Sensitivity of fish embryos to weathered crude oil: Part I. Low-level exposure during incubation causes malformations, genetic damage, and mortality in larval pacific herring (Clupea

481

Environmental Toxicology and Chemistry, Vol. 18, No. 3, pp. 481–493, 1999Printed in the USA

0730-7268/99 $9.00 1 .00

SENSITIVITY OF FISH EMBRYOS TO WEATHERED CRUDE OIL: PART I. LOW-LEVELEXPOSURE DURING INCUBATION CAUSES MALFORMATIONS, GENETIC DAMAGE,

AND MORTALITY IN LARVAL PACIFIC HERRING (CLUPEA PALLASI)

MARK G. CARLS,*† STANLEY D. RICE,† and JO ELLEN HOSE‡†National Oceanic and Atmospheric Administration, National Marine Fisheries Service, Alaska Fisheries Science Center,

Auke Bay Laboratory, 11305 Glacier Highway, Juneau, Alaska 99801, USA‡Occidental College, Los Angeles, California 90041, USA

(Received 29 October 1997; Accepted 15 June 1998)

Abstract—Pacific herring eggs were exposed for 16 d to weathered Alaska North Slope crude oil. Exposure to an initial aqueousconcentration of 0.7 parts per billion (ppb) polynuclear aromatic hydrocarbons (PAHs) caused malformations, genetic damage,mortality, and decreased size and inhibited swimming. Total aqueous PAH concentrations as low as 0.4 ppb caused sublethalresponses such as yolk sac edema and immaturity consistent with premature hatching. Responses to less weathered oil, which hadrelatively lower proportions of high molecular weight PAH, generally paralleled those of more weathered oil, but lowest observedeffective concentrations (LOECs) were higher (9.1 ppb), demonstrating the importance of composition. The LOEC for moreweathered oil (0.4 ppb) was similar to that observed in pink salmon (1.0 ppb), a species with a very different development rate;by inference, other species may be similarly sensitive to weathered oil. Our methods simulated conditions observed in PrinceWilliam Sound (PWS) following the Exxon Valdez oil (EVO) spill. Biological effects were identical to those observed in embryolarvalherring from PWS in 1989 and support the conclusion that EVO caused significant damage to herring in PWS. Previous demonstrationby our laboratory that most malformed or precocious larvae die corroborates the decreased larval production measured after thespill.

Keywords—Herring larvae Morphological damage Genetic damage Polynuclear aromatic hydrocarbons ExxonValdez oil spill

INTRODUCTION

On March 24, 1989, the T/V Exxon Valdez grounded inPrince William Sound (PWS), Alaska; the resultant oil spillwas the largest in U.S. history. In the short-term aftermath ofthe spill, damage attributed to oil was measured in severalspecies of fish, including herring, salmonids, and rockfish [1–4]. Herring were just beginning to spawn when the spill oc-curred, and about half of the spawned biomass was exposedto Exxon Valdez oil (EVO) [1]. Exposure of herring eggs topetroleum hydrocarbons frequently results in small, abnormallarvae with poor survival potential [5–7]. Herring larvae col-lected throughout PWS (May–July 1989) exhibited conditionsassociated with oil exposure, including morphologic malfor-mations, genetic damage, and small size [8]. It is likely thatsuch severely affected larvae die because site-specific instan-taneous mortality rates were highest near oiled beaches [1].One Natural Resource Damage Assessment (NRDA) study cal-culated a 52% loss in larval herring productivity for PWS in1989 as a result of the spill [1].

However, the existence of toxic effects in herring eggs andlarvae due to oil exposure remains contentious because a studyfunded by Exxon did not find oil-related impacts in PWS [9].Differences in methodology may primarily explain the diver-gent conclusions. This investigation was undertaken to mimic

* To whom correspondence may be addressed([email protected]).

The research described in this paper was supported by the ExxonValdez Oil Spill Trustee Council. However, the findings and conclu-sions presented by the authors are their own and do not necessarilyreflect the view or position of the Trustee Council.

post-EVO spill conditions and determine the effects of oil onherring development in a laboratory setting free from the con-founding variables inevitable in field studies. We studied theresponses of herring eggs exposed to weathered oil similar incomposition and concentration to that observed in PWS andcompared the effects on larval survival, morphology, and de-velopment with the effects reported in the NRDA studies. Spe-cific experimental objectives were to determine if exposure ofincubating eggs to polynuclear aromatic hydrocarbons (PAHs)leached from spilled oil would cause mortality, morphologicaleffects, behavioral effects, and changes in hatch timing; ifexposure of incubating herring eggs to PAH would cause ge-netic defects in somatic cells, effective PAH concentrationsand exposure durations; and if oil weathering influenced tox-icity. Experimental results support the conclusion that EVOcaused significant damage to herring eggs and larvae in PWS.

MATERIALS AND METHODS

Experimental overview

Total PAH (TPAH) concentrations in water were chosen tobracket maximum mean TPAH concentrations observed inopen water of PWS following the EVO spill (EVOS) (up to6.24 6 0.63 ppb [10]). Water was contaminated by passagethrough oiled gravel; eggs were exposed to PAH dissolvedfrom oil, not oil droplets. To test weathering effects, the gravelused to produce less-weathered oil (LWO) was reused in thesecond experiment, resulting in more-weathered oil (MWO).The two consecutive experiments each consisted of a gradedseries of doses plus multiple exposure times at the middleconcentration (34 ppb for LWO and 0.7 ppb for MWO). Fertile

482 Environ. Toxicol. Chem. 18, 1999 M.G. Carls et al.

Fig. 1. Representative composition of polynuclear aromatic hydro-carbons (PAHs) in treatment water compared to Exxon Valdez crudeoil. In this example, composition is shown for the mid-oil treatmenton days 0 and 16 in less weathered (LWO) and more weathered oil(MWO) experiments. Total PAH (TPAH) concentrations are listed foreach set. Also listed are total percentages of the following homologouschemical groups (left to right): naphthalenes, fluorenes, dibenzothio-phenes, phenanthrenes, and chrysenes; numbers in parentheses in thefirst panel are homolog percentages for the Alaska North Slope crudeoil used in this study.

herring eggs, attached to glass slides, were incubated in cleanor contaminated seawater. We controlled for possible maternalgenetic variation within each of these experiments by includingeggs from each female in each treatment.

Collection of eggs

Prespawn herring were collected from two locations insoutheast Alaska. Herring for the LWO experiment were col-lected near Cat Island (55.08N latitude, 131.28W longitude) onApril 11, 1995. Fish for the MWO experiment were collectedin Seymour Canal (57.58N latitude, 133.88W longitude) onMay 13, 1995. Herring were spawned as rapidly as possibleafter capture in Ketchikan or Juneau. Ovarian membranes werecut longitudinally, and eggs were removed with a stainlesssteel spatula. From each female, eggs were deposited in asingle layer on nine 25 3 75 mm glass slides placed in ambientseawater at the bottom of a shallow glass dish. Deposition wasaccomplished with a gentle swirling motion; approximately100 to 150 eggs were deposited per slide. Slides containingeggs from each female were placed in a staining rack andsuspended in separate beakers of seawater. Within 30 min ofdeposition, a few milliliters of milt, pooled from three males,was added to the eggs in each beaker. After 5 min, slides witheggs were removed and gently rinsed in seawater. Eggs from15 (LWO) or 21 (MWO) females were fertilized in three sep-arate spawning events; a total of 9 males contributed spermin each experiment. Fertilized eggs (LWO) were transportedin ambient seawater from Ketchikan to Juneau. All fertilizedeggs were incubated 1 d in clean seawater before experimentaltreatment.

