Interannual variability in fish biomarkers in a contaminated temperate urban estuary

ARTICLE IN PRESS

Ecotoxicology and Environmental Safety 62 (2005) 53–65 www.elsevier.com/locate/ecoenv

Interannual variability in fish biomarkers in a contaminated temperate urban estuary Diane Webba, Marthe Monique Gagnona,, T.H. Roseb a

Department of Environmental Biology, Curtin University of Technology, GPO Box U1987, Bentley Campus, Perth, WA 6845, Australia b Water & Rivers Commission, Hyatt Centre, East Perth, WA 6892, Australia Received 2 December 2003; received in revised form 29 October 2004; accepted 1 December 2004 Available online 2 February 2005

Abstract During the past decade the Swan–Canning estuary, Western Australia, has shown signs of stress which has been attributed to high nutrient inputs. There is little information on the effect of nonnutrient contaminants on biota inhabiting the estuary. A suite of biomarkers was measured on black bream (Acanthopagrus butcheri) to determine whether annual variations in fish biomarkers exist in the wet (winter) and dry (summer) seasons. Serum sorbitol dehydrogenase showed no significant differences between years, indicating that measured mixed-function oxygenase (MFO) enzyme activities were not affected by annual variations in hepatic tissue damage. Ethoxyresorufin-O-deethylase activity was lower in female black bream than in male fish while ethoxycoumarin-Odeethylase activity was not influenced by gender. Biomarker levels measured at various sites confirm that major roads and drains are significant contributors of MFO-inducing chemicals into the Swan–Canning estuary. No consistent upstream or downstream gradient in biomarker response was identified. The ratio of naphthalene-type to benzo(a)pyrene-type biliary metabolites was linked to runoff from urban areas into the estuary. There was high annual variability in all biomarkers in both seasons, suggesting that biannual monitoring is required to evaluate the effect of contaminants on the biota in the estuary. r 2004 Elsevier Inc. All rights reserved. Keywords: Bile metabolites; Biomarker; Biomonitoring; Black bream; ECOD; EROD; PAH; SDH; Swan–Canning estuary

1. Introduction In the early 1990s, the Swan–Canning estuarine basin was showing signs of stress with algal blooms, fish deaths, and toxic blue-green bloom events occurring on a frequent basis. In 1994, the Government of Western Australia launched the Swan–Canning Cleanup Program to study the situation and develop a management program (Swan River Trust, 1998). The main emphasis of this program has been on nutrient inputs to the river system from rural and semirural land in the upper coastal catchment, which fuels highly visible algal blooms. Very little about the impacts of urban inputs of potential, low-level contaminants into the estuary from road runoff and stormwater drainage is known. Corresponding author. Fax: +61 8 9266 2495.

E-mail address: [email protected] (M.M. Gagnon). 0147-6513/$ - see front matter r 2004 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2004.12.003

The Swan and Canning rivers flow through the heart of metropolitan Perth, the capital of Western Australia with a population of 1.4 million people. The total estuarine portion of the Swan–Canning system occupies an area of around 55 km2 and is an important historical, recreational, and economic focus for Western Australia (Swan River Trust, 1998). Residential areas, golf courses, parks, light industry, landfills, and old liquid disposal sites combined with an extensive road network that bound the estuary may be point sources of pesticides, hydrocarbons, solvents, and organic and inorganic chemicals into the estuary. Sealed road surfaces, car parks, and airport tarmacs deflect water runoff to drains carrying petroleum hydrocarbons, which eventually reach the estuary following heavy rains. Conventional methods of environmental monitoring of these inputs have relied on chemical analysis of sediments, water, and fish flesh. However, these

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chemical analyses do not reveal the sublethal effects of contaminants on the estuarine biota. Biochemical measurements using biomarkers of health link the bioavailability of compounds of interest with their concentration in target organs and intrinsic toxicity (van der Oost et al., 2003). Detoxification of xenobiotics is a function of the liver and involves enzymes that metabolize compounds to facilitate their excretion from the body. Specific forms of mixedfunction oxygenase (MFO) enzymes, such as cytochrome P450, are induced by exposure to a variety of lipophilic compounds such as organochlorines, polychlorinated dibenzodioxins, polychlorinated dibenzofurans, petroleum aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs) (Hodson et al., 1991; Connell et al., 1999). Fish exposed to PAHs, PCBs, and organochlorine pesticides show increased levels of MFO enzyme activity in the liver (Funari et al., 1987; Collier and Varanasi, 1991; Holdway et al., 1994; Arcand-Hoy and Metcalfe, 1999). Because these contaminants are present in the Swan–Canning River system (Gerritse et al., 1995, 1998; Swan River Trust, 1999), the induction of MFO enzyme activity has the potential to be a suitable biomarker of contaminant bioavailability. The measurement of MFO induction, as a biomarker of exposure, can be complemented with the measurement of a biomarker of effect such as the activity of the liver enzyme sorbitol dehydrogenase (SDH). The presence of this enzyme is normally negligible in the blood stream and an elevated serum SDH (s-SDH) activity indicates that hepatocellular injury has occurred (Ozretic and Krajnovic-Ozretic, 1993). Fish livers with cellular injuries related to xenobiotic exposure are less capable of MFO induction than are noninjured livers (Holdway et al., 1994). Importantly, s-SDH activity is specific to liver damage and is not altered by conditions often affecting the induction of ethoxyresorufin-Odeethylase (EROD), such as reproductive status (Dixon et al., 1987). Hepatocellular injury confounds interpretation of MFO induction and supports the parallel use of a marker of liver damage such as s-SDH activity. Holdway et al. (1994) found elevated s-SDH activity in association with reduced EROD induction in fish collected from Port Phillip Bay, Victoria, which had been exposed to high contaminant levels. Following metabolism of petroleum hydrocarbons by MFO enzymes, metabolites are mainly excreted via the biliary route (Connell et al., 1999). The presence of metabolites of PAHs in the bile indicates that the compounds are bioavailable and that absorption and metabolism of contaminants has taken place (Gagnon et al., 1999). PAH exposure cannot be reliably determined by measuring fish tissue levels (van der Oost et al., 2003); however, PAHs biliary metabolites display strong and characteristic fluorescent properties. Bile metabolites are measured by fixed-wavelength fluorescence (FF), with

the metabolites classified into naphthalene-, pyrene-, or benzo(a)pyrene (B(a)P)-types of metabolites. The simplicity of the method makes it possible to measure PAH exposure on a large number of fish at a low cost (Lin et al., 1996; Aas et al., 1998). In 2000, a preliminary laboratory study on the bioavailability of contaminants to fish inhabiting the estuary was conducted to assess whether the biota were exposed to potentially harmful PAHs. Black bream (Acanthopagrus butcheri Munro) was selected from among four native fish species as a suitable candidate for field investigations (Webb and Gagnon, 2002a). The objective of the present study was to determine whether patterns of responses within the estuary were sufficiently strong to overcome natural annual variability. Data from sampling black bream from four sites in late winter 2000 and late winter 2002 in addition to five sites during summer 2001 and summer 2002 are compared. Conclusions on the interannual variability of a suite of fish biomarkers in a contaminated temperate, urban estuary are drawn.

