs did not

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exhibit developmental effects at exposure concentrations causing mortality (Table S5). 5-d LC50s are

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reported in Table 1 and concentration-responses for embryo mortality are shown in Figure S4. Where 12 ACS Paragon Plus Environment


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sufficient concentration-responses were observed, EC50s for sub-lethal effects were also calculated

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(Table S6). However, reliable estimates were in some cases not possible due to reduced sample sizes

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from embryo mortalities during test exposures. With the exception of biphenyl and naphthalene, where

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the most sensitive 2 d developmental effect EC50 values were two or three fold lower than the 5 d LC50,

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respectively. EC50s were similar to LC50 estimates (phenanthrene and pyrene) or higher for the

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remaining compounds (Table 1).

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The 5 d LC50 for zebrafish recently reported using PD for phenanthrene by Vergauwen24 of 310 µg/L is in

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close agreement with the value of 334 µg/L obtained in this study. These authors also report that after

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5 d exposure to phenanthrene at 362 µg/L 100 and 70% of surviving embryos exhibited tail curvature or

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pericardial or yolk sac edemas, respectively, which seems consistent with the 2 d EC50 of 415 µg/L for

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these endpoints reported in Table 1. A range of other sublethal effects were also investigated in this

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study and swim bladder inflation was reported to be the most sensitive endpoint with a 5 d EC50 of 59

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µg/L. These authors noted this is a common sub-lethal effect observed for many substances in different

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classes and varying modes of action. This endpoint, which may reflect either a delay or lack of swim

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bladder inflation, was not evaluated in the present study. Direct comparison of developmental fish

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embryo abnormalities observed in this work to other literature is difficult given the uncertainty in

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exposure concentrations, varying exposure durations, alternative fish species and different sub-lethal

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endpoints investigated in earlier studies.

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In contrast, these authors report sub-lethal effects on swim bladder and other morphological effects on

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developing embryos at six-fold lower exposure concentrations than we observed. These results suggest

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the potential differences in inter-lab reproducibility of lethal versus sub-lethal embryo endpoints even

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when PD methods are used.

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No acute or sub-lethal effects were noted at the highest, single limit concentration tested for chrysene

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and benzo(a)pyrene (Table 1, Table S5). The lack of acute effects reported for these two substances is

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consistent with the findings of Seiler et al. (2014)22 who also showed zebrafish embryo mortality during

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2 d tests was not significantly different from controls at the highest concentration attainable using a PD

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test design. Further, pyrene was the most hydrophobic AH compound tested by Seiler that caused

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embryo mortality consistent with the findings of this study (Table 1).

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The 2 - 5 d LC50s (Table S7, Table 1) were fit to the TLM using equation (2) in order to derive CTLBB

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estimates. The CTLBBs for zebrafish were estimated to be 158 ± 6 and 81 ± 10 µmol/g octanol,

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respectively. These values fall within the range of 24 to 500 µmol/g octanol reported previously for

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different test species and acute endpoints.29 Additional acute zebrafish toxicity data for AHs and other

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nonpolar organic chemicals including aliphatic hydrocarbons and alcohols, halogenated solvents, and

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oxygen and sulfur containing AHs were compiled from the literature (Table S7). Data for 2-5 d test

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exposures are plotted separately in Figure 2A and B along with the TLM model derived from our study

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data. Literature data based on conventional dosing techniques (i.e. semi-static, static, flow through test

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substance renewals, etc.) include both nominal (unfilled) and measured (filled symbols) exposures.

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Tests showing no acute effects at the highest exposure concentration investigated are shown as a plus

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symbol. The relationship between sub-cooled solubility and Log Kow described earlier, are used to

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generate the dashed lines corresponding to CAs of 0.01, 0.1 and 1.0 on these figures.

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Figure 2 indicates acute LC50s, adjusted using the relevant TLM class correction factor to express results

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in terms of baseline toxicity, for water soluble substances with Log Kow < 3 fall within approximately a

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factor of three above or below the TLM prediction with LA50s generally within the 0.01 to 1.0 range.