Eggs were examined for fertilization success within 5 d ofspawning. Excess eggs along slide margins and those in areasmore than one layer deep were removed. Processing was ac-complished in clean seawater with a minimum of emersion.Remaining fertile, infertile, and dead eggs were counted. Forall treatments combined, fertility was 79.0% 6 0.8 in LWOexperiments and 93.1% 6 0.4 in MWO experiments.

Oil exposure conditions and chemical analysis

Seawater was contaminated by contact with oiled gravel.Water percolated upward through gravel in vertical 30-cm di-ameter 3 122-cm polyvinyl chloride cylinders fitted with atrap to prevent slick overflow [11]. Water flowed at 5 L/minfrom these oil generators via glass manifold to the bottom offour 40 to 50-L treatment tanks; one cylinder was used perdose. Water flowed through oiled gravel for 5 d before anyeggs were exposed. Each cylinder was loaded with 45 kg ofoiled gravel (or nonoiled control gravel), except the high-oiltreatment generator was loaded with 90 kg gravel. Gravel,approximately 1 to 11 mm in diameter (Mdf 5 5.2 mm, Q1–Q3 5 4.2–6.6 mm), was washed, dried, and then contaminatedby spraying measured quantities of artificially weathered Alas-ka North Slope crude oil on tumbling gravel.

To simulate the composition of oil stranded on PWS beach-es in 1989 after the EVO spill, the oil applied to gravel wasartificially weathered by heating it to 708C overnight until itsinitial mass was reduced by approximately 20% [12]. Theresultant PAH composition was dominated by the naphtha-lenes, followed by phenanthrenes, dibenzothiophenes, fluo-renes, and chrysenes. In general, homologs with 2-methyl sub-stitutions (C2) had the highest concentrations followed by C1,C3, and then the parent compounds (Fig. 1).

Eggs from each female were incubated 16 d in each of five

treatments (control, trace, low, mid-, and high oil). Additionaleggs (also from each female) were exposed in the mid-oiltreatment for 1, 2, 4, and 8 d to generate a complete 0 to 16d time series; controls were reused as zero exposure time theseanalyses. Eggs were first exposed to oil 1 d after fertilization.After 16 d, all remaining eggs were transferred to clean sea-water. Eggs were isolated by female and treatment 1 d later;each slide was individually suspended in a 1-L jar filled withseawater. Temperature was controlled by placing jars in a flow-ing seawater bath. To facilitate oxygen exchange, slides weresuspended from mobile racks. Mean temperature was 5.18C(4.0–6.78C) in the LWO experiment and 6.28C (5.6–7.18C) inthe MWO experiment. Mean salinity was 32 6 0.3 ppt. At thebeginning of treatment, and every fourth day thereafter, 1 to3 replicate 3.8-L water samples per dose were extracted andanalyzed for PAH by gas chromatography-mass spectrometry(GC-MS) [13].

Analysis of PAH samples by GC-MS was performed at theNational Marine Fisheries Service, Auke Bay Laboratory [13].After addition of six internal standards, all samples were ex-

Sensitivity of herring embryos to weathered crude oil Environ. Toxicol. Chem. 18, 1999 483

Fig. 2. Total polynuclear aromatic hydrocarbon (TPAH) concentra-tions in water by treatment as a function of time in two consecutiveexperiments. Treatments were control (C), trace (T), low (L), mid(M), and high oil (H). Most points represent single observations;where present, vertical bars indicate SE.

tracted with dichloromethane. Isolation and purification of cal-ibrated and uncalibrated compounds was completed by silicagel/alumina column chromatography followed by size-exclu-sion high-pressure liquid chromatography (HPLC) and frac-tionation; seawater samples were not fractionated by HPLC.Extracts of PAH were separated and analyzed by GC equippedwith a mass selective detector. Calibrated PAH were identifiedby retention time and two mass fragment ions characteristicof each PAH and quantified using a five point calibration curve.Uncalibrated PAH homologs (which included alkyl-substitutedisomers of naphthalene, fluorene, dibenzothiophene, phenan-threne, and chrysene) were identified by retention time andthe presence of a single characteristic mass fragment ion. Un-calibrated PAHs were quantified by using calibration curvesof their respective parent homologs. Experimentally deter-mined method detection limits (MDLs) depended on sampleweights and generally were 1 ppb in tissue and 1 to 8 pptr(nanograms per liter) in water. Concentrations below MDLwere treated as 0. Tissue concentrations are reported on a wet-weight basis, but wet to dry weight ratios were measured bydehydrating 1 g wet samples for 24 h at 608C and weighingthe remaining mass. The accuracy of the hydrocarbon analyseswas about 615% based on comparison with National Instituteof Standards and Technology values, and precision expressedas coefficient of variation was usually less than about 20%,depending on the PAH. Total PAH concentrations were cal-culated by summing concentrations of individual PAH. Rel-ative PAH concentrations were calculated as the ratio of PAHconcentration to the TPAH concentration.

Gravel used in the LWO experiment was reused in thesubsequent MWO experiment. Water flow through gravel wasrestarted 1 d before MWO egg exposures; thus, the oil wasmore weathered in MWO than in LWO, and TPAH concen-trations were lower (Fig. 2). Weathering parameter, w, wasestimated from oiled gravel using methods of Short and Heintz[14], where w represents the extent of first-order kinetic lossesof PAH from petroleum and summarizes the exposure historyof the sample (w 5 0 in unweathered samples) and increaseswith weathering. Weathering was dose dependent and rangedfrom 0.53 to 0.04 in LWO (low- and high-oil treatments, re-spectively, on day 2), and 1.28 to 0.96 at the end of the MWOexperiment (low- and high-oil treatments, respectively).

Egg and larval measurements

Hatch timing and success, larval viability, and larval ab-normalities were observed daily during peak hatch. Obser-vation frequency was reduced to 2- or 3-d intervals duringperiods of low hatch, and larvae were collected only if five ormore were present per jar. Living larvae were assessed forswimming ability and gross morphological deformities, anes-thetized with tricaine methanesulfonate, and preserved in 5%buffered formalin. Dead larvae were enumerated and discard-ed. Without exposing remaining eggs to air, water was ex-changed in each jar after completion of these observations.After hatch was complete, the remaining eggs and dead em-bryos were inspected and enumerated.

The swimming ability of live larvae was categorized aseffective, ineffective, or incapable. Effective swimmers wereactive, frequented the water column, and avoided capture. In-effective swimmers were generally more lethargic and weremore likely to be found on jar bottoms. Incapable larvae wereunable to swim in a straight line and were often only capableof spasmodic twitching.