2. Methods 2.1. Chemicals All chemicals used in this work were purchased from Sigma Chemical Co., USA, unless otherwise indicated. 2.2. Sampling stations Black bream were captured from the Swan River, between 6 and 40 km from the estuary mouth, and from the Canning River, 9.5 km upstream from its confluence with the Swan River (Fig. 1). None of the sites could be considered true reference sites, as all areas in the Swan–Canning estuary have been impacted by human activity (Swan River Trust, 1998). The following sites were sampled (Fig. 1): Helena: This site is about 40 km upstream from the Swan River estuary mouth, but downstream from the major tributaries of Bennett Brook and the Helena River. The area is surrounded by developed land with p15% natural vegetation. Drainage from roads, bridges, residential properties, hobby farms, light industry, and parks enter this site. Ascot: This site is approximately 6 km downstream from the Helena site and is bounded by a racecourse and major residential development. This station receives drainage from the International and the Domestic Airports to the south, and a major drain bringing stormwater and road drainage from up to 7 km north of the estuary enters nearby. An important arterial road crosses the river at this point.

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collection area are two bridges carrying road traffic across the river. The site receives drainage directly from surrounding urban roads, gardens, and light industry. 2.3. Fish and sample collection

Fig. 1. Field collection sites within the Swan–Canning estuary were Helena, Ascot, Barrack Street, and Freshwater Bay in the Swan River and Riverton in the Canning River (adapted from Swan River Trust, 1998).

Barrack Street: This site is 20 km from the mouth near the northern bank of a section of the estuary known as Perth Water and has the Perth Central Business District (CBD) on its northern shore and significant residential and commercial development along the southern banks. Perth Water has undergone many changes since European settlement with river reclamations, channel deepening, and riverbank modifications and is regularly dredged for navigational purposes. The banks are bound by major roads and a freeway interchange carrying heavy vehicular traffic. The site has jetties from which river ferries operate, moor, and conduct maintenance. Barrack Street receives stormwater drainage from the CBD including air conditioning discharges, road runoff from its immediate surrounds, and a major drain, which collects road and stormwater drainage from heavy urban development up to 8 km north of the estuary. The Swan River Trust annually reports spillages occurring near the Barrack Street Jetty (e.g., 600 L diesel in January 2002 and 30 kl sewage in February 2002, Swan River Trust, 2002). Freshwater Bay: This site is the most downstream site on the Swan River, 6 km from the estuarine mouth, and has two large yacht clubs with slipping facilities servicing both sail and motorized watercraft. The banks are bounded by long established suburban residential development with roads, parks, and gardens, which drain directly into the estuary. Riverton: This site is situated in the Canning River about 9.5 km upstream from its confluence with the Swan River and has river sediments consisting of thick, greasy black mud. Slightly downstream from the

Black bream were collected by a commercial fisherman, during the late winter months of August and September 2000 and 2002, from the four sites in the Swan River (N ¼ 69 and 76 for each year, respectively). During late summer (April–May) in 2001 and 2002, the same four Swan River sites (N ¼ 84 and 64 for each year, respectively) were sampled together with an additional site at Riverton in the Canning River (N ¼ 16 and 21 for each year, respectively) (Fig. 1). Sampling occurred when barometric pressure, moon phases, and prevailing winds were favorable for black bream commercial fishing. A 120-m, 100-mm monofilament haul net was used to capture adult black bream. Collection took place between 2200 and 2400 h when the fish entered shallow water for feeding. Water samples, collected at a depth of approximately 2 m, were measured for temperature, pH, and salinity. Upon capture, the fish were maintained alive on board in a fiberglass tank with carbon-filtered river water. On return to shore, the fish were transferred to a 1000-L aerated vat filled with river water for transport back to the nearby laboratory. Fish were sacrificed within 2 h of capture. Fish were weighed, standard, fork, and total lengths were recorded, and an external examination for any signs of abnormalities or parasitic infestation was conducted. A sample of blood from the caudal vein was taken using a vaccutainer. The blood samples were allowed to clot on ice for 15 min and then centrifuged for 10 min at 3000 rpm, and the serum was collected. Each fish was then killed by a spike through the brain (Iki Jimi method) and dissected, and the bile was removed from the gall bladder using a 1 mL syringe and needle. Livers were detached, quickly examined for any anomaly, weighed, and rinsed in ice-cold KCl, and a 1-g sample was placed in a cryovial for subsequent analysis. All samples were immediately immersed in liquid nitrogen and then transferred to a 80 1C freezer until analysis. Gonads were removed, examined for anomalies, weighed, and assigned to one of the following maturity stages: (1) undeveloped; (2) early development; (3) maturing; (4) pre spawning; (5) spawning; (6) spent; using Nikolskii’s (1969) scale of gonad development. The gonads and remaining abdominal organs were discarded and the fish weighed to record the carcass weight. The condition factor (CF), liver somatic index (LSI), and gonadosomatic index (GSI) were calculated according to the equations (1) CF ¼ [(BWGW)/TL3]  100,

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(2) LSI ¼ (LW/CW)  100, and (3) GSI ¼ (GW/CW)  100, where BW is the total body weight, GW is the gonad weight, TL is the total length, LW is the liver weight, and CW is the carcass weight. The CF is based on gonad-free weight to avoid any bias due to variations in sexual maturation, and the LSI and GSI are based on CW to avoid bias due to variable levels of fat in the gonads and intestines and due to variable GW (Hodson et al., 1991).

Ltd., USA)–NaOH buffer, pH 10.4, was added. The fluorescence of the glycine–NaOH-buffered supernatant was read at excitation/emission wavelengths of 380/ 452 nm (10 ex/10 em). ECOD activity was expressed as picomoles of 7-hydroxycoumarin produced per milligram of total protein per minute (pmol H mg Pr1 min1).

