OCS Study 

•

MMS 2006-008 

Use of Biological Endpoints in Flatfish 

to Establish Sediment Quality Criteria 

for Polyaromatic Hydrocarbon 

Residues and Assess Remediation 

Strategies 

Final Technical Summary 

Final Study Report 

 

U.S. Department of the Interior 

Minerals Management Service

Pacific OCS Region 

OCS Study 

MMS 2006-008 

Use of Biological Endpoints in Flatfish 

to Establish Sediment Quality Criteria 

for Polyaromatic Hydrocarbon 

Residues and Assess Remediation 

Strategies 

Final Technical Summary 

Final Study Report 

Authors 

Luke A. Roy, Scott Steinert, Steve M. Bay, Darrin Greenstein, Yelena 

Sapozhnikova, Ola Bawardi, Cherlynn Seruto, Ira Leifer, Daniel Schlenk 

Principal Investigator 

Daniel Schlenk 

Prepared under MMS Cooperative 

Agreement No. 14-35-0001-31063 

by 

Coastal Marine Institute 

Marine Science Institute 

University of California 

Santa Barbara, CA 93106-6150 

U.S. Department of the Interior 

Minerals Management Service Camarillo

Pacific OCS Region February 2006
Disclaimer 

This report has been reviewed by the Pacific Outer Continental Shelf Region, Minerals 

Management Service, U.S. Department of the Interior and approved for publication. The opinions, 

findings, conclusions, or recommendations in this report are those of the authors, and do not 

necessarily reflect the views and policies of the Minerals Management Service. Mention of trade 

names or commercial products does not constitute an endorsement or recommendation for use. This 

report has not been edited for conformity with Minerals Management Service editorial standards. 

Availability of Report

Extra copies of the report may be obtained from: 

U.S. Dept. of the Interior 

Minerals Management Service 

Pacific OCS Region 

770 Paseo Camarillo 

Camarillo, CA 93010 

Phone: 805-389-7621 

A PDF file of this report is available at: 

http://www.coastalresearchcenter.ucsb.edu/CMI/ 

Suggested Citation 

The suggested citation for this report is: 

Schlenk, D. Use of Biological Endpoints in Flatfish to Establish Sediment Quality Criteria for 

Polyaromatic Hydrocarbon Residues and Asses Remediation Strategies. MMS OCS Study 2006-

008. Coastal Research Center, Marine Science Institute, University of California, Santa Barbara, 

California. MMS Cooperative Agreement Number 14-35-0001-31063. 39 pages. 
Table of Contents 

FINAL TECHNICAL SUMMARY........................................................................ 1 

FINAL STUDY REPORT....................................................................................... 5 

 1. Chapter 1 ........................................................................................................ 5 

 1.1 Introduction................................................................................................... 5 

 1.2. Materials and Methods.................................................................................. 6 

 1.2.1. Recirculating exposure system ................................................ 6 

 1.2.2. Sediment chemical and water analysis .................................... 8 

 1.2.3. Biochemical endpoint measurements ...................................... 8 

 1.2.4. Statistical analysis.................................................................... 9 

 1.3. Results........................................................................................................... 9 

 1.3.1. Sediment PAHs........................................................................ 9 

 1.3.2. Biochemical markers of exposure............................................ 11 

 1.4. Discussion..................................................................................................... 14 

 1.5. Conclusion .................................................................................................... 17 

 

 2. Chapter 2 ........................................................................................................ 18 

 2.1. Introduction................................................................................................... 18 

 2.2 Materials and Methods.................................................................................. 19 

 2.2.1. Exposure system ...................................................................... 19 

 2.2.2. Chemical evaluation of PAHs in sediments............................. 20 

 2.2.3. Biochemical endpoint measurements ...................................... 20 

 2.3 Results..................................................................................................... 21 

 2.4 Discussion..................................................................................................... 23 

 2.5 Conclusion .................................................................................................... 25 

Acknowledgements ............................................................................................ 26 

References........................................................................................................... 27 

 

 
List of Tables

Table 1. Concentrations of high molecular weight PAHs......................................... 9 

Table 2. Concentrations of low molecular weight PAHs.......................................... 10 

Table 3. E2 Concentrations (pmol/ml) in Hornyhead turbots................................... 14 

List of Figures 

Figure 1. Recirculating system for exposing fish to sediments .............................. 7 

Figure 2. High molecular weight PAHs measured in sediments ............................. 10 

Figure 3. Low molecular weight PAHs measured in sediments.............................. 11 

Figure 4. Equivalents of naphthalene, phenanthrene, and benzo(a)pyrene (μg/ml) 

in bile of hornyhead turbot (Pleuronichthys verticalis)....................................... 12 

Figure 5. Percent DNA damage in comet tail .......................................................... 13 

Figure 6. Tail Moment (TM) values for the blood and liver.................................... 13 

Figure 7. Hepatic cytochrome P450 1A (CYP 1A) expression ............................... 21 

Figure 8. FAC’s in halibut bile ................................................................................ 22 

Figure 9. Plasma steroid concentrations between treatment groups ........................ 23 

 

 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

1

FINAL TECHNICAL SUMMARY 

STUDY TITLE: Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

Criteria for Polyaromatic Hydrocarbon Residues and Assess Remediation Strategies. 

REPORT TITLE: Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

Criteria for Polyaromatic Hydrocarbon Residues and Assess Remediation Strategies. 

CONTRACT NUMBER: 14-35-0001-31063 

SPONSORING OCS REGION: Pacific 

APPLICABLE PLANNING AREA(S): Southern California 

FISCAL YEAR(S) OF PROJECT FUNDING: FY 01, FY 02 

COMPLETION DATE OF REPORT: December 2005 

COSTS: FY 01 - $64,746; FY 02 - $64,746; FY 03 - no cost 

CUMULATIVE PROJECT COST: $129,492 

PROJECT MANAGER: Russell J. Schmitt

AFFILIATION: Coastal Research Center, Marine Science Institute, University of 

California, Santa Barbara, CA 93106-6150 

ADDRESS: University of California, Santa Barbara, CA 93106-6150 

PRINCIPAL INVESTIGATOR: Daniel Schlenk 

ADDRESS: Department of Ecotoxicology and Environmental Toxicology, Science 

Laboratories I 316, University of California, Riverside, CA 92521 

KEY WORDS: Polycyclic aromatic hydrocarbons, Pleuronichthys verticalis, Fluorescent 

aromatic compounds, DNA damage, PAHs, oil seep, CYP1A, FACs, GSI, sex steroids

BACKGROUND: Input of Polyaromatic hydrocarbons (PAHs) occurs through 

anthropogenic and natural mechanisms. Toxicities resulting from chronic exposure include 

immune suppression, reproductive dysfunction and carcinogenesis. As most PAHs tend to be 

lipophilic, there is a high propensity for accumulation in organisms and sediments with high 

organic content. While analytical measurement of specific compounds has been shown to be 

a relevant indicator of exposure within invertebrates, rapid biotransformation prevents 

accurate assessments of exposure in vertebrates such as fish. 
Final Technical Summary – Schlenk 

2

OBJECTIVES: The following study attempted to utilize biochemical and physiological 

indicators in flatfish to estimate a threshold concentration which could be used in risk 

assessment paradigms to evaluate sediments contaminated with PAHs. 

DESCRIPTION: Two species of flatfish were exposed to various dilutions of sediments 

collected from the natural oil seep off the coast of Santa Barbara, CA. In contrast to other 

studies carried out in anthropogenically contaminated areas, the predominant PAHs observed 

in the sediments were of low molecular weight. Hepatic cytochrome P450 1A (CYP1A), 

bilary fluorescent aromatic compounds (FACs), plasma steroid concentrations, gonadal 

somatic indices, and in some cases, hepatic DNA damage was utilized in each species 

exposed for either 7 (Hornyhead turbot) or 30 days (California Halibut) to diluted sediments. 

Attempts were made to generate dose-response curves which could be calibrated against 

reproductive function (GSI, sex steroids) for the estimation of sediment threshold 

concentrations. 

SIGNIFICANT CONCLUSIONS: Of all of the endpoints measured, diminished plasma 

concentrations of 17β-estradiol was the most sensitive endpoint in either species. With the 

exception of hepatic DNA damage in Hornyhead turbot following the 7 day exposure, none of 

the endpoints, in particular, the reproductive endpoints, displayed dose-response relationships 

preventing comparisons between indicators. Comparisons of a potential threshold at the 1% 

COP sediment dilution for E2 diminishment with literature values for anthropogenic PAHs, 

demonstrated significant recalcitrance toward reproductive dysfuntion in California Halibut. 

Reasons for the discrepancy include a differing profile of specific PAHs, with low-molecular 

weight PAHs likely being less toxic than higher molecular weight PAHs. 