Uptake of PAH by eggs was determined on days 4, 8, and16, plus days 1 and 2 in the mid-oil treatment of MWO andthe high-oil treatment of LWO. Eggs pooled from several fe-males were spawned on nylon plankton netting, fertilized withmilt from several males, and attached to mobile racks. (Thesepooled eggs were used only for hydrocarbon samples.) Eggdensity was high (often multiple layers). Nets were suspendedin incubation tanks separate from but identical to those witheggs on slides. Uptake was measured in all treatments inMWO, but only in control and high-oil treatments in LWO.In LWO, eggs remaining after 16 d were transferred to cleanwater, and depuration was followed through day 24. For eachsample, $10 g of eggs were scraped at random from the net-ting, frozen, and processed using methods of Short et al. [13].Visual inspection revealed no evidence of oil-coated eggs inany treatment. To estimate variance, three replicate high-oiltreatment samples were analyzed on day 16 in each experi-ment. Reported concentrations of TPAH in tissue are basedon wet weight.

For morphological measurement (total body length, yolkvolume, and spinal curvature), five preserved larvae were ran-domly subsampled from each female in each treatment group.Lateral views were digitized with a video camera and a framegrabber. To minimize variance, specimens were rotated to aligneyes before image capture and viewed at 6 or 123 for totallength measurement and 503 for yolk measurement. Totallength was measured from snout to tip of notochord, using asmany line segments as necessary to approximate larval pos-ture. Spinal curvature was estimated from the same line seg-ments as the sum of (1808 2 segment angle). Yolks weregenerally elliptical; the major axis (yolk length) was measuredparallel to the notochord, and the minor axis (yolk height) wasperpendicular to the body axis. Yolk volume was estimatedfrom length and height measures [15]. Muscle width, measuredfor staging purposes, was immediately dorsal to the yolk heightmeasurement.

Edema and jaw size were scored from digital images. Yolksac edema was indicated if the anterior margin of the yolkmembrane was bounded by an area of clear fluid. Pericardialedema was indicated when the pericardium was convex ven-trally or had an unusually large clear area. Jaws were classifiedas small if posterior to the anterior margin of the eye or absent.


Genetic and additional morphologic responses in MWOtests were evaluated using blind review. Ten larvae, randomlysubsampled from 15 randomly selected females, were exam-ined at 303 magnification. The severity of skeletal, cranio-facial, and finfold defects was scored using the graduated se-verity index (GSI) method [16], where 0 5 no effect, 1 5slight defect in structure or size, 2 5 moderate defect in struc-ture or size, and 3 5 severe defect in structure or size. Typesof skeletal defects evaluated were kyphosis or lordosis of thenotochord and stunting. Craniofacial abnormalities consistedof reduction, malformation or absence of one or both jaws,ocular and otic capsule defects, and microcephaly. Reductionsin the width of the dorsal and ventral finfold were recorded.Pectoral fin development was categorized as a bud (no fin raysvisible) or containing fin rays. Yolk sac edema was also scoredblind as a confirmatory observation.

During blind review, both pectoral fins were removed,placed onto a glass microscope slide, and stained with aceto-orcein [16]. All mitotic figures in a fin were enumerated at1,0003 magnification, and anaphase-telophase (AT) configu-rations were visually assessed for evidence of chromosome/chromatid breakage (bridges and attached fragments), multi-polar spindles, and aneuploidy (unequal masses of chromatinin daughter cells). Anaphase aberration rate was calculated foreach dose treatment by dividing the total number of aberrantATs by the total number of ATs present. The number of mitosesper fin was recorded for each larva.

Data analysis

The percentage of infertile eggs was based on the initialnumber of eggs. Percentages of eggs that hatched or died werebased on the total hatched larvae plus the number of dead eggsdetermined at the endpoint. Hatched larvae were categorizedas live, moribund, and dead, and percentages were based onthe total number hatched. Swimming ability of live larvae wascategorized as effective, ineffective, or incapable (moribundand dead larvae were not evaluated), and percentages werebased on the number of live larvae. The percentage of larvaewith spinal aberrations was based on the total of live andmoribund larvae. Peak and median hatch were estimated byfemale parent and averaged.

Scored data (e.g., skeletal defects) were translated into in-cidence of defective larvae; larvae were either considered nor-mal (score 0) or abnormal (score .0). Additionally, scoreddata were analyzed directly with the Kruskal-Wallis nonpara-metric test.

To determine if responses in specific treatments were sig-nificant, analysis of variance (ANOVA) techniques were used.The experimental design was a randomized block; where theoverall test was significant, treatments were compared to con-trols with a priori multiple comparisons. Before ANOVA, per-centage data were arc-sine transformed and corrected for smallsample size where necessary [17]. (The same general conclu-sions were reached with untransformed data). In cases wheremultiple larval observations occurred for each female (e.g.,larval lengths), larvae were nested by female to avoid pseu-doreplication.

Median effective concentrations (EC50s) were estimatedfrom logit fits by female (x 5 ln(initial TPAH)) and averaged.Analyses were restricted to cases where response in controlsand highest treatments spanned 50%, and r2 $ 0.45. Resultsof these analyses produced EC50 values comparable to graph-

ical estimates. Comparison of EC50s was accomplished withANOVA.

Biological responses were regressed against time. In LWOtests, percentage data were analyzed with logistic regression.In MWO tests, linear regression was more appropriate.

To describe changes in PAH composition as a function oftime, percentages of homologous families (naphthalenes, fluo-renes, dibenzothiophenes, phenanthrenes, and chrysenes) wereregressed against the total number of days water passedthrough gravel. Models tested were ladder of powers (x-trans-formations from linear through 21/x3) and y 5 aebx. No singlemodel fit all data adequately.

RESULTS

Oil exposure

Aqueous TPAH concentrations declined over time duringthe two consecutive experiments (Fig. 2). Initial aqueousTPAH concentrations in high-, mid-, low-, and trace-LWOtreatments were 85.9, 34.3, 9.11, and 1.72 ppb, compared withconcentrations of 0.04 to 0.05 ppb in control treatments. Aque-ous TPAH concentration in the high-oil treatment declined75% to 21.4 ppb by the end of the 16-d exposure period.Concentrations of the lower LWO treatments declined morethan 90% to final concentrations that ranged from 1.92 to 0.14ppb. Concentration declines continued in the MWO experi-ment. Aqueous TPAH concentration in the high-oil MWOtreatment declined an additional 84% from 7.6 ppb to 0.84ppb during the 16-d exposure period. Initial concentrations inthe lower MWO treatments ranged from 0.72 to 0.14 ppbTPAH and generally declined during the exposure period.Aqueous concentrations reported throughout the remainder ofthe article refer to the initial TPAH concentration.

