contaminant pathways in port curtis: final report · contaminant pathways in port curtis: final...

Contaminant pathways in Port Curtis: Final report

Simon Apte, Leonie Andersen, John Andrewartha, Brad Angel, Damon Shearer, Stuart Simpson, Jenny Stauber and Vicky Vicente-Beckett

May 2006

Contaminant pathways in Port Curtis: Final report Copyright © 2006: Cooperative Research Centre for Coastal Zone, Estuary and Waterway Management Written by:

Simon Apte Leonie Andersen John Andrewartha Brad Angel Damon Shearer Stuart Simpson Jenny Stauber Vicky Vicente-Beckett Published by the Cooperative Research Centre for Coastal Zone, Estuary and Waterway Management (Coastal CRC)

Indooroopilly Sciences Centre 80 Meiers Road Indooroopilly Qld 4068 Australia

www.coastal.crc.org.au

The text of this publication may be copied and distributed for research and educational purposes with proper acknowledgment. Photos cannot be reproduced without permission of the copyright holder. Disclaimer: The information in this report was current at the time of publication. While the report was prepared with care by the authors, the Coastal CRC and its partner organisations accept no liability for any matters arising from its contents.

National Library of Australia Cataloguing-in-Publication data Contaminant pathways in Port Curtis: Final report QNRM06215 ISBN 1 921017 30 9 (print) ISBN 1 921017 31 7 (online)

Contaminant pathways in Port Curtis: Final report Simon Apte1, Leonie Andersen2, John Andrewartha3, Brad Angel4, Damon Shearer2, Stuart Simpson1, Jenny Stauber1 and Vicky Vicente-Beckett5 1 CSIRO Energy Technology 2 Central Queensland University, Gladstone Campus 3 CSIRO Marine 4 University of Wollongong 5 Central Queensland University, Rockhampton Campus The report should be cited as: Apte, SC, Andersen, LE, Andrewartha, JR, Angel, BM, Shearer, D, Simpson, SL., Stauber, JL & Vicente-Beckett, V (2006) Contaminant pathways in Port Curtis: final report. CRC for Coastal Zone, Estuary and Waterway Management, Brisbane.

iv

Acknowledgments

The Contaminant Pathways team thank the following people for their contributions:

Helen Morrison (Central Queensland University), Niels Munksgaard (Charles Darwin University) and Gary Hancock (CSIRO Land and Water) for assistance with the sediment studies. Merrin Adams, Leigh Hales, Ian White (CSIRO Energy Technology) and Dianne Jolley (University of Wollongong) for their assistance with trace metals analysis and the biological pulses studies. Andrew Davis (Central Queensland University) for his excellent field work support. Andrew Revill (CSIRO Marine), Andrew Storey (University of Western Australia), Karen Boundy, Jill Campbell Larelle Fabbro, Lee Hackney, Felicity Melville, Clayton Plummer, Kirsty Small and Rebecca Hendry (all Central Queensland University), Scott Wilson (Australian Catholic University), William Siu, Eric Ching, C.T. Kwok and Paul Lam (City University, Hong Kong) for their contributions to the biological monitoring studies. John Parslow and Mike Herzfeld (CSIRO Marine) for their contributions to the hydrodynamic model evaluation.

The authors also thank the Australian Institute of Nuclear Science and Engineering for providing financial assistance (Award No. AINGRA05082) to enable work to obtain electron microscope images of algal cells used in this study. Finally we thank Maria Vandergragt of the Coastal Zone CRC for her admirable leadership and strong support over the last three years.

Contaminant pathways in Port Curtis: Final report Executive summary

v

Executive summary

Background The Port Curtis Estuary has a well-developed and expanding industry within its

catchment. It is also one of Australia�s leading ports and is located adjacent to

the World Heritage-listed Great Barrier Reef Marine Park. As a consequence of

increasing population and industrial activities, the Port Curtis Estuary is expected

to receive increasing quantities of contaminant inputs from diffuse sources

(e.g. urban runoff) and point source discharges (e.g. industrial effluents).

Sources of chemical stressors are many, and multiple contaminants are likely

to be transported to the estuary by air and/or water. The challenge for coastal

management within the region is the long-term sustainable management of

further port and industrial development, related population growth, and the

management of potentially significant impacts on coastal resources.

The release, fate and impacts of contaminants generated within the region by

industrial and urban activities are issues of obvious concern. When the

Cooperative Research Centre for Coastal Zone, Estuary and Waterway

Management (Coastal CRC) first started its activities in Port Curtis in 1998, there

were few published studies describing contaminant distributions in Port Curtis.

During the first phase of its activities, the CRC undertook the Port Curtis screening

level risk assessment (SLRA) (Apte et al. 2005) which employed a rigorous, risk-

based approach to identify and prioritise contaminant issues of potential concern.

While there were no issues of regulatory concern, the SLRA identified some

contaminant-related issues worthy of further investigation which included tributyltin

(TBT) in waters, the anomalous bioaccumulation of metals by biota from Port

Curtis and slightly elevated concentrations of arsenic, TBT and naphthalene in

sediments. Recommendations were made for future investigations. A separate

CRC project developed a pilot-scale hydrodynamic model of Port Curtis which

enabled water movement to be predicted. The model has clear applications to the

prediction of contaminant movement, especially point source discharges

associated with industrial activities. Contaminant Pathways in Port Curtis was part

of Phase 2 of the CRC�s activities in Port Curtis and focused on some of the key

issues that were identified in the SLRA.


vi

Objectives Following extensive scoping activities which included discussions with local

stakeholders and CRC members, a number of specific goals for the Port Curtis

Contaminant Pathways project were developed. The goals comprised a mixture of

scientific investigations with direct linkages to Port Curtis and some frontier

research activities which would advance the ability to assess contaminant impacts.

These goals were as follows:

(i) Identification of the sources of dissolved and particulate contaminants

to Port Curtis

(ii) Further development of the hydrodynamic model and its application to

contaminant management

(iii) Development and trial of sublethal biomarker tests for assessing

organism health in Port Curtis (including an investigation on the

biological effects of TBT)

(iv) Characterisation of metal bioaccumulation pathways affecting key

organisms residing in Port Curtis

(v) Examination of the effect of contaminant pulse events on biological

responses. It was recognised that pulse exposure to contaminants is

probably a more realistic scenario for Port Curtis than steady-state,

continuous exposure.

The Contaminant Pathways study commenced in June 2003 and finished in

October 2005. This report summarises the outputs of the study.

Conclusions and recommendations

Water quality

The Contaminant Pathways study has produced the first accurate data on

dissolved trace metal concentrations in the coastal waters of Central Queensland

and in close proximity to the Great Barrier Reef. In the offshore coastal waters,

dissolved metal concentrations were extremely low and were comparable to those

measured at open Pacific Ocean and New South Wales coastal water locations.

Intensive surveying of Port Curtis has confirmed the presence of elevated metal

concentrations within the harbour. The Narrows region was found to have the

highest concentrations of dissolved copper and nickel and this could be attributed

to natural geological sources. The Fitzroy River is a source of dissolved metals to

the local coastal region. In particular, the Fitzroy River contains elevated dissolved

nickel concentrations. Under some flow conditions, the Fitzroy River plume may

enter The Narrows region and supply dissolved metals to Port Curtis. There were


vii

no conspicuous sources of trace metals within Port Curtis. The trace metal

distributions in Port Curtis are likely to reflect a subtle mixture of metal inputs

including industrial and other anthropogenic discharges, inputs from unidentified

sources in The Narrows and the Fitzroy River plume. Survey measurements

showed that trace metal inputs to Port Curtis which contribute to the observed

dissolved metal concentrations are most likely to be delivered in solution form and

not by release of metals from particulates.

Sediment quality

Using multiple lines of evidence, it was shown that the concentrations of

particulate arsenic, chromium and nickel in the benthic sediments of Port Curtis

are elevated because of the local geology and not because of metal contamination

from anthropogenic sources. This important factor needs to be taken into account

when applying the ANZECC/ARMCANZ (2000) sediment quality assessment

framework to this region. Polycyclic aromatic hydrocarbon (PAH) contaminants in

sediments were highest around the industrial area of Gladstone; however

concentrations at all locations were below ANZECC trigger values.

Several types of PAHs characteristic of combustion sources were detected at the

middle harbour largely at the Clinton Coal Facility, along Calliope River and South

Trees Inlet�Boyne River, but again concentrations were considered relatively low.

Relatively high proportions of the naturally-occurring PAH perylene were found in

sediments from The Narrows and Munduran Creek. At least the top 28 cm of

sediments at intertidal and subtidal sites were estimated to have been deposited

since 1958 in Port Curtis, which is roughly the start of the industrialisation of

Gladstone. The rate of sediment deposition was at least 0.6 cm/y. The sediment

depositional zones identified were: the northern Narrows, lower Calliope River and

South Trees Inlet�Boyne River areas and therefore these areas are potentially

sinks for metal deposition.

Hydrodynamic modelling

Some field data problems were encountered which did not allow a full evaluation

of model performance. The comparison of modelled and field data for a modest

flood event did, however, show that the model under-predicted salinity. This was

most likely due to inputs of fresh water occurring during the flood event that were

not included in the model (e.g. freshwater flow from the Fitzroy via The Narrows).