STUDY PRODUCTS: 

Abstract – Chapter 1: 

Concentrations of serum/plasma estradiol, biliary fluorescent aromatic compounds (FACs), 

levels of hepatic CYP1A expression, and DNA damage were measured in sexually mature 

hornyhead turbot (Pleuronichthys verticalis) exposed in the laboratory for 7 days to a gradient of 

sediments collected from a natural petroleum seep in the Santa Barbara channel. Coal Oil Point 

(COP) sediments were homogenized and divided into 4 treatments containing 0% (sediment 

from the Orange County Sanitation District’s reference location), 33%, 66%, and 100% (COP) 

sediments. Sediment concentrations of twenty PAHs ranged from below the detection limit for 

the 0% COP sediment treatments to 105 μg/g in the 100% treatments with lower molecular 

weight compounds predominating. Concentrations of biliary FACs were not linear with COP 

treatment but levels of hepatic DNA damage increased linearly with increasing concentrations of 

high molecular weight PAHs. Hepatic CYP1A expression was elevated only in the 100% 

treatments. A reduction of plasma estradiol in male and female fish was observed in all COP 

exposures. These results demonstrate that acute sediment-only exposure of flatfish to naturallyderived PAHs elicits alterations in biochemical endpoints indicative of PAH bioavailability and 

adverse effects with different sensitivities.
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

3

Abstract - Chapter 2: 

Coal Oil Point (COP) is a natural oil seep off the coast of Santa Barbara, California. Although 

most studies examining the fate and effects of petroleum have focused upon urbanized or 

anthropogenic sources of inputs, few have examined the effects of PAHs derived from natural 

seeps. In order to evaluate the effects of polyaromatic hydrocarbons (PAHs) derived from COP 

on marine fish populations, hatchery-reared California Halibut (Platichthys californicus) were 

exposed for 30 days to seven dilutions of sediments collected from COP. Hepatic cytochrome 

P450 1A (CYP1A), biliary fluorescent aromatic compounds (FACs), gonadal somatic indices, 

and plasma steroid concentrations. Sixteen USEPA priority PAHs were targeted fro analysis in 

each sediment dilution. In general, biochemical responses were somewhat recalcitrant to doseresponse relationships and were less sensitive than literature values established for the same 

indicators following exposure to urbanized PAHs. Trends toward reductions in plasma 

17β−estradiol concentrations were observed, but reductions in gonadal somatic indices were not 

observed. FAC values for naphthalene, benzo(a)pyrene, phenanthrene-related compounds 

reached maximums at 33% COP sediment, but declined at higher concentrations. The resulting 

insensitivity may be unique for exposure to “natural” petroleum due to a higher concentration of 

lower molecular weight PAHs or uncharacterized inhibitors. 

Publications: 

Roy, L.A., S. Steinert, S. Bay, D. Greenstein, Y. Sapozhnikova, O. Bawardi, I. Leifer, and D. 

Schlenk. 2003. Biochemical effects of petroleum exposure in hornyhead turbot 

(Pleuronichthys verticalis) exposed to a gradient of sediments collected from a natural 

petroleum seep in CA, U.S.A. Aquatic Toxicology 65:159-169. 

Seruto, C., Y. Sapozhnikova, and D. Schlenk. 2004. Evaluation of the relationships between 

biochemical endpoints of PAH exposure and physiological endpoints of reproduction in 

male California halibut (Paralichthys californicus) exposed to sediments from a natural 

oil seep. Marine Environmental Research (submitted). 
Final Study Report – Schlenk et al. 

4
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

5

FINAL STUDY REPORT 

Chapter 1.

Biochemical effects of petroleum exposure in hornyhead turbot (Pleuronichthys verticalis) 

exposed to a gradient of sediments collected from a natural petroleum seep in CA, USA. 

(Published in Aquatic Toxicology (2003), 65:159-169)

Luke A. Roy, Scott Steinert, Steve M. Bay, Darrin Greenstein, Yelena Sapozhnikova, Ola 

Bawardi, Ira Leifer, Daniel Schlenk. 

1. Introduction 

The area surrounding the coastal region of Santa Barbara, California is well known for its natural 

sources of petroleum (Spies et al., 1980). The natural petroleum seeps in the Santa Barbara 

Channel, such as Coal Oil Point (COP), represent a continuous source of hydrocarbon exposure 

and contamination independent of effects caused by urban areas (Spies et al., 1996). Seepage of 

petroleum at the Isla Vista seep, one of the seeps around Coal Oil Point, has been estimated to 

have been occurring for thousands of years (Simoneit and Kaplan, 1980). An estimated 9500-

19000 liters/day of petroleum are released from the Isla Vista seep (Allen et al., 1970). Seep 

sediment hydrocarbon concentrations have been reported as high as 1000 μg/g (Straughan, 1976; 

Reed et al., 1977; Stuermer et al., 1982). 

Since fish tend to degrade PAHs rapidly, biochemical markers of exposure have been used 

extensively to evaluate biological effects in feral fish exposed to sediment PAHs (Stein et al., 

1992; Myers et al., 1994; Collier et al., 1995; French et al., 1996; Spies et al., 1996). Biliary 

fluorescent aromatic compounds (FACs) have been used extensively to monitor the metabolites 

of PAHs in feral fish (Krahn et al., 1984; Malins et al., 1987; Collier and Varanasi, 1991; Krahn 

et al., 1991; Beyer et al., 1996; Lin et al., 1996; Spies et al., 1996; Aas et al., 2000b) and have 

proven one of the most direct indicators of PAH exposure. DNA modifications including strand 

breaks, DNA base modifications, cross-linkages, and depurination have been observed to be 

associated with exposure to a number of contaminants, including PAHs (Padrangi et al., 1995; 

Steinert, 1996; Steinert et al., 1998). 

In this study, hornyhead turbot (Pleuronichthys verticalis) were collected from a reference 

station and exposed to sediments collected from COP. The primary objective of this study was 

to determine chemical concentrations of selected PAHs in sediment capable of eliciting 

biochemical responses in organisms following sediment-only exposure. A gradient of exposure 

regimes (0%, 33%, 66%, and 100% COP sediments) were utilized to determine exposure using 

biochemical markers. Hepatic CYP1A, serum estradiol, biliary fluorescent aromatic compounds 

and DNA damage in blood and hepatic cells were used to compare organism responses to 

sediment PAH concentrations. 
Final Study Report – Schlenk et al. 

6

2. Materials and Methods 

2.1. Recirculating exposure system 

The experiment was conducted in the Aquatic Toxicology laboratory at the Southern California 

Coastal Water Research Project (SCCWRP) facility in Westminster, California. The exposure 

system (Figure 1) consisted of eight 10-gallon glass aquaria, divided into four separate treatment 

groups. The first treatment group, consisting of two separate tanks, contained 0% COP sediment 

and was comprised of sediment collected from the reference station (118 05.199' Longitude; 33 

36.055' Latitude) used by Orange County Sanitation District (OCSD) (OCSD, 1999). The total 

organic carbon of the sediments were 0.42 % + 0.004. The composition was 73.8 % sand; 21.3% 

silt, and 4.90 % clay with a median grain size (phi) of 3.68 (see OCSD 1999). 

The second, third, and fourth treatments contained 33%, 66%, and 100% COP sediments, 

respectively (119. 53.428 Longitude; 34. 24.370 Latitude). Because of consistency restraints and 

logistics, no measures of TOC, composition or grain size were available for these sediments. 

The tanks containing 33% and 66% COP contained 67% and 34% reference sediments, 

respectively. Sediments were homogenized and divided prior to their introduction into the 

exposure system. Two liters of sediment were placed in each of the 10-gallon aquaria. Two 

sexually mature hornyhead turbot were placed in each tank. 

Fish were collected from Santa Monica Bay (118.5008 Longitude; 33.9039 Latitude).and held 

for 1 month prior to sediment exposure. Average length and weight of turbot were 20.2 cm (± 

2.1 standard deviation) and 116.7 grams (± 30.6 standard deviation), respectively. Gender was 

not differentiated prior to exposure. Water from the recirculating system was filtered through a 

40 L carbon filter (Figure 1). Flow rates were set at 300 ml/min for all tanks and water 

temperature was maintained at 13.8 °C. Light-dark cycles were set at 16 hours of light and 8 

hours of darkness. 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

7

Figure 1. Recirculating system for exposing fish to sediments. 

For quality control purposes, water samples for PAH analyses were collected twice throughout 

the 7-day experiment (day 0 and day 7) to monitor the effectiveness of the filter. Ammonia 

levels were also monitored throughout the experiment and revealed no impacts on the fish. 

After the sediments were introduced into the tanks, the water was permitted to circulate through 

the system for four days before the hornyhead turbot were introduced into the tanks to allow for 

complete settling of the sediments. Prior to and during the experiment hornyhead turbot were 

fed lug worms (Neanthes sp.) obtained from a local bait store in Seal Beach, CA. These animals 

possessed no detectable levels of the 20 PAHs that were analyzed and were fed to the fish twice 

a week. Feeding organisms throughout the course of the experiment has been suggested to 

portray more actual environmental conditions (Aas et al., 2000a). In a similar study, English 

sole (Pleuronectes vetulus) not fed during exposure had increased retention of PAH metabolites 

in the bile, up to seven times higher, than reference organisms fed throughout the experiment 

(Collier and Varanasi, 1991). 

At the end of the experiment, hornyhead turbot were sacrificed by severing the backbone of the 

fish. Length, weight and gender were recorded for each fish. Liver and gall bladder samples 

were removed and stored in dry ice until transport to a -80° C freezer. Samples for DNA damage 

were preserved and transported in liquid nitrogen to the laboratory. 

100

% 

COP 

Tank/Carbon 

filter

100

% 

COP

66% 

COP

66% 

COP

33% 

COP

33% 

COP 

Con 2 

0% 

COP

Con 1 

0% 

COP

Seawater 

Source
Final Study Report – Schlenk et al. 

8

2.2 Sediment chemical and water analysis 

Sediment and water samples were stored in a refrigerator (4°C) until analysis. PAHs were 

analyzed as described in EPA method 8100 (USEPA, 1996). For water samples separatory 

funnel liquid-liquid extraction was employed. Water samples were extracted with 

dichloromethane three times for 2 minutes and then passed through sodium sulfate with 

subsequent evaporation of solvent. Sediment samples were extracted with hexane. An 

ultrasonic disruptor was used and extracts were passed through sodium sulfate, combined and 

solvent evaporated. Cleanup was performed with fully activated silica gel (8 g), preluted with 

hexane, and PAHs were collected from the column with 25 ml of methylene chloride /hexane 

(2:3) (v:v). A GC-FID (flame ionization detector) with a capillary column (DB-5) was used for 

analysis and quantification. The oven temperature was 40°C, ramped to 160°C with 40°/min, 

and up to 300°C with 5°/min. The recoveries were 91-100% with a standard deviation of 4-13% 

and method detection limit (MDL) of 6-41 ng/g for water. In sediment, recoveries were 30-111% 

with a standard deviation of 2-15% and MDLs ranging from 11-53 ng/g sediments. 