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However, as the Log Kow increases, available literature data shows more scatter that likely reflects the 14 ACS Paragon Plus Environment


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greater uncertainty in the actual exposure concentrations associated with conventional dosing methods

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of sparingly soluble and sometimes volatile test substances. In fact, some tests indicate LA50s that are

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above the theoretical limit of unity thus questioning test reliability. In a recent study that investigated

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algal toxicity to fifty organic chemicals ranging in Log Kow from 1 to 5 and expected to act via baseline

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toxicity, LA50s fell within a similar range of 0.1 to 1.0 and increased with Log Kow consistent with the

331

zebrafish data in this study (Figure 2) and predictions from equation (6).62

332 333

Further comparison of acute zebrafish toxicity data included in Table S7 for nitrogen containing AHs to

334

the TLM frameworks indicates toxicity is systematically underestimated by an order of magnitude

335

(Figure S5). Previous work on benthic invertebrates has shown that nitrogen containing heterocyclic AHs

336

can exhibit greater toxicity than the corresponding homocyclic AHs.63 However, if the correction term in

337

equation 1 is adjusted to account for differential partitioning of this compound class to target lipid,

338

revised TLM predictions are in good agreement with empirical data (Figure S5). This provides an

339

example of how interrogating toxicity data sets using the TLM framework can lead to improved models

340

for toxicity prediction.

341 342

Chronic Exposures and ACRs

343

Less than 10 percent control mortality was observed in all chronic exposures. Cumulative embryo larval

344

mortality over 30 days was used to estimate LC10 values (Table 1). Mortality showed a similar temporal

345

trend across substances. Once fish matured past the elutheroembryo developmental stage (days 5-7)

346

most fish survived (Figure S6). Further, based on measured lengths of surviving larvae at test

347

termination, no effects on larval growth were observed for 1-methylnaphthalene and

348

octahydrophenanthrene. For phenanthrene, while a statistically significant 12 % reduction in growth

349

was observed at 105 µg/L, cumulative embryo larval survival was a more sensitive endpoint than larval


Page 16 of 32

Page 17 of 32


350

growth. The acute sub-lethal embryo developmental endpoints evaluated in this study appeared less

351

sensitive than the chronic test endpoints (Table 1). Consistent with acute toxicity results, chronic values

352

decreased with increasing logKOW (Table 1). The ACRs calculated from these tests (5.8 to 7.6) fall within

353

the range of values (1 to 13) reported previously for AHs with other test species29 as well as data

354

compiled for a broader class of nonpolar organic chemicals (median ACR=4.4; range 1-30, n= 19).64 No

355

chronic effects were observed for chrysene and benzo(a)pyrene at mean measured exposure

356

concentrations of 1.4 and 1.7 μg/L, which appears to be the limit of saturation in this test system. These

357

exposures are close to the reported aqueous solubilities of 1.6 and 1.5 μg/L, respectively.65

358 359

To gain further insights on the chronic effects of AHs to fish ELS, three data sets generated from testing

360

multiple AH compounds were compiled from the literature (Table S8). Literature chronic values are

361

plotted for comparison with our study data (red symbols) in Figure 3. Unfilled symbols denote test data

362

where chronic effects were not observed at the maximum exposure concentration tested. The chronic

363

TLM line was derived from the relationship from our acute data after dividing by the mean acute to

364

chronic ratio of 6.9 (Table 1). This is equivalent to assuming a CTLBB for chronic effects by application of

365

the median ACR determined in the present study of 81/6.9 = 12 µmol/g octanol. To facilitate comparison

366

of effects data in terms of CA, dashed lines ranging 0.001 to 0.1 are also shown.

367 368

The first data set was obtained from a recent review summarizing unpublished chronic zebrafish ELS

369

tests performed using static renewals with conventional dosing methods. 66 Like our chronic tests, both

370

survival and growth endpoints were investigated. While results were expressed in terms of mean

371

measured exposure concentrations, significant declines in exposure were noted thereby introducing

372

uncertainty in the chronic values reported. Results from this study are shown as gray symbols in Figure

373

3 and decrease with increasing Log Kow consistent with TLM predictions until 5.5, above which, three of



374

the four AHs exhibit no chronic effects based on either endpoint. Both chrysene and benzo(a)pyrene

375

that were evaluated in this study was not chronically toxic in 42 d tests at mean concentrations of 0.9

376

and 4 μg/L, respectively, in support of our study findings. In contrast, benzo(k)fluoranthene was

377

reported to cause chronic effects on growth and survival at mean concentrations of 0.62 and 0.17 μg/L,

378

respectively (Table S8). Further, for all AHs that exhibited chronic effects, larval growth appeared more

379

sensitive in these studies than survival in contrast to our findings (cf. squares and diamonds in Figure 3).