Nevertheless, there are grounds to be optimistic that the model represents tracer

transport reasonably well. The transport regime in the estuary is predominantly

tidally driven, and the distribution of passive tracer will reflect this dominant

forcing. The model reproduced tidal elevation satisfactorily.


viii

Sublethal indicators of contaminant exposure

Imposex was detected in mulberry whelk specimens collected from Port Curtis,

confirming a sublethal, biological response to TBT exposure. Although related to

local shipping intensity, the frequency and grade of the imposex condition were

not severe in comparison to surveys of other ports in Australia and overseas.

Globally, the condition is likely to slowly improve with the introduction of further

restrictions on the use of TBT in 2008.

The concentrations of stress biomarkers (glutathione, glutathione-s-transferase,

catalase and lipid peroxidase) in field-deployed oysters were quite variable, and

few consistent trends were observed that could be related to contaminant

exposure. No firm conclusions could be drawn regarding the suitability of these

biomarkers for biomonitoring in Port Curtis.

Contaminant foodweb dynamics

A food web including mud crabs, other crustaceans, fish, molluscs and a variety of

plants was characterised in Port Curtis. In general, the food web was not unlike

those established for other estuarine embayments. It appears that very few

species rely on mangroves as a predominant food source but are more dependent

on benthic organic matter and algae. Mud crabs were identified as one of the

dominant predators in the food chain. Carbon isotope measurements suggested

that prawns were feeding either directly or indirectly on a blue green algal bloom

(Lyngbya majuscula) and this finding was supported by observations of pigment

from the algae being visually evident in the prawns.

The finding may have consequences for consumers should the toxin produced by

the algae follow similar uptake pathways to the pigment and accumulate in the

prawn muscle tissue. Although there were very few significant between-site

differences in metal bioaccumulation, organisms from inner harbour sites tended

to be more enriched in metals than those from the reference site outside the

harbour. The findings of this study indicate that for the majority of organisms the

uptake of metals through food pathways is likely to be complex and integrated,

particularly for those in higher trophic positions and those that have the ability to

regulate metal accumulation.

Pulse exposure to contaminants

Contaminant pulse studies were conducted in the laboratory using the marine alga

Phaeodactylum tricornutum as the model organism and copper as the model


ix

contaminant. These studies suggest that, at least for microalgae over short

exposure times (up to 72 h), bioaccumulation of copper from pulse exposure is no

greater than bioaccumulation from continuous exposure. Copper bioaccumulation

measurements indicated that P. tricornutum did not have an effective mechanism

for eliminating copper from cells, rather the intracellular copper decreased as a

result of dilution by cell division. If predictive models were developed for key

organisms in Port Curtis, this would allow better assessment of pulse exposure.

This approach is currently the best practicable approach to solving this complex

problem.

Future directions After six years of activity, the Coastal CRC has left a lasting legacy in Port Curtis.

There is an increased awareness amongst stakeholders of contaminant issues

based on good quality data. The CRC study was the first to adopt a whole-of-port

approach to understanding contaminants in Port Curtis. With a few exceptions, the

majority of previous research had either not focussed on contaminants or their

effects, or had been limited to studies of particular receiving environments.

Specific project outputs, including reports, press releases and research papers,

are listed in the appendix. A considerable database of accurate contaminant

distributions is now available for utilisation by local industry, researchers and

regulators alike. The �report card� for contaminants in Port Curtis is generally quite

good, although a recent oil spill event illustrates the sensitivity of the ecosystem

and the need for reliable baseline information and strong environmental

management.

In the future, is it envisaged that the Port Curtis Integrated Monitoring Program

(PCIMP) and the Centre for Environmental Management (CEM) at the Gladstone

campus of Central Queensland University (CQU) will carry on the legacy of the

CRC. PCIMP is a consortium of members from 14 bodies representing industry,

government (both local and state), research institutions and other stakeholders to

develop a cooperative, integrated program for monitoring the ecosystem health of

Port Curtis. A strong, long-term, annual monitoring program, building on the initial

groundwork established by the CRC, is being formulated by PCIMP members in

consultation with local stakeholders. Research will focus on water and sediment

quality�particularly bioavailable contaminants�and on mangrove ecosystems.

Results will be presented to the community in the form of an Ecosystem Health

Report Card for Port Curtis.

Based on the CRC studies of the last six years, we suggest the following

directions for future contaminants management and research:


x

(i) Risk-based management

The SLRA illustrated the utility of using a risk-based approach to contaminant

management. Owing to limited resources and more time-critical research priorities,

the CRC was not able to assess the risks posed by many organic contaminants.

We recommend that this issue now be covered, but with a staged approach. The

CRC research has shown the value of using contaminant bioaccumulation as an

indicator of ecosystem health. A first stage would therefore be the measurement of

organic contaminants in indicator organisms in Port Curtis. Further investigations

may be necessary if bioaccumulated organic contaminants prove to be significant.

Mercury in piscivorous fish such as barramundi can be elevated owing to food

chain biomagnification. This is an issue of regional importance and should not be

forgotten given the large recreational and commercial fishing industries present in

Port Curtis and surrounding regions. The characterisation of mercury

bioaccumulation and biomagnification through food webs over the coastal region

of the whole Central Queensland region is appropriate.

(ii) Improved monitoring

The Contaminant Pathways study and the SLRA have shown the value of �good

quality data�. It is recommended that future monitoring adopt and enforce rigorous

quality assurance protocols to ensure quality is maintained.

As noted earlier, contaminant concentrations may fluctuate over various time

scales in Port Curtis. Such variations are not easily discerned with a discrete

sampling approach. Time-integrated monitoring such as biomonitoring using

deployed organisms (e.g. oysters) and chemical surrogates such as diffusive

gradients in thin films (DGT) for metals and solvent-filled dialysis cells for organics

(passive samplers) is therefore recommended. Seasonal monitoring may also be

considered to determine if there are any seasonal fluctuations to contaminant

loads in Port Curtis.

(iii) Ability to predict contaminant concentrations and effects

A key tool in sustainable management is the ability to predict impacts. We

recommend that the further development of the hydrodynamic model in

conjunction with contaminant data and a subsequent suite of predictive models is

pursued. The model should be expanded to include The Narrows region and some

of the major estuaries in Port Curtis to assist in understanding the contribution of

metal load from point sources, and flows from the Fitzroy River to Port Curtis.

Predictive models of metal bioaccumulation and biological impact (or establishing

if environmental harm has occurred) are also worthy of consideration. The effect of


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pulse versus continuous discharges on bioaccumulation should also be pursued.

Information could result in management changes for a more favourable controlled

release of contaminants at point source discharges.

(iv) Future concerns

It is hard to predict the priorities for contaminant management over the next

decade. It is necessary to keep a watching brief on emerging issues. Unforseen

events such as oil spills and resulting PAH contamination will always be a threat in

a busy commercial port, and stringent protocols should be in place to manage and

subsequently assess the impacts of such events. Ross (2002) recently reported a

survey of acid sulfate soils (ASS) in the Central Queensland coast and found high

occurrence of these soils on the coastal plain along the Curtis and Capricorn

coasts, Shoalwater Bay and Broadsound. When exposed to air, sulfides are

oxidised, producing sulfuric acid and can also release iron, aluminium, and other

heavy metals. This is an issue of potential concern in Port Curtis especially in

areas where development results in the aerial exposure of sulfide-containing

sediments.

The reduced flushing of the estuary highlighted by the hydrodynamic model also

brings into question the resilience of the estuary in terms of its ability to cope with

increased contaminant loads from new industries or current industry expansions.

A number of new industries are proposed in the near future including a nickel

refinery and aluminium smelter, in addition to the expansion of already existing

industries. Bioaccumulation of metals in the inner harbour area has already been

demonstrated indicating that the harbour has potentially a limited threshold for

contaminant loads. Ecosystem health should continue to be monitored over the

long term to ensure the threshold is not exceeded resulting in a decline in the

current state of the harbour.