2.3 Biochemical endpoint measurements 

FACs: Florescent biliary metabolites of benzo[a]pyrene (BAP), naphthalene (NAP), and 

phenanthrene (PHN) were analyzed in fish bile using fluorescence detection. The assays were 

conducted using a variation of previously reported methods (Krahn et al., 1984; Krahn et al., 

1986). Fluorescence was measured with a Shimadzu fluorescence detector (RF-10 AXL) at 

380/430 nm, 256/380 nm, and 290/335 nm, excitation/emission for BAP, PHN, and NAP, 

respectively. 

DNA Damage: Preservation of blood was achieved by gently mixing and freezing a small 

volume (<100 μl) in 1 ml of ice cold cryopreservation solution, phosphate buffered saline 

(PBS)/10% DMSO. Small sections of liver were placed in 1 ml of ice-cold cryopreservation 

solution. Within 20 minutes, all samples were frozen in liquid nitrogen. Samples were 

transported to the Comet Analysis Laboratory and transferred to a -80°C freezer. Samples were 

evaluated for DNA damage using Comet analysis as previously described (Steinert et al. 1998). 

CYP1A: Hepatic CYP1A levels were quantified using Western blotting techniques followed by 

semi-quantitative measurements of densitometry in optical density units (ODU) as previously 

described (Schlenk et al 1996). The CYP1A protein was assayed utilizing monoclonal mouse 

anti-peptide CYP1A IgG as the primary antibody (Rice et al. 1998). Microsomal protein was 

measured using the Pierce kit (Pierce Inc. Rockford, Il, USA) method with bovine serum 

albumin as the standard. Protein samples were normalized to contain 50 micrograms of protein 

per lane. 

Serum/Plasma Estradiol: Approximately, 0.5 ml of blood was collected from the dorsal aorta 

of the fish prior to euthanasia. Blood was immediately centrifuged at 750 x g for 2 minutes at 

room temperature. Plasma/Serum was removed and stored in liquid nitrogen until processed for 

estradiol measurements. Estradiol was measued using enzyme-linked immunosorbant assay kits 

purchased from Cayman Chemical Co (Ann Arbor, MI, USA) following the manufactures 

guidelines. 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

9

2.4 Statistical analysis 

Analysis of variance (ANOVA) was conducted to determine possible difference between 

treatment groups. Bartlett’s test of homogeneity was conducted to verify homogeneity of 

variance. If homogeneity of variance was not obtained values were log transformed prior to 

ANOVA analysis. When significance (p<0.05) was obtained, Tukey’s Post Hoc test was used 

to single out treatments responsible for the statistical differences. Linear regressions were 

performed to determine relationships between biochemical effects and sediment concentrations. 

Pearson correlations were used to explore relationships between biomarkers. 

3. Results 

3.1. Sediment PAHs 

Sediment PAHs ranged from below detection limits in the control treatments, to a mean of 31.9 

μg/g in the 100% COP treatment groups for high molecular weight PAHs (Table 1). Low 

molecular weight PAHs ranged from 0.22 μg/g in the control groups to a mean of 61.19 μg/g in 

the 100% COP treatment groups (Table 2). 

Table 1. Concentrations of high molecular weight PAHs (μg/g dry weight) in COP sediments diluted with reference 

sediments for fish exposures. 

HMW PAHs control 1 control 2 33% 1 33% 2 66% 1 66% 2 100% 1 100% 2 

fluorene < 0.029 < 0.029 2.9 3.2 4.05 7.2 9.08 12.4 

phenanthrene < 0.042 < 0.042 1.48 2.09 1.69 1.06 4.64 4.23 

anthracene < 0.048 < 0.048 0.34 0.59 3.1 1.4 1.74 2.14 

fluoranthene < 0.038 < 0.038 0.77 1.13 0.8 1.5 2.01 0.78 

pyrene < 0.048 < 0.048 0.25 0.27 0.54 0.4 1.7 1.94 

Benz(a)anthracene < 0.030 < 0.030 0.56 0.36 0.33 0.3 1.54 1.16 

chrysene < 0.032 < 0.032 0.15 ND 0.21 0.14 3.98 4.35 

Benzo(ghi)perylene < 0.025 < 0.025 0.52 0.24 3 2.8 0.9 0.8 

Benzo(b)fluoranthene < 0.028 < 0.028 0.39 0.32 0.26 0.3 0.5 0.9 

Benzo(k)fluoranthene < 0.022 < 0.022 < 0.022 < 0.022 < 0.022 < 0.022 < 0.022 < 0.022

Benzo(e)pyrene < 0.043 < 0.043 0.34 0.36 0.21 0.73 3.72 3.35 

Benzo(a)pyrene < 0.021 < 0.021 0.78 0.58 0.79 2.2 0.28 0.45 

dibenz(ah)anthracene < 0.020 < 0.020 1.88 4.76 2.3 1.4 8.35 5.98 

Benzo(j)fluoranthene < 0.011 < 0.011 0.09 0.07 0.11 0.22 0.14 0.06 

Total <DL <DL 10.46 13.97 17.38 19.64 38.58 38.54 
Final Study Report – Schlenk et al. 

10

Table 2: Concentrations of low molecular weight PAHs (μg/g dry weight) in COP sediments diluted with reference 

sediments for fish exposures. ND = not determined. 

Low Molecular Weight 

PAHs control 1 control 2 33% 1 33% 2 66% 1 66% 2 100% 1 100% 2 

indene 0.1 < 0.053 0.76 0.65 0.28 0.2 0.89 0.81 

naphthalene 0.08 < 0.027 1.25 0.8 0.19 0.2 4.46 4.4 

acenaphthylene 0.1 < 0.038 4.42 6.6 7.93 9.16 26.18 21.47 

acenaphthene 0.06 0.1 11.8 17.88 6.3 8.76 36.61 23.15 

perylene < 0.028 < 0.028 0.05 ND 0.05 0.02 0.15 0.33 

indeno(123cd)pyrene < 0.030 < 0.030 2.23 1.89 2.28 5.24 0.76 3.15 

Total 0.34 0.1 20.52 27.82 17.03 23.58 69.05 53.33 

High molecular weight PAHs increased linearly from the 0% COP treatment through the 100% 

treatment (Figure 2). Low molecular weight PAHs increased from 0.22 μg/g in the 0% COP 

group to the 24.2 μg/g in the 33% COP treatments, then were diminished for the 66% treatments 

(22.43 μg/g ), and were elevated again in the 100% (61.2 μg/g ) treatment groups (Figure 3). 

Analysis of water samples before and after the 7-day exposure indicated the presence of only 

indene at detectable levels. Aqueous indene concentrations in the exposure tanks and returning 

from the carbon filter ranged from 0.35 – 0.62 μg/g . All other PAHs were beneath the detection 

limit. 

0

5

10

15

20

25

30

35

40

45

Control 33% 66% 100%

Treatment

High Molecular Weight PAHs 

(mg/kg)

Figure 2. Total high molecular weight PAHs measured in sediments used in fish exposures. Control represents 0% 

COP.
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

11

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

Control 33% 66% 100%

Treatment

Low Molecular Weight PAHs 

(ppm)

Figure 3. Total low molecular weight PAHs measured in sediments used in fish exposures. 

3.2. Biochemical markers of exposure 

Analysis of biliary FACs revealed increased levels of equivalents of naphthalene (NAP), 

phenanthrene (PHN), and benzo[a]pyrene (BAP) from 0% COP treatment through the 66% COP 

treatment and then a decrease at the 100% treatment groups (Figure 4). The 66% treatment 

group was statistically different from the control group. Equivalents of BAP (p=0.015), PHN 

(0.0068), and NAP (0.011) were log transformed prior to ANOVA analysis. The 66% treatment 

group was approximately 400%, 350%, and 300% higher than the 0% COP treatment group for 

NAP, PHN, and BAP, respectively. Mean high molecular weight sediment PAH concentrations 

of 18.5 μg/g produced a statistically significant increase in biliary FACs for equivalents of 

benzo[a]pyrene and phenanthrene, while mean low molecular weight PAH concentrations of 

20.3 μg/g produced a significant increase in equivalents of naphthalene. 
Final Study Report – Schlenk et al. 

12

0

500

1000

1500

2000

2500

3000

3500

4000

4500

5000

Control 33%COP 67%COP 100% COP

FAC (ug/ml bile)

NAP

PHE

BaP

*

*

*

Figure 4. Equivalents of naphthalene, phenanthrene, and benzo(a)pyrene (μg/ml) in bile of hornyhead turbot 

(Pleuronichthys verticalis) exposed to a gradient of petroleum contaminated sediments for 7 days. Values represent 

mean and standard deviations of measurements. Asterisks (*) denote statistically significant difference from control 

(p<0.05).

Levels of hepatic DNA damage increased from the 0% COP sediments to the 100% COP 

sediments (Figure 5). The mean tail moment (TM) was 26%, 93%, and 187% higher than the 

control for the 33%, 66%, and 100% treatments, respectively (Figure 6). DNA damage in the 

blood revealed no significant relationships with any other metric. 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

13

0

5

10

15

20

25

30

Controls 0.33 0.66 1

Treatment

% DNA in Comet Tail

Blood

Liver

Linear (Liver)

Figure 5. Percent DNA damage in comet tail for the blood and liver of hornyhead turbot (Pleuronichthys verticalis) 

exposed to a gradient of PAH contaminated sediments.

0

1

2

3

4

5

6

7

Controls 33% 66% 100%

Treatment

Mean TM

Blood

Liver

Linear (Liver)

Figure 6. Tail Moment (TM) values for the blood and liver of hornyhead turbot (Pleuronichthys verticalis) exposed 

to a gradient of PAH contaminated sediment. 
Final Study Report – Schlenk et al. 

14

Hepatic CYP1A expression was only observed in fish receiving the 100% COP treatment and 

was not detected in fish from any of the other treatments. Plasma estradiol was observed slightly 

above detection in male and female fish exposed to all sediment exposures having PAHs. 

Because of the low sample number of males and females (n=1-2) no statistical comparisons 

could be carried out between gender. However, when males and females were combined, a 

significant reduction in serum/plasma estradiol was observed following exposure to all of the 

COP dilutions (Table 3). 

