www.elsevier.com/locate/marpolbul

Marine Pollution Bulletin 49 (2004) 23–32

Toxicity assessment of sediments associated with variousland-uses in coastal South Carolina, USA,using a meiobenthic copepod bioassay

Adriana C. Bejarano a,*, Keith A. Maruya b, G. Thomas Chandler a

a Department of Environmental Health Sciences, Arnold School of Public Health, University of South Carolina,

Columbia, SC 29208, USAb Skidaway Institute of Oceanography, Savannah, GA 31411, USA

Abstract

Coastal urbanization supplies surrounding estuarine environments with urban-related contaminants such as polycyclic aromatic

hydrocarbons (PAHs), metals and pesticide mixtures. In our study, adult female and male copepods (Amphiascus tenuiremis) were

chronically exposed to 18 sediment samples collected from sites influenced by different land-uses. Sediment samples were collected

from three major suburban areas (Hilton Head, Kiawah Island and the Okatee River watershed) and a pristine site (North Inlet) in

coastal South Carolina. Three-sediment bioassays (six sites per bioassay) were conducted by culturing copepods for 14 days in

quadruplicate test sediments under flow-through conditions at 20 �C and 12:12 LD cycles. Adult survival and copepod reproductive

output were quantified. Sediment samples were also analyzed by GC–MS for low and high molecular weight PAHs. Minimal adult

mortality was observed in most sediment samples. However, sediments from Hilton Head Island and the Okatee River showed

significant effects on copepod reproductive output (i.e., nauplii, copepodites and clutch size). Thus, we determined that reproductive

endpoints rather than adult copepod survivorship were more sensitive to effects of contaminated sediments on A. tenuiremis.

Furthermore, six (33%) of the 18 sites had a >25% reduction in copepod bioassay endpoints relative to controls, suggesting a high

risk to long term A. tenuiremis population maintenance.

� 2004 Elsevier Ltd. All rights reserved.

Keywords: Sediments; Copepod Amphiascus tenuiremis; PAHs; Urbanization; ERMQ; Chronic exposure

1. Introduction

The southeastern US seaboard has experienced tre-

mendous urban expansion and population growth over

the past 40 years. In South Carolina, urban growth,

coupled with development of residential neighborhoods

near major transportation routes and along estuarinemarshes, is 6.2 times faster than population growth

(Berkeley–Charleston–Dorchester Council of Govern-

ments, (BCDCOG), pers. comm.). The proximity of

various land-uses (i.e., shopping centers, parking lots,

bridges, roads and golf course communities) to estuarine

salt marshes and tidal creeks places enormous pressure

on these coastal ecosystems. These areas provide nurs-

*Corresponding author. Tel.: +1-803-777-6452; fax: +1-803-777-

3391.

E-mail address: acbejara@mailbox.sc.edu (A.C. Bejarano).

0025-326X/$ - see front matter � 2004 Elsevier Ltd. All rights reserved.

doi:10.1016/j.marpolbul.2004.01.004

eries for economically valuable fisheries as well as nest-

ing and migratory resting sites for birds, while serving as

filters for human-related contaminants. With urbaniza-

tion comes an increase in impervious surface coverage

resulting in increased storm water runoff, a major source

of non-point source pollution to coastal and estuarine

environments (Fulton et al., 1993). Urban runoff carriesa wide variety of toxicants including polycyclic aromatic

hydrocarbons (PAHs), metals and mixtures of pesticides

that could cause deleterious effects on benthic fauna.

Many studies have shown the usefulness of acute

and chronic toxicity bioassays to evaluate the effects

of estuarine contaminated sediment on a variety of

organisms (McGee et al., 1999; Long, 2000). Toxicity

bioassays using marine/estuarine organisms such as thesea urchin (Paracentrotus lividus), heart urchin (Echino-

cardium cordatum), oyster (Crassostrea gigas), amphi-

pod (Corophium volutator), and bacterium Vibrio

fischeri (Geffard et al., 2001; Stronkhorst et al., 2003)

24 A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32

have been successful in identifying contaminated sed-

iments with potential toxicological effects. This study

was done to assess risks posed by estuarine sediments

influenced by various land-uses to estuarine meioben-thic copepod survival, development and reproduc-

tion.

2. Materials and methods

2.1. Test organisms

The diosaccid harpacticoid copepod Amphiascus

tenuiremis inhabits the top 0.5 cm of muddy inter- and

subtidal estuarine sediments. This cosmopolitan cope-

pod, distributed from the Baltic Sea to the southern

Gulf of Mexico (Lang, 1948), has been mono-specifically

cultured (Chandler, 1986) in contaminant-free flow-

through sediment microcosms at 30 ppt, 23 �C, and a

12 h:12 h Light:Dark cycle. With a rapid generationtime of 21 days at 20 �C (Chandler and Green, 1996),

discrete life-stages, sexually dimorphic adults, and easily

quantifiable reproductive outputs, A. tenuiremis is an

ideal model organism for the assessment of sex specific

survival and reproductive effects (Chandler and Green,

1996).