2.4. S9 postmitochondrial supernatant preparation

s-SDH activity was analyzed using the Sigma Diagnostic Kit (procedure No. 50-UV). The change in absorbance was read on a Pharmacia UV–Visible Spectrophotometer at 340 nm. An SDH control sample was run every 10 samples to ensure that the assay was performing within acceptable limits. s-SDH levels, in the serum of black bream, are expressed as milli-International Units (mU).

Individual liver samples were thawed on ice and then homogenized in N-2-hydroxyethylpiperazine-N0 -2-ethanesulfonic acid (Hepes), pH 7.5, using a Heidolph DIAX 900 homogenizer. The homogenate was centrifuged (Jouan CR3i centrifuge) at 9000g for 20 min at 4 1C and the postmitochondrial supernatant (PMS) collected for immediate use. Protein content of the PMS was determined according to Lowry et al. (1951). 2.5. Ethoxyresorufin-O-deethylase assay EROD activity was measured using a modified method of Hodson et al. (1991). The reaction mixture contained Hepes, pH 7.8, MgSO4, bovine serum albumin, b-nicotinamide adenine dinucleotide phosphate (NADPH), reduced form solution, and PMS. The reaction was initiated by adding ethoxyresorufin and incubating at room temperature for 2 min and then terminated by adding HPLC-grade methanol. Resorufin standards (0.000–0.085 mM) and samples were centrifuged to precipitate proteins and the fluorescence of the supernatant was read on a Perkin–Elmer LS-45 Luminescence Spectrometer at excitation/emission wavelengths of 535/585 nm (slit 10 ex/10 em). EROD activity was expressed as picomoles of resorufin produced per milligram of total protein per minute (pmol R mg Pr1 min1). 2.6. Ethoxycoumarin-O-deethylase assay Ethoxycoumarin-O-deethylase (ECOD) activity was assessed using an amended method of Holdway et al. (1998). The reaction mixture contained 0.1 M Tris (hydroxymethyl) aminomethane buffer, pH 7.4, KCl, MgCl2, 0.125 mM NADPH solution, and PMS. This mixture was incubated for 2 min in a water bath at 30 1C and then the reaction was initiated by adding 2 mM ethoxycoumarin. After further incubation at 30 1C for 10 min, the reaction was terminated by adding 5% ZnSO4 and saturated Ba(OH)2. Umbelliferone (7hydroxycoumarin) standards (0.000–0.093 nM) and samples were centrifuged to precipitate proteins. Then 1 mL of the resultant supernatant was transferred to a test tube and 0.5 M glycine (Bio-Rad Laboratories Pty

2.7. Serum sorbitol dehydrogenase assay

2.8. Determination of bile metabolites Three biliary metabolite-types (naphthalene, pyrene, and B(a)P) were measured by FF. The term metabolitetype is used because the method detects groups of compounds that fluoresce at specific wavelengths; e.g., nearly all naphthalene metabolites fluoresce at ex 290/ em 335 nm (Lin et al., 1996). This method offers the advantage of a very sensitive detection of a group of metabolites originating from a common parent compound (Lin et al., 1996). Consequently, the results are expressed as naphthalene-type metabolites as one group, with other groups of bile metabolites being of pyrenetype or B(a)P-type. The concentrations of bile metabolites are measured as metabolite equivalents to their respective standards. This represents the amount of a metabolite that would be present if the group of metabolites originated from a parent compound and is a relative term rather than an absolute concentration (Krahn et al., 1986). Bile samples were thawed on ice and diluted to 1:2000 in 50% HPLC-grade methanol/H2O for FF determination of pyrenol-type metabolites. Metabolites of pyrenetype were measured by FF at excitation/emission wavelengths of 340/380 nm (10 ex/10 em) (Aas et al., 1998), and B(a)P-type metabolites were read at 380/ 430 nm (10 ex/10 em) (Lin et al., 1996). The bile was further diluted to 1:5000 for the determination of naphthalene-type metabolites by FF at 290/335 nm (10 ex/10 em) (Lin et al., 1996). The concentration of PAH metabolites in the bile was determined using 1-OH pyrene (hydroxypyrene (98%), Aldrich Chemical Co., USA), for pyrene- and B(a)P-type, and 1-naphthol, for naphthalene-type metabolites, as external standards. The fluorescence reading of bile was converted to 1-OH pyrene or 1-naphthol equivalents from linear regression curves.

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Collier and Varanasi (1991) have shown that the normalization for protein concentration in the bile can, to a large extent, account for changes in the level of PAHs due to differences in the feeding status of some fish. The protein content of the bile was determined using the method of Lowry et al. (1951) to take into account the amount of water in the bile when collected from the gall bladder. FF analysis of bile measures groups of PAH metabolites with FF290/335 mainly composed of twoand three-ring structures, FF340/380 mainly four-ring structures, and FF380/430 mainly five-ring structures (Aas et al., 2000). PAH contamination originating from petrogenic sources show a dominance of two- and threering compounds over four- and five-ring compounds (Aas et al., 2000) while pyrogenic PAH is dominated by four- and five-ring compounds (Neff, 1990; Aas et al., 2000). Therefore, ratios between FF290/335, FF340/380, and FF380/430 can be used as indicators of the source of PAH contamination (Aas et al., 2000). Higher ratios indicate exposure to naphthalene-rich petroleum products such as unburnt fuels, while lower ratios indicate exposure to petroleum compounds of pyrolytic origin such as combusted petroleum hydrocarbons. Biliary PAHs are reported based on the ratio of naphthalene-type to B(a)P-type biliary metabolites (FF290/335:FF380/430). 2.9. Statistical analysis For each biomarker, the data were tested for normality and homoscedasticity and, where necessary, log transformed to achieve normality. Data from different years were compared using one-way analysis of variance

57

(ANOVA) on SPSS 10 for Macintosh. t-Tests were carried out to identify differences in sex where applicable. Where differences between the sexes were identified (Pp0.05), data for each sex were treated separately. Data are presented as mean7standard error (SE).