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Table of contents Acknowledgments .......................................................................................................................iv

Executive summary................................................................................................................... v Background........................................................................................................................... v Objectives .............................................................................................................................vi Conclusions and recommendations......................................................................................vi Future directions ...................................................................................................................ix

Glossary of terms, acronyms and abbreviations ..................................................................... xviii

Chapter 1. Introduction............................................................................................................. 2 1.1. Background .......................................................................................................................... 2 1.2 Summary of the screening level risk assessment................................................................. 3 1.3 Study objectives .................................................................................................................... 6 1.4 References ............................................................................................................................ 7

Chapter 2. Concentrations and sources of dissolved trace metals in Port Curtis and surrounding coastal waters ..................................................................................................... 8 2.1 Introduction............................................................................................................................ 8 2.2 Experimental ......................................................................................................................... 9 2.3 Results and discussion........................................................................................................ 10

2.3.1 Dissolved metal concentrations ................................................................................... 10 2.3.2 Temporal variations in metal concentrations ............................................................... 11 2.3.3 Particulate metal concentrations in suspended solids................................................. 11 2.3.4 Salinity and pH............................................................................................................. 16 2.3.5 Sources of dissolved metals ........................................................................................ 18

2.4 Conclusions......................................................................................................................... 18 2.5 References .......................................................................................................................... 20

Chapter 3. Metal and polycyclic aromatic hydrocarbon contaminants in benthic sediments of Port Curtis......................................................................................................... 22 3.1 Introduction.......................................................................................................................... 22 3.2 Experimental ....................................................................................................................... 23 3.3 Results and discussion........................................................................................................ 25

3.3.1 Metals in surficial sediments and sediment cores ....................................................... 25 3.3.2 Estimates of background metal concentrations in sediments...................................... 26 3.3.3 Sediment geochronology ............................................................................................. 27 3.3.4 Stable lead isotope ratios (PbIRs) ............................................................................... 29 3.3.5 PAHs in sediments....................................................................................................... 31

3.4 Conclusions......................................................................................................................... 37 3.5 References .......................................................................................................................... 38 Appendix 3.1. Sediment samples (2003�2005) ........................................................................ 41

Chapter 4. Port Curtis hydrodynamic model evaluation ..................................................... 44 4.1 Background ......................................................................................................................... 44 4.2 Field program ...................................................................................................................... 45 4.3 Model description and development ................................................................................... 47 4.4 Model forcing....................................................................................................................... 48 4.5 Model trials .......................................................................................................................... 48 4.6 Evaluation results ................................................................................................................ 49 4.7 Conclusions......................................................................................................................... 52 4.8 References .......................................................................................................................... 53

Chapter 5. Metal bioaccumulation through foodweb pathways in Port Curtis ................. 54 5.1 Introduction.......................................................................................................................... 54 5.2 Experimental ....................................................................................................................... 55 5.3 Results and discussion........................................................................................................ 58

5.3.1 Foodweb elucidation .................................................................................................... 58 5.3.2 Metal distributions and relation to food web structure ................................................. 60



xiv

Chapter 6. Occurrence of imposex in Port Curtis................................................................ 68 6.1 Introduction.......................................................................................................................... 68 6.2 Imposex in marine gastropods ............................................................................................ 68 6.3 Experimental ....................................................................................................................... 69 6.4 Results and discussion........................................................................................................ 71 6.5 References .......................................................................................................................... 76

Chapter 7. Antioxidant enzymes as biomarkers of environmental stress in oysters in Port Curtis ................................................................................................................................ 80 7.1 Introduction.......................................................................................................................... 80 7.2 Experimental ....................................................................................................................... 81 7.3 Results and discussion........................................................................................................ 84

7.3.1 Oyster metal concentrations ........................................................................................ 84 7.3.2 Oyster biomarker concentrations................................................................................. 86 7.3.3 Laboratory bioasssay................................................................................................... 89


Chapter 8. Effect of pulse events on biological responses to contaminants ................... 96 8.1 Background ......................................................................................................................... 96 8.2 Experimental ....................................................................................................................... 97

8.2.1 Chemical analysis ........................................................................................................ 97 8.2.2 Algal bioassay procedure............................................................................................. 97 8.2.3 Pulsed exposures to dissolved copper ........................................................................ 98 8.2.4 Intracellular and extracellular copper determinations .................................................. 99 8.2.5 Modelling bioassay response with fluctuating copper concentrations....................... 100

8.3 Results and discussion...................................................................................................... 101 8.3.1 Continuous exposure ................................................................................................. 101 8.3.2 Pulsed copper exposures .......................................................................................... 101 8.3.3 Copper uptake ........................................................................................................... 102 8.3.4 Copper elimination ..................................................................................................... 104 8.3.5 Modelling effects of pulsed copper exposures on algal growth ................................. 105

8.4 Conclusions....................................................................................................................... 107 8.5 References ........................................................................................................................ 109

Chapter 9. Conclusions and future directions ................................................................... 112 9.1 Conclusions....................................................................................................................... 112 9.2 Future directions................................................................................................................ 114 9.3 References ........................................................................................................................ 117

Appendix. Specific project outputs ..................................................................................... 118


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List of figures 2.1 Port Curtis estuary and surrounding waters showing positions of SLRA

sampling sites during Survey 1 and Survey 2 ......................................................7

2.2 Dissolved copper, nickel and zinc concentrations (ng/L) in Port Curtis estuary and surrounding waters........................................................................................ 13

2.3 Extent of dissolved metal fluctuations with time in The Narrows, 3 km south of Ramsays Crossing (a), and near Fisherman�s Landing (b) ................................. 14

2.4 Water pH and salinity in the Port Curtis region .................................................... 16

3.1 Distribution of sediment with particle size fraction < 60 µm (% mud) .................. 22

3.2 Variation of metal concentrations with depth (Calliope River mouth) .................. 24

3.3 Variation of metal concentrations with depth (Targinnie Creek) .......................... 19

3.4 137Cs activity in sediment cores from Port Curtis ................................................ 27

3.5 210Pb activity in sediment cores from Port Curtis ................................................. 27

3.6 Lead isotope ratios in Port Curtis sediments and other samples......................... 29

3.7 Total PAHs in Port Curtis ..................................................................................... 32

3.8 Naphthalene in benthic sediments ....................................................................... 32

3.9 Benzo[b+k]fluoranthene in benthic sediments ..................................................... 33

3.10 Perylene in benthic sediments ............................................................................. 33

3.11 Depth profile of PAHs in Munduran Creek ........................................................... 34

4.1 Typical output from the MECO model showing dispersion of a conservative tracer released from Fisherman�s Landing........................................................... 43

4.2 Field program sampling sites ............................................................................... 45

4.3 Field program transect sampling sites ................................................................. 45

4.4 Measured sea level compared with modelled sea level for a site near South Trees during one spring neap tidal cycle ............................................................. 47

4.5 Time series comparisons of salinity from the transect measurements and the model for the half-length river............................................................................... 49

4.6 Time series comparisons of salinity from the transect measurements and the model for the short river ....................................................................................... 50

5.1 Location of organism sampling sites in Port Curtis .............................................. 54

5.2 Examples of organisms collected as part of the foodweb study .......................... 56

5.3 Relationship of δ13C and δ15N of selected primary producers and consumers in a Port Curtis food web ...................................................................................... 58

5.4 Blue-green algae (Lyngbya majuscula) demonstrating released pigment and the same pigment observed in the hepatopancreas of a banana prawn from the same site ........................................................................................................

59

5.5 Mean aluminium concentrations in biota at three inner harbour sites in Port Curtis and an outer harbour reference site .......................................................... 61

5.6 Mean arsenic concentrations in biota at three inner harbour sites in Port Curtis and an outer harbour reference site ......................................................... 62


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List of figures (continued) 6.1 The mulberry whelk, Morula marginalba .............................................................. 69

6.2 Collection sites for M. marginalba at ten sites in Port Curtis in relation to the shipping channel ................................................................................................. 70

6.3 Imposex in M. marginalba with penis bud ........................................................... 72

6.4 Imposex frequency in female M. marginalba at ten sites in Port Curtis in 2003 .. 73

6.5 Frequency of imposex in M. marginalba at ten sites in Port Curtis...................... 74

7.1 Location of Sites 1 and 2 for oyster field experiments in Port Curtis Harbour ..... 81

7.2 Individual bags of oysters attached to buoys ready for deployment .................... 82

7.3 Oysters in treatment tanks in copper bioassay .................................................... 82

7.4 Mean concentration of biomarkers in oysters from Site 1 and Site 2................... 87

7.5 Accumulation in copper-exposed oysters from the five treatment concentrations ......................................................................................................

89

7.6 Regression of mean GSH concentration in gills against time .............................. 89

8.1 Copper exposure scenarios tested ...................................................................... 98

8.1 Measured and predicted effect of dissolved copper concentrations on algae cell biomass.......................................................................................................... 100

8.3 Relationships between the exposure time and extra-cellular and intra-cellular copper concentrations in P. tricornutum cells ...................................................... 102

8.4 Copper efflux after placing P. tricornutum cells in clean sea water ..................... 104

8.5 Measured and predicted effect of copper exposure scenarios tested with equivalent copper �dose�, but varying duration and magnitude............................ 105

8.6 Transmission electron microscopy of P. tricornutum ........................................... 105

A.1 Example of a contaminant pathways article published in the Gladstone Observer, 2005..................................................................................................... 119


xvii

List of tables 2.1 Dissolved and particulate metal concentrations: Survey 1................................ 11

2.2 Dissolved and particulate metal concentrations: Survey 2................................ 12

2.3 Concentration of trace metals in waters around the world ................................ 14

3.1 Metal concentrations in benthic sediments from various locations ................... 26

3.2 List of PAHs studied and their abbreviations..................................................... 31

3.3 PAH ratios and origins ....................................................................................... 35

A.3.1 Sediment samples (2003�2005)........................................................................ 40

A.3.2 Particulate metal concentrations and other parameters.................................... 41

A.3.3 Polycyclic aromatic hydrocarbons in Port Curtis ............................................... 42

5.1 Organisms collected for the foodweb study....................................................... 55

5.2 Mean trace metal concentrations for the organisms collected .......................... 60

6.1 Site locations and shipping intensity in Port Curtis............................................ 69

6.2 Imposex grading system for M. marginalba....................................................... 70

6.3 Field data for M. Marginalba in Port Curtis ........................................................ 73

7.1 Mean concentration of metals in oysters at Sites 1 and 2................................. 84

7.2 Concentrations of antioxidant enzymes in oysters ............................................ 86

7.3 Correlations between metal concentrations and enzyme concentrations in gills and hepatopancreas of oysters in Sites 1 and 2 ........................................