2.2. Sediment collection and processing

Three 14-day partial life cycle bioassays with A. ten-

uiremis were conducted using salt marsh sediments col-

lected from sites adjacent to several land-uses within

three major suburban areas in coastal South Carolina,

USA. These areas included the relatively new (<10 yr)

suburban developments of the Okatee River watershed,

the somewhat older (<20 yr), less densely populatedKiawah Island resort, and the older (>40 yr) residential

Hilton Head Island area. The collection sites included

retention ponds and creeks within golf course commu-

nities and adjacent to impervious surfaces (i.e., paved

roadways and parking lots). Bioassay 1 was conducted

using sediments from Hilton Head Island (sites HTA,

PD2, PD3; Table 2) and Kiawah Island (sites TCE,

TCW, OR; Table 2), SC, USA. Bioassays 2 and 3 usedsediments from Hilton Head Island (sites HT1, HT2,

HT3, OM T1-P T2-H; Table 2) and the Okatee River

watershed (sites OCB, OCF, T3TG, T3C, HTB, BP;

Table 2). Within each site, 2-cm surface scrapes were

collected from �1 m2 of sediments and composited into

a single homogeneous sample. Composite samples were

placed in clean acetone-rinsed glass jars and stored at 4

�C for a maximum of 30 days until assayed. Within 5days prior to test initiation, ‘‘clean’’ salt marsh control

sediments were collected from a pristine, salt marsh

estuarine preserve little influenced by urbanization in

North Inlet, SC, USA (Sanger et al., 1999).

2.3. 14-day chronic copepod bioassay

Chronic sediment toxicity testing was performed

following procedures described by Chandler and Green(1996). Exposure chambers consisted of 50 ml Teflon

flasks with two 1-cm2 windows on opposite sides cov-

ered with a 63 lm Nitex mesh netting, allowing dripping

water flow through the flask at a rate of 5 ml/h.

Chambers were loaded with 20 ml-filtered (0.45 lm)

aerated artificial seawater (30 ppt, >90% DO satura-

tion), and 10 ml well homogenized and press-sieved

(6 250 lm fraction) sediments (four chambers per site).Chambers were loaded with laboratory cultured

non-gravid female and male copepods (25/each) and

placed in a flow-through system held in an incubator at

20 �C and 12:12 LD cycles. Every third day, a 1:1:1

algae mixture (Dunaliella terticolecta:Isochrysis gal-

bana:Phaeodactylum tricornutum) was added to a con-

centration of �107 cells/chamber (Coulter� Multisizer,

Coulter Electronics Inc., Hialeah, FL, USA). Waterquality (pH, dissolved oxygen, salinity) was monitored

every third day in the flow-through system, and

ammonia (NH3) measured in the exposure chambers at

the end of the bioassay.

After 14 days of culturing, parent copepods and off-

spring were collected by gently rinsing test sediments

through a 63 lm sieve. After checking for live/dead

individuals, samples were stained and fixed with 0.5%Rose-Bengal in 10% buffered formaldehyde, and indi-

viduals counted/staged manually using a microscope.

A total of 10 bioassay endpoints including sex spe-

cific survival, reproductive output (i.e., % gravid

females, clutch size per female, nauplius and copepo-

dites), indices of reproductive success (total offspring

(nauplii + copepodites), and realized production ((nau-

plii + copepodites)/surviving female)) as well as devel-opmental rate (copepodite-to-naupliar ratio) were

obtained for each site.

To assess effects of contaminated sediments on

copepod survival and reproduction, and to allow for

comparisons of among all sites, data from individual

replicates per site were referenced to mean endpoint

values from within bioassay control sediments. Values

are presented as percent change (%) relative to controls.Using this approach, bioassay endpoints were catego-

rized based on the number of test sediments (sites) in

which values of individual bioassay endpoints were re-

duced by >10% or >25% relative to controls. Similarly,

sites were categorized based on the number of bioassay

endpoints in which values were reduced by >10% or

>25% relative to controls. Based on our knowledge of

the natural history of A. tenuiremis, a >25% reduction inadult survival or reproductive output could potentially

influence long-term population growth and mainte-

nance, resulting in rapid population declines (Bejarano

and Chandler, 2003).

A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32 25

2.4. Sediment PAH analysis

Freeze-dried composite sediment samples (10 g) per

site (15 sites total) were homogenized with Hydro-matixTM and extracted with dichloromethane using an

Accelerated Solvent Extraction System (ASE 200, Dio-

nex Corp., Salt Lake City, UT, USA) under elevated

temperature and pressure. After exchanging the solvent

to hexane, each extract was treated with acid-activated

copper to remove elemental sulfur. Extracts were sepa-

rated into fractions of increasing solvent polarity (F1

and F2) using silica gel (2.0 g activated at 130 �C for >4h) packed into a glass column. All fractions were ex-

changed to hexane and reduced to 1.0 ml using a

TurboVap II (Zymark, Hopkinton, MA, USA).

The extracts containing polycyclic aromatic hydro-

carbons (PAH F1 and F2) were analyzed using a HP

6890 digital flow control GC coupled to a 5973 mass

selective detector (GC–MS), operating in the full scan

(50–350 amu at 1 cycles/s) electron ionization mode.Source and transfer line temperatures were 230 and 280

�C, respectively. The injector was programmed to track

the column oven temperature. The GC column oven was

programmed as follows: 60 �C (1 min hold); ramp to 150

�C at 10 �C/min; ramp to 300 �C at 4 �C/min (11 min

hold). Total run time was 60 min. A fused silica column

(30 m (L) · 0.25 mm (i.d.)) coated with 0.25 lm DB-

XLB (Agilent Technologies, Folsom, CA, USA) wasused to separate target PAH analytes.