3. Results In the Swan–Canning estuary, low water levels are associated with the predominance of high barometric pressures and offshore winds with successive ebb phases dominant (Spencer, 1956). This event prevents the black bream from moving into mussel beds and reaching barnacles on jetty pylons, their preferred feeding grounds (K. Littleton, pers. commun.). At these times, black bream feed on macrophytes, polychaetes, shrimps, small fish, and other prey items in the deeper channels. Severe weather conditions (heavy rains and strong southerly winds) in mid-August and early September 2000 delayed sampling at most sites that year while during 2002 unusually low tides throughout the estuary restricted the black bream from moving into shallow feeding sites, resulting in delays in collection. Salinity of the estuarine water in the winter 2000 was low, ranging from 2.6 to 9.2 ppt, however, in winter 2002 salinities ranged from 2.8 upstream to 21.6 ppt downstream (Table 1). Summer salinities in 2001 ranged from 26.4 to 38.2 ppt and in 2002 from 7.6 to 34.0 ppt (Table 1). An overall total of 327 black bream were collected with the following annual breakdown (Table 2): winter 2000, 22 male, 47 female; winter 2002, 29 male, 47 female; summer 2001, 38 male, 59 female; and summer

Table 1 Water parameters at sites during collection times Site

Winter (August–September) 2000 Temp. (1C)

Helena Ascot Barrack Street Freshwater Bay

14 15 16 15

2002 Salinity (ppt) 2.6 2.7 4.6 9.2

pH

Temp. (1C)

Salinity (ppt)

pH

7.3 7.3 7.5 8.4

16 18 18 17

2.8 3.9 9.1 21.6

8 7.7 7.9 8.1

Summer (April–May) 2001

Helena Ascot Barrack Street Freshwater Bay Riverton

2002

Temp. (1C)

Salinity (ppt)

pH

Temp. (1C)

Salinity (ppt)

pH

17 18 20 17 18

26.4 29.8 37.1 38.2 35.3

8 7.7 7.9 8.1 7.8

20 23 20 21 17

18.9 24.3 34 31.5 7.6

7.9 8.2 7.6 7.7 7.6

Note: Measurements were taken in the water column at approximately 2 m depth.

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58 Table 2 Collection data for sampling sites Site

Date sampled

Sex

N

Length7SE (cm)

Total weight7SE (g)

Gutted weight7SE (g)

Reproductive stagea

Helena

28 Aug. 2000

Male Female Male Female Male Female Male Female

8 7 11 10 9 9 8 12

30.670.5 30.370.5 30.970.5 30.870.5 30.570.9 31.070.7 27.670.7 28.870.9

550748 524735 606732 584733 576761 563743 458737 536749

488744 467731 556728 532728 522754 506738 393732 479743

4–5 3–4 2 2–3 1 1–3 4–5 3–4

Male Female Male Female Male Female Male Female

2 17 5 15 6 7 4 16

31.970.0 31.370.6 32.971.2 32.270.4 29.370.4 31.670.6 29.670.6 29.970.5

64778 622740 670775 606726 487717 613737 544739 567728

565714 543735 606772 549723 438713 543734 473737 490725

4–5 3–4 2–3 2–3 2 2–3 4–5 3–5

Male Female Male Female Male Female Male Female

7 8 8 12 8 13 10 10

29.770.7 27.971.0 36.971.3 36.270.8 35.070.9 32.870.7 28.870.7 28.470.6

482745 407745 9087107 885752 814763 668736 503739 480733

430739 362741 8437101 805744 741757 604733 421730 415728

4–5 3–4 2 2–3 2 and 5 2–4 5 4

Male Female Male Female Male Female Male Female

5 15 6 14 3 9 7 9

30.370.7 31.370.6 35.571.6 34.971.1 30.870.7 34.671.0 30.270.7 31.370.4

509723 629750 8047101 802799 548740 809780 582754 627724

439717 541742 749794 750794 503738 638793 490743 531722

5 3–4 2 2 2 and 5 2–4 4–5 3–4

Male Female Male Female

8 8 5 16

32.17.06 32.97.07 34.872.0 34.870.7

570731 617727 8087139 839756

514731 552726 7357128 764752

1–2 2 2–5 2–3

8 May 2001 15 May 2002 18 Sept. 2002 Ascot

30 Aug. 2000 9 Apr. 2001 5 May 2002 10 Sept. 2002

Barrack Street

15 Aug. 2000 1 Apr. 2001 28 Apr. 2002 25 Sept. 2002

Freshwater Bay

21 Sept. 2000 29 Apr. 2001 8 May 2002 30 Sept. 2002

Riverton

30 Apr. 2001 12 May 2002

a

According to Nikolskii (1969).

2002, 31 male, 54 female fish. Based on length data (Table 2), the age of all black bream collected was estimated to be 43 years (Sarre and Potter, 1999). 3.1. Condition factor Winter: Pooled data for each sampling period revealed sex-related differences in CF during winter 2000 with male CF significantly lower (P ¼ 0:01) than female CF. In winter 2002, although male CF was again lower than female CF, this was only marginally different (P ¼ 0:05). The CF of male black bream was significantly lower in winter 2000 than in winter 2001 at Helena (P ¼ 0:003), Barrack Street (P ¼ 0:03), and Freshwater Bay (P ¼ 0:004) but there was no difference at Ascot (P ¼ 0:75). Similarly, female CF was lower in

winter 2000 than in winter to 2001 at Helena (Po0:001) and Barrack Street (P ¼ 0:004) but not at Ascot (P ¼ 0:10) or Freshwater Bay (P ¼ 0:76) (Table 3). Summer: No sex-related difference was observed in pooled CF in either summer 2001 (P ¼ 0:50) or summer 2002 (P ¼ 0:89). CF was significantly higher in summer 2002 than in summer 2001 at Ascot (P ¼ 0:003), Freshwater Bay (P ¼ 0:02), and Riverton (Po0:001) while Helena (P ¼ 0:01) was lower in 2001 than in 2002. No difference between years was recorded at Barrack Street (P ¼ 0:14) (Table 4). 3.2. Liver somatic index Winter: The LSI of male black bream was significantly lower than the LSI of female fish (Po0:001)

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Table 3 Mean (7SE) male and female physiological indices of black bream collected (winter) August–September 2000 and 2002 from four sites in the Swan–Canning estuary Site