Page 18 of 32

380 381

Two recent studies have also investigated ELS toxicity to a different fish species (Japanese medaka) using

382

PD for a range of parent and alkylated AHs.20,27 Previous work has shown that the acute CTLBB for this

383

species is 127 ± 57 µmol/g octanol suggesting similar sensitivity to zebrafish.29 Given the limited use of

384

PD to characterize ELS toxicity with AHs, we have compared results from these studies to zebrafish data

385

derived in this study. In both studies, toxicity was expressed in terms of a 17 d EC50 that incorporated

386

several embryo pathology endpoints that were scored to provide an effect index. In the first study

387

phenanthrene and corresponding methyl, ethyl and dimethyl homologs were tested. All test

388

compounds were found to produce EC50s at concentrations below aqueous solubility limits. In the

389

second study, chrysene and benz(a)anthracene and six alkylated homologs were investigated. All

390

compounds, except chrysene, were found to produce sub-lethal effects on embryos at the solubility

391

limit including benzo(a)pyrene (Table S8, Figure 3) but only retene (i.e. 1-methyl-7-

392

isopropylphenanthrene ), which served as the positive control, benzo(a)anthracene and 7-methyl

393

benzoanthracene yielded sufficient concentration responses below the solubility limit to allow EC50s to

394

be calculated. Further, none of these compounds, except the positive control, reduced embryo-larval

395

survival after 17 d to exposures corresponding to the solubility limit. Other studies have reported similar

396

findings. For example, no significant differences in either survival or hatching success from the control


Page 19 of 32


397

were reported for rainbow trout eggs exposed 36 d post-fertilization to measured concentrations of

398

benzo(a)pyrene up to 3 μg/L although developmental effects were noted.67

399 400

EC50s for medaka embryo pathology also decrease with increasing log Kow. The zebrafish chronic TLM

401

prediction is shown to be protective (i.e. data points fall above model line) for this endpoint (Figure 3).

402

Further, chronic effects generally corresponded to CA > 0.01. In a number of tests, chronic effects were

403

not observed at CA > 0.1. This suggests that lack of chronic effects cannot only be attributed to an

404

insufficient chemical activity resulting from solubility constraints. One alternative explanation is that

405

toxicokinetics may limit internal residues and preclude potential toxicity of AHs.9 While this may help

406

explain a lack of effects observed in short term exposures, this hypothesis seems less plausible for

407

chronic exposures where equilibrium with tissues would be achieved. A second explanation to account

408

for no chronic effects at elevated CA is biotransformation processes may limit exposure of higher log Kow

409

AHs or related metabolites in embryo-larvae tissues to concentrations that do not invoke chronic

410

effects. Biotransformation rates in fish have been shown to increase with AH hydrophobicity.68,69

411

Further, biotransformation processes may also help explain why certain hydrophobic AHs exhibit chronic

412

toxicity at much lower CAs. For certain AHs, specific metabolites may be produced that enhance chronic

413

effects. Recent bioconcentration studies of benzo(a)anthracene with zebrafish embryos have

414

documented formation of a polar metabolite that might help explain the high chronic toxicity reported

415

(Table S8)70 However, uncertainties in exposure concentrations in available ELS toxicity studies with

416

benzo(a)anthracene warrant further research to determine if previously reported chronic values can in

417

fact be replicated using controlled exposures with PD methods.

418 419

Further investigation is also needed to assess the chronic effects of 4+ ring AHs on reproductive

420

endpoints. In one study, 71 no effects on zebrafish survival was reported after a 56 d exposure to



421

benzo(a)pyrene at two treatment levels with mean reported concentrations of 1.63±0.32 and 3.35±0.91

422

µg/L consistent with other studies (Table S8) . However, a statistically significant 50% reduction in egg

423

production was observed after 49 d in both treatments. Altered transcription of genes important in

424

regulating reproduction was also demonstrated and hypothesized as an underlying mechanism of

425

benzo(a)pyrene reproductive toxicity.