Twenty-four PAH analytes (2–6 ring structures) were

quantified in all samples, except for the Kiawah Island

sites, based on calibration solutions of an authentic

standard mixture (SRM2260; National Institute of

Standards and Technology (NIST), Gaithersburg, MD,

USA). Target analytes included low molecular weight

(LMW; naphthalene [Naph], 2-methylnaphthalene,1-methylnaphthalene [1mN], biphenyl, 2,6-dimethyl-

naphthalene, acenaphthylene, acenaphthene [Ace-Nap], 2,

3,5-trimethylnaphthalene, fluorine [Fluor], phenanthrene

[Phen], anthracene [Anth] and 1-methylphenanthrene

[1mP]) and high molecular weight (HMW; fluoranthene

[Flu], pyrene [Pyr], benz[a]anthracene [B(a)A], chrysene

[Chry], benzo[b+k]fluoranthene [B(bk)F], benzo[e]py-

rene, benzo[a]pyrene, perylene [Per], indeno[1, 2,3-cd]py-rene, dibenz[a,h]anthracene and benzo[g,h,i]perylene)

PAHs. Total sediment PAH concentrations at the refer-

ence site (North Inlet, NI) have been measured previously

and are consistently low (<116 ng/g dw; Sanger et al.,

1999).

2.5. Sediment effects assessment

Based on the PAH levels at each of the sites, the

general risk of multiple, co-existing PAH exposures was

assessed following suggestions by Long et al., 1995) and

Hyland et al. (1999). For individual PAHs, Long et al.

(1995) published effects guidelines for sediment con-

centrations resulting in adverse effects for 10% (effects

range-low, ERL) and 50% (effects range median, ERM)

of the exposed organisms. For each site PAH-ERL andERM ratios were calculated by dividing PAH concen-

trations by the respective guideline. The ERM quotient

(ERM-Q) was then calculated by dividing the sum of all

ERM ratios per site by the number of analytes.

According to Hyland et al. (1999), the predictive risk of

observing degraded benthic invertebrate communities

is assumed low for ERMQ<0.02, moderate for

ERMQ >0.02)<0.077 and high for ERMQ>0.077.

2.6. Statistical analysis

All endpoints were tested for normality and homo-

geneity of variances using the Shapiro-Wilk’s ‘‘goodness-

of-fit test’’, and the Levene test, respectively. Variables

failing normality were transformed via logarithmic (i.e.,

Log10ðxþ 1Þ, square root ðpðxþ 0:5ÞÞ or angular (i.e.,Arcsine ðpxÞ) transformations where appropriate. Dif-

ferences in adult survival and reproductive output within

bioassays were determined by Analysis of Variance

(ANOVA; PROC GLM, SAS Institute Inc., Cary, NC,

USA). Within bioassays, multiple comparisons between

contaminated sediment samples and control sediments

were done by Dunnett’s procedure. All tests for sig-

nificance were performed using at a ¼ 0:05.The strength of association between bioassay end-

points and low and high molecular weight sediment PAH

(LMW and HMW) concentrations, and ERMQ values

was evaluated using the non-parametric Spearman Rank

correlation analysis (SPEARMAN; PROC FREQ, SAS

Institute Inc., Cary, NC, USA). The p-value of the cor-

relation tests the hypothesis of no association between

variables (correlation¼ 0), with rho ðqÞ as the measure ofcorrelation. This statistical analysis does not demon-

strate cause–effect relationships.

3. Results

3.1. 14-day chronic copepod bioassay

Water quality during the 14-day exposure and at

bioassay termination was consistent across bioassays,

and fell within acceptable limits for chronic sediment

bioassays (ASTM, 1992; Table 1). Dissolved oxygen

saturation (DO%) in the system was >95% during the

14-day exposure for all bioassays, and 70–85% in test

chambers at the end of the exposure.

In sediment samples from sites on Hilton Head andKiawah Island (Bioassay 1), adult survival and gravid

female abundances were not statistically different from

control sediments (ANOVA, p > 0:05; Table 2). How-

ever, reproductive outputs were severely reduced relative

Table 1

Water quality parameters (mean± 1 SD) in bioassay flow-through system during each 14-day chronic bioassay, and in exposure chambers at the end

of each bioassay

Bioassay Flow-through system Experimental chamber

Salinity (ppt) pH DO (%) Salinity pH DO (%) NH3 (lg/l)

1 30.0± 0.7 8.02± 0.08 96.1± 0.3 30.7± 0.5 7.99± 0.05 79.6± 5.0 62.9± 63.4

2 30.0± 0.1 8.09± 0.03 97.8± 0.8 30.6± 0.5 8.05± 0.02 74.7± 2.6 15.7± 19.6

3 30.0± 0.1 8.05± 0.09 95.7± 1.7 30.0± 0.1 7.93± 0.04 75.2± 5.4 20.9± 13.6

(1) Hilton Head and Kiawah Island sediments; (2,3) Hilton Head Island and Okatee River watershed sediments.