Year

CFa Male

LSIb Female

GSIc

Male

Female

Male

Female

Helena

2000 2002

1.8170.04 1.9870.03

1.8070.06 2.1070.04

1.26 70.08 1.19a70.10

1.67 70.11 1.67a70.10

5.0070.48 8.8771.18

4.01a70.57 3.56a70.36

Ascot

2000 2002

1.90a70.05 1.95a70.09

1.91a70.03 1.98a70.03

1.31b70.14 1.07b70.29

1.9170.08 1.6170.06

5.72a71.23 8.35a71.25

4.0370.23 6.1170.69

Barrack Street

2000 2002

1.7370.04 1.8770.04

1.7870.03 1.9270.03

1.32c70.21 0.93c70.05

1.34b70.09 1.51b70.08

4.2670.47 11.7371.35

2.8770.42 8.1270.47

Freshwater Bay

2000 2002

1.6970.04 1.9270.04

1.88b70.05 1.91b70.05

0.94d70.07 1.38d70.28

2.38c70.13 2.10c70.24

8.61b71.79 9.62b71.60

6.64b70.57 7.91b71.06

a

a

Note: Within columns, values with the same superscript letter have no significant differences within a site (PX0:05). N as indicated in Table 2. a CF ¼ [(BWGW)/TL3]  100. b LSI ¼ (LW/CW)  100. c GSI ¼ (GW/CW)  100.

Table 4 Mean (7SE) male and female physiological indices of black bream collected summer (April–May) 2001 and 2002 from five sites in the Swan–Canning estuary Site

Year

CFa Male

LSIb Female

GSIc

Male

Female

Male

Female

Helena

2001 2002

2.0270.04 1.9470.04

1.9670.04 1.8370.04

1.09 70.07 0.92a70.04

1.24 70.07 1.22a70.11

1.05 70.24 1.60a70.51

1.46a70.18 1.67a70.32

Ascot

2001 2002

1.8270.05 1.9270.05

1.7670.04 1.9170.03

1.23b70.08 1.05b70.13

1.46b70.08 1.62b70.15

2.32b70.77 1.37b70.39

1.51b70.22 1.41b70.14

Barrack Street

2001 2002

1.75a70.06 1.85a70.02

1.82a70.04 1.86a70.03

1.00c70.08 0.94c70.07

1.31c70.06 1.15c70.04

0.3670.04 1.5770.07

1.39c70.19 1.74c70.42

Freshwater Bay

2001 2002

1.7470.03 1.8570.04

1.7670.03 1.8770.06

0.77d70.05 0.67d70.05

0.93d70.06 2.36d71.28

0.67c70.11 1.00c70.31

1.1970.04 3.5971.85

Riverton

2001 2002

1.7070.01 1.8270.02

1.7270.05 1.9270.03

1.19e70.08 1.10e70.09

1.43e70.07 1.22e70.05

1.0270.27 2.1070.41

1.20d70.05 1.71d70.91

a

a

a

Note: Within columns, values with the same superscript letter have no significant differences within a site (PX0:05). N as indicated in Table 2. a CF ¼ [(BWGW)/TL3]  100. b LSI ¼ (LW/CW)  100. c GSI ¼ (GW/CW)  100.

between years with the exception of Barrack Street in 2000 where no difference was identified. There was no significant difference in male LSI between winter 2000 and winter 2002 (P ¼ 0:28). Female LSI was significantly higher in 2000 than in 2002 at Ascot (P ¼ 0:004) but not at Helena (P ¼ 0:96), Barrack Street (P ¼ 0:18), or Freshwater Bay (P ¼ 0:19) (Table 3). Summer: Male LSI was significantly lower than female LSI in summer (Po0:001). No significant differences between the years 2001 and 2002 for either male (P ¼ 0:07) or female (P ¼ 0:64) black bream were found (Table 4).

3.3. Gonadosomatic index Winter: Male GSI was significantly higher than female GSI (P ¼ 0:000) each year. Winter GSI male black bream was lower in 2000 than in 2002 at Helena (P ¼ 0:004) and Barrack Street (P ¼ 0:000) but not at Ascot (P ¼ 0:23) or Freshwater Bay (P ¼ 0:77). Likewise, female black bream had significantly lower GSI in winter 2000 than in winter 2002 at Ascot (P ¼ 0:006) and Barrack Street (P ¼ 0:000); however; no difference was detected at Helena (P ¼ 0:54) or Freshwater Bay (P ¼ 0:34) (Table 3).

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60

Summer: Significant sexual differences were detected in the GSI in both years (Po0:001), which had an erratic pattern (Table 4). Summer 2002 was higher than summer 2001 in male GSI at Barrack Street (P ¼ 0:01) and Riverton (P ¼ 0:04) but no difference was noted at Helena (P ¼ 0:24), Ascot (P ¼ 0:27), or Freshwater Bay (P ¼ 0:24). Although no statistical difference was detected at Ascot, the trend in male GSI at the other sites was reversed. Female black bream GSI was higher in 2002 than in 2001 at Freshwater Bay (P ¼ 0:03) and Riverton (P ¼ 0:046), with the other sites (Helena (P ¼ 0:66), Ascot (P ¼ 0:93), and Barrack Street (P ¼ 0:59)) recording no differences (Table 4).

Bay (Po0:001). No annual difference in female EROD activity was found at Ascot (P ¼ 0:06) (Fig. 2). Summer: EROD activity was significantly higher in summer 2002 than in summer 2001 at Riverton (P ¼ 0:01) in male black bream and at Helena (P ¼ 0:03) and Riverton (P ¼ 0:000) in female black bream (Fig. 2). Male EROD activity did not differ between years at Helena (P ¼ 0:55), Ascot (P ¼ 0:23), Barrack Street (P ¼ 0:61) or Freshwater Bay (P ¼ 0:71). EROD activity in the female black bream did not differ between years at Ascot (P ¼ 0:10), Barrack Street (P ¼ 0:95), or Freshwater Bay (P ¼ 0:99) (Fig. 2).

3.4. Ethoxyresorufin-O-deethylase activity

3.5. Ethoxycoumarin-O-deethylase activity

Male EROD activity was significantly higher than female EROD activity in the black bream in both winter (Po0:001) and summer (P ¼ 0:001) samplings at all sites with the exception of Helena in winter where no difference was detected (P ¼ 0:71). Winter: No significant interannual differences between 2000 and 2002 were found in male black bream EROD activity at Barrack Street (P ¼ 0:36); however, 2002 levels were higher than 2000 levels at Helena (P ¼ 0:01), Ascot (P ¼ 0:01), and Freshwater Bay (Po0:001) (Fig. 2). Female black bream EROD activity was significantly higher in 2002 than in 2000 at Helena (P ¼ 0:01), Barrack Street (P ¼ 0:001), and Freshwater

No significant difference between males and females in ECOD activity was found (winter, P ¼ 0:17; summer, P ¼ 0:96) in any year, so the results for both sexes were pooled. Winter: Significant annual differences in ECOD activity were detected at all sites (Helena, Ascot, Freshwater Bay (Po0:001), and Barrack Street (P ¼ 0:02)) (Fig. 3). Summer: 2002 ECOD activity was higher than ECOD activity 2001 at Helena (Po0:001), Barrack Street (P ¼ 0:03), and Riverton (Po0:001). Ascot (P ¼ 0:85) and Freshwater Bay (P ¼ 0:06) showed no significant annual differences (Fig. 3).