Table 3. E2 Concentrations (pmol/ml) in Hornyhead turbots exposed to Coal Oil Point Sediments. Each value 

represents the mean of each value in parentheses + SD. 

 Reference 33% 67% 100% 

Female 1750 + 900 (3) 104 + 12 (2) 102 + 8 (2) 116 + 60 (2) 

Male 17.6 (1) ND (2) ND (2) ND (2) 

Linear regression analyses revealed a significant relationship (r2

: 0.66; p = 0.014) between DNA 

damage (Tail Moment) and high molecular weight PAHs. A significant relationship was also 

observed between DNA damage (Tail Moment) and low molecular weight PAHs (r2: 0.56; p = 

0.032). Significant relationships were observed between sediment BAP concentrations and 

biliary PHN equivalents (r: 0.99, p = 0.0084), sediment BAP and biliary NAP equivalents (r: 

0.99; p = 0.011), as well as sediment PHN and NAP equivalents (r: 0.99; p = 0.0022). 

4. Discussion 

Previous studies have indicated the value of biochemical markers in exploring effects due to 

PAH contamination in marine systems (Collier et al., 1995; Spies et al., 1996). The use of 

multiple biomarkers in a laboratory setting provides insight into the processes and mechanisms 

that govern field responses of organisms exposed to PAH contaminated sediments. This study 

sought to explore biochemical effects and dose-response relationships using hornyhead turbot 

(Pleuronichthys verticalis). Hornyhead turbot are generally associated with soft mud and sand 

bottoms and are found throughout southern California coastal waters (Kramer et al., 1995; Love, 

1996) at depths ranging from 10-90 meters, but have been observed at depths of up to 201 meters 

(Allen, 1982; Love, 1996). Their diet consists almost entirely of worms, although they also feed 

on clam siphons (Allen, 1982; Love, 1996). The demersal ecology of this flatfish species and its 

resilience in captivity make it a sound experimental model for biochemical endpoint studies. 

Only 6 low molecular weight hydrocarbons were measured and compared to 14 high molecular 

weight PAHs. Earlier studies have demonstrated the occurrence of numerous alkylated PAHs as 

well as branched and cyclic alkanes (Spies et al. 1980; Stuermer et al. 1982). Concentrations of 

low molecular weight compounds in COP sediments were higher (61.2 μg/g at the 100% COP 

treatment) than the high molecular compounds (38.6 μg/g in the 100% COP treatment). This is 

consistent with previous studies showing that natural petroleum seeps tend to have higher 

concentrations of lower molecular weight aromatic hydrocarbons (Spies et al., 1980). 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

15

The use of biliary FACs as a biochemical endpoint has proved an effective method of 

determining exposure to xenobiotic compounds (Krahn et al., 1987; Beyer et al., 1996; Aas et al., 

2000a). Our results indicate a clear increase in BAP, PHN, and NAP equivalents in bile with 

increasing sediment PAH concentrations up to the 100% COP treatment group, which possessed 

reduced levels of PAH equivalents when compared to the 66% treatment group. Other studies 

have demonstrated that increasing levels of PAH contaminants in sediments, diet and water 

result in higher levels of biliary FACs in fish exposed to these contaminants. Aas et al. (2000a) 

reported an increase in biliary PAH metabolites over a 30-day period in a laboratory experiment 

with Atlantic cod (Gadus morhua) exposed to crude oil. Hellou and Upshall (1995) exposed 

winter flounder (Pleuronectes americanus) to PAH contaminated sediments and found a strong 

relationship between PAH metabolites in bile and sediment PAH contaminants. Spies et al. 

(1996) observed compounds fluorescing at naphthalene and phenanthrene wavelengths in 

rainbow surfperch (Hypsurus caryi) to be significantly higher from a petroleum seep in close 

proximity to COP with high levels of PAH contaminants compared to a reference location. In 

the same study, a significant difference between compounds fluorescing at phenanthrene 

wavelengths from the petroleum seep when compared to the reference for rubberlip surfperch 

(Rachochilus toxodes) was also observed. An earlier study by Collier and Varanasi (1991) 

observed an excellent dose-response relationship for FAC levels in bile of English sole 

(Pleuronectes vetulus) exposed to an increasing concentrations of sediment benzo[a]pyrene. 

The lower FAC response at 100% COP was intriguing and warrants further studies to determine 

the potential mechanism(s) for this phenonmenon. 

Hepatic CYP1A induction was only observed in animals receiving the 100% COP treatment. 

Compared to the serum estradiol, or FAC responses, the CYP1A response was as the least 

sensitive. These results contrast other studies which have demonstrated coordinate increases in 

CYP1A with FACs in field samples (Collier et al. 1991; Spies et al. 1996). Reasons for the 

insensitivity of CYP1A in hornyhead turbot include the short duration of a sediment-only 

exposure (7 days), the predominance of low molecular weight PAHs (limited Ah-receptor 

binding), and/or unique species-specific peculiarities in CYP1A expression. Little is known 

regarding the CYP1A response in hornyhead turbot, so further calibration studies are necessary 

to determine whether unique species effects are present. 

Although other indicators of PAH uptake were only induced in fish receiving 67 and 100% COP 

treatments, serum/plasma estradiol concentrations were reduced by all treatments in both 

genders. Reductions in circulating steroid concentrations in fish following exposure to PAHs 

have been observed in earlier studies. Female flounder (Platichthys flesus) exposed to dietary 

PAHs chronically for 12 weeks demonstrated significant reductions in plasma steroid levels 

(Monteiro et al. 2000a). Atlantic croaker dietarily exposed to BaP (Thomas, 1988), and English 

sole (Parophrys vetulus) captured in areas contaminated with PAHs and PCBs (Johnson et al. 

1988) also possessed lower levels of serum/plasma estradiol. Many field studies are often 

confounded by the occurrence of other reproductive toxicants in “contaminated sediments”. 

However, in the current study, PAHs are the predominant, if not exclusive, chemicals in the 

sediments that were used for the fish exposures. The mechanism(s) for PAH-mediated 

antiestrogenicity are complex and appear to involve not only reductions in estrogen receptor 

levels (Chaloupka et al. 1992), but also inhibition of steroidigenic enzymes (Monteiro et al., 

2000b). Since FAC or CYP1A values failed to correlate with either group of PAHs, it is possible 
Final Study Report – Schlenk et al. 

16

that other hydrocarbons or PAHs not measured analytically may be involved in the repression of 

estradiol in the exposed fish. As it was not possible to differentiate gender prior to exposure, 

future studies implementing greater numbers of organisms are necessary to obtain adequate 

sample size for each gender to determine threshold responses for steroid reduction and other 

reproductive anomalies due to PAH exposure. 

Johnson et al. (1993; 2002) observed relationships between levels of FACs indicative of PAH 

exposure an inhibition of ovarian development in English sole females from sediment 

contaminated sites in Eagle Harbor and the Duwamish Waterway in Puget Sound, WA. The 

National Marine Fisheries Program of the National Oceanic and Atmospheric Administration has 

recommended a sediment quality criteria of 1 μ/g of total PAH for adverse reproductive and 

hepatic lesions in English sole (Johnson et al. 2002). Although obtained from a non-urban 

source, the concentrations observed in this study significantly exceed these recommendations. 

Further studies on organisms residing in this location are necessary to better understand the 

adaptive mechanisms that allow these animals to survive in this ecosystem. 

The assessment of DNA damage has proven effective in monitoring the effects of xenobiotic 

chemicals in sediment (Padrangi et al., 1995; Steinert et al., 1998; Aas et al., 2000a). Our results 

indicate a clear dose-response relationship in hepatic DNA damage from the 0% COP treatment 

up to the 100% COP treatment, indicating that increased concentrations of sediment PAHs may 

be responsible for an increase in levels of hepatic DNA damage in hornyhead turbot. Levels of 

DNA damage in the control fish were identical to those from fish obtained at the OCSD 

reference site where sediments were collected (Roy et al. 2003). French et al. (1996) also 

reported a linear increase in DNA damage (hepatic DNA adducts) in English sole with increasing 

concentrations of sediment PAHs. Levels of DNA damage in bullheads (Ameriurus nebulosus) 

collected at PAH contaminated sites (Big Creek, Hamilton Harbour, and the Detroit River) were 

elevated when compared to reference locations (Padrangi et al. 1995). Stein et al. (1992) 

reported increased levels of DNA damage (PAH adducts) in English sole and starry flounder 

(Platichthys stellatus) sampled from the PAH contaminated Duwamish Waterway when 

compared to a reference station near Puget Sound, WA. Levels of DNA damage (hepatic DNA 

adducts) were also found to be significantly elevated in Atlantic cod and corkwing wrasse 

(Symphodus melops) at a PAH contaminated site in the Karmsund Straight, Norway (Aas et al., 

2001). As with FAC and CYP1A analyses, failure of DNA damage to inversely correlate with 

estradiol diminishment indicates other compounds or mechanisms may be involved in steroid 

reduction and warrants further study on uncharacterized hydrocarbons in petroleum. 

Coal Oil Point seep sediment PAH concentrations were much higher than concentrations 

typically found in most other areas of southern California coastal waters (Anderson et al., 1999; 

Schiff et al. 2000). It is generally accepted that anthropogenic sources and emissions of PAHs 

into the southern California marine environment have been steadily decreasing since 1970 

(Schiff et al., 2000). As a result, concentrations of PAHs in sediments have also been declining 

(Schiff et al., 2000). Improved wastewater treatment since the 1970’s has allowed a recovery of 

the benthic infaunal community and demersal fish populations surrounding outfall areas, 

including hornyhead turbot and California tonguefish (Symphurus atricauda) that were impacted 

by chemical contamination several decades ago (Schiff et al., 2000). Although the sediment 

PAH concentrations in this study were higher than in other areas of southern California that are 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

17

recovering from historical impacts, this information could prove useful in monitoring areas 

immediately surrounding oil platforms where concentrations of PAHs tend to be higher or to 

verify remediation efforts. 