Table 2

Bioassay endpoints (mean± 1 SD) (4 replicates per site) from copepod chronic exposures to sediments collected adjacent to various land-uses

Bio-

assay

Site Land-use type Bioassay endpoints

Female

survival (%)

Male survival

(%)

Copepodites Nauplii Realized

offspring

production

Clutch size

1 NI Non-urban creek

(control)

79± 16 79± 15 319± 46 603± 65 48± 9 17± 6

HHI HTA Golf course––creek 82± 7 70± 10 89± 31� 375± 75� 23± 5� 13± 4�

HHI PD2 Golf course––Ret.

pond

82± 9 84± 13 102± 7� 349± 28� 22± 4� 11± 4�

HHI PD3 Suburban––Ret.

pond

80± 12 67± 8 90± 32� 343± 11� 22± 5� 10± 4�

KI TCE Golf course––Ret.

pond

87± 7 74± 3 396± 84 703± 91 50± 2 17± 5

KI TCW Suburban––creek 71± 8 84± 3 226± 47 431± 87� 37± 7 14± 4

KI OR Suburban––creek 73± 13 68± 14 308± 51 646± 59 54± 9 11± 4�

2 NI Non-urban creek

(control)

56± 14 31± 10 85± 27 216± 57 22± 5 13± 4

HHI HT1 Golf course––creek 52± 19 49± 13 72± 7 129± 46� 19± 9 12± 4

HHI HT2 Golf course––creek 40± 7 28± 12 85± 44 165± 24 25± 6 13± 4

HHI HT3 Golf course––creek 60± 9 48± 26 9± 6� 89± 34� 6± 2� 10± 5

ORW OM Outlet Mall––Ret.

pond

6± 2� 10± 7� 3± 2� 21± 12� 15± 6 5± 6�

ORW T1-P Suburban––creek 51± 7 50± 13 36± 18� 148± 40 15± 4 11± 4

ORW T3-H Hospital––Ret.

pond

2± 2� 5± 9� 0± 0� 0± 0� 0± 0� 0± 0�

3 NI Non-urban creek

(control)

81± 7 78± 7 3± 3 151± 23 8±1 8± 2

ORW OCB Highway––creek 78± 10 65± 14 3± 4 9± 6� 1± 0� 7± 1�

ORW OCF Suburban––creek 81± 4 72± 12 2± 1 15± 6� 1± 0� 7± 1�

ORW T3GC Golf course––Ret.

pond

68± 18 58± 14 2± 1 53± 3� 4± 1� 7± 1

ORW T3C Golf course––creek 83± 4 81± 8 6± 3 47± 13� 2± 1� 8± 3

HHI HTB Golf course––creek 78± 8 67± 19 12± 5� 327± 31� 17± 2� 11± 3�

B BP Golf course––creek 73± 11 72± 19 21± 7� 205± 7� 13± 3� 7± 2

(�) Represents significant difference from North Inlet sediments (NI, control sediments; a ¼ 0:05). HHI: Hilton Head Island, KI: Kiawah Island,

ORW: Okatee River watershed and B: City of Beaufort. Starting male and non-gravid female copepod abundances in each test chamber were 25

each.

26 A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32

to controls. For example, copepodite production was

reduced by 55–85% for all Hilton Head sites (HTA,

PD2, PD3) (176–271 copepodites; ANOVA, p <0:0001), while nauplii production was reduced by 27–56% (163–338 nauplii; ANOVA, p < 0:001) relative to

controls (including the Kiawah Island (TCW) site).

Except for sites TCE and TCW on Kiawah Island,

clutch size of surviving females was reduced by 12–54%

(2–9 embryos; ANOVA, p < 0:002). Consequently, real-ized offspring production in all Hilton Head samples

was 38–66% (ANOVA, p < 0:0002) lower than controls.

Moreover, copepodite-to-naupliar ratios were 21–72%(ANOVA, p < 0:008) lower than controls, indicating

either a significant reduction in naupliar production or

a reduction in numbers of nauplii successfully develop-

ing into copepodites.

Bioassay endpoints

%GF %F %M F/M C N C/N TO RO CS

Num

ber o

f tes

t sed

imen

ts

0

3

6

9

12

15

18 >10%>25%

50% test sediments

Fig. 1. Number of test sediments ðn ¼ 18Þ, in which bioassay endpoints

were reduced by >10% or >25% relative to controls. %GF¼% Gravid

female; %F¼% surviving females; %M¼% surviving males;

F/M¼ surviving female-to-male ratio; C¼ copepodites; N¼ nauplii;

C/N¼Copepodite-to-naupliar ratio; TO¼ total offspring; RO¼ rea-

lize offspring; and CS¼ clutch size.

A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32 27

In Bioassay 2 we tested sediments from Hilton Head

Island and the upper Okatee River watershed near

Bluffton. Adult copepod survival was consistently <30%

in OM and T3-H sediments (<15 adult copepods; AN-OVA, p < 0:0001) relative to controls. Adult copepods

exposed to T3-H sediments showed 100% reproduc-

tive failure. Compared to controls, sites OM, T1-P and

T3-H showed a reduction in copepodite production

>40% (>26 copepodites; ANOVA, p < 0:0001). Simi-

larly, most sites (except for HT1 and T1-P) showed a

reduction in naupliar production >40% (>87 nauplii;