2000 2002

7

3

8

5

n

8

rto

ck

Ri

sh

ve

8

Fr e

Fr e

sh

Ba

6

55

(d)

12 13

13

9

*6.9 8

16

n

8.4 7.5

7

to er

en el H

hw es Fr

15

a

er at

ck Ba

rra

t sc o A

en

9

Ri v

10

er

9

at

11

*4.8 15

12.2 10.9

hw

*4.8 8 10

*17.9 9.6

es

17

0

el H

17

*15.0 13.3 *10.3

Fr

12

22

ck

0

7

*11.3

ot

*4.7

*12.6

11.0

sc

11

8.1

A

*12.8

33

rra

22

44

Ba

33

EROD activity

44

a

EROD activity

5

(c)

55

(b)

9

rra

rra

sc A

(a)

11

0

Ba

ck

ot

a en el H

7

*12.6

11.3

ot

10

sc

7

A

4

el

2

H

8

16.3 13.7

14.7 11.9

11

*5.4 5

er

8 0

12.2 13.0

at

11

16.6

22

er

*9.2

at

*16.7

*11.4

33

a

22

16.1

w

22.6 *21.2

*37.4

44

en

*26.2

33

EROD activity

44

2001 2002

55

w

EROD activity

55

Fig. 2. EROD activity (mean7SE) in pmol R mg Pr1 min1 of black bream collected in the Swan–Canning estuary: (a) winter, male; (b) winter, female; (c) summer, male; and (d) summer, female. Significant differences between years are represented by *. Numbers within the bars represent number of fish.

ARTICLE IN PRESS D. Webb et al. / Ecotoxicology and Environmental Safety 62 (2005) 53–65

2000 2002

8

70

*6.60

s-SDH activity

*5.03

54.6

*3.88 4

*2.94

*1.09

20

19

15

20

20

15

16

el

en

a

Fr es hw at er

Ba rra ck

el

en

a

A sc ot

H 70 60

*5.48 3.88

*3.86 *3.00

s-SDH activity

6 3.65

3.32

*2.85 2.12

20

60.7

*6.41

4

19

15

20

(a)

2001 2002

8

20

*2.28

2

20

16

0

0

H

20 10

20

(a)

30

Fr es hw at er

15

40

Ba rra ck

2

50.8

50

A sc ot

ECOD activity

53.2 57.4

48.1

*5.30

*0.96

ECOD activity

2000 2002

60.6 54.4 55.5

60

*6.10 6

61

55.5

55.3

2001 2002 57.1 50.2

48.2

50

49.8

46.0

51.2 51.1

40 30 20 10 21 18

16 21

1

20

21

20 12

16 21

Ri ve rto n

Fr es hw at er

Ba rra ck

A sc ot

a en el H

(b) 1

20 13

0

Ri ve rto n

21 12

Fr es hw at er

(b)

20 21

Ba rra ck

H

el

en

a

0

20 13

A sc ot

20 18

Fig. 3. ECOD activity (mean7SE) in pmol H mg Pr min of black bream collected in the Swan–Canning estuary: (a) winter 2000 and 2002; (b) summer 2001 and 2002. Significant differences between years are represented by *. Numbers within the bars represent number of fish.

Fig. 4. SDH activity (mean7SE) in mU in the serum of black bream collected in the Swan–Canning estuary: (a) winter 2000 and 2002; (b) summer 2001 and 2002. Significant differences between years are represented by *. Numbers within the bars represent number of fish.

No significant correlation between ECOD activity and EROD activity exists (r2 ¼ 0:004 (winter) and 0.17 (summer)).

4. Discussion

3.6. Serum sorbitol dehydrogenase No significant differences between sexes were found so the results were pooled. With all data pooled, no significant differences were found in s-SDH activity between years in either winter or summer samplings (P40.05) (Fig. 4). 3.7. Bile metabolites No significant differences between males and females were found (P40:05); therefore, results were pooled to include both sexes. The ratio FF290/335:FF380/430 was significantly higher in winter 2002 than in winter 2000 at all sites (Po0:001) (Fig. 5). In the summer sampling, 2002 was significantly higher than 2001 at Barrack Street (P ¼ 0:02) but not at the remaining sites (Helena (P ¼ 0:42), Ascot (P ¼ 0:05), Freshwater Bay (P ¼ 0:05), and Riverton (P ¼ 0:12)) (Fig. 5).

The region experiences a Mediterranean-type climate with cold wet winters and hot dry summers. The estuary is extremely seasonal (Twomey and John, 2001). During the wet winter months freshwater discharge dominates any tidal influence on the movement of marine waters throughout the estuary, while during the dry summer months freshwater flow ceases and marine conditions prevail. Estuarine conditions can extend inland up to 60 km from the mouth of the Swan River and up to the Kent Street Weir on the Canning River (Stephens and Imberger, 1996) (Fig. 1). During winter, the lower estuary is well mixed with intruding marine water driven by tides and barometric pressure (Twomey and John, 2001). Surface temperatures recorded for each year are within expected ranges for the Swan–Canning estuary in both winter and summer (Twomey and John, 2001). Total rainfall in the southwest of Western Australia, between May and October, has fallen significantly since the mid-20th century with the decline in rainfall being associated with a decrease in the intensity and frequency of extreme rainfall events (IOCI, 2002). Low surface

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62

2000 2002

3000

*235

FF290/335:FF380/430

2500 *205

*199

2000

*175

1500 1000 500

*37

*39

*32

15

20

18

20

13

*36 20

19

16

Fr es hw at er

Ba rra ck

H

el

en

a

A sc ot

0

(a)

*275

2500

2001 2002 220

174 146

1500

154

153 128

1000

20 21

20 12

Ba rra ck

Fr es hw at er

el H

16 21

Ri ve rto n

19 13

en

20 18

A sc ot

500 0

(b)