The relative abundances of aqueous PAH changed through-out the two experiments. The predominant PAH initially pre-sent were the smaller and less substituted PAH (Fig. 1). Rel-ative PAH abundances progressively shifted to larger and moresubstituted PAH during the LWO and MWO exposure periods.Thus, less substituted naphthalene homologs were the mostabundant PAH initially, but more substituted phenanthrene ho-mologs were most abundant by the end of the MWO exposuresas smaller PAH were exhausted from the oiled gravel (Fig. 1).The percentage of naphthalenes consistently declined withtime (20.97 # r # 20.92, p , 0.001), whereas percentagesof other homologs consistently increased, particularly of phen-anthrenes and chrysenes (0.81 # r # 0.99, p # 0.002). SmallerPAH were also exhausted from gravel more rapidly at pro-gressively lower oil treatments. Consequently, aqueous con-centrations of smaller PAH were greater in LWO tests than inMWO tests, but concentrations of larger PAH were similar.Specifically, concentrations of C3-fluorenes, C3-dibenzothio-phenes, C3- and C4-phenanthrenes, and chrysenes in the mid-and high-oil treatments of MWO were roughly the same as incorresponding LWO treatments.

Maximum TPAH concentrations in eggs were higher inLWO treatments than in MWO treatments (Fig. 3), and therelative abundance of PAH accumulated by eggs resembledthat in contaminated water. In LWO, TPAH concentrations ineggs increased throughout the 16-d exposure and peaked at13,700 ppb (high-oil treatment). When contaminated eggswere transferred to clean water (days 16–24), PAH depuratedexponentially from egg tissue. Naphthalene homologs consis-tently accounted for more than 80% of the TPAH concentrationin eggs throughout LWO exposure and depuration. In contrast,


Fig. 3. Total polynuclear aromatic hydrocarbon (TPAH) concentra-tions in eggs as functions of exposure time. Treatments are control(C), low (L), mid- (M), and high oil (H). In less weathered oil (LWO),tissue concentrations were only measured in control and high-oil treat-ments. For clarity, the trace-oil concentration of more weathered oil(MWO) was not displayed. Arrows indicate when eggs were trans-ferred to clean water. Variance was estimated in high-oil treatmentson day 16 (n 5 2); error bars are 6SE.

Fig. 4. Hatch timing and mortality as functions of initial total poly-nuclear aromatic hydrocarbon (TPAH) concentration. Hatch timingwas measured by peak and median hatch day. Observation of larvalmortality was generally completed within 24 h of hatch. Data dis-played are means 6 SE. Solid symbols indicate significant differencesfrom controls.TPAH concentrations in eggs exposed to the mid- and high-

MWO treatments peaked at 226 and 1,020 ppb on day 4 ofthe exposure period and then declined. Naphthalene homologsconsistently accounted for less than 60% of TPAH accumu-lated by eggs exposed to the high-MWO treatment and forless than 35% of TPAH accumulated during exposure to themid-MWO treatment. Arithmetic mean TPAH concentrationin low-MWO eggs (22 ppb) was twice that in control eggsand was significantly greater than in control eggs (p 5 0.010).All differences in PAH composition between tissue and waterphases were less than 17% and were usually less than 10%.

Biological response to more weathered oil

Exposure of eggs to MWO significantly reduced incubationtime (p , 0.001). Peak and median hatch time were signifi-cantly shorter in the highest three treatments (0.4–7.6 ppbTPAH) than in controls (p # 0.037), and median hatch timewas also significant in the 0.1-ppb MWO treatment (Fig. 4).Mean incubation time was reduced 2.6 d in the high-oil treat-ment.

Egg and larval mortality increased significantly as exposureconcentration increased (p , 0.001). Egg mortality was sig-nificantly elevated at 7.6 ppb TPAH (p , 0.001), and larvalmortality was significantly elevated at TPAH concentrations$0.7 ppb (p # 0.011) (Fig. 4). Differences between the controland high-oil treatment response were 12% for dead eggs and2% for moribund plus dead larvae.

Exposure of eggs to low levels of MWO during incubationcaused morphological abnormalities in larvae. Percentages oflarvae with yolk sac edema were significantly elevated atTPAH concentrations $0.4 ppb (p # 0.002) (Fig. 5). Blindassessment of yolk sac edema produced similar results; per-centages of larvae with yolk sac edema were significantly el-evated in the 0.4- and 0.7-ppb TPAH treatments (p # 0.012);the 7.6-ppb treatment was not analyzed. Percentages of larvaewith abnormally small jaws were significantly higher than incontrols following exposure to TPAH concentrations $0.4 ppb(p , 0.042). Percentages of larvae with pericardial edema orspinal defects were significantly elevated TPAH concentrations$0.7 ppb (p , 0.001). Mean differences in response betweenhigh-oil treatment and control were 91% (yolk sac edema),88% (small lower jaws), 43% (pericardial edema), and 58%(spinal defects).

Larval swimming ability was significantly reduced by ex-posure of eggs to MWO during incubation (p , 0.001) (Fig.5). Percentages of effective swimmers were significantly re-duced relative to controls at TPAH concentrations $0.7 ppb(p , 0.001). The difference in response between the controland the 7.6-ppb TPAH treatment was 67%.

Skeletal and craniofacial defects, finfold defects, and failureto develop pectoral fin rays were semiquantitatively assessedin the control, 0.4-, and 0.7-ppb TPAH treatments of MWO


Fig. 5. Incidence of abnormalities and swimming ability as functionsof initial total polynuclear aromatic hydrocarbon (TPAH) concentra-tion. Incidence of yolk sac edema was verified by blind assessmentin the more weathered experiment. Curve fits were estimated withlogistic regression. Data displayed are means 6 SE; r equals estimatedcorrelation coefficients. Solid symbols indicate significant differencesfrom controls.

Fig. 6. Incidences of morphological deformities and genetic aberra-tion as functions of initial total polynuclear aromatic hydrocarbon(TPAH) concentration in a random subsample of larvae from the moreweathered of two consecutive experiments. Curve fits were estimatedwith logistic regression. Data displayed are means 6 SE; r equalsestimated correlation coefficient. Solid symbols indicate significantdifferences from controls.

(Fig. 6). Incidences of defects were significantly elevated at0.7-ppb TPAH in every case (p , 0.001). The most frequentlyobserved skeletal defects were kyphosis or lordosis of thenotochord followed by stunting. Craniofacial abnormalitiesconsisted of the reduction or absence of the lower jaw and,less frequently, the upper jaw. Occasionally smaller brain size(microcephaly) or reduced retinal pigmentation was also pre-sent in severely affected larvae. Most individuals with yolksac edema also had reductions in the width of the dorsal andventral finfolds. Incidences of larvae without fin rays (onlypectoral fin buds present) were significantly elevated after ex-posure to 0.4- and 0.7-ppb TPAH (p # 0.034). Severity ofdefects, estimated nonparametrically from GSI scores, alsoincreased significantly as oil concentration increased.

Exposure of eggs to TPAH concentrations $0.7 ppb inMWO significantly reduced larval length and increased spinalcurvature and yolk sac volume (p # 0.006) (Fig. 7). Comparedto that of controls, larval length in the 7.6-ppb TPAH treatmentdecreased by 2.4 mm, spinal curvature increased by 1008, andyolk sac volume increased by 0.12 mm3.

Genetic response to more weathered oil

Chromosomal damage was evaluated in control, 0.4-, and0.7-ppb treatments of MWO (Fig. 6). The anaphase aberrationrate was significantly elevated to 10% at 0.7 ppb TPAH com-pared to 6% in controls (p 5 0.008). The number of mitosesper pectoral fin averaged 11 and was not significantly affectedby exposure to TPAH concentrations #0.7 ppb.