88

7.4 Concentration of biomarkers in gill and hepatopancreas of copper-exposed oysters ...............................................................................................................

90

8.1 Pulse exposure scenarios and the biomass inhibition observed at 72 h......... 101

8.2 Extra- and intra-cellular copper determined following the different 72 h copper pulse exposure scenarios....................................................................

102


xviii

Glossary of terms, acronyms and abbreviations Al: Aluminium

Algae: Comparatively simple chlorophyll-bearing plants, most of which are aquatic, and microscopic in size

ANOVA: Analysis of variance

ANZECC: Australian and New Zealand Environment and Conservation Council

ANZFA: Australian and New Zealand Food Authority

Aquatic ecosystem: Any water environment from small to large, from pond to ocean, in which plants and animals interact with the chemical and physical features of the environment

ARMCANZ: Agriculture and Resource Management Council of Australia and New Zealand

As: Arsenic

ASS: Acid sulfate soils

Assessment endpoint: Explicit expression of the environmental value that is to be protected that links the risk assessment to management concerns

Benchmark: A standard or point of reference

Benthic: Referring to organisms living in or on the sediments of aquatic habitats

Bioaccumulation: A general term describing a process by which chemical substances are accumulated by aquatic organisms from water directly or through consumption of food containing the chemicals

Bioavailable: Able to be taken up by organisms

Bioconcentration: A process by which there is a net accumulation of a chemical directly from water into aquatic organisms, resulting from simultaneous uptake (e.g. by gill or epithelial tissue) and elimination

Biodiversity: The variety and variability of living organisms and the ecological complexes in which they occur

Biomagnification: The result of the processes of bioconcentration and bioaccumulation by which tissue concentrations of bioaccumulated chemicals increase as the chemical passes up through two or more trophic levels. The term implies an efficient transfer of chemicals from food to consumer so that the residue concentrations increase systematically from one trophic level to the next

Bloom: An unusually large number of organisms of one or a few species, usually algae, per unit of water

BSL: Boyne Smelters Ltd

Cd: Cadmium

CDI: Chronic daily intake

CEM: Centre for Environmental Management (part of Central Queensland University based at Gladstone campus)

Clean: Denotes a site, piece of equipment, sediment or water that does not contain concentrations of test materials under consideration in the study

Community composition: All the types of taxa present in a community

Community: Assemblage of organisms characterised by a distinctive combination of species occupying a common environment and interacting with one another

Concentration: The quantifiable amount of a substance in water, food or sediment

Conceptual model: Diagrammatic tool to identify important pathways, sources and uncertainty


xix

Contaminants: Biological or chemical substances or entities, not normally present in a system, capable of producing an adverse effect in a biological system, seriously injuring structure or function

Contaminated sediment: A sediment containing chemical substances at concentrations above background concentrations and above the ANZECC/ARMCANZ guideline values

COPC: Contaminant of potential concern

COPEC: Contaminant of potential ecological concern

CQU: Central Queensland University

Cr: Chromium

CRC: Cooperative Research Centre

Criteria (water quality): Scientific data evaluated to derive the recommended quality of water for different uses

Cu: Copper

Detection limit: The concentration of a substance that, when processed through the complete analytical method, produces a signal that has a 99% probability of being different from the blank

DGT (diffusive gradients in thin films): A relatively new technique for measuring in situ labile metal ions in water

DO: Dissolved oxygen

DOC: Dissolved organic carbon

EEC: Expected environmental concentration

Environmental values: Particular values or uses of the environment that are important for a healthy ecosystem or for public benefit, welfare, safety or health and that require protection from the effects of contaminants, waste discharges and deposits. Several environmental values may be designated for a specific water body

ERA (ecological risk assessment): A process that evaluates the likelihood that adverse ecological effects are occurring or will occur as a result of exposure to one or more stressors

Eutrophication: Enrichment of waters with nutrients, primarily phosphorus, causing abundant aquatic plant growth and often leading to seasonal deficiencies in dissolved oxygen

Fate: Disposition of a material in various environmental compartments (e.g. soil or sediment, water, air, biota) after transport, transformation and degradation

Fe: Iron

Guideline trigger levels: The concentrations (or loads) for each water quality parameter, below which there exists a low risk that adverse biological (or ecological) effects will occur. They are the levels that trigger some action, either continued monitoring in the case of low-risk situations or further ecosystem-specific investigations in the case of high-risk situations

Guideline: Numerical concentration limit or narrative statement recommended to support and maintain, for example, a designated water use

Hg: Mercury

HHRA (Human health risk assessment): A process that determines the level of risk of harm to humans from exposure to stressors

HQ: Hazard quotient

Hypothesis: Supposition drawn from known facts, made as a starting point for further investigation

IMO: International Maritime Organisation


xx

Indicator: Measurement parameter or combination of parameters that can be used to assess, for example, the quality of water

Invertebrates: Animals lacking a dorsal column of vertebrae or a notochord

ISQG: Interim sediment quality guideline

Level of protection: The acceptable level of change from a defined reference condition

LOEC: Lowest observable effects concentration

Management goals: Long-term management objectives that can be used to assess whether the corresponding environmental value is being maintained. They should reflect the desired levels of protection for the aquatic system and any relevant environmental problems

Measurement parameter: Any parameter or quantifiable variable that is measured to find something out about an ecosystem

NATA: National Association of Testing Authorities of Australia

NHMRC: National Health and Medical Research Council

Ni: Nickel

Organism: Any living animal or plant; anything capable of carrying on life processes

Overlying water: The water above the sediment at a collection site or in a test chamber

PAHs: Polycyclic aromatic hydrocarbons

Pb: Lead

PCBs: Polychlorinated biphenyls

PCIMP: Port Curtis Integrated Monitoring Program

Percentile: Interval in a graphical distribution that represents a given percentage of the data points

pH: The intensity of the acidic or basic character of a solution, defined as the negative logarithm of the hydrogen ion concentration of a solution

POM: Particulate organic matter

Pore water: The water that occupies the space between and surrounds individual sediment particles in an aquatic sediment (often called interstitial water)

QA (Quality assurance): The implementation of checks on the success of quality control (e.g. replicate samples, analysis of samples of known concentration) (See also QC, below)

QAL: Queensland Alumina Ltd QAL–RMDO: Queensland Alumina Ltd red mud dam outlet QC (Quality control): The implementation of procedures to maximise the integrity of

monitoring data (e.g. cleaning procedures, contamination avoidance, sample preservation methods) (See also QA, above)

QCL: Queensland Cement Ltd

Reference condition: An environmental quality or condition that is defined from as many similar systems as possible (including historical data) and used as a benchmark for determining the environmental quality or condition to be achieved and/or maintained in a particular system of equivalent type


xxi

Risk: A statistical concept defined as the expected frequency or probability of undesirable effects resulting from a specified exposure to known or potential environmental concentrations of a material, organism or condition. A material is considered safe if the risks associated with its exposure are judged to be acceptable. Estimates of risk may be expressed in absolute or relative terms. Absolute risk is the excess risk due to exposure. Relative risk is the ratio of the risk in the exposed population to the risk in the unexposed population

Salinity: The presence of soluble salts in water or soils

Se: Selenium

Sediment: Unconsolidated mineral and organic particulate material that has settled to the bottom of aquatic environments

SHOC (Sparse Hydrodynamic Ocean Code): General-purpose hydrodynamic model that allows distributed processing on super-computer platforms

SLERA: Screening level ecological risk assessment

SLRA: Screening level risk assessment

Speciation: Measurement of different chemical forms or species of an element in a solution or solid

Species: Generally regarded as a group of organisms that resemble each other to a greater degree than members of other groups and that form a reproductively isolated group that will not normally breed with members of another group. (Chemical species are differing compounds of an element)

SPWC: Spillway Creek

Stakeholder: A person or group (e.g. an industry, a government jurisdiction, a community group, the public, etc.) that has an interest or concern in something

Standard, e.g. water quality standard: An objective that is recognised in environmental control laws enforceable by a level of government

Stressors: The physical, chemical or biological factors that can cause an adverse effect on an aquatic ecosystem as measured by the condition indicators

Sublethal: Involving a stimulus below the level that causes death

TBT: Tributyltin

Threshold toxicity values: Values below which toxicity is unlikely

Tissue residue guideline: Concentration of a contaminant in tissue linked to potential adverse effects in that organism

TOC: Total organic carbon

Trigger value: A guideline value that if exceeded triggers further investigations

Trophic level: A notional stage in the food chain that transfers matter and energy through a community; primary producers, herbivores, carnivores and decomposers each occupy a different trophic level

TSS: Total suspended solids

UCC: Upper continental crust

Water quality guideline: A numerical concentration limit for a water quality parameter

Water quality standard: A legally enforceable water quality guideline

Zn: Zinc

Contaminant pathways in Port Curtis: Final report 1: Introduction

1


2

Chapter 1 Introduction

1.1. Background The Port Curtis Estuary has a well-developed and expanding industry within its

catchment. It is also one of Australia�s leading ports and is located adjacent to

the World Heritage-listed Great Barrier Reef Marine Park. As a consequence of

increasing population and industrial activities, the Port Curtis Estuary is expected

to receive increasing quantities of contaminant inputs from diffuse sources

(e.g. urban runoff) and point source discharges (e.g. industrial effluents).