208 *217

212 2000

a

FF290/335:FF380/430

3000

Fig. 5. Ratio naphthalene-type to B(a)P-type biliary metabolites in black bream collected in the Swan–Canning estuary: (a) winter 2000 and 2002; (b) summer 2001 and 2002. Significant differences between years are represented by *. Numbers within the bars represent number of fish.

salinity levels reflect winter freshwater discharge (Spencer, 1956). Average rainfall and river flows are low during the dry summer months of November–April (rainfall p40 mm month1, flow p20 million m3 month1). Historically, winter rains commence in May, peak in late July to early August (X200 mm month1), and then decline throughout September and October. River flow follows the same pattern, peaking in late July at ffi160–180 million m3 month1, with flow from the Swan River ffi15 times greater than that from the Canning River (Swan River Trust, 1998). Salinity and river flow, with the associated transport of sediments and suspended particulate matter, have the potential to affect significantly the bioavailability of contaminants, fish movements, and, consequently, monitoring using fish biomarkers. Salinity readings recorded in this study in winter indicate that freshwater discharge in the lower Swan estuary was greater in 2000 than in 2002. Very dry conditions were experienced in May and June, marking a late start to the 2000 winter wet season; however,

heavy rainfall in July and August (BOM, 2001) resulted in strong river flow by mid-August. A study of the Swan River Trust Weekly Reports for 2002 shows that surface salinity remained around ffi3 ppt in the upper estuary and ffi15–25 ppt in the lower estuary throughout the winter season of 2002 (Swan River Trust, 2003). The higher salinity reading throughout the lower estuary in 2002 was an indication of lower rainfall and reduced river flow in 2002 compared to those in 2000. Salinity of the estuarine waters was higher in summer 2001 than in summer 2002 with both Barrack Street (37.1 ppt) and Freshwater Bay (38.2 ppt) being hypersaline in 2001. Rainfall in summer 2001 was well below average for the period (BOM, 2001). During the 2002 sampling period several severe thunderstorms passed across the southwest Australian coast in May, bringing heavy rain and strong winds (BOM, 2003). From late on the 10 May to early 11 May 2002 a strong frontal system passed through Perth, resulting in localized heavy surface runoff in some areas. This resulted in the lower surface salinity at Riverton (sampled 12 May 2002) in 2002 compared to that in 2001. The CF indicates that, overall, the black bream were in reasonable condition. Annual variations in the CF were significant in black bream for a majority of the sites which could relate to the following: (1) differences in food source/availability due to the low water levels throughout the estuary in 2002 the gut contents of all fish collected in winter 2000 consisted of mussel shells while those in winter 2002 were either empty or contained small quantities of a macrophyte, polychaetes, and shrimp and (2) time differences in collection days of black bream each season (e.g., there was a 5week difference in sampling dates at Barrack Street for winter 2000 and winter 2002). The general trend in the LSI was that the female black bream had higher values than the males with the exception of Barrack Street in winter 2000. In female fish, vitellogenin is synthesized in the liver in response to estradiol stimulation. During vitellogenesis dramatic structural changes occur in the female’s liver, resulting in an increase in liver size (Bun Ng and Idler, 1983). It is likely that the higher LSI in female black bream is due to the influence of estradiol in the females. Male black bream LSI did not differ between years while the only difference for female black bream occurred at Ascot in the winter months with females captured in 2000 having a higher LSI than females captured in 2002, which cannot be explained by sampling date differences and is most likely related to natural interannual variability. There was an increase in the GSI in both male and female black bream over time during both the 2000 and the 2002 winter sampling periods. Black bream gonadal development in the Swan–Canning estuary rises sharply from July–August to a well-defined peak in October and then falls progressively until February of the following

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year (Sarre and Potter, 1999). The increase in GSI is clearly related to sampling dates. For example, in 2000, fish were collected at Barrack Street in mid-August with males at reproductive stages 4–5 and females at stages 3–4 while, in 2002, fish were not collected until late September with all males at milting stage and all females at stage 4. It is likely that the rapid gonadal development of black bream at this time of year has the potential to obscure the detection of contaminant-related reduced reproductive performance in the estuary. Summer GSI ratios were generally higher in female black bream than in male black bream in both 2001 and 2002 with some exceptions. Male GSI was higher at Barrack Street and Riverton in 2002 than in 2001. At both these sites, in 2001, male black bream either had undeveloped gonads (stage 1) or were in early gonad development (stage 2), while in 2002 most males were in early gonad development but each site also had two milking (stage 5) males in the respective samples. One milking male was also found at Freshwater Bay in 2002 and although 2002 male GSI is higher at this site than 2001 male GSI the difference was not significant. Milking males at this time of year are unexpected although Sarre and Potter (1999) reported that 25% of male black bream in the upper estuary had stage 4 gonad development in May and June 1994. Female GSI was also higher in summer 2002 than in summer 2001 at many sites but this was significant only at Freshwater Bay. As with the male black bream, the female black bream had a small number of individuals showing more advanced gonad development than expected. At this time of year black bream gonad development in both sexes is expected to be between stages 1 and 3 but it is believed that an increase in water temperatures may trigger spawning activity (Sarre and Potter, 1999). Water temperatures throughout the Swan–Canning estuary were marginally higher in summer 2002 than in summer 2001 (Table 1). The lack of interannual and intersite differences in s-SDH activity indicates that there is no bias in the MFO activities measured relative to hepatic tissue damage. With the exception of Helena in summer 2002, the EROD activity was higher in males than in females in each year. Inhibition of EROD in female fish due to competition between inducible enzymes and estradiol has been reported in several studies (Stegeman et al., 1982; Hodson et al., 1991; Eggens et al., 1995; Lange et al., 1999). In the present study, maximum induction of EROD observed in males (37.42712.34, n ¼ 5) and females (17.8971.66, n ¼ 16) was at Riverton during summer 2002 following the first heavy rains for the year (Fig. 2). EROD activity was three-fold higher at this site compared to that in summer 2001. Black bream are a mobile fish but schools tend to stay in areas over a period of several days to weeks before moving