Comparison of biological response to less and moreweathered oil

Biological response to LWO was the same as that to MWO,but higher initial TPAH concentrations were necessary to elicita significant response. As LWO exposure concentration in-creased, incubation time decreased, and egg and larval mor-tality increased (p , 0.001) (Fig. 4). Morphological abnor-malities increased as LWO concentration increased, and swim-ming ability was reduced (p , 0.001) (Fig. 5). Exposure ofeggs to LWO significantly reduced larval length, and spinalcurvature and yolk sac volume were significantly elevated (p# 0.029) (Fig. 7).

Lowest observed effective concentrations (LOECs) forMWO were generally more than an order of magnitude lowerthan in LWO. Significant differences in biological responsesbetween the two experiments as a function of concentrationis clearly evident in Figures 4, 5, and 7. Responses using LWOwere always significant at concentrations $34 ppb and oc-casionally significant at 9 ppb.

The magnitude of biological response to PAH was generallygreater in LWO tests than in MWO tests (Figs. 4, 5, and 7),except that the incidence of pericardial edema was about thesame or slightly less than after exposure to MWO (Fig. 5).However, pericardial edema induced by LWO tended to bemasked by prominent downward rotation of the head.

Malformed larvae were less able to swim normally, andspinal condition was the most important predictor of swimmingability (partial r2 5 0.90, p , 0.001). Tested with stepwise


Fig. 7. Larval length and curvature, and yolk volume as functions ofinitial total polynuclear aromatic hydrocarbon (TPAH) concentration.Data displayed are means 6 SE. Solid symbols indicate significantdifferences from controls.

Table 1. Median concentrations (EC50) of TPAH in less weathered (LWO) and more weathered (MWO) experimentsa

Response

LWO

LOEC EC50 SE n

MWO

LOEC EC50 SE n

DeathEggLarvae

34.334.3

53.3NC

3.6NC

90

7.610.72

NCNC

NCNC

00

AbnormalitiesSkeletalSpinalCraniofacialFinfoldYolk sac edemaYolk sac edema (blind)Pericardial edema

—34.3——9.1—

34.3

—33.5——

19.6—NC

—3.1——1.6—NC

—13——3

—0

0.720.720.720.720.410.410.72

0.273.600.330.360.770.433.53

0.060.550.050.070.160.080.45

617

75

139

12

Developmental rateSmall jawAbsence of fin rays

34.3—

22.3—

1.4—

2—

0.410.41

1.000.53

0.210.05

134

GeneticAnaphase aberration — — — — 0.72 NC NC 0

BehavioralEffective swimmers 34.3 18.4 1.1 15 0.72 2.44 0.29 21

a SE, standard error; n, number of calculable observations; NC, not calculable; —, not tested. Lowest observed effective concentrations (LOEC)were determined by analysis of variance (with 15 or 21 observations per treatment in LWO and MWO, respectively).

regression, lower jaw size was the second predictor (partial r2

5 0.03, p , 0.001), and yolk sac edema was third (partial r2

, 0.01, p 5 0.092); correlation with pericardial edema wasnot significant. Although data from both LWO and MWO wereincluded in this analysis, logical subgroupings produced com-parable results.

Median effective concentrations

Median effective concentrations were calculable for somebut not all biological responses. The EC50 values estimatedfrom LWO exposures were substantially greater than those inMWO. However, EC50s were based on initial TPAH concen-trations and did not account for other variables, such as dif-ferential weathering and rate of concentration decline (Table1). In LWO, EC50s ranged from 18 to 53 ppb. In MWO, EC50sranged from 0.3 to 3.6 ppb; spinal defects, pericardial edema,and estimated swimming ability produced significantly higherEC50 estimates than did other responses. Biological responsesin MWO were frequently significantly different from controlresponses at concentrations below the estimated EC50 (Table1).

Exposure time effects

As exposure time (to 34 ppb of LWO, and 0.7 ppb of MWO)increased, incubation time decreased, mortality and abnor-malities increased, larval length decreased, and swimmingability declined (Fig. 8). In general, impacts of MWO on eggsand larvae were linearly related to exposure time; pregression ,0.001 for all sublethal responses except yolk volume (p 50.019). Egg and larval death were linearly related to time inLWO (p , 0.001) but not in MWO (p 5 0.628 and 0.265,respectively). However, larval mortality in the mid-oil treat-ment of MWO was significantly elevated after a 16-d exposure(Fig. 4). (In LWO, logit fits appeared to be better descriptorsof response than linear fits for effective swimmers, spinal de-fects, yolk sac edema, and reduced lower jaw responses.)

Embryos reacted more quickly to LWO than MWO. In


Fig. 8. Biological responses to more weathered oil (MWO) as functions of time. Not shown are curvature, pericardial edema, and egg and larvaldeath. The only two responses not significantly related to time in the MWO experiment were egg and larval death. All p , 0.001 for graphedresponses; r is the correlation coefficient.

MWO, exposures as brief as 4 d elicited significant responsesin several cases (spinal defects, reduced jaw size, reducedswimming ability, and yolk sac edema [blind assessment only])(Fig. 8). For most other responses, an 8-d exposure causedsignificant differences, except in skeletal deformity, spinal cur-vature, and yolk volume. The significant 2-d peak hatch re-sponse was inconsistent with other measurements. However,consistently significant 2-d responses were observed in LWO.

DISCUSSION

This is the first report of biologically significant, dose-related sublethal effects and mortality in teleost embryos con-sistently occurring after exposure to aqueous PAHs in the lowparts-per-billion range. Exposure of herring eggs to low con-centrations of MWO (0.7–7.6 ppb) during incubation causedmortality as well as a number of sublethal effects, includingmalformations, reduced swimming ability, and genetic dam-age. Test results were similar with LWO, but LOECs werehigher (9–34 ppb).

Laboratory oil exposure

Embryonic herring consistently responded to a suite ofaqueous PAH in the low parts-per-billion range that resembledthe PAH composition encountered in PWS after the EVOS[14]. Time-to-hatch declined with increasing TPAH concen-tration; exposed larvae were physiologically more immatureand smaller than unexposed larvae. Teratogenic effects in-

cluded dose-related increases in the incidence and severity ofskeletal, craniofacial, fin, cardiovascular, and yolk sac mal-formations. Most of these adverse responses have been re-ported elsewhere following exposure to higher oil concentra-tions [e.g., 6,7,16]. However, it is clear from our study thatenvironmental persistence of the larger, more toxic PAHs canbe injurious at very low concentrations. In the discussion thatfollows, we refer to concentrations observed in the MWOexperiment unless otherwise stated. Observed skeletal, car-diovascular, and yolk sac malformations appear sufficient todecrease survival [6]. The larval mortality threshold appearsto be between mean egg TPAH concentrations of 22 ppb (noobserved effect concentration) and 108 ppb of the MWO(LOEC).