426 427

Mixture Exposures

428 429

The effects on zebrafish ELS are summarized in Table 2. Cumulative embryo-larval mortality exhibited

430

same trend over time as individual compound tests (Figure S6). Measured exposure concentrations

431

(Table S9A) were translated into both ∑ TUs (Table S9B) and ∑ CA (Table S9C) as alternative exposure

432

metrics. Sub-lethal embryo effects were observed after 5 days in the lowest treatment group

433

corresponding to 0.1 acute ∑TUs. However, statistically significant effects on survival after 5 d were only

434

observed at the highest treatment in which a 30% embryo mortality corresponded to an acute ∑TU = 1.2

435

(Table 2). This observed mortality is less than predicted based on concentration addition of mixture

436

components (i.e. 50% response predicted at ∑TU = 1) but within typical variability noted previously

437

based on application of the additive model (~2-fold).29

438 439

This treatment corresponds to an estimated 5 d EA30 of 0.319. However, if the CA contribution of the

440

two non-aromatic hydrocarbons (perhydrophenanthrene and perhydropyrene) which have estimated

441

log Kow’s of 6.76 and 7.54 are ignored, the 5 d EA30 is reduced to 0.033. This estimate is similar to the PD

442

test results of Seiler22 who derived a 2 d LA50s of 0.059 to 0.087 by comparing observed zebrafish

443

mortality at the maximum chemical activity achievable for different AHs.22 These results suggest that

444

these compounds do not contribute to acute ELS effects at CA of 0.09 and 0.18, respectively (Table S9C).


Page 20 of 32

Page 21 of 32


445 446

Cumulative mortality was the most sensitive endpoint when exposures were extended 30 d to include

447

the larval stage. Although fish length was quantified at test termination no effects on growth were

448

observed (Table 2). If the mixture components are additive, a chronic ∑ TU = 1 should correspond to 10%

449

mortality since the LC10 served as the chronic effect endpoint. Adjusting for control performance, 0 and

450

25% mortality was observed at chronic ∑ TU of 0.54 and 3.76, respecƟvely which brackets the expected

451

value of unity (Table 2). Assuming, the mixture components contribute additively to toxicity, the TLM

452

appears to provide a reasonable basis for estimating the chronic effects of the mixture investigated.

453

If the mixture test is expressed in terms of ∑ CA, the LA10 would fall within the range of 0.016 to 0.171

454

(Table 2) which corresponds to CAs associated with chronic effects for individual AHs (c.f. Figure 3).

455

Unlike acute tests, these results suggest that the two more hydrophobic non-aromatic test substances

456

contributed to observed chronic effects. The relative contribution of these two components to chronic

457

∑ TUs and ∑ CA at the low and mid dose treatments ranged from 56-79% and 86-95%, respectively Table

458

S9).

459 460

This analysis highlights the comparative use ∑ TU and ∑ CA as exposure metrics for interpreting mixture

461

toxicity studies. The use of ∑ TU provides the advantage of allowing differences in species and endpoint

462

sensitivity to be explicitly taken into account. Further, since expression of toxicity results in terms of CA

463

even for a given test organism and endpoint is not chemical independent, the use of ∑ TU is

464

hypothesized to provide a more precise exposure parameter for predicting toxicity across mixtures with

465

different compositions.

466 467

This study also demonstrates the utility of PD in maintaining constant exposures that allow reliable

468

chronic fish ELS effects data to be developed for individual AHs and related mixtures. While we



469

investigated several sub-lethal embryo developmental effects that were included in a draft test

470

protocol, these endpoints are not included in the final OCED test guideline.72 Future research is

471

therefore needed to better understand the inter-laboratory precision of various sub-lethal embryo

472

endpoints including impairment of swim bladder inflation and the relationship of these effects to

473

chronic effects on survival, growth and reproduction which serve as the key endpoints used in chemicals

474

management73. This work also describes the linkage between TLM and CA models and illustrates how

475

both frameworks can be practically applied to characterize the acute and chronic toxicity of individual

476

substances and mixtures. Additional effort is needed to objectively compare the relative performance

477

of ∑ TU and ∑ CA as exposure metrics for predicƟng AH mixture toxicity.


Page 22 of 32

Page 23 of 32

478


Table 1. Summary of zebrafish embryo-larval acute, sub-lethal and chronic toxicity endpoints.