ANOVA, p < 0:001). Despite the depressed naupliarproduction, realized offspring production was reduced

by >70% only for two sites (HT3 and T3-H), while co-

pepodite-to-naupliar ratios were reduced by >40% in

three sites (HT3, OM and T3-H; ANOVA, p < 0:0004).Bioassay 3 evaluated sediments from the Okatee

River watershed (OCB, OCF, T3GC, T3C), Hilton

Head (HTB) and the City of Beaufort retention pond

(BP). As with Bioassay 1, adult survival and incidence ofgravid females were not statistically different from con-

trols for any sample (p > 0:05; Table 2). Although

copepod production was unusually low in control sedi-

ments, in all Okatee River sites (OCB, OCF, T3TG and

T3C) nauplii and realized production were reduced by

>60% (>90 nauplii) and >68%, respectively. HTB and

BP showed significantly higher reproductive output than

those of controls (ANOVA, p ¼ 0:001). Due to lowcopepodite counts in controls, copepodite-to-naupliar

ratios in contaminated sediments were not statistically

different from controls (ANOVA, p > 0:05).In order to determine (1) how bioassay endpoints were

influenced by sediments from sites adjacent to several

land-uses; and (2) which sites posed the greatest risk to

A. tenuiremis, we pooled all of the bioassay results by

calculating changes in endpoints relative to controlsediment responses within bioassays. Specifically, the

number sites in which values of individual bioassay

endpoints were reduced by >10% or >25% relative to

control sediments (Fig. 1), were tallied, as were the

number of bioassay endpoints per site reduced by >10%

or >25% relative to within bioassay controls (Fig. 2).

Sites with >10% or >25% reduction in bioassay end-

points were categorized as of moderate or high risk to A.

tenuiremis, respectively.

Using this approach, male survival was reduced by

>10% relative to controls in 50% (9 of 18) of test sedi-

ments (Fig. 1). Copepodite and naupliar production,

and clutch size were reduced by >10% in 75% (12 of 18),

88% (14 of 18) and 75% (12 of 18), respectively; while

copepodite and naupliar production were reduced by

>25% in 50% (9 of 18) and 75% (12 of 18) of the sedi-ment tested, respectively. Total and realized production

were reduced by >10% in 88% (14 of 18) and 81% (13 of

18) of test sediments, and >25% in 75% (12 of 18) and

69% (11 of 18) of test sediments, respectively. Thus it is

apparent that reproductive endpoints to A. tenuiremis

(clutch size, naupliar and copepodite production), rather

than adult survivorship, are more sensitive indicatorsof chronic effects associated with exposure to these

sediments.

Similarly, 11 of the 18 sediments (HTA, PD2, PD3,

TCW, HT3, OM, T1-P, T3-H, OCB, OCF, T3GC)

showed a >10% reduction in more than half of the

bioassay endpoints (Fig. 2) relative to controls. Six sites

(PD2, PD3, HT3, OM, T1-P, T3-H) showed a >25%

reduction in the same bioassay endpoints (Fig. 2). Theseresults indicate that 39%, 28% and 33% of the sites

respectively may pose a low, moderate or high risk,

to A. tenuiremis-like meiobenthic copepod populations

(Table 3).

3.2. Sediment chemistry

Sanger et al. (1999) reported a total PAH concen-tration (

P23 PAHs) of 116 ng/g (dw) for North Inlet

sediments (NI, control site). Normalization of the data

from Sanger et al. (1999) results in an estimated 37 ng/g

(dw) for low molecular weight PAHs (LMW; 2–3 rings)

and 26 ng/g (dw) for high molecular weight PAHs

(HMW; 4–6 rings; Table 3).

Sediment samples from PD2, PD3, OM and T3-H

had non-detectable levels of either LMW or HMWPAHs. In site HTA (Bioassay 1) Phen accounted for

100% of the LMW PAHs, while Flu, Pyr, and B(bk)F

accounted for �50% of the total HMW PAHs. Hilton

Head sites HT3, HT2, HT1 (Bioassay 2) had detectable

LMW and HMW PAHs. In HT3 sediments Phen ac-

counted for 100% of the LMW PAHs, while the only

HMW PAH detected wereP

Flu, Pyr, B(a)A, Chry,

B(bk)F. HT2 sediments had detectable levels of the

SitesHTA PD2 PD3

TCETCW OR HT1HT2 HT3 OMT1-PT3-H OCBOCFT3GC T3C HTB BP

Bio

assa

y en

dpoi

nts

0

2

4

6

8

10>10%>25%

50% bioassayendpoints

Fig. 2. Number of bioassay endpoints (i.e., % male, female and gravid females, and sex specific survival; copepodite, nauplii, copepodite-to-naupliar

ratios, clutch size and realized offspring) per site ðn ¼ 18Þ reduced by >10% or >25% relative to controls. Values represent mean (±1 SD) (4 replicates

per site).