There were also significant sublethal effects (yolk sac ede-ma and both physical and temporal evidence of prematurehatching) after exposure to only 0.4 ppb TPAH (MWO) thatwere not accompanied by increased mortality. Mean TPAHconcentrations in these eggs were 22 ppb, significantly higherthan in controls (11 ppb). The observed reduction in jaw andfin maturity probably reflected precocious hatching [18] andmight be reversible with continued development. The con-current induction of yolk sac edema demonstrates that expo-sure to 0.4-ppb TPAH causes toxicity, not merely a hormeticor arguably beneficial effect from reduced incubation time[19]. Acites (edema) was also observed in pink salmon (On-


corhynchus gorbuscha) larvae similarly exposed to aqueousPAH [20,21].

Of the observed abnormalities, edema appeared to be re-sponsible for most of the larval mortality. Edema was inducedin larvae exposed to initial TPAH concentrations as low as0.4 ppb (MWO) and was apparent after only a 4-d exposureto 0.7 ppb TPAH. Middaugh [22] reported that cardiovascularmalformation (edema or tube hearts) in Menidia beryllina wastheir most sensitive measured response and was significantlyelevated at the lowest dose of water soluble fraction (WSF)of Alaska North Slope crude oil tested, 75 ppb. These con-ditions resulted in reduced cardiac output and cessation ofcirculation. The incidence and severity of edema in our studywere directly related to exposure duration and dose. In mildcases, slight edema was visible only in the yolk sac; moresevere cases were accompanied by increased fluid within thepericardial cavity and the ventricular spaces of the nervoussystem. Edema in the embryolarval stages apparently resultsfrom ionic disturbances and does not require metabolic acti-vation of PAH by the cytochrome P450 system [19].

Two important consequences of the observed sublethal ef-fects would be impairment of swimming and feeding in oil-exposed larvae. Although spinal deformation appeared to bethe dominant factor, observation of live larvae suggested thatedema also adversely affected swimming and would inhibitprey capture. Larvae with enlarged yolk sacs had greater dif-ficulty orienting in the water column than normal larvae andswam more slowly, even with apparent vigorous exertion. Mostindividuals with yolk sac edema also had reductions in thewidth of the dorsal and ventral finfolds, which are the respi-ratory surfaces in pelagic fish larvae [23]. Physical and phys-iological changes accompanying edema (decreased blood flowto tissues, interference with nervous system function, and in-creased energy expenditures), reduced finfold surface area, andretarded pectoral fin development undoubtedly contributed todecreased larval swimming ability. As spinal curvature becamemore pronounced, larvae were less able to swim normally; atthe extreme, larvae were only capable of swimming in a circleor spasmodic twitching—movement that resulted in no di-rected motion. The development of yolk sac edema, spinaldeformities, and swimming problems were correlated tempo-rally and in terms of dose, implying that these abnormalitieshad a causal role in the loss of swimming ability. Oil-exposedeggs hatched early, and embryonic growth was retarded, asindicated by shorter body length, larger yolks, smaller lowerjaws, and immature pectoral fins. Because swimming abilityimproves with maturity, immature larvae are less capable ofprey capture and predator avoidance than mature counterparts[23]. However, older larvae with residual edema are less likelyto feed successfully [24]. Recent studies with dioxin, a com-pound that induces malformations in fish larvae identical tothose of oil exposure, have demonstrated reduced functioningof the digestive system at levels lower than those that induceedema [25] as well as decreased protein synthesis in both fishlarvae and cultured cells [19].

Genetic damage was also induced in oil-exposed larvaefollowing exposure to an initial TPAH concentration of 0.7ppb (MWO). The genotoxicity endpoint (anaphase aberrationrate) measured microscopically visible chromosome/chromatidbreaks and bridges during the later stages of mitosis. Thismethod has been previously used to demonstrate genotoxicityin field studies of the Exxon Valdez [1,16] and Argo Merchantoil spills [26], the New York Bight [27], and sediment from

the San Francisco Bay [28]. In the first two cases, mitoticaberrations were quantitatively or spatially related to oil. Inthe sediment toxicity study and in a previous experiment usingan oil-water dispersion (OWD) of Prudhoe Bay crude oil [7],the anaphase aberration rate was significantly correlated toTPAH concentration.

The consequences of the low level of genetic damage ob-served in this study cannot be predicted with any certainty butmight include reductions in successful cell division andgrowth. A previous study found that TPAH concentrations ofapproximately 130 ppb wet weight (range: 34–844 ppb) wereassociated with increased defects and mortality of early winterflounder embryos [29]. In that study, the only PAH that yieldedstrong statistical correlations with mitotic effects and mortalitywas 1-methylphenanthrene with a mean concentration of 1 ppbwet weight; this PAH is highly toxic in its unmetabolized form[30]. In our study, maximum 1-methylphenanthrene concen-trations in egg tissue exceeded 1 ppb in the upper two treat-ments of MWO and peaked at 32 ppb (wet weight) in the 7.6-ppb treatment of MWO. (In the 86-ppb treatment of LWO, 1-methylphenanthrene concentrations in eggs ranged from 20 to74 ppb.)

Mechanisms of polynuclear aromatic hydrocarbon toxicity

The induction of deleterious sublethal effects and mortalityin herring eggs by exposure to low concentrations of oil ap-pears to conflict with prior studies in our laboratory that in-dicated far greater concentrations of WSF were required tocause toxicity. However, differences in hydrocarbon compo-sition explain toxicity differences. The LOEC for egg mortalitywas between 1,000 and 1,400 ppb WSF of Cook Inlet crudeoil ([31], concentration reanalyzed with the authors’ permis-sion) with a composition of 82% mono- and 18% diaromatichydrocarbons. Larval effects were evident at WSF concentra-tions in the parts-per-million range [31]. In contrast, the oilused in this study had higher proportions of multiring PAHand alkyl-substituted homologs, compounds that are more tox-ic to fish eggs than the mono- and diaromatics [32]. The com-position of PAH accumulated by herring eggs generally mir-rored that in water, but concentration in tissue was roughlytwo orders of magnitude greater than in water.

Comparison of biological responses to MWO and LWOsupports the contention that differential toxicity resulted fromalteration in PAH composition. For example, comparing thetwo oil treatments that had very similar TPAH concentrations(9 ppb in LWO and 7.6 ppb in MWO), the MWO, whichcontained proportionately more high molecular weight PAHs,proved more toxic. It was clear, particularly at higher concen-trations, that the higher molecular weight PAHs were morerefractory and were present in water passed through the MWOat concentrations roughly equal to those in water passedthrough LWO. Increased alkyl-substitution also increasedcompound persistence. Increases in toxicity of PAH with in-creasing molecular size and alkyl-substitution has been doc-umented in other studies and summarized [33].

Although differences in PAH composition appear to bestexplain sensitivity differences between LWO and MWO tests,we recognize that differences in parental stock, collectiontimes, temperature, and other uncontrolled factors could po-tentially confound comparison. Increased temperature, for ex-ample, reduced overall incubation time in the MWO test withrespect to the LWO test (Fig. 4), but within-test dose responseswere consistent, and between-test response patterns were sim-


ilar. The conclusion that MWO is more toxic than LWO re-mains unchanged if we postulate (based on poorer fertilityrate) that the eggs in the LWO experiment were more suscep-tible to toxic effects than eggs in the MWO experiment. Con-versely, there was no evidence to suggest that eggs in the MWOexperiment were more vulnerable to oil than eggs in the LWOexperiment. In support of the conclusion that differences inPAH composition primarily explain differences between LWOand MWO tests, similar differences in sensitivity betweenLWO and very weathered oil were observed in pink salmonembryos originating from the same homogenous stock [21].