Sources of chemical stressors are many, and multiple contaminants are likely

to be transported to the estuary by air and/or water. The challenge for coastal

management within the region is the long-term sustainable management of

further port and industrial development, related population growth and the

management of potentially significant impacts on coastal resources.

The release, fate and impacts of contaminants generated within the region by

industrial and urban activities is an issue of obvious concern. When the Coastal

Zone CRC first started its activities in Port Curtis in 1998, there were few

published studies describing contaminant distributions in Port Curtis. During the

first phase of its activities, the CRC undertook the Port Curtis screening level risk

assessment (SLRA) (Apte et al. 2005) which employed a rigorous, risk-based

approach to identify and prioritise contaminant issues of potential concern. The

SLRA identified some contaminant-related issues which included the anomalous

bioaccumulation of metals by biota from Port Curtis and elevated concentrations

of some contaminants in sediments. Recommendations were made for future

investigations. A separate CRC project developed a pilot-scale hydrodynamic

model of Port Curtis which enabled water movement to be predicted. The model

has clear applications to the prediction of contaminant movement, especially point

source discharges associated with industrial activities.

Contaminant Pathways in Port Curtis is part of Phase 2 of the CRC�s activities in

Port Curtis and focusses on some of the key issues that were identified in the

SLRA. Before describing the outcomes of the Contaminant Pathways project in

detail, a brief summary of the Port Curtis SLRA and its key findings are given

below.


3

1.2 Summary of the screening level risk assessment The objectives of the Port Curtis screening level risk assessment project were to:

(i) Review and collate the existing physical, chemical and biological data

with particular reference to contamination of sediments, waters and biota

(ii) Identify data gaps and collect new contaminant data for waters,

sediments and biota

(iii) Identify contaminants that pose a risk to humans and the environment, by

undertaking a screening level ecological risk assessment (SLERA) and a

human health risk assessment (HHRA) using the combined data sets

(iv) Outline future research needs.

A risk assessment framework formulated by the US Environmental Protection

Agency (USEPA 1998) was utilised. Conceptual models were developed to assist

with planning and design of the study. Following a systematic screening process,

aluminium, arsenic, cadmium, copper, chromium, iron, mercury, nickel, lead,

selenium, zinc, fluoride, cyanide, polycyclic aromatic hydrocarbons (PAHs) and

tributyltin (TBT) were the contaminants examined in detail. Contaminant

concentrations were measured in water and sediments and in biota including

seagrass (Zostera capricorni), oysters (Saccostrea spp.), and mud whelks

(Telescopium telescopium). For the HHRA, concentrations of aluminium, arsenic,

cadmium, copper, chromium, iron, mercury, nickel, lead, selenium, and zinc were

measured in fish and shellfish likely to be consumed by humans, namely,

barramundi (Lates calcarifer), sea mullet (Mugil cephalus), mud crab (Scylla

serrata) and banana prawns (Penaeus merguiensis). Tributyltin was measured in

the edible flesh of mud crab.

The compiled data for water, sediment and biota concentrations was assessed

against assessment endpoints. For waters and sediments, the chosen endpoints

were the latest Australian water and sediment quality guidelines

(ANZECC/ARMCANZ 2000). The ability of biota to integrate fluctuating

concentrations of metals over time and to reflect exposure via dietary uptake

allowed a thorough investigation of the exposure of biota to contaminants in the

Port Curtis Estuary. The observed contaminant concentrations in biota at study

sites were compared with benchmarks derived from concentrations in biota at

control sites and from the literature.

For the screening level HHRA, chronic daily intakes (CDIs) of contaminants by

adults and children consuming seafood from this region were compared to

threshold toxicity values set by regulatory agencies (ATSDR 2005). To account


4

for additivity of other chemicals, a hazard quotient greater than 0.1 indicated a

contaminant of potential concern (COPC).

The following contaminants of potential ecological concern were identified:

• TBT in waters

• Arsenic, TBT and naphthalene (based on limited historical data) in

sediments.

Particulate arsenic and naphthalene may be derived from natural sources within

Port Curtis (e.g. oil shale deposits). The main sources of tributyltin were

commercial shipping and the leisure boats that historically utilise the area.

TBT contamination is a problem affecting all large commercial ports. TBT

concentrations are expected to decline in Port Curtis over the next decade as

this antifoulant is completely phased out worldwide.

The concentrations of dissolved metals in waters of the Port Curtis Estuary were

below levels of regulatory concern. However, the concentrations of dissolved

copper, nickel, lead and zinc were elevated relative to concentrations at pristine

coastal water sites in Australia. The reasons for these elevated concentrations

may be industrial discharges or natural inputs of metals from local geological

formations.

The concentrations of metals in sediments were generally below levels of

regulatory concern. However, arsenic, chromium and nickel concentrations were

consistently above the ANZECC/ARMCANZ (2000) low interim sediment quality

guideline trigger values at many sites, which does not necessarily imply

deleterious effects but is a trigger for further investigations. The concentrations

of arsenic, chromium and nickel were comparable to those at control sites,

suggesting natural sources.

The concentrations of aluminium, arsenic, copper, chromium, iron, mercury, nickel,

selenium and zinc were significantly enriched in marine biota sampled within Port

Curtis relative to organisms at reference sites. This indicates that marine

organisms living in Port Curtis are exposed to higher metal concentrations (as

compared to pristine coastal locations). This did not necessarily imply adverse

effects resulting from exposure to elevated concentrations. It was noted that

further studies were required to investigate whether organism health is impaired

by these increased body burdens of metals.

The spatial analysis conducted as part of the SLERA indicated that the Calliope

River and mid-harbour regions of Port Curtis contain the highest concentrations of

contaminants.


5

The HHRA identified mercury concentrations in large barramundi as a potential

concern for the health of adult and child populations likely to consume fish from

this area. It should be noted that this is a general public health issue affecting

most regions of Australia and is not exclusive to Port Curtis.

The study flagged a number of areas that required further study and possible

management actions:

(i) The sources of contaminants that are bioaccumulated by organisms in

Port Curtis should be identified. Further field surveys conducted over a

wider geographical area and scenario modelling of contaminant

dispersion using the recently developed hydrodynamic model of Port

Curtis (Herzfeld et al. 2003) may allow the differentiation of natural

versus anthropogenic sources of metals and help resolve these issues.

(ii) The sources of particulate arsenic and naphthalene in benthic

sediments should be elucidated. It is highly likely that both

contaminants originate from natural sources.

(iii) The ecological health of organisms that have increased metal burdens

should be evaluated. This may be achieved by measuring sublethal

stress indicators such as enzyme biomarkers in selected organisms.

(iv) The impact of butyltin antifoulants on Port Curtis should be evaluated

by measuring the incidence of imposex in gastropods. This is the most

reliable and sensitive indicator of exposure.

(v) A screening level risk assessment of organic contaminants which were

not covered in this study should be conducted. In particular, the risks

associated with dioxins and poly-chlorinated biphenyls (PCBs), which

were identified as potential chemical stressors in this study, should be

evaluated.

(vi) It was recognised that the role of pulse events (e.g. storms and

dredging) which may result in periodic introduction of contaminants and

sediments from the surrounding catchment area should be evaluated.

Current risk assessment protocols are only directed at understanding

the effects of steady-state contaminant exposure on organisms.

(vii) Further work is required to understand the factors leading to the

bioaccumulation of mercury by barramundi (e.g. sources of mercury,

fish size and age). Additional survey work is required to determine if

other piscivorous fish are also high accumulators of mercury. It was

noted that mercury bioaccumulation is an issue of interest along the

entire Queensland Coast and not just isolated to Port Curtis.


6

1.3 Study objectives Following extensive scoping activities which included discussions with local

stakeholders and CRC members, a number of specific goals for the Port Curtis

Contaminant Pathways project were developed. The goals comprised a mixture of

scientific investigations with direct linkages to Port Curtis and some frontier

research activities which would advance the ability to assess contaminant impacts.

These goals were as follows:

(i) Identification of the sources of dissolved and particulate contaminants

to Port Curtis

(ii) Further development of the Port Curtis hydrodynamic model and its

application to contaminant management

(iii) Development and trial of sublethal biomarker tests for assessing

organism health in Port Curtis (including an investigation on the

biological effects of TBT)

(iv) Characterisation of metal bioaccumulation pathways affecting key

organisms residing in Port Curtis

(v) Examination of the effect of contaminant pulse events on biological

responses. It was recognised that pulse exposure to contaminants is

probably a more realistic scenario for Port Curtis than steady-state,

continuous exposure.