Table 3

Measured total low and high molecular weight polycyclic aromatic hydrocarbon (LMW and HMW-PAH; ng/g dw) concentrations in sediments

collected adjacent to several land-uses

Sites LMW (ng/g dw) HMW (ng/g dw) HMW/LMW ERMQ PAH-riska Copepod-riskb

NIc 37.0 26.0 0.70 0.002 Low na

HTA 1,060 12,400 12 0.629 High Moderate

PD2 nd nd na na Low High

PD3 nd nd na na Low High

HT1d 4,170 24,600 5.9 1.020 High Low

HT2 99.0 1,380 14 0.060 Moderate Low

HT3 15.0 105 7.0 0.008 Low High

OM nd nd na na Low High

T1-P nd 149.0 0.005 Low High

T3-H nd nd na na Low High

OCB 18.0 131.0 7.3 0.005 Low Moderate

OCF 16.0 51.0 3.2 0.004 Low Moderate

T3GC 7.0 29.0 4.1 0.003 Low Moderate

T3C 22.0 89.0 4.0 0.004 Low Low

HTBe 616 6,950 11.3 0.245 High Low

BP 126 1,190 9.4 0.048 Moderate Low

nd¼ not detected (<�10 ng/g). na¼not applicable. Average effects range median quotient (ERMQ) value represents the combined effect of multiple

co-existing PAH concentrations. Predicted risk based on PAH-ERMQ values and copepod bioassay endpoints.a ERMQ low<0.02, moderate >0.02–<0.077, and high> 0.077 risk of observing deleterious effects on benthic communities (Long et al., 1995;

Hyland et al., 1999).bRisk based on the number of bioassay endpoints per site in which values were reduced by <10% (low risk), >10% (moderate risk) and >25% (high

risk) compared to controls.c Sanger et al. (1999).d #ERL exceedances¼ 10, #ERM exceedances¼ 5.e #ERL exceedances¼ 7.

28 A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32

LMW PAHs Phen, Anth and 1mP-PAHs with Phen

accounting for 68% of the LMW PAHs. In these sedi-

ments Flu, Pyr and B(bk)F accounted for 48% of the

total HMW PAHs (P

all HMW PAHs detected). In

HT1 sediments, although several LMW PAHs (P

Naph,

1mN, Ace-Nap, Fluor, Phen, Anth and 1mP) were de-

tected, Phen accounted for 74% of the total LMW

PAHs. Flu, Pyr and B(bk)F accounted for 51% of the

total HMW PAHs (P

all analytes¼ 24,629 ng/g). Sedi-

ments from HT1 had the greatest PAH concentrations

found in this study (P

Total PAHs¼ 28.8 lg/g, dw).

Okatee River sediments (sites T3GC, T3C, OCB and

Table 4

Spearman rank correlation ðqÞ of the bioassay endpoints versus total low and high molecular weight polycyclic aromatic hydrocarbons (LMW and

HMW-PAHs, ng/g dw), and average effects range median quotient (ERMQ) values ðn ¼ 72ÞBioassay endpoints LMW HMW ERMQ

% gravid females 0.36 (0.002) 0.18 (>0.05) 0.21 (>0.05)

% female survival 0.001 (>0.05) 0.03 (>0.05) 0.06 (>0.05)

% male survival )0.02 (>0.05) 0.02 (>0.05) 0.06 (>0.05)

Female-to-male ratio )0.01 (>0.05) )0.1 (>0.05) 0.13 (>0.05)

Copepodites 0.43 (0.0002) 0.1 (>0.05) 0.14 (>0.05)

Nauplii 0.48 (<0.0001) 0.08 (>0.05) 0.13 (>0.05)

Copepodite-to-nauplii ratios 0.3 (0.01) 0.15 (>0.05) 0.18 (>0.05)

Clutch size 0.37 (0.002) 0.27 (0.02) 0.35 (0.003)

Parenthesis represents the p-value of the correlation.

A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32 29

OCF; Bioassay 3) contained <22 ng/g LMW PAHs with

Phen accounting for 37–100%. HMW PAH ranged from

as low as 29 ng/g (T3GC) to 131 ng/g (OCB), with Flu,

Pyr, B(it bk)F and Per accounting for 63–80%. In con-

trast, BP and HTB had elevated LMW PAHs with Phen

accounting for 56% and 74% of the total LMW PAHs,

respectively. Although all HMW PAHs were detectedin these two sites, Flu, Pyr, Chry and B(bk)F accounted

for 54% and 57% of the HMW PAHs, respectively.

The distribution of PAH in environmental samples

may affect the toxic potential and thus pose risk to

benthic estuarine infauna. For example, the mutage-

nicity/carcinogenicity of HMW compounds such as

benzo[a]pyrene are well documented. Conversely, the

increased bioavailability via the aqueous phase of LMWPAH such as naphthalene and phenanthrene may pose a

greater risk due to narcosis. In contrast to the control

(North Inlet) sediment, all test sediments were enriched

with HMW PAH (Table 3). This distribution is consis-

tent with sediments from other coastal urban estuaries

(Kucklick et al., 1997; Kim et al., 1999) containing

weathered combustion or petroleum source PAH (Prahl

and Carpenter, 1983).Based on measured PAH concentrations and fol-

lowing the criteria developed by Hyland et al. (1999), the

estimated ERMQ (Table 3) indicates that sediments

from HTA, HT1 and HTB, may pose a high risk

(ERMQ >0.077) to associated sediment communities.

Similarly, sediments from HT2 and BP with ERMQ

values ranging >0.02 to <0.077, indicates that these

sediments may pose a moderate risk to associated sedi-ment communities. Sediments from the remaining 11

sites have ERMQ levels <0.02, indicating a low risk to

benthic assemblages. No apparent relationship between

HMW/LMW PAH and copepod rise were observed.