5. Conclusion 

The most sensitive sublethal response observed in a sediment-only exposure scenario was 

serum/plasma estradiol which was reduced in male and female animals at the lowest 

concentration of COP sediments (approximately 36 μg/g of the twenty PAHs measured) after 7 

days of exposure. Significant increases in biliary FACs and hepatic CYP1A were only observed 

in fish receiving the 67 and 100% COP treatments. Although no significant increases were 

observed, the relationship between hepatic DNA damage and the increasing concentration 

gradient of PAH contaminated sediments indicated reliability of this biochemical endpoint in 

demonstrating PAH uptake and bioavailability exposure in an acute exposure scenario. This 

study also demonstrates that a dietary component for PAH contamination may not be necessary 

to impair reproduction in flatfish species. Although a proposed threshold for hepatic damage 

related to total PAHs in sediment has been suggested, further studies are necessary to determine 

threshold PAH concentrations for “natural” seep sediments that impair plasma/serum steroid 

concentrations and, likely, reproduction in flatfish. 
Final Study Report – Schlenk et al. 

18

Chapter 2.

Evaluation of the Relationships Between Biochemical Endpoints of PAH Exposure and 

Physiological Endpoints of Reproduction in Male California Halibut (Paralichthys californicus) 

Exposed to Sediments from a Natural Oil Seep. (Submitted to Marine Environmental Research) 

Cherlynn Seruto, Yelena Sapozhnikova, and Daniel Schlenk. 

1. Introduction 

Input of PAHs into the marine environment continue to increase (Brown, McCain et al. 1998; 

Matthiessen and Law 2002). Although anthropogenic sources are thought to dominate 

environmental input, natural oil seeps can also contribute a significant amount to local 

environments (Reed et al 1977, Blumer 1972, McKirdy 1986, Wijayarante and Means 1984). 

Seepage from the natural oil seep near Coal Oil Point, off the coast of Goleta/Santa Barbara, 

California has been estimated to have been occurring for thousands of years at 95000 to 19000 

liters/day of petroleum (Simoneit and Kaplan 1980, Allen et al 1970). In contrast to urbanized 

PAHs, low molecular weight hydrocarbons tend to be the primary constituents of the seep 

material (Spies et al. 1980). 

Indicators of PAH bioavailability in several studies around the world in various fish species 

include a variety of biochemical, histological and physiological endpoints (Petersen and 

Kristensen 1998; Arinc and Sen 1999; Kirby, Matthiessen et al. 1999; Miller, Mills et al. 1999; 

Aas, Baussant et al. 2000; Monteiro, Reis-Henriques et al. 2000; Teles, Pacheco et al. 2003). 

These indicators of exposure have been linked to various effects such as carcinogenicity, hepatic 

lesions, precocious and inhibited gonadal development, alterations in egg size and number, 

inhibited spawning, morphological abnormalities, and reduced egg and larval viability (Johnson 

et al 1999, 1997, 1993, 1988, Casillas et al 1991, Collier et al 1992, Incardona 2003). However, 

most of these studies evaluated urbanized multi-ring PAHs resulting from anthropogenic inputs. 

Previous studies in feral fish collected from the COP area demonstrated enhanced hepatic 

cytochrome P4501A (CYP1A) expression and biliary fluorescent aromatic compounds, 

indicative of PAH bioavailbility (Spies et al. 1996; Roy et al. 2003). Seven day exposure studies 

by Roy et al. (2003) examining biochemical responses in naïve Hornyhead Turbot 

(Pleuronichthys verticalis) exposed to COP sediments indicated enhanced CYP1A and FAC 

concentrations, but at sediment PAH concentrations significantly higher than sediment 

thresholds suggested in urbanized areas (Collier et al. 2002). In contrast, serum estradiol 

concentrations were significantly reduced following sediment-only exposures at the lowest PAH 

dilution which was approximately 12 ug/g of 16 PAHs listed as USEPA priority pollutants (Roy 

et al. 2003). 

To determine the potential effects of diminished steroid concentrations on higher level responses 

in flatfish, cultured male California halibut (Paralichthys californicus) were exposed to 7 

concentrations of COP sediments for 30 days. Gonadal somatic indices, plasma steroid 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

19

concentrations, hepatic CYP1A and biliary FACs were measured to determine whether 

relationships existed between biochemical and physiological endpoints. California Halibut are 

benthic flatfish which inhabit estuaries and nearshore coastal areas as larvae and juveniles, then 

increase in spatial distribution offshore with increasing age. California Halibut are also an 

economically important species, being on average the 6th most valuable vertebrate catch off the 

California coast. Revenue for 2002 was $2,800,000 for California Halibut, with an average 

increase of about 10% per year for the past three years (http://www.st.nmfs.gov/; Personal 

communication from the National Marine Fisheries Service, Fisheries Statistics and Economics 

Division, Silver Spring, MD). Exposures were conducted with the aim of determining dose 

response concentrations which may be used in helping to set sediment thresholds for adverse 

effects and help in assessing locations undergoing remediation. 

2. Materials and Methods 

2.1 Exposure System 

PAH contaminated sediment was collected from the natural marine oil seep Coal Oil Point 

(COP) in Goleta (Santa Barbara) (119.53.428 Longitutde; 34.24.370 Latitude), California as 

previously described (Roy et al. 2003). Exposure sediments contained 0, 0.33, 0.66, 1, 33, 66, 

and 100 percent COP sediments diluted to their respective nominal concentrations with reference 

sediment from Orange County, California (118.05.1999 Longitude; 33.36.055 Latitude). Due to 

consistency restraints, total organic carbon (TOC) and grain size could not accurately be 

determined from COP sediments. The TOC of the reference sediments were 0.42 + 0.004 %. 

The composition was 73.8% sand; 21.3% silt and 4.9% clay with a median grain size of 3.68 

(Roy et al. 2003). Exposures were conducted at The SEA Laboratory in Redondo Beach, 

California using filtered ambient seawater (18 + 1 oC) from inflow pipes from the open ocean. 

Aquaria were housed in shaded outdoor areas receiving indirect sunlight in June of 2003 under 

natural light:dark conditions. 

Male California halibut (56.5 + 5.0 cm; 550 + 52 g) were generously donated by Hubbs 

SeaWorld Research Institute (Carlsbad, CA). Fish were acclimated to the facilities at The SEA 

Laboratory for over two months in a circular tank of approximately 12 m3

. Prior to the addition 

of fish, approximately 2 L of sediment were transferred into 40 liter glass tanks under flowthrough conditions and the sediment was allowed to settle. Two fish were placed in each aquaria, 

with two replicates per exposure. Fish were fed pellet food daily and exposures were conducted 

for thirty days. 

Following 30 days, fish were transferred to containers having 10 L of seawater with 10g/L of MS 

222. Following anesthesia, the fish were weighed and the lengths measured. Blood samples 

were collected from the spinal caudal artery with a syringe, and immediately centrifuged at 750 x 

g to separate serum/plasma. Serum/plasma samples were placed in separate tubes and frozen on 

dry ice. Length measurements were taken from the tip of the lower mouth to the end of the 

tailbone where the back fin begins to scale differently than the body scales. Backbones were 

severed with a knife, and liver and bile samples were removed, placed in cryogenic tubes, and 

immediately placed in dry ice for transport to a -80 freezer. 
Final Study Report – Schlenk et al. 

20

2.2 Chemical evaluation of PAHs in sediments 

Chemical analyses of sediment were carried out as reported in Roy et al (2003). Samples were 

stored at 4 °C until analysis. Sixteen USEPA priority PAHs were measured as described in EPA 

method 8100 (USEPA, 1996). Sediment samples were extracted with hexane using an ultrasonic 

disruptor (550 Sonic Dismembrator, Fisher Scientific, Pittsburgh, PA). Cleanup was performed 

with fully activated silica gel (8 g), conditioned with hexane, and PAHs were collected from the 

column with 25 ml of methylene chloride/hexane (2:3, v:v) A GC-FID (flame ionization 

detector) with a capillary column (DB-5) was used for analysis and quantification. The oven 

temperature was 40 °C, ramped to 160 °C with 40 °C/min, and up to 300 °C with 5 °C/min. The 

recoveries were 30–111% with a S.D. of 2–15% and MDLs ranging from 11–53 ng/g sediments. 

Selected PAHs were only a small representative of the total PAHs present. 

2.3 Biochemical Endpoint Measurements 

Blood steroid analysis- 17β−Estradiol (E2) and Testosterone (T) concentrations were analyzed 

spectrophotometrically using immunoassay kits from Cayman Chemical Company (Ann Arbor, 

MI) following the manufacturers instructions. Approximately 50 uL of serum/plasma was used 

for triplicate analyses. Testosterone measurements did not include 11-keto testosterone. 

FAC analysis- Gall bladders were thawed on ice, and bile removed. Samples were diluted in 

99% HPLC- grade methanol and fluorescence measured with a Shimadzu fluorescence detector 

(RF-10 AXL) at 380/430 nm, 256/380 nm, and 290/335 nm excitation/emission for 

benzo(a)pyrene (BAP), phenanthrene (PHN) and naphthalene (NAP) like compounds, 

respectively. Concentrations were calculated using standard curves developed from PHN, NAP, 

and BAP standards. 

CYP1A analysis- Liver microsomal fractions were obtained by ultracentrifugation as previously 

described (Roy et al. 2003). Total protein concentrations for these fractions were attained by 

utilizing the Pierce kit BSA standards. CYP1A protein was measured using 50 ug of protein per 

sample and conducting western blots. Western blot detection was carried out using a primary 

anti-fish CYP1A monoclonal antibody, C10-7 from Biosense Laboratories (Bergen Norway). 