The Contaminant Pathways study commenced in June 2003 and finished in

October 2005. This report summarises the outputs of the study. Further detailed

information of the specific studies conducted may be found in the reports and

publications which are referenced at the end of each chapter.


7

1.4 References ANZECC/ARMCANZ (2000) Australian and New Zealand guidelines for fresh and

marine water quality, Volume 1: The guidelines. Australian and New Zealand

Environment and Conservation Council (ANZECC) and Agriculture and

Resource Management Council of Australia and New Zealand (ARMCANZ).

Apte, S.C., Jones, M.-A., Simpson, S.L., Stauber, J.L., Vicente-Beckett, V.,

Duivenvoorden, L., Johnson, R. and Revill, A. (2005) Contaminants in Port

Curtis: screening level risk assessment. Technical Report No. 25, CRC for

Coastal Zone, Estuary and Waterway Management, Brisbane.

ATSDR (2005) ATSDR Minimal Risk Levels (MRLs for hazardous substances)

[Online]. Agency for Toxic Substances and Disease Registry.

<http://www.atsdr.cdc.gov/mrls.html>. Last accessed 8 July 2006.

Herzfeld M., Parslow J., Andrewartha J.R., Sakov P. and Webster I.T. (2003)

Numerical modelling of the Port Curtis region. Technical Report No. 7,

CRC for Coastal Zone Estuary and Waterway Management, Brisbane.

Ross, D.J. (2002) Acid sulfate soils, Tannum Sands to St Lawrence, Central

Queensland. Queensland Department of Natural Resources & Mines,

Rockhampton.

USEPA (1998) Guidelines for ecological risk assessment. Draft, US Environmental

Protection Agency. Washington DC.

Contaminant pathways in Port Curtis: Final report 2: Dissolved trace metals

8

Chapter 2 Concentrations and sources of dissolved trace metals in Port Curtis and surrounding coastal waters

2.1 Introduction The screening level risk assessment (SLRA) of contaminants in Port Curtis

indicated that the concentrations of dissolved metals were in the low- or sub- parts

per billion range and below levels of regulatory concern (Apte et al. 2005). Trace

metal concentrations were, however, generally elevated relative to other coastal

Australian waters. This indicated additional sources of metals to the water column

within Port Curtis which may be related to local industry or regional geology.

This study involved a detailed investigation of cadmium, copper, nickel, lead and

zinc concentrations in waters and suspended particulates collected in the Port

Curtis Estuary and surrounding coastal waters (Figure 2.1).

Figure 2.1. Port Curtis estuary and surrounding waters showing positions of SLRA

sampling sites during Survey 1 (!) and Survey 2 (")

Survey 1 Survey 2 Survey 1 and 2

The Narrows

Curtis Island

Facing Island

Hummock Hill Island

Great Keppel Island

Fitzroy River

Great Keppel Bay

Rodds Bay

Gladstone

Ramsays Crossing

Fishermans Landing

StudyArea

Kilometres


9

A much larger geographical area was investigated than in the SLRA in order to

identify potential sources of trace metals to Port Curtis. Two sampling trips were

undertaken, first along axial transects extending away from possible point sources

within the Harbour and second, through The Narrows, and targeted sampling

along transects and up selected waterways, including the Fitzroy River. Periodic

sampling at two fixed locations was undertaken to determine possible temporal

variations in dissolved metal concentrations. The influence of small pH variations

and sediment resuspension on metal release was also investigated.

2.2 Experimental Two extensive field surveys were carried out in successive years: Survey 1

(1�3 December 2003) and Survey 2 (6�9 December 2004). In Survey 1, water

samples were collected from 49 sites approximately 4 km apart along four

transects: (i) north-west of Gladstone and through The Narrows, (ii) south-east of

Gladstone to Hummock Hill Island then north-east into coast water, (iii) north-east

from the seaward side of Facing Island, and (iv) north-east from Keppel Bay to

beyond Great Keppel Island (maximum distance from shore was 55 km). In

Survey 2, water samples were collected from 51 sites and included: (i) a repeat

transect through The Narrows, (ii) targeted sampling in the larger inlets and creeks

of the northern section of Port Curtis and through The Narrows to the Fitzroy

Delta, (iii) larger inlets and creeks of Fitzroy delta, the Fitzroy mouth and up the

Fitzroy River.

Temporal sampling to assess possible short-term fluctuations in dissolved metals

concentrations was undertaken at a point 3 km south of Ramsay�s Crossing

(7�8 December 2004, hourly sampling for 5 h) and Fisherman�s Landing

(8�9 December 2004, sampling over a 29 h period).

Since the waters of Port Curtis are well mixed, only surface water samples were

collected. Ultratrace sampling techniques described by Apte et al. (2002) were

employed. All trace metal sample analyses were undertaken in a trace metals

clean room. Dissolved copper, cadmium, nickel, lead and zinc were analysed

using a dithiocarbamate complexation/solvent extraction, and GF-AAS detection

procedure described elsewhere (Apte et al. 1998). For all analyses, spiking

recovery tests, duplicate determinations, sea water certified reference material

(CRM) CASS-4 (National Research Council of Canada) and blanks were

processed as part of routine quality control procedures.


10

2.3 Results and discussion Quality control data indicated satisfactory performance with reference material

values (cadmium, copper, nickel, lead and zinc) ranging from 86�99% of the

certified values. Spike recoveries ranged from 88�102% for the metals of interest.

2.3.1 Dissolved metal concentrations

The dissolved and particulate metals results for the two surveys are summarised

in Tables 2.1 and 2.2. Figure 2.2 shows the dissolved concentrations of copper,

nickel and zinc along the sampling transects of Survey 1 and during Survey 2

when the Fitzroy River region was also investigated. Dissolved concentrations of

copper and zinc were typically in the low parts per billion range (µg/L) and

dissolved concentrations of cadmium, nickel and lead were typically in the low

parts per trillion range (ng/L). The lowest metal concentrations occurred in the

open water sites north-east of Facing Island (dissolved cadmium, copper, nickel,

lead, and zinc were <1.5, <19, 118, <11 and <31 ng/L, respectively). Dissolved

concentrations were typically higher in the mid-harbour close to Gladstone, in

the middle of The Narrows near Ramsays Crossing, and up the Fitzroy River

(Figure 2.2). Nevertheless, these concentrations were well below the Australian

water quality guidelines that apply for marine waters (Table 2.3).

Cadmium concentrations ranged from <1.5 to 38 ng/L, with the highest

concentration measured in The Narrows during Survey 2. Cadmium

concentrations were typically <1.5 ng/L in the open ocean waters off Facing Island

and Great Keppel Island and were generally 5�20 ng/L closer to Gladstone and in

The Narrows. Dissolved copper concentrations were typically <40 ng/L in the open

ocean waters off Facing Island and Great Keppel Island and were generally in the

400�800 ng/L range closer to Gladstone and in The Narrows. A dissolved copper

maximum occurred in the southern Narrows during Survey 1 and in the mid-

harbour close to Fisherman�s Landing in Survey 2.

The highest measured concentrations of dissolved copper occurred in the Fitzroy

River with concentrations of 650 and 694 ng/L near the mouth of the estuary and

1290, 1200, and 1410 ng/L near the city of Rockhampton. Dissolved nickel

concentrations were above 100 ng/L at all sites, with typical concentrations of

300�700 ng/L in the harbour and Narrows, 1000�2000 ng/L measured in The

Fitzroy River, and as high as 535 ng/L in Great Keppel Bay. A dissolved nickel

maximum (800�900 ng/L) occurred in the middle of The Narrows in both surveys.

The highest measured dissolved nickel concentrations occurred in the Fitzroy

River, and were 982 and 1080 ng/L near the mouth of the estuary and 1760, 1450

and 1570 ng/L at river sites adjacent to the city of Rockhampton. Dissolved zinc


11

concentrations were typically higher closer to Gladstone, in the middle of The

Narrows, and up the Fitzroy River (Figure 2.2).

Dissolved metals in Port Curtis, The Narrows and the adjacent coastal sites are

compared to data for other coastal water locations in Table 2.3. Open water

coastal sites adjacent to Port Curtis had low dissolved metal concentrations similar

to those measured at other uncontaminated coastal sites in Australia (Apte et al.

1998; Apte & Day 1998; Mackey et al. 2002).

2.3.2 Temporal variations in metal concentrations

Time series samples were taken at two locations to investigate the variability of

dissolved metals with tidal state (Figure 2.3). The data showed remarkably little

variation of metal concentration over the relatively short time period of the study.

There was little evidence of pulsed inputs of metals, for example, from industrial

sources or release from sediments.

2.3.3 Particulate metal concentrations in suspended solids

Total suspended solids (TSS) concentrations and particulate metal concentrations

in the water samples are shown in Table 2.1. The highest TSS concentrations

occurred close to the mouth of the Fitzroy River (22�89 mg/L). The particulate

metal concentrations of suspended sediments are reported in Tables 2.1 and 2.2.