The results from the Spearman’s rank correlation

analysis for all reproductive output variables versus

both total LMW and HMW PAHs, and ERMQ values

are presented in Table 4. None of the endpoints (exceptclutch size) show a correlation with either HMW PAHs

or ERMQ values ðp > 0:05Þ. Surprisingly, percent gra-vid females, copepodites, nauplii, and copepodite-to-

naupliar ratios were all positively correlated with LMW

PAHs ðp < 0:05Þ. Clutch size was the only endpoint

positively correlated with both LMW and HMW PAHs,

and ERMQ values ðp < 0:05Þ.

4. Discussion

Concentrations and composition of PAHs in coastal

environments are reflective of land-use. Highly alkylated

and low molecular weight PAHs (LMW, naphthalenes)

dominate suburban and forested watersheds (Kucklick

et al., 1997; Sanger et al., 1999), while pyrogenic high

molecular weight PAHs (HMW, 4–6 rings) dominate

urban runoff (Kucklick et al., 1997; Menzie et al., 2002).

In our sediments, phenanthrene [Phen] was the mostcommonly detected LMW PAH, while fluoranthene

[Flu], pyrene [Pyr], chrysene [Chry], benzo [b+ k], fluo-

ranthene [B(bk)F] and perylene [Per] were the most

abundant among HMW PAHs.

Comparing measured total PAHs concentrations of

the current study with those reported by Sanger et al.

(1999) for South Carolina tidal creeks and salt marshes,

the majority of sites fell within PAH concentrationsreported for forested/reference sites. Sites with moderate

(HT2, BP) to high (HTA, HT1, HTB) total PAH con-

centrations were similar to coastal South Carolina’s

suburban and industrial/urban areas, respectively. Only

three of these sites (HTA, HT1, HTB) had total PAH

concentrations exceeding the highest concentration

(12,300 ng/g dw) found in estuarine sediments along the

Carolinian Province (Hyland et al., 1998). The pre-dominance of Phen, Flu, Pyr, Chry and B(bk)F is con-

sistent with the PAH combustion product signature

found in runoff water from Murrells Inlet, SC, USA

(Ngabe et al., 2000) and in sediment samples from

Charleston Harbor, SC (Kucklick et al., 1997).

The partial life cycle bioassay with the copepod A.

tenuiremis indicated that reproductive endpoints were

the most sensitive indicators of chronic effects followingexposure to sediments associated with coastal land-uses.

Reproductive outputs were depressed in 11 sediment

sample sites, with the most dramatic effects observed in

sediments from Hilton Head (HTA, PD2, PD3 and

30 A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32

HT3) and the Okatee River watershed (OM, T3-H, OCB

and OCF). Sites from Kiawah Island (TCE, TCW, OR)

did not cause any reproductive effects to A. tenuiremis.

Furthermore, based on all the bioassay endpoints, 33%of the sites (PD2, PD3, HT3, OM, T1-P and T3-H) may

pose a high risk to long term A. tenuiremis population

maintenance. In this study, we found a general lack of

agreement between the chronic toxicity of PAH con-

taminated sediments to A. tenuiremis and the cumulative

risk estimate based on effects range median quotient

(ERMQ) values for PAHs (Long et al., 1995 and Hyland

et al., 1999). Sediments from PD2, PD3, OM, HT3 andT3-H were found to pose a high risk to A. tenuiremis, yet

ERMQ for these sites place them in a low risk category.

Likewise, the ERMQ categorized HTA, HT1, and HTB

sediments as high potential risks to benthic fauna; yet

copepod bioassay results placed two of these sediments

(HT1, HTB) in low, and one (HTA) in the moderate risk

category. Furthermore a few sediment samples, with

PAH concentrations near or below instrument detectionlimits (PD2, PD3, OM and T3-H), and with ERMQ

values indicating low risk (the above plus HT3, T1-P,

T3-H, OCB and OCF), in some cases produced sig-

nificant effects on reproductive output. These results

suggest that observed deleterious effects on copepod

reproduction could not predicted by the ERMQ ap-

proach.

LMW PAHs are indicative of non-anthropogenicsources (Kucklick et al., 1997) and/or relatively fresh

petroleum contamination (Prahl and Carpenter, 1983).

LMW PAHs are also generally considered less toxic

than HMW PAHs (Eisler, 1987). Interestingly, a strong

positive correlation was seen between LMW PAHs and

percent gravid, copepodites, nauplii, and the develop-

mental rate (i.e., copepodite-to-naupliar ratios). Clutch

size was also positively correlated with LMW as well asHMW PAHs, which is consistent with a previous study

that found larger clutch sizes in female A. tenuiremis

chronically exposed to contaminated sediments (Fulton

et al., 1996). An increase in clutch size might be a

compensatory mechanism by which females reduce

PAH bioavailability. Another study with A. tenuiremis

found that female copepods chronically exposed to

chrysene had an increased yolk (lipovitellin) depositionto eggs relative to control females (Chandler and Volz,

in press). Also, female grass shrimp Paleomonetes pugio

exposed to pyrene (63 ppb) showed elevated vitellin

production, as well as increased embryo mortality

(Oberd€orster et al., 2000a). Thus, enhanced yolk pro-

duction, and potentially increased clutch size, may serve

as a mechanism for depuration of highly lipophilic

compounds such as PAHs, rendering these compoundsless biologically available to females (Oberd€orster et al.,2000b; Chandler and Volz, in press), but more available

to offspring. An alternative explanation for this positive

association is a mild bio-stimulatory effect due to low

level petroleum contamination, which in turn may

nourish higher trophic levels of the tidal creek benthic

food web.