Although multiring PAHs may require activation via thecytochrome P450 system to exert toxicity [29,34], alkyl-sub-stituted homologs have received less study and may be directlytoxic [30]. The absolute activity of the cytochrome P450 sys-tem appears to be negligible until soon after hatching, whena burst of dose-dependent inducible activity occurs [34]. Thus,it is likely that the embryonic effects observed here (prematurehatching and defects) instead resulted from cellular oxidativedamage and membrane destabilization [25].

Accumulation of TPAH by herring eggs is consistent withfirst-order kinetics and is determined by the rate of aqueousTPAH concentration decline during exposure. Individual PAHconcentrations decline in oil-contaminated seawater accordingto a first-order kinetic process [14]; thus, the rate of TPAHconcentration decline is also approximately first order [21].Similarly, rates of PAH accumulation and depuration by her-ring eggs follow first-order kinetics. The solution of the dif-ferential equation describing these combined processes indi-cates that the time of maximum TPAH concentration in eggsdepends on the rate of aqueous TPAH concentration decline[21]. Consequently, the more rapid decline of the aqueousTPAH concentrations during the MWO experiment (Fig. 2)explains why maximum TPAH concentration in exposed eggswas earlier in the MWO test than in the LWO test (Fig. 3).This pattern of TPAH accumulation was, therefore, comparableto that observed for similarly exposed pink salmon eggs [21],and both were approximately consistent with the assumed first-order kinetic processes.

Enhanced toxicity of PAHs as a result of photoactivationby UV light has been well documented [e.g., 35], but ourapparatus was arranged to minimize the possibility of acti-vation, and the intensity of ambient UV light probably did notactivate a significant fraction of the PAH molecules. However,photoactivation in PWS after the EVOS was more likely be-cause of direct exposure of PAH to sunlight. Thus, the LOECswe measured (0.4 ppb, MWO) may actually be conservativecompared with conditions existing after the EVOS. Photoac-tivation definitely did not enhance toxicity in a similar pinksalmon study [21] because incubation was accomplished indarkness; yet, LOECs (1 ppb) were similar to those in ourstudy.

In the only other published study demonstrating adverseembryolarval effects at low parts-per-billion oil concentrations[6], similar effects were produced, although the exposure meth-od differed from that reported here. In that study, herring eggswere exposed for 24 h to mechanically dispersed, fresh, Prud-hoe Bay crude oil. Aqueous exposure to concentrations $4.4ppb total hydrocarbons caused larval abnormalities, but tox-icity was attributed to adherence of microdroplets to the eggsand interference with gas exchange. In contrast, eggs in ourexperiment were exposed to PAH dissolved in the water (in-dicated by enrichment of PAH relative to that in parent oil),

and eggs were not physically coated by oil, demonstrating thatchemical toxicity alone may cause larval abnormalities. Sim-ilarly, another researcher concluded that tissue uptake of PAHby pink salmon eggs and alevins from oil-coated gravel wasmediated by dissolution of oil in water; significant biologicaleffects were observed at 4.4 ppb [20]. These effects includedmortality, spinal deformation, ascites, and opercular hypolasia.In a follow-up experiment, salmon eggs and alevins were ex-posed to only aqueous PAH and thus were not in direct contactwith the oil-coated gravel [21]. Results of this aqueous ex-posure were the same as those of simultaneous direct exposuresand confirmed that toxicity was mediated by dissolution ofPAH [21].

The consistency of response in two dissimilar species, Pa-cific herring and pink salmon, suggests that these results maybe generally applicable to other species of fish. Although bothspecies spawn intertidally, herring eggs are much smaller (1.2–1.5 mm diameter) and adhere to surfaces [36]. Pink salmoneggs are 4.0 to 7.9 mm in diameter and are buried an averageof 20 to 30 cm in rocky substrate [37]. Development to hatchin Pacific herring is complete in approximately 27 to 32 d at5.18C, and larvae are planktonic. In contrast, development ofpink salmon is much slower, 171 d at 5.68C (mean tempera-ture), and hatched larvae remain buried in gravel for about142 d; oil exposure continued through emergence in the pinksalmon experiment [21]. Despite large differences in devel-opment time, habitat, and exposure time, LOECs for sublethalresponses to MWO were remarkably similar (0.4–0.7 ppb inherring, 1.0 ppb in pink salmon [21]). Many of the observedabnormalities likely began early in development in both spe-cies. The incidence of abnormalities in herring was consis-tently elevated within 8 exposure days or less, but time tocause abnormalities was not recorded in the pink salmon ex-periment [21]. As was the case for herring, previously reportedsensitivity of pink salmon alevins to WSF (mono- and dia-romatics) was in the low parts-per-million range [e.g., 38].Furthermore, as in herring, PAH from MWO exerted greatereffects than PAH from LWO at similar concentrations. Forexample, 1.3 ppb TPAH from relatively unweathered oil hadno discernable effects on pink salmon embryos, whereas waterwith 1.0 ppb TPAH from weathered oil resulted in reducedsurvival and growth [21]. It is apparent that the compositionof the PAH to which developing eggs is exposed is of primeimportance, and it is likely that other species are similarlysensitive to high molecular weight or alkyl-substituted PAH.We encourage further investigation to explore this hypothesis.

Relationship to 1989 Exxon Valdez oil spill results

The intent of this study was to expose herring eggs to oilof similar composition and at concentrations in the range en-countered in PWS following the EVOS; thus the oil chosenfor study (Alaska North Slope crude oil) and the toxicant de-livery method were designed to mimic conditions observed inPWS. After the EVOS, TPAH concentrations in open seawaterranged up to 6.24 ppb, and concentrations up to 1.59 ppb werepresent 5 weeks after the spill [10]. It is likely that concen-trations were even higher in intertidal spawning areas for theherring where oil was stranded on beach substrate and resus-pended by each tide or wave turbulence. In our MWO exper-iment, TPAH concentrations slightly exceeded values reportedby Short and Harris [10] only in the highest treatment (7.6ppb TPAH). In addition, the concentrations we report werebased on initial TPAH concentrations. Because laboratory con-


centrations declined with time, mean exposure concentrationswere lower; for example, the 7.6-ppb treatment (MWO) hada mean concentration of 3 ppb. First-order loss rate kineticsaccounted for PAH weathering in the laboratory and for thedominant PAH weathering processes in PWS [14]. During nat-ural weathering of EVO in PWS, shifts in composition similarto those found here were repeatedly observed in water andsediment [e.g., 14,39]. The weathering state of MWO oil ongravel substrate in our experiment (w 5 0.04–0.5 for LWOand 1.0–1.3 for MWO) overlapped the weathering range ofbiologically available oil as determined from mussel tissue inareas of PWS where herring spawned in 1989 (0.7 # w #5.7) [40]. Thus, correspondence between the experimentallygenerated MWO and spilled oil underscores the environmentalrelevance of the biological effects we observed at low con-centrations of MWO.