Proteins were initially separated in SDS PAGE, then transferred overnight to nitrocellulose. The 

nitrocellulose was incubated with blocking buffer (0.5 g nonfat dried milk in 50ml phosphate 

buffer solution) in a heat-sealed bag for one hour, then incubated with primary antibody followed 

by secondary anti-mouse alkaline phosphatase, each for one hour. The primary and secondary 

antibodies were each diluted 1000X in blocking buffer. Four fifteen minute rinses with PBS were 

completed in between each incubation. Following colorimetric detection, band optical densities 

were quantified using QuantityOne software from BIORAD (Hercules, CA). Proteins were 

measured using the Pierce reagent (Pierce Inc., Rockford, IL). 

Statisical analysis- Bartletts’s test of homogeneity was conducted to verify homogeneity of 

variance. If data was homogeneous, and treatment groups possessed an N of at least 3, one-way 

analysis of variance (ANOVA) was conducted to determine possible differences between groups. 

A Bonferonni post-hoc test was utilized with p < 0.05 employing GraphPad Prism 3 software 

(San Diego, CA). 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

21

3. Results 

Concentrations of selected PAHs ranged from 0.180-99.8 ug/g (Table 1). In the 0.66 to 66 % 

COP sediments, acenapthene was the dominant PAH with an average concentration of 30 ug/g in 

the 100% COP treatment. Other PAHs with high relative concentrations included 

acenaphthylene as well as fluorine. Other notable concentrations included fluoranthene with an 

average concentration of 1.4 ug/g, followed by dibenzo(ah)anthracene at 7.2 ug/g, and 

phenanthrene at 4.4 ug/g in the 100% COP sediment. 

Hepatic cytochrome P450 1A (CYP1A) expression was similar to the control values to the 33 % 

COP sediments, with significant induction at the 66 and 100% COP sediments (Figure 7). Pvalues were 0.05 for the 66% sediment, and 0.01 for the 100% sediment exposure. 

0

0.5

1

1.5

2

2.5

0 0.33 0.67 1 33 67 100

% COP SEDIMENT

OPTICAL DENSITY CYP/ug protein

a

a

a a

a

b

c

Figure 7. Hepatic cytochrome P450 1A (CYP 1A) expression in COP sediment. 

FAC accumulation in bile was variable in the 33, 66 and 100% PAH sediment exposures, with a 

trend towards increasing FACs in the bile at the higher sediment concentrations (Figure 8). 
Final Study Report – Schlenk et al. 

22

[FAC's] IN HALIBUT BILE

0

1

2

3

4

5

6

0 0.33 0.66 1 33 66 100

% CONTAMINATED PAH SEDIMENT

LOG10 FAC VALUE

(ug/ml for NAP and PHN, ng/ml for 

BAP)

PHN NAP BAP

Figure 8. FAC’s in halibut bile. 

Values for BAP ranged from 5 to 30 ug/ml, PHE from 33 to 2,300ug/ml, and NAP from 500 to 

14,000ug/ml. The lower concentrations of contaminated sediment (0.33%, 0.66%, and 1%) 

showed a slight trend for increasing PAHs with increases in sediment exposure. Increases in 

FACs did not correspond directly to increases in PAHs detected in the sediment particularly in 

the 33 and 66% COP sediments. Metabolites were highest following exposure to the 33% COP 

sediment, while PAHs were highest at the 100% sediment exposures. Much greater amounts of 

PAHs were found in the 100% sediments versus the 33% and 66%, and this did not result in a 

similar increase in FACs. FACs fluorescing at the PHN and NAP wavelengths were highly 

correlated, with an r2

 of 0.97. BAP values were less related to PHN and NAP, with r2 values of 

0.72 and 0.71 respectively. 

No significant differences in plasma steroid concentrations between treatment groups and 

controls were observed (Figure 9). However, a trend toward reduction of estradiol levels was 

observed at the 1% COP treatment. E2 concentrations were between 347 to 849 pg/ml from the 

control to the 0.66% COP sediments, and from the 0.1 to the 100% COP sediments the values 

ranged from 21 to 494 pg/ml. T levels were lower than estradiol at all concentrations (including 

control). GSI values did not show any significant response to treatment. 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

23

0

100

200

300

400

500

600

700

800

900

0 0.3 0.67 1 33 67 100

%COP SEDIMENT

STEROID LEVEL (ng/ml)

0

0.001

0.002

0.003

0.004

0.005

0.006

GSI

ESTRADIOL

TESTOSTERONE

GSI

Figure 9. Plasma steroid concentrations between treatment groups. 

4. Discussion 

Benthic flatfish are included often in sediment contamination studies due to their continual 

exposure to sediments during their life histories. Threshold levels of contaminant exposure are 

needed in marine organisms to conduct accurate environmental risk assessments. California 

Halibut is an important marine fisheries resource economically, and they are also exposed to 

natural PAHs through seeps in the ocean floor along the California Coast as well as in their 

nurseries (estuaries). This study investigated the effects of natural PAH exposure to halibut by 

conducting lab exposures to several concentrations of natural PAH contaminated sediments from 

the oil seep Coal Oil Point in Goleta/Santa Barbara, California. 

Attempts were made to measure 16 priority PAHs in the COP sediments and in treatment 

dilutions of COP sediments. The concentrations of PAHs of the 3 lowest dilutions (33, 67, and 

100% COP) were similar to those found by Roy et al. (2003) who only evaluated 3 dilutions. 

The sediments for that study were collected more than a year prior to the sediments utilized for 

this study from the same location. Thus, there appears to be relatively little temporal variability 

in the concentrations of the 16 PAHs measured in sediments of the Coal Oil Point seep. 

Although fish were exposed to sediments possessing 7 distinct PAH concentrations, no 

biochemical or physiological metric demonstrated a classic dose-response relationship. The 
Final Study Report – Schlenk et al. 

24

LOEC of the 16 compounds for enhanced FAC responses was approximately 36.3 ug/g and the 

NOEC was 1.9 ug/g. Since only targeted PAHs were measured, it is likely the NOEC and LOEC 

values of the total PAHs in the sediments are much greater than our estimates, and significantly 

higher than the 1 ug/g thresholds determined by Collier et al. (2002). Although FACs 

accumulated in the bile of Halibut following exposure to the PAH contaminated sediments, the 

response was not associated with PAH concentrations with increases only observed at the 33% 

COP treatment which were essentially unchanged in animals exposed to the 67 and 100% COP 

treatments. An 18 fold increase in the sediment PAH profiles was noted between the 1 and 33% 

treatments which corresponded to a 7-200-fold increase in various FACs. However, as PAH 

concentrations were approximately 3-fold higher in the 100% COP sediments, FAC 

concentrations remained unchanged. A similar “saturation” effect was noted previous in 

Hornyhead turbot exposed to 33, 67, and 100% COP sediments with saturation occurring at 67 

rather than 33% (Roy et al. 2003). FACs have been used as indicators of PAH exposure in other 

species of demersal flatfish (Krahn et. al.1986, Ariese et. al. 1993, Huggett et al 2003). FACs in 

fish bile accumulate with longer durations of exposure, with saturation that may or may not be 

species or location specific (French et al 1996; Roy et al. 2003). Biliary FACs and CYP1A 

activity have been shown to be casually related with higher level effects such as neoplasia related 

liver lesions (Myers et al 1998; 2003). Contrasting studies showing direct relationships between 

FACs and PAH exposure, Aas et al. (2000) actually found an inverse relationship in flatfish 

between FACs and PAHs. Since FACs are highly dependent upon biotransformation and 

conjugation of PAHs to biliary metabolites, any alteration of biotransformation may have 

significant impacts on FAC formation. Numerous compounds have been shown to be substrates, 

inhibitors and/or inducers Phase I and Phase II enzymes. Fluoranthene has been shown to be a 

relatively strong inhibitor of CYP1A catalyzed oxidation of BaP (Willett et al. 2001). It is 

worthwhile to note that most previous studies have primarily focused their efforts on evaluating 

anthropogenic PAHs, whereas the PAHs in this study are un-urbanized. This difference in itself 

may be responsible for the lower than expected CYP induction. Consequently, the occurrence of 

complex mixtures and uncharacterized PAHs within sediments which may also act as substrates 

or alter metabolic processes within fish could conceivably be responsible for the lack of 

correlation between sediment PAH concentrations, biliary FACs and recalcitrant CYP1A 

induction. 

Whereas FAC levels were relatively high at the 33% COP treatment, induction of CYP1A 

determined by protein quantification with western blots, was significant in the 66% and the 

100% COP treatments. Several studies have shown marked increases in CYP1A induction in 

response to PAH exposure (for review see: Stegeman and Hahn 1994). This can typically be 

measured by optical density analysis of western blots (as completed in our study), or 

measurement of the catalytic activity of 7-ethoxyresorufin O-deethylase (EROD). Western blots 

carried out by Roy et al (2003) in Hornyhead Turbot (Pleuronichthys verticalis) showed similar 

recalcitrance with induction only in the 100% COP sediment group compared to the lower 

sediment exposures, which showed no presence of the protein. A study in Norway analyzing 

flounder (Platichthys flesus) and Atlantic cod (Gadus morhua) from various field locations found 

significant differences from reference and PAH-impacted areas in EROD as well as CYP activity 

(Beyer, Sandvik et al 1996). The insensitivity of each of these responses in two separate studies 

using two separate fish species suggests consistent reduction in the metabolic capabilities due to 

exposure to COP sediments. Since the chronic toxicity of PAHs has been suggested to be 
Use of Biological Endpoints in Flatfish to Establish Sediment Quality 

25

mediated through bioactivation and metabolic conversion to reactive intermediates, the lack of 

response by naïve fish would indicate potentially lower levels of risk. However, other adverse 

effects, particularly targeting the endocrine and reproductive systems, may not be mediated 

through bioactivation. Consequently, gonadal indices and steroid concentrations were evaluated 

following treatment. 

The gonad-somatic indexes (ratio of gonad weight to total body weight) were not associated with 

priority PAH concentrations or dilution treatments. Similarly, plasma steroid levels were 

unaltered by treatment, although a trend toward reduced E2 concentrations may be apparent. 