The absence of a strong statistical correlation be-tween sediment PAHs (LMW, HMW, ERMQ) and

toxicity may be related to the presence of other

unmeasured anthropogenic contaminants (i.e., pesti-

cides, heavy metals) or unaccounted for physical (i.e.,

grain size) and chemical (i.e., organic matter content and

quality) sediment properties, which may have influenced

the nutritional status of test copepods as well as the

bioavailability of PAHs. A similar study (McGee et al.,1999) using PAH-laden field sediments found no sig-

nificant Spearman correlation between PAH concen-

trations and amphipod survival (10-day sediment

bioassay) and field densities. In this study, however,

amphipod survival was correlated with PAH ERMQ

values.

Despite a lack of inverse correlation between copepod

and amphipod bioassay endpoints and PAH concen-trations, PAHs can cause adverse effects on sediment

infauna. Although most PAHs exert narcotic toxicity

at environmental concentrations, the toxicity of PAHs

such as fluoranthene, pyrene, benzo[a]anthracene, anth-

racene and benzo(a)pyrene (all of which were detected

in some our sediment samples) can be exacerbated by

ultraviolet (UV)-photoactivation (Kosian et al., 1998;

Newsted and Giesy, 1987; Ankley et al., 1997). Bioas-says in this study were conducted under cool-fluorescent

lighting, so UV impacts would not be expected.

Other studies have shown the potential interaction of

PAHs with the arthropod ecdysteroid hormone system

(Oberd€orster et al., 1999, 2000b). Oberd€orster et al.

(1999) showed that PAHs enhance the in-vitro effects of

the hormone 20-Hydroxyecdysone (20HE; responsible

for crustacean molting, development and reproduction)on ecdysteroid-responsive imaginal disc cell lines by

reducing cell proliferation. Furthermore, PAHs could

interfere with the metabolism of ecysteroids in molting

crustaceans (Oberd€orster et al., 2000a,b). In the present

study, A. tenuiremis cultured in field contaminated sed-

iments from Hilton Head (HT1, HT3) and the Okatee

River (OCB, OCF) showed reduced titers of 20HE rel-

ative to controls (Bejarano, unpublished). For example,in late stage embryos microdissected from females ex-

posed to sediments from HT1 and HT3, 20HE titers

were 65% lower (11.7 ± 2.5 and 16.3 ± 8.4 fmol/embryo,

respectively, n ¼ 7) than 20HE in embryos from con-

trols (35.2 ± 7.8 fmol/embryo; n ¼ 7). Similarly, 20HE

titers in males exposed to sediments from OCB and

OCF were �60% lower (9 ± 2.2 and 9.1 ± 1.8 fmol/male,

respectively, n ¼ 10) compared to males exposed tocontrol sediments (24.9 ± 7.5 fmol/male, n ¼ 10). Re-

duced 20HE titers in HT1 and HT3 embryos might be

related to depressed naupliar production and develop-

mental rates in these sediments, since in this species late

A.C. Bejarano et al. / Marine Pollution Bulletin 49 (2004) 23–32 31

stage embryos have elevated 20HE titers prior to

hatching into the naupliar stage (Block et al., 2003).

Conversely, the relationship between reduced 20HE in

males exposed to OCB and OCF sediments and repro-ductive output is less straightforward. As suggested in a

previous study, elevated 20HE in male A. tenuiremis

may be associated with spermatogenesis (Block et al.,

2003). In other arthropods (i.e., ticks) ecdysteroids have

been found to stimulate spermatocyte-DNA synthesis

during early spermatogenesis (Oliver and Dotson, 1993).

However, the link between 20HE titers and offspring

production under exposure to contaminants such asPAHs is currently not well understood for any crusta-

cean.

This study did not elucidate relationships among

sediment PAH concentrations and distributions, and

reproductive and developmental responses in A. tenui-

remis. Thus, determining the potential risk to indigenous

copepod populations inhabiting areas adjacent to vari-

ous land-use coastal environments is more complexthan considering PAH alone as a major pollutant.

Further research should focus on the analysis of multi-

ple anthropogenic-related contaminants (PAHs, heavy

metals, legacy and newly registered pesticides, PCBs),

and the addition of endocrine/hormonal endpoints in

toxicological models to better assess risks of contami-

nated sediments to estuarine meiobenthos.

Acknowledgements

The authors would like to thank T.L. Cary, S. Klo-

sterhaus and J.J. Rafalowski for technical assistance

with bioassay setup and sediment analyses, and two

anonymous reviewers for their useful comments on the

manuscript. This research was funded by the SouthAtlantic Bight Land Use-Coastal Ecosystem Study (LU-

CES) under an award from the Coastal Ocean Pro-

gram (COP), National Ocean Service (NOS), National

Oceanic and Atmospheric Administration (NOAA),

through the South Carolina Sea Grant Consortium,

Grant number NA960PO113. This publication does not

constitute an endorsement of any commercial product

or intend to be an opinion beyond scientific or otherresults obtained by NOAA. This research has not been

subject to either agencies’ peer or policy review, and

no endorsement should be inferred.

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