Exposure of herring eggs in PWS probably occurred viatwo routes, from dissolved PAH and from direct contact withoiled substrates. This experiment, as well as recent studiesusing Alaska North Slope crude oil [21,22] and soot extracts[19], demonstrate that embryolarval toxicity can result fromexposure to dissolved PAH. In 1989, dissolved PAH werebioavailable within the PWS oil trajectory to a depth of 25 m[41]. Early life stages are also sensitive to very low concen-trations of dissolved or particulate oil with significant biolog-ical effects occurring at 4.4 ppb in both herring [6] and pinksalmon [20] and at 0.4 ppb TPAH in this study.

Sublethal effects identical to those described here were con-sistently observed in herring larvae of different ages from oiledareas within PWS during spring 1989 [1,8,16]. These includedtemporal and physical evidence of premature hatching, smallsize, malformations including edema, and genetic damage. Inoiled areas of PWS in 1989, herring larvae hatched approxi-mately 3 d earlier, at a more immature stage, than in unoiledareas [1]. Although altered hatching dynamics, stage at hatch,and certain malformations can also be caused by natural stress-ors such as extreme salinity, temperature, or desiccation[42,43], environmental conditions during 1989 were not un-usual and were well within optimal ranges for herring devel-opment. Edema could not be evaluated in larvae hatched inthe laboratory [16], but residual edema was present in field-collected larvae at 1 to 3 weeks posthatch [24]. In newlyhatched larvae, genetic damage was significantly correlated tothe EVO–PAH concentration in adjacent mussels in a dose-dependent fashion [16]. Incidences of genetic damage and oth-er malformations were significantly elevated in oiled areasrelative to unoiled sites and decreased over time until no mea-surable differences remained in 1990 [1,8,24].

We expect that most of the abnormal and precocious herringlarvae observed in PWS in 1989 did not recover, but ratherdied prematurely. In a study where similarly deformed pollocklarvae were observed for a longer period of time, abnormalitiesbecame more pronounced with time, and the majority of ab-normal larvae died [44]. Field observations in PWS also sug-gested premature death of severely affected larvae, becauseincidence and severity of malformations decreased in larvalpopulations sampled over time. Newly hatched larvae morefrequently displayed skeletal bends, absent jaws, and micro-cephaly than did pelagic larvae [16,24]. Jaws were absent onlyin the smallest, least mature larvae, and it is assumed thatthese larvae died after their yolk reserves were utilized. It isalso likely that larvae with severe skeletal bends and cranio-facial defects would succumb to predators. Manifestations of

retarded development (immature pectoral fins and jaws) weregenerally the only types of morphological defects present inolder larvae [8,45]. Larvae with reduced jaws might later re-cover their ability to feed because jaws continue to growthroughout the larval period [46]. However, Marty et al. [24]found that larvae from oiled areas in PWS, which had higherincidences of ascites (the histological equivalent of edema),also grew less and had significantly less food within theirgastrointestinal tract than did larvae from unoiled areas.Growth rates of larval herring from PWS were extremely low(from 0.10 to 0.17 mm/d), bordering on rates found in star-vation studies [8,24,47].

Although the existence of toxic effects in PWS herring eggsand larvae has been disputed [9], TPAH concentrations in eggsmeasured at many locations in 1989 appear sufficient to elicittoxicity, based on the LOECs estimated here (22–108 ppb inegg tissue). Herring eggs collected from some oiled sites inPWS in 1989 and 1990 had TPAH concentrations in excessof 100 ppb [9]. In contrast to Pearson et al. [9], who did notfind any consistent, significant correlations between morpho-logical responses and oil exposure, the Natural Resource Dam-age Assessment studies [1,16] observed morphologic and ge-netic responses identical to ours among herring larvae of dif-fering ages that decreased with distance and time from thespill. Although it is impossible to directly relate EVO–TPAHconcentrations in ambient seawater, mussel tissue, and herringeggs, it is likely that biologically effective oil concentrationswere exceeded at many sites within the oil trajectory.

Following the spill, egg to larval mortality was three to sixtimes higher at oiled sites in PWS than at unoiled sites [47].Although the numbers of eggs deposited were approximatelyequal at oiled and unoiled sites, estimated larval productionwas 99.9% less in oiled areas [1]. Because oiled sites tendedto be more subject to wave turbulence than unoiled sites, how-ever, oil exposure may not have been the only significant factorthat affected mortality rates. Using these differential mortalityrates, an overall 52% loss in larval productivity was estimatedin PWS following the EVOS [1]. Based on the number of eggsdeposited, prediction was that the 1989 year class would bestrong, but it was one of the smallest cohorts observed in PWSin over 20 years [1]. However, high natural variability in re-cruitment precludes a conclusion that the population was af-fected by oil. Conversely, the same variability also precludesthe conclusion that the oil spill did not have an effect on theherring population.

In summary, the findings that very low aqueous TPAHconcentrations (0.4 ppb) are detrimental to herring eggs andlarvae, that low concentrations of EVO in PWS caused larvalmortality, and that another teleost species (pink salmon) issimilarly sensitive [21] suggests that current water quality stan-dards are not adequate to protect sensitive early life stages.The State of Alaska standard, 10 ppb, is the lowest state stan-dard in the United States for total aqueous aromatics [48] andwas established at 1% of the lowest dose known to kill fishand invertebrates in short-term or chronic exposures [49]. Incontrast, the U.S. Environmental Protection Agency acute wa-ter quality criterion is 300 ppb TPAH. Results of this studydemonstrate that PAH composition is as important as totalconcentration and should be considered in regulatory guide-lines. Composition and weathering of PAH in our experimentswas consistent with a major oil spill as shown by Short andHeintz [14]; this suggests general applicability to other spillsituations. Mean aqueous TPAH concentrations from 0.9 to


6.2 ppb were present at heavily oiled beaches within 2 weeksof the EVOS [10]; concentrations well above those capable ofeliciting both acute and sublethal effects in Pacific herring andpink salmon. Our results underscore the necessity to updatestate and national standards for dissolved aromatics to reflectrealistic exposures to critical life stages.

Acknowledgement—We thank J. Campbell, J. Clark, M. Drew, D.Fremgen, C. Hanson, D. Love, X. Luan, N. Magaret, J. Millstein, C.Pohl, J. Regelin, J. Sage, M. Sleeter, and J. Stekoll for technicallaboratory assistance and L. Holland, M. Larsen, and staff for theexcellent quality hydrocarbon analyses. We also thank the AlaskaDepartment of Fish & Game Commercial Fisheries Division for helpin collecting herring and for providing temporary laboratory space.Research was supported by the Exxon Valdez Oil Spill Trustee Coun-cil. However, the findings and conclusions presented by the authorsare their own and do not necessarily reflect the view or position ofthe Trustee Council.

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Documents

Sensitivity of fish embryos to weathered crude oil: Part I. Low-level exposure during incubation causes malformations, genetic damage, and mortality in larval pacific herring (Clupea