Test Substance

2 to 4 d EC50 embryo development

(µg/L)1

488 489 490

Slope 2 Parameter

30-d LC10 embryo-larval survival

30-d EC10 larval growth

(µg/L)

(µg/L)

(µg/L)

Acute to Chronic 3 Ratio

6309 2.73 NT NT CNC (5888-6760) 1013 141 (1361-methylnaphthalene >1716 1.88 >277 7.2 (1182-1703) 144) 973 1548 biphenyl 2.04 NT NT CNC (971-976) (1148-2290) 415 334 phenanthrene 1.57 44 (22-65) 88 7.6 (CNC) (184-1493) 1,2,3,4,5,6,7,8 52 >682 0.76 9 (CNC-13) >88 5.8 octahydrophenanthrene (39-80) 91 96 pyrene 3.16 NT NT CNC (34-147) (87-102) chrysene >1.4 >1.4 N/A >1.4 >1.4 N/A benzo(a)pyrene >1.7 >1.7 N/A >1.7 >1.7 N/A Values in parenthesis indicate confidence intervals NT= not tested; N/A = not applicable as no effects observed at maximum concentration tested; CNC =could not calculate since chronic testing not performed 1 slope parameter of logistic equation that define lines plotted in Figure S4 2 acute to chronic ratio calculated as 5 d LC50 / lowest 30 d chronic value naphthalene

479 480 481 482 483 484 485 486 487

5-d LC50 embryo mortality

2343 (2372-2358)

Table 2. Acute and chronic effects observed during 30-days exposure to a defined mixture of aromatic and saturated cyclic hydrocarbons. Acute Exposure 5 d Embryo Acute Effects Chronic Exposure 30 d ELS Chronic Effects Toxic Chemical PE TC Mortality Toxic Chemical Mortality Length ± SD (cm) Units Activity (%) (%) (%) Units Activity (%) 0 0 0 0 0 0 0 5 1.57 ± 0.22 0.10 0.015 0 26* 0 0.47 0.016 5 1.59 ± 0.12 0.46 0.151 26* 35* 5 3.76 0.171 30* 1.46 ± 0.10 1.21 0.309 35* 7 30* 10.52 0.374 90* 1.75 ± 0.35 *Statistically significant (p=0.05) PE=pericardial edema; TC= tail curvature; for embryo developmental endpoints the percentage refers to the proportion of the test population affected that were alive



A

Page 24 of 32

B

1-Methylnaphthalene

Phenanthrene 1000

Cmeasured (ug/L)

1000

100

100

60 ug/L

277 ug/L

105 ug/L

32 ug/L

1227 ug/L

423 ug/L

10 0

10

20

30

0

10

20

C

D

1,2,3,4,5,6,7,8-Octahydrophenanthrene

Chrysene & Benzo(a)Pyrene

30

10

Cmeasured (ug/L)

100

1 10

25 ug/L

5 ug/L

1.4 ug/L Chrysene

88 ug/L

1

1.7 ug/L BaP

0.1 0

10

20

30

0

Time (days)

10

20

30

Time (days)

491 492 493 494 495 496

Figure 1. Measured aqueous concentrations in chronic tests. Each symbol represents the mean of two replicates plus the standard deviation (often not visible due to low variability). Solid lines denote predicted nominal concentrations. Concentrations reported for each compound at the bottom of plot is the 30 day mean value. Analysis of the high treatment level in 1-methylnaphthalene and phenanthrene tests (Figure 1, A and B) was terminated before 30 d due to complete mortality.

497


Page 25 of 32


498

(b)

499 500 501 502 503 504 505 506 507 508

Figure 2. Zebrafish LC50s as a function of substance Log Kow. Empirical data are adjusted using reported class-specific correction factors (e.g., logLC50 - Δc, by re-arranging Eqn 2) to allow direct comparison to the target lipid model shown as the solid line; dashed lines indicate approximate chemical activities corresponding to 0.01, 0.1 and 1.0 from bottom to top respectively; Literature data based on conventional dosing techniques (i.e. semi-static, static, flow through test substance renewals, etc.) include both nominal (unfilled) and measured (filled symbols) exposures. Data from this study are represented as filled red squares. Data points denoted as “+” indicate no acute toxicity was observed at the maximum concentration indicated; (a) 2 d LC50s (b) 4 and 5 d LC50s; Data in this plot are provided in Table S7.

509



510 511 512 513 514 515 516

Figure 3. Chronic effect values for aromatic hydrocarbons a function of substance Log Kow. Empirical data are adjusted using reported class-specific correction factors (e.g., logLC50 - Δc, by re-arranging Eqn 2) to allow direct comparison to the target lipid model shown as the solid line; dashed lines indicate approximate chemical activities corresponding to 0.001, 0.01 and 0.1 from bottom to top respectively; Data in this plot are provided in Table S8.


Page 26 of 32

Page 27 of 32

517


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Assessing Aromatic-Hydrocarbon Toxicity to Fish Early Life Stages

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