Previous studies in Hornyhead turbot observed significant reductions in circulating E2 

concentrations following 7 days of treatment to 33% COP sediments (Roy et al. 2003). 

Response differences may be due to species sensitivities as concentrations of sex steroids in 

Hornyhead turbot are significantly different from California halibut as well as English Sole 

(Schlenk et al., submitted). For example, serum E2 concentrations were higher than testosterone 

concentrations in controls as well as treated animals. Whether this observation is an artifact of 

the method utilized to measure steroids (i.e. lack of specificity with regard to the E2 antibody) is 

unclear as there are no published reports of “normal” steroid concentrations in California Halibut 

in different developmental stages such as post-spawing. E2 concentrations in female flounder 

(P. flesus) have been shown to decrease with exposure to phenanthrene (Monteiro, ReisHenriques et al. 2000), and similar steroid inhibition has been linked to inhibition in vivo of 

ovarian steroidogenesis in goldfish (Evanson and Van der Kraak 2001). 

5. Conclusion 

Biochemical and physiological metrics evaluated in the current study failed to demonstrate doseresponse relationships preventing the estimation of accurate sediment thresholds for biomarker 

responses. With the possible exception of depressed E2 concentrations at 1% COP sediments, no 

significant adverse reproductive effects were noted. Significant induction of CYP1A and a trend 

toward enhanced biliary FACs was observed after 30 days of exposure to 67 and 33% COP 

sediments, respectively. In comparison to anthropogenic PAH values seen in other studies, 

halibut had diminished responses in regard to FAC accumulation in bile. CYP1A induction was 

also insensitive with regard to the amounts of PAHs in naturally contaminated sediment from oil 

seeps. This may be due to a species specific low induction response, or more likely due to unique 

mixtures of compounds which repress or inhibit Phase I or Phase II biotransformation pathways. 

Studies are currently underway to explore this possibility. 
Final Study Report – Schlenk et al. 

26

Acknowledgements 

The authors wish to acknowledge the help of Jeff Armstrong of Orange County Sanitation 

District and Ann Martin-Dalkey of the Bureau of Sanitation from the City of Los Angeles for the 

provision of sediments and fish, respectively. This research was supported by the Minerals 

Management Service, U.S. Department of the Interior, under MMS Agreement No. 14-35-0001-

31063. The views and conclusions contained in this document are those of the authors and 

should not be interpreted as necessarily representing the official policies, either express or 

implied, of the U.S. Government. 
27

References 

Aas, E., T. Baussant, L. Balk, B. Liewenborg, and O.D. Andersen. 2000a. PAH metabolites in 

bile, cytochrome P4501A, and DNA adducts as environmental risk parameters for 

chronic oil exposure: a laboratory experiment with Atlantic cod. Aquatic Toxicology

51:241-258. 

Aas, E., J. Beyer, A. Goksøyr. 2000b. Fixed wavelength fluorescence (FF) of bile as a 

monitoring tool for polyaromatic hydrocarbon exposure in fish: an evaluation of 

compound specificity, inner filter effect and signal interpretation. Biomarkers 5:9-23. 

Aas, E., J. Beyer, G. Jonsson, W.L. Reichert, and O.K. Andersen. 2001. Evidence of uptake, 

biotransformation and DNA binding of polyaromatic hydrocarbons in Atlantic cod and 

corkwing wrasse caught in the vicinity of an aluminum works. Marine Environmental 

Research 52:213-229. 

Allen, A.A., R.S. Schlueter, P.G. Mikoloj. 1970. Natural oil seepage at Coal Oil Point, Santa 

Barbara, California. Science 170: 974-977. 

Allen, M.J., 1982. Functional structure of soft-bottom fish communities of the Southern 

California Shelf. Ph.D. Dissertation. University of California, San Diego. CA. pp. 1-

577. 

Anderson, J.W., E.Y. Zeng, and J.M. Jones. 1999. Correlation between response of human cell 

line and distribution of sediment polycyclic aromatic hydrocarbons and polychlorinated 

biphenyls on Palos Verdes Shelf, California, USA. Environmental Toxicology and 

Chemistry 18:1506-1510. 

Arinc, E. and A. Sen. 1999. Hepatic cytochrome P4501A and 7-ethoxyrcsorufin O-deethylase 

induction in mullet and common sole as an indicator of toxic organic pollution in Izmir 

Bay, Turkey. Marine Environmental Research 48(2): 147-160. 

Ariese F., S.J. Kok, M. Verkaik, C. Gooijer, N.H. Velthorst, and J.W. Hofstraat. 1993. 

Synchronous fluorescence spectrometry of fish bile: A rapid screening method for the 

biomonitoring of PAH exposure. Aquatic Toxicology 26:273-286. 

Beyer, J., M. Sandvik, K. Hylland, E. Fjeld, E. Egaas, E. Aas, J.U. Skare, and A. Goksøyr. 1996. 

Contaminant accumulation and biomarker responses in flounder (Platichthys flesus L.) 

and Atlantic cod (Gadus morhua L.) exposed by caging to polluted sediments in 

Søfrjorden, Norway. Aquatic Toxicology 36: 75-98. 

Blumer M., and J. Sass. 1972. Indigenous and petroleum-derived hydrocarbons in a polluted 

sediment. Marine Pollution Bulletin 3(6):92-94. 

Brown, D.W., and B.B. McCain. 1998. Status, correlations and temporal trends of chemical 

contaminants in fish and sediment from selected sites on the Pacific coast of the USA. 

Marine Pollution Bulletin 37(1-2): 67-85. 
28

Casillas E. D.A. Mistiano, L.L. Johnson, L.D. Rhodes, T.K. Collier, J.E. Stein, B.B. McCain, 

and U. Varanasi. 1991. Inducibility of spawning and reproductive success of female 

English sole (Parophrys vetulus) from urban and nonurban areas of Puget Sound, 

Washington. Marine Environmental Research 31:99-122. 

Chaloupka, K., Krishnan, V., Safe, S. 1992. Polynuclear aromatic hydrocarbon carcinogens as 

antiestrogens in MCF-7 human breast cancer cells: Role of the Ah receptor. 

Carcinogenisis 13: 2233-2239. 

Collier, T.K., J.P. Meador, et al. 2002. Fish tissue and sediment effects thresholds for 

polychlorinated biphenyls, polycyclic aromatic hydrocarbons, and tributyltin. Aquatic 

Conservation-Marine and Freshwater Ecosystems 12(5): 489-492. 

Collier, T.K., J.E. Stein, A. Goksøyr, M. Myers, and J.W. Gooch. 1995. Biomarkers of PAH 

exposure in oyster toadfish (Opsanis tau) from the Elizabeth River, Virginia. Marine 

Environmental Research 39: 348-349. 

Collier T.K., J.E. Stein, H.R. Sanborn, T. Hom, M.S. Myers, and U. Varanasi. 1992. Field 

studies of reproductive success in English sole (Parophrys vetulus): Correlations with 

bioindicators of maternal contaminant exposure. Science of the Total Environment 116:

169-185. 

Collier, T.K., and U. Varanasi. 1991. Hepatic activities of xenobiotic metabolizing enzymes and 

biliary levels of xenobitics in English sole (Parophrys vetulus) exposed to environmental 

contaminants. Archives of Environment Contamination and Toxicology 20: 462-473. 

Evanson, M. and G.J. Van der Kraak. 2001. Stimulatory effects of selected PAHs on 

testosterone production in goldfish and rainbow trout and possible mechanisms of action. 

Comparative Biochemistry and Physiology C-Toxicology & Pharmacology 130(2):249-

258. 

French, B.L., W.L. Reichert, T. Hom, M. Nishimoto, H.R. Sanborn, and J.E. Stein. 1996. 

Accumulation and dose-response of hepatic DNA adducts in English sole (Pleuronectes 

vetulus) exposed to a gradient of contaminated sediments. Aquatic Toxicology 36:1-16. 

French B.L., W.L. Reichert, T. Hom, M. Nishimoto, H.R. Sanborn, and J.E. Stein. 1996. 

Associations between metabolites of aromatic compounds in bile and the occurrence of 

hepatic lesions in English sole (Parophrys vetulus) from Puget Sound, Washington. 

Archives of Environment Contamination and Toxicology 15:61-67. 

Hellou, J., and C. Upshall. 1995. Monocyclic aromatic hydrocarbons in bile of flounder exposed 

to a petroleum oil. International Journal of Environment Analytical Chemistry 60:101-

111. 

Incardona J.P., T.K. Collier, and N.L. Scholz. 2004. Defects in cardiac function precede 

morphological abnormalities in fish embryos exposed to polycyclic aromatic 

hydrocarbons. Toxicology and Applied Pharmacology. 196:191-205. 
29

Johnson, L.L., T.K. Collier, and J.E. Stein. 2002. An analysis in support of sediment quality 

thresholds for polycyclic aromatic hydrocarbons (PAHs) to protect estuarine fish. Aquatic 

Conservation: Marine Freshwater Ecosystems 12:517-538. 

Johnson L.L., S.Y. Sol, G.M. Ylitalo, T. Hom, B. French, O.P. Olson, and T.K. Collier. 1999. 

Reproductive injury in English Sole (Pleuronectes vetulus) from the Hylebos Waterway, 

Commencement Bay, Washington. Journal of Aquatic Ecosystem Stress and Recovery. 6: 

289-310. 

Johnson L.L., S.Y. Sol, D.P. Lomax, G.M. Nelson, C.A. Sloan, and E. Casillas. 1997. Fecundity 

and egg weight in English sole (Pleuronectes vetulus) from Puges Sound, WA: influence 

of nutritional status and chemical contaminants. Fishery Bulletin 95:231-249. 

Johnson, L.L., E. Casillas, T.K. Collier, B.B. McCain, and U. Varanasi. 1988. Contaminat 

effects on ovarian development in English sole (Parophrys vetulus) from Puget Sound, 

Washington. Canadian Journal of Fish and Aquatic Toxicology 45:2133-2146. 