At most sites, suspended particulate copper ranged between 10�20 µg/g and

suspended particulate zinc ranged between 30�80 µg/g. These values compare to

mean benthic sediment concentrations in Port Curtis of 18 ± 12 and 32 ± 29 µg/g

for copper and zinc respectively (Apte et al. 2005). By comparison, particulate

copper and zinc concentrations in suspended particulate matter from Sydney

Harbour, a system receiving numerous contaminant inputs, are typically 100 and

700 µg/g, respectively (Hatje et al. 2001).

Clearly, the concentrations of particulate metals both in the suspended sediment

and benthic sediments in Port Curtis do not suggest gross contamination or

geological enrichment of metals. Mass balance calculation using the combined

data set from both surveys indicate that 73 ± 14% and 19 ± 12 % of total copper

and zinc, respectively, were present in the dissolved phase. Owing to the low

sediment metals load, it is therefore unlikely that desorption of copper and zinc

from suspended sediments and/or release of metals from benthic sediments are

significant sources of metals to the water column. Laboratory experiments are

currently being conducted to confirm this important issue. It appears that inputs

of copper and zinc to the system are predominantly in dissolved forms.


12

Table 2.1. Dissolved and particulate metal concentrations: Survey 1

Site TSSThe Narrows Ni Cd Cu Pb Zn Al Mn Fe Cu Zn (mg/L) SalinityA1.1 South of 334 4.5 523 33 128 9190 917 15400 24 46 3.7 37.5A1.2 The Narrows 360 15.0 606 133 197 8700 882 14800 22 56 6.1 36.4A1.3 334 2.0 551 <11 133 7780 748 12400 16 30 6.7 37.2A1.4 466 5.0 623 21 209 8370 952 13700 14 61 5.5 37.2A1.5 519 3.0 637 25 138 9310 936 14900 19 55 5.7 38.4A1.6 623 3.0 607 <11 91 8360 665 12500 17 57 6.3 38.0A1.7 Ramsays 781 3.5 510 15 92 9590 671 16200 14 55 11.5 37.9A1.8 Crossing 842 4.0 548 52 134 8450 643 13900 9 43 12.5 37.4A1.9 905 5.8 557 17 76 8430 555 11000 6 44 11.1 39.0A1.10 875 5.8 582 47 102 7810 456 9720 9 32 18.0 37.9A1.11 700 3.5 539 20 55 6810 356 9520 9 26 17.5 38.7A1.12 611 6.5 516 82 152 4320 235 7040 8 19 39.5 37.0A1.13 North end of 504 3.5 431 30 57 7030 408 11900 11 22 32.4 36.3A1.14 The Narrows 458 3.0 384 17 61 7000 424 12000 13 26 21.2 37.4Gladstone to Hummock Hill Island and then to open oceanA2.1 Gladstone 348 4.5 504 16 164 6310 603 10700 15 29 11.1 37.1A2.2 305 15.0 455 445 234 4480 480 8236 10 26 15.1 35.9A2.3 282 6.3 411 11 129 8830 771 14000 20 61 5.1 36.3A2.4 225 15.0 295 422 189 2160 187 3600 46 27 11.7 36.3A2.5 196 2.0 250 32 70 6370 467 8590 14 88 2.8 35.8A2.6 195 3.8 185 96 64 2420 172 2900 3 34 5.8 34.8A2.7 187 3.0 215 54 57 8110 - - - - 2.2 35.2A2.8 170 10.0 66 228 142 - - - - - 0.8 35.2A2.9 Hummock 136 1.8 84 28 61 6440 354 7460 - 71 2.9 34.5A2.10 Hill Island 124 3.0 70 65 92 2560 126 2520 16 19 6.6 34.3A2.11 148 <1.5 69 <11 42 490 85 595 4 53 4.6 34.9A2.12 114 3.0 60 106 95 600 90 729 - 44 5.0 34.5A2.13 124 <1.5 41 17 68 370 51 506 - 51 4.3 34.9A2.14 163 2.0 39 <11 138 - - - - - 0.8 35.7A2.15 Ocean 164 <1.5 63 19 41 - - - - - 0.6 33.6Facing Island to open oceanA3.1 Facing Island 168 <1.5 118 <11 61 - - - - - 0.8 35.5A3.2 150 1.5 68 14 <31 - - - - - 0.6 34.0A3.3 145 1.5 38 <11 41 - - - - - 1.0 34.9A3.4 118 <1.5 51 12 128 230 160 400 - - 1.9 34.9A3.5 141 <1.5 19 12 <31 420 280 640 - - 1.4 36.4A3.6 130 <1.5 <19 <11 37 - - - - - 0.7 34.9A3.7 143 <1.5 30 21 67 - - - - - 0.9 34.3A3.8 137 1.5 41 120 43 - - - - - 0.4 34.4A3.9 142 <1.5 22 <11 <31 - - - - - 1.0 35.5A3.10 Ocean 161 <1.5 35 <11 <31 - - - - - 0.8 33.6Great Keppel Bay to Beyond Great Keppel IslandA4.1 Great 535 3.0 429 14 48 9260 413 13700 14 31 19.5 36.8A4.2 Keppel Bay 435 3.0 373 <11 64 4570 233 7140 7 26 43.1 36.6A4.3 341 2.0 272 23 46 7960 510 12200 10 68 7.7 35.8A4.4 247 <1.5 194 11 82 8280 462 10700 10 82 4.7 35.6A4.5 256 <1.5 174 <11 <31 7270 397 7640 8 80 2.7 36.7A4.6 256 <1.5 160 11 <31 4210 458 4480 7 108 1.3 34.7A4.7 185 <1.5 118 <11 <31 1710 360 2280 - 101 1.3 36.4A4.8 Beyond 172 4.0 85 112 103 420 82 507 - 18 6.2 35.2A4.9 Great Keppel 161 <1.5 61 11 <31 1690 255 1960 - 69 1.1 35.1A4.10 Island 183 <1.5 67 39 33 - - - - - 0.1 36.0

Dissolved metals (ng/L) Suspended particulate metals (µg/g)


13

Table 2.2. Dissolved and particulate metal concentrations: Survey 2

Site TSS Ni Cd Cu Pb Zn Al Mn Fe Cu Zn (mg/L) Salinity pH

Middle Port Curtis and The Narrows (repeat transect)B1.1 South of 147 2.3 179 <7 62 - - - - - - 34.7 8.16B1.2 The Narrows 195 2.3 272 9 94 - - - - - - 35.8 8.16B1.3 275 4 520 14 153 3850 275 6320 12 78 12 36.3 8.16B1.4 296 12 578 14 306 4600 298 7610 11 74 16 36.5 8.10B1.5 371 6 652 9 262 4850 389 8640 13 78 21 36.7 -B1.6 478 13.3 636 <7 154 - - - - - - 37.1 7.99B1.7 526 5.4 637 11 168 4190 326 6780 11 77 13 37.3 8.01B1.8 Ramsays 704 11.5 598 <7 101 - - - - - - 37.7 7.95B1.9 Crossing 798 10.6 610 <7 183 5780 408 9650 11 68 24 37.9 8.04B1.10 803 14.8 564 <7 83 - - - - - - 37.5 8.01B1.11 696 4.8 467 10 87 4650 286 6930 8 75 15 37.1 8.09B1.12 644 38.3 451 9 89 - - - - - - 36.7 8.06B1.13 531 4.1 401 <7 61 4440 283 6160 7 75 14 36.3 8.15B1.14 North end of 390 3 348 8 79 - - - - - - 35.9 -B1.15 The Narrows 382 9.6 332 14 76 5160 253 7840 9 69 14 35.9 8.20Creeks and inlets in direction of Port Curtis to The Narrows through to the Fitzroy DeltaB2.1 Calliope 1 330 5.6 670 12 343 5340 369 10100 17 78 25 36.1 8.08B2.2 Calliope 2 429 9.2 725 <7 496 - - - - - - 22.8 7.95B2.3 Boat Creek 481 5.6 768 <7 136 - - - - - - 36.2 8.01B2.4 Fisher. Landing 411 5 737 11 215 - - - - - - - 8.12B2.5 Gully 446 4 712 8 167 4790 468 8800 14 72 18 37.1 8.12B2.6 Targinnie 644 6 606 <7 122 - - - - - - 38.1 7.99B2.7 GC1 511 3.9 631 <7 94 - - - - - - 37.4 8.11B2.8 GC2 599 4.2 632 <7 138 - - - - - - 38.4 8.00B2.9 Black Swan 789 4 557 <7 84 - - - - - - 37.8 8.11B2.10 TS average 771 6 509 10 101 5960 365 9990 10 55 30 37.2 8.08Fitzroy delta to Fitzroy mouth and up the Fitzroy RiverB3.1 DP 673 5.2 433 8 119 4490 295 6610 8 74 15 36.8 8.17B3.2 Connors 664 16.0 436 16 93 - - - - - - 36.4 8.12B3.3 Port Alma 756 6.8 533 <7 90 5570 255 8140 12 67 22 36.3 8.24B3.4 Casuarina 1 710 7 538 8 261 7430 434 12700 15 55 31 36.2 8.18B3.5 Casuarina 2 767 19.9 566 13 118 - - - - - - 36.4 8.12B3.6 Cattle Point 469 8.8 413 11 124 6050 256 9000 10 71 41 35.9 8.18B3.7 Lower Fitzroy 1 1080 7.2 694 18 96 6720 437 11300 13 55 89 36.5 8.11B3.8 Lower Fitzroy 2 982 8.9 650 23 139 - - - - - - 36.6 8.12B3.9 Upper Fitzroy 1 1570 4.1 1198 18 363 8490 2520 14900 22 157 23 11.9 7.73B3.10 Upper Fitzroy 2 1430 2 1214 8 143 10200 2720 17500 29 234 8 0.2 8.12B3.11 Upper Fitzroy 3 1760 4.3 1291 30 582 - - - - - - 11.0 7.80