63

significant distances (K. Littleton, pers. commun.). The salinity readings at Riverton indicate significant runoff following localized rainfall. Since EROD activity reflects exposure to contaminants within the previous few days (Arcand-Hoy and Metcalfe, 1999; Gagnon and Holdway, 2000), the elevated 2002 EROD activity at Riverton suggests that road and stormwater drainage carried significant EROD-inducing contaminants from surrounding roads, parks, and residential properties. No differences in EROD activities in male black bream were noted at any other site in summer 2001 and summer 2002. However, female EROD activity was significantly higher in summer 2002 than in summer 2001 at both Helena and Riverton. During the winter sampling most sites recorded significant differences, with 2002 higher than 2000. This suggests that lower winter rainfall in 2002 compared to that in 2000 (i.e., reduced river flow) resulted in contaminants from road runoff and stormwater drainage remaining within the estuary for a longer period in 2002. Since dietary factors can affect EROD induction (Goksøyr and Fo¨rlin, 1992), the higher EROD activity in winter 2002 compared to that in winter 2000 could also be the result of different food source availability to the black bream due to the predominant ebb tide conditions throughout the estuary in winter 2002. In 2002, the gut contents suggest a diet high in polychaetes from the deeper channels in the estuary. No upstream or downstream gradient in EROD activity was observed in any year. Arukwe and Goksøyr (1997) and Lange et al. (1999) recorded differences between sexes in ECOD activity in turbot (Scophthalmus maximus) and dab (Limanda limanda), respectively. However, there was no sexual difference in the ECOD activity of the black bream in any year during this study, which agrees with the findings of Holdway et al. (1994) in sand flathead (Platycephalus bassensis) and Machala et al. (1997) in carp (Cyprinus carpio). The lack of correlation between EROD and ECOD activities supports the hypothesis that ECOD activity represents a cytochrome P450 pattern different from that in EROD activity (Machala et al., 1997; Stegeman et al., 1997). Lange et al. (1999) demonstrated that each activity exhibits species- and xenobiotic-specific differences indicating that EROD and ECOD activities are catalyzed by more than one protein. In addition, it has been suggested that diesels, oils, isoafrol, and piperonylbutoxide, compared with b-naphthoflavone and 3methylcholanthrene, can differentially induce EROD and ECOD activities (Goksøyr and Fo¨rlin, 1992). Contrary to EROD activity, a gradient of ECOD activity was observed within the Swan River arm of the estuary in winter; however, the direction was not the same in each year. Winter 2000 ECOD activity was highest at the upstream site of Helena (5.307 0.68, n ¼ 18) and gradually reduced downstream to

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D. Webb et al. / Ecotoxicology and Environmental Safety 62 (2005) 53–65

Freshwater Bay (2.9470.36, n ¼ 20). The opposite trend was observed in winter 2002 with Helena (0.9670.16, n ¼ 20) having the lowest ECOD activity which increased downstream to Freshwater Bay (6.6070.79, n ¼ 16) (Fig. 3). A gradient in ECOD activity was not observed in summer 2001; however, in summer 2002, a trend similar to that of winter 2000, declining from a high at Helena (5.4870.66, n ¼ 18) to a low at Freshwater Bay (3.3270.52 n ¼ 12) was apparent (Fig. 3). The inconsistency in the gradient of ECOD activity in different years reflects the nature of runoff into the river from different areas, the patchy nature of rainfall events, and, as a result, the different patterns of input of ECOD-inducing chemicals into the estuary. Consequently, data from any single year cannot be used to make assumptions about inputs or trends of input of ECOD-inducing chemicals into the Swan–Canning estuary. Comparison of the ratio of FF290/335 (naphthalenetype) to FF380/430 (B(a)P-type) can determine the source of PAH inputs into an estuary (Neff, 1990; Aas et al., 2000). A high ratio shows a dominance of two- and three-ring compounds over five-ring compounds, indicating exposure to naphthalene-rich petroleum products (petrogenic) such as unburnt fuels, while lower ratios indicate exposure to combusted petroleum compounds of pyrolytic origin (burnt fuels). The ratios in winter 2000 in black bream from the Swan–Canning estuary were in the range of 323–396 and were proportionally similar throughout the estuary, indicating the dominance of PAHs originating from burnt fuels (e.g., motor vehicles) from winter runoff into the estuary. Higher ratios were detected in summer 2001 (1539–2204), summer 2002 (1281–2759), and winter 2002 (1752–2356), indicating that naphthalene-type compounds of petrogenic sources from unburnt fuels (outboard motors, fuel spills, or leakage from industry) were more dominant in the estuary compared to those in winter 2000. The difference in winter ratios in 2000 and 2002 reflects the reduced rainfall, runoff, and river flow in winter 2002. PAHs with a low molecular weight (i.e., naphthalene) tend to dissipate faster in ecosystems than the larger PAHs (Neff, 1990). No downstream enrichment in large PAH compounds in the river exists, suggesting that there are multiple sources of inputs along the estuary. The results show that there is high interannual variability in biomarker response in the black bream in the Swan–Canning estuary with no consistent pattern of exposure or upstream/downstream trends. There was a bias toward female black bream being caught, particularly in the downstream sites in each year. Suppression of EROD activity in the female black bream makes interpretation of this biomarker difficult. Although the suppression of ECOD activity in female black bream does not occur, the lack of correlation with

EROD activity suggests that the ECOD activity is not a suitable substitute for EROD for determination of PAH exposure. Of the biomarkers measured, biliary metabolites are the most suitable biomarker of exposure to PAH compounds. ECOD activity has the potential to be a suitable biomarker for exposure for a mixture of contaminants; however, baseline levels have not been determined in black bream and no reference site exists within the estuary. Webb and Gagnon (2002b) suggested that biomarker response in the lower estuary was moderated by tidal influences during the winter months but this is not apparent in the winter of 2002. It is clear that assumptions from single year cannot be used to predict conditions within the estuary in subsequent years. The level of runoff into the estuary is determined by the onset and finish of winter rains and the number of significant rainfall events. The strength of freshwater outflow influences the level of input of potential contaminants into the estuary and their residence time within the system. The proximity of major roads and drains into the Swan–Canning estuary is the most significant factor influencing biomarker response in the black bream; however, yearly rainfall patterns will have a strong influence on the pattern and strength of that response.

Acknowledgments This study was supported by a Curtin University Research Grant and ARC SPIRT funding to M.M.G. and by the Water and Rivers Commission, Perth, Western Australia. The authors extend special thanks to Mr. Kerry Littleton, commercial fisherman, for his assistance in fish collection. The treatment of animals was in accordance with Curtin University Animal Experimentation Ethics N-25-00.

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