Johnson, L.L., E. Casillas, S. Sol, T. Collier, J. Stein, and U. Varanasi. 1993. Contaminant 

effects on reproductive success in selected benthic fish. Marine Environment Research

35:165-170. 

Kirby, M.F., P. Matthiessen, et al. 1999. Hepatic EROD activity in flounder (Platichthys flesus) 

as an indicator of contaminant exposure in English estuaries. Marine Pollution Bulletin 

38(8):676-686. 

Krahn, M.M., D.G. Burrows, W.D. MacLeod Jr., D.C. Malins. 1987. Determination of 

individual metabolites of aromatic compounds in hydrolyzed bile of English sole 

(Parophrys vetulus) from polluted sites in Puget Sound, Washington. Archives of 

Environment Contamination and Toxicology 16:511-522. 

Krahn, M.M., M.S. Myers, D.G. Burrows, and D.C. Malins. 1984. Determination of metabolites 

of xenobiotics in the bile of fish from polluted waterways. Xenobiotica 14:633-646. 

Krahn, M.M., L.D. Rhodes, M.S. Myers, L.K. Moore, W.D. MacLeod, and D.C. Malins. 1986. 

Associations between metabolites of aromatic compounds in bile and the occurrence of 

hepatic lesions in English sole (Parophrys vetulus) from Puget Sound, Washington. 

Archives of Environment Contamination and Toxicology 15:61-67. 

Krahn, M.K., G.M. Ylitalo, J. Joss, and C. Sin-Lam. 1991. Rapid, semi-quantitative screening of 

sediments for aromatic compounds using sonic extraction and HPLC/Fluorescence 

analysis. Marine Environment Research 31:175-196. 

Kramer, D.E., W.H. Barss, B.C. Paust, and B.E. Bracken. 1995. Guide to Northeast Pacific 

Flatfishes. Alaska Sea Grant Program. University of Alaska, Fairbanks. Pp. 1-104. 

Law, R. J. 2000. The analysis of polycyclic aromatic hydrocarbons in marine samples. 

International Journal of Environment and Pollution 13(1-6): 262-283 

Lin, E.LC, S.M. Cormier, and J.A. Torsella. 1996. Fish biliary polycyclic aromatic hydrocarbon 

metabolites estimated by fixed-wavelength fluorescence: comparison with HPLCflourescent detection. Ecotoxicology and Environment Safety 35:16-23. 
30

Love, M. 1996. Probably more than you wanted to known about the fishes of the Pacific coast. 

Really Big Press. Santa Barbara, CA. pp. 1-381. 

Miller, H. C., G.N. Mills, et al. 1999. Induction of cytochrome P4501A (CYP1A) in 

Trematomus bernacchii as an indicator of environmental pollution in Antarctica: 

assessment by quantitative RT-PCR. Aquatic Toxicology 44(3):183-193. 

Malins, D.C., B.B. McCain, D.W. Brown, M.S. Myers, M.M. Krahn, and C. Sin-Lam. 1987. 

Toxic chemicals, including aromatic and chlorinated hydrocarbons and their derivatis, 

and liver lesions in white croaker (Genyonemus lineatus) from the vicinity of Los 

Angeles. Environmental Science and Technology 21: 765-770. 

Monteiro, P.R.R., M.A. Reis-Henriques, and J. Coimbra. 2000a. Plasma steroid levels in female 

flounder (Platichthys flesus) after chronic dietary exposure to single polycyclic aromatic 

hydrocarbons. Marine Environment Research 49:453-467. 

Monteiro, P.R.R., M.A. Reis-Henriques, and J. Coimbra. 2000b. Polycyclic aromatic 

hydrocarbons inhibit in vitro ovarian steroidogenesis in the flounder (Platichthys flesus). 

Aquatic Toxicolology 48:549-559. 

Myers, M.S., C.M. Stehr, O.P. Olson, L.L. Johnson, B.B. McCain, S.L. Chan, and U. Varanasi. 

1994. Relationships between toxicopathic hepatic lesions and exposure to chemical 

contaminants in English Sole (Pleuronectes vetulus), starry flounder (Platiichthys 

stellatus), and white croaker (Genyonemus lineatus) from selected marine sites on the 

Pacific Coast, USA. Environmental Health Perspectives 102:2-17. 

OCSD (Orange County Sanitation District). 1999. Annual Report, July 1998-June 1999. Marine 

Monitoring, Fountain Valley, CA. 

Padrangi, R., M. Petras, S. Ralph, and M. Vrzoc. 1995. Alkaline single cell gel (Comet) assay 

and genotoxicity monitoring using bullheads and carp. Environmental Molecular 

Mutagenesis. 26:345-256. 

Phillips, C.R., M.I. Venkatesan, and R. Bowen. 1997. Interpretation of contaminant sources to 

San Pedro Shelf sediments using molecular markers and principal component analysis. 

Pages 242-260, In: Molecular Markers in Environmental Geochemistry. RP Eganhouse 

(ed.) American Chemical Society Symposium Series No. 671. 

Petersen, G. I. and P. Kristensen. 1998. Bioaccumulation of lipophilic substances in fish early 

life stages. Environmental Toxicology and Chemistry 17(7):1385-1395. 

Reed, W.E., I.R. Kaplan, M. Sandstrom, and P. Mankiewicz. 1977. Petroleum and 

anthropogenic influence on the composition of sediments from the southern California 

Bight. In: Proceedings 1977 Oil Spill Conference. American Petroleum Institute. 

Washington, D.C. pp. 183-188. 

Rice, C.D., D. Schlenk, J. Ainsworth, and A. Goksøyr. 1998. Cross-reactivity of monoclonal 

antibodies against peptide 277-294 of rainbow trout CYP1A1 with hepatic CYP1A 

among fish. Marine Environment Research 46:87-91. 
31

Roy, L.A., J.L. Armstrong, K. Sakamoto, S. Steinert, E. Perkins, D.P. Lomax, L.L. Johnson, and 

D. Schlenk. 2003. The relationships of biochemical endpoints to histopathology and 

population metrics in feral flatfish species collected near the municipal wastewater outfall 

of Orange County, CA. Environmental Toxicology and Chemistry (in press). 

Schiff, K.C., M.J. Allen, E.Y. Zeng, S.M. Bay. 2000. Southern California. Marine Pollution 

Bulletin 41:76-93. 

Schlenk, D., E.J. Perkins, G. Hamilton, Y.S. Zhang, W. Layher. 1996. Correlation of hepatic 

biomarkers with whole animal and population-community metrics. Canadian Journal of 

Fisheries and Aquatic.Science 53:2299-2309. 

Simoneit, B.R.T., I.R. Kaplan. 1980. Triterpenoids as molecular indicators of paleo seepage in 

recent sediments of the Southern California Bight. Marine Environment Research 3:113-

128. 

Spies, R.B., P.H. Davis, and D.H. Stuerner. 1980. Ecology of a petroleum seep off the 

California coast: Marine Environmental Pollution, edited by R. Geyer. Elsevier, 

Amsterdam, pp.229-263. 

Spies, R.B., J.S. Felton, and L. Dillard. 1982. Hepatic mixed-function oxidases in California 

flatfishes are increased in contaminated environments and by oil and PCB ingestion. 

Marine Biology 70:117-127. 

Spies, R.B., J.J. Stegeman, D.E. Hinton, B. Woodin, R. Smolowitz, M. Okihiro, and D. Shea. 

1996. Biomarkers of hydrocarbon exposure and sublethal effects in embiotocid fishes 

from a natural petroleum seep in the Santa Barbara Channel. Aquatic Toxicology 34:195-

219. 

Stein, J.E., T.K. Collier, W.L. Reichert, E. Casillas, T. Hom, and U. Varnasi. 1992. 

Bioindicators of contaminant exposure and sublethal effects: Studies with benthic fish in 

Puget Sound, Washington. Environmental Toxicology and Chemistry 11:701-714. 

Steinert, S.A. 1996. Contribution of apoptosis to observed DNA damage in mussel cells. 

Marine Environment Research 42:253-259. 

Steinert, S.A., R. Streib-Montee, J.M. Leather, and D.B. Chadwick. 1998. DNA damage in 

mussels at sites in San Diego Bay. Mutation Research 399:65-85. 

Straughn, D. 1976. Sublethal effects of natural chronic exposure to petroleum in the marine 

environment. Publications of the American Petroleum Institute 4280:1-120. 

Stuermer, D.H., R.B. Spies, P.H. Davis, D.J. Ng, C.J. Morris, and S. Neal. 1982. The 

hydrocarbons in Isla Vista marine petroleum seep environment. Marine Chemistry

11:413-426. 

Teles, M., M. Pacheco, et al. 2003. Anguilla anguilla L. liver ethoxyresorufin O-deethylation, 

glutathione S-tranferase, erythrocytic nuclear abnormalities, and endocrine responses to 

naphthalene and beta-naphthoflavone. Ecotoxicology and Environmental Safety 55(1):98-

107. 
32

Thomas, P. 1988. Reproductive endocrine function in female Atlantic croaker exposed to 

pollutants. Marine Environment Research 24:179-183. 

United States Environmental Protection Agency (USEPA). 1996. Test methods for evaluating 

solid waste, physical/chemical methods, SW-846, 3rd edtion, proposed update II. 

Washington, DC, USA. Office of Solid Waste and Emergency Response. 

Varanasi, U., J.E. Stein, and M. Nishimoto. 1989. Biotransformation and disposition of PAH in 

fish. In: Metabolism of polycyclic aromatic hydrocarbons in the aquatic environment. 

ed. U. Varanasi. CRC Press. Boca Raton, FL. pp. 93-150. 

van der Oost, R., J. Beyer, et al. 2003. Fish bioaccumulation and biomarkers in environmental 

risk assessment: a review. Environmental Toxicology and Pharmacology 13(2):57-149. 

Wijayaratne R.D, and J.C. Means. 1984. Sorption of polycyclic aromatic hydrocarbons by 

natural estuarine colloids. Marine Environment Research 11:77-89.
33

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