Dissolved metals (ng/L) Suspended particulate metals (µg/g)


14

Zinc, ng/L

Nickel, ng/L

Copper, ng/L

Survey 1 Survey 2

Survey 1 Survey 2

Survey 1 Survey 2

Figure 2.2. Dissolved copper, nickel and zinc concentrations (ng/L) in Port Curtis estuary and surrounding waters


15

0

200

400

600

800

1000

0 1 2 3 4 5 6Time (h)

Dis

solv

ed c

once

ntra

tion

(ng/

L) NiCuZn

A

0

200

400

600

800

1000

0 5 10 15 20 25 30Time (h)

Dis

solv

ed c

once

ntra

tion

(ng/

L)

NiCuZn

B

(a) (b)

Figure 2.3. Extent of dissolved metal fluctuations with time in The Narrows, 3 km south of Ramsays Crossing (a), and near Fisherman�s Landing (b)

Table 2.3. Concentration of trace metals in waters around the world

Metal concentration, ng/L Reference Location Cd Cu Ni Pb Zn

Port Curtis (average) 6 496 407 76 163 This study

The Narrows (average)

7 512 536 21 124 This study

Central Queensland Coastal waters (average)

1 42 147 13 34 This study

Lower Fitzroy River (saline)

8 672 1030 21 118 This study

NSW coast 2.4 31 180 9 <22 Apte et al. 1998

North Pacific 1.1 38 120 � � Mackey et al. 2002

North Atlantic 0.7 68 � 136 � Kremling & Pohl 1989

Port Jackson, Australia

6�104 932�2550 175�1610 � 3270�9660 Hatje et al. 2003

Torres Straight and Gulf of Papua

<1�29 36�986 940�4600 � Apte & Day 1998

Humber estuary, UK 50�450 1800�10100 2500�12000 3000�20500 Comber et al. 1995

Mersey estuary, UK 10�110 800�4950 2000�10500 6500�28000 Comber et al. 1995

San Francisco Bay estuary

22�123 315�2230 140�2410 � 160�1960 Sanudo-Wilhelmy et al. 1996

Guideline values (95% species protection)

55000 1300 70000 44000 15000 ANZECC/ ARMCANZ 2000


16

2.3.4 Salinity and pH

Salinities in Port Curtis, The Narrows and at coastal sites ranged between

33.6�39.0 � (Table 2.1, Figure 2.4). Freshwater inputs resulted in lower salinities

being measured at upstream sites in the Calliope and Fitzroy Rivers. In the tidal

part of the Fitzroy Estuary the salinity ranged between 11.0�36.5 �, and above

the weir at Rockhampton the salinity was negligible. In general, the waters of Port

Curtis were highly saline with little evidence of freshwater inputs. In both surveys

the salinities in Port Curtis and The Narrows were slightly higher than the adjacent

ocean waters. Analysis of the major ions contributing to salinity showed no

individual major ion was responsible for this increase. The elevated salinity

conditions may therefore be a result of evaporative losses of water occurring in

the more enclosed areas of the estuary. Low salinity groundwater inputs are not

significant in the area as these would cause a drop in salinity or a change in the

ratio of major ions to salinity.

The pH measured at sample sites in Survey 2 ranged between 7.73�8.24

(Table 2.2, Figure 2.4). The lowest pH values were measured at sites receiving

freshwater inputs (Upper Fitzroy and Calliope Rivers). In The Narrows, the pH

ranged between 7.95 and 8.2, with a minimum occurring near Ramsays Crossing.

Decreases in pH similar to those observed in Port Curtis and The Narrows have

been reported for other mangrove systems (Clark et al. 1998; Kristensen 2000;

Van Cappellen and Wang 1996; Wang and Van Cappellen 1996).

The breakdown of organic matter is responsible for lower pH values within

mangrove-lined systems, as the organism-facilitated aerobic oxidation of organic

matter results in a net increase in the concentration of H+, thus lowering the pH of

the sediment pore waters. Humic and fulvic acids formed during decomposition

also contribute to the lowering of pH. Abiotic oxidation of reduced species

(i.e. Fe, Mn and sulfides), which probably occurs during periods of low tide, also

contributes to the release of H+, and the lowering of pH. These processes have a

greater effect in mangrove systems because there is usually a greater volume of

organic matter available for oxidation, and there are often more surfaces exposed

to oxygen at low tide, resulting in greater release of H+. Acid sulfate soils are

common along the east coast of Queensland and northern New South Wales and

may also contribute to the observed pH decrease in The Narrows (Powell &

Martens 2005).


17

!

!

!

! ! !

!

! !

!

! ! ! !

!

! ! ! ! !

! ! ! ! !

! ! !!

! ! !!

!!

! ! ! ! ! ! ! !

! ! ! ! ! ! ! ! ! ! ! !

! ! ! ! ! ! ! ! ! ! ! !

!

!

!

! ! ! ! ! ! ! ! ! ! ! !

! ! ! ! ! ! !

!! ! ! ! ! ! ! !

!

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! ! ! ! ! ! ! !

! ! !

! ! ! ! ! ! !

!

! ! ! ! !

! ! ! !

! !

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!

!

!

! !

!

!

!

!

!

!

!

!

!

pH

7.79 -

7.95

7.95 -

8.06

8.06 -

8.13

8.13 -

8.17

8.17 -

8.20

8.20 -

8.22

8.22 -

8.24

8.24 -

8.29

8.29 -

8.35

8.35 -

8.46

!

!

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!!

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!! ! ! ! ! ! ! !

!

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! ! ! !

! !

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!

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!

!

!

Salinity (ppt)

35.79

- 36.0

1

36.01

- 36.1

8

36.18

- 36.3

0

36.30

- 36.3

9

36.39

- 36.4

6

36.46

- 36.5

5

36.55

- 36.6

7

36.67

- 36.8

4

36.84

- 37.0

6

37.06

- 37.3

6

Figure 2.4. Water pH and salinity in the Port Curtis region

The higher salinities and lower pH values measured in The Narrows will influence

the partitioning of metals towards the dissolved phase. The Narrows is also likely

to contain greater quantities of dissolved organic matter than the harbour due to a

greater mangrove surface area to water volume ratio in this area. Complexation of

metals (especially Cu) by natural organic ligands is therefore expected in these

areas.


18

2.3.5 Sources of dissolved metals

A comparison of the dissolved metals concentrations measured in Port Curtis, The

Narrows and coastal waters is presented in Table 2.3. Based on this summary

data and the transect plots shown in Figure 2.2, it is clear that The Narrows region

is elevated in trace metals�especially nickel and copper. During Survey 1,

dissolved copper and nickel concentration maxima occurred in The Narrows and

were 637 and 905 ng/L, respectively. The dissolved nickel maxima occurred

further (~8 km) north in The Narrows than dissolved copper, which may reflect

different sources of these metals. Dissolved lead and zinc concentrations,

however, were higher in Port Curtis (Table 3) and may reflect the importance of

industrial inputs. As shown in Table 2.3, the Fitzroy River plume is particularly

enriched in dissolved nickel and to a lesser extent dissolved copper. This may act

as a potential source of dissolved metals to the north of The Narrows.

The maxima in dissolved copper and nickel, occurring in The Narrows do not

necessarily imply that this region is a source of these metals to Port Curtis. The

volume of water in The Narrows is small compared to the volume of water in the

harbour and may not be a significant source to Port Curtis. Mass balance

calculations using the hydrodynamic model are required to explore this issue.

2.4 Conclusions 1. This study has provided the first accurate data on dissolved trace metal

concentrations in the coastal waters of Central Queensland and in close

proximity to the Great Barrier Reef. It is somewhat surprising that the

accurate measurement of dissolved metal concentrations in this sensitive

ecological region has been previously overlooked. This is probably a

reflection of the technical difficulties in making such measurements at part per

trillion concentration ranges. In the offshore coastal waters, dissolved metal

concentrations were extremely low and were comparable to concentrations

measured at open Pacific Ocean and New South Wales coastal water

locations. Trace metal limitation rather than trace metal contamination is

likely to be more of an issue for organisms inhabiting these waters.

2. Intensive surveying of Port Curtis has confirmed the presence of elevated

metal concentrations within the harbour. The Narrows region was found to

have the highest concentrations of copper and nickel. This was previously

thought to be a relatively pristine area.


19

3. Surveying of the Fitzroy River to above Rockhampton indicated that this major

river is a source of dissolved metals to the local coastal region. In particular,

the Fitzroy contains elevated nickel concentrations which are consistent with

sediment studies conducted by the Coastal Zone CRC. Under some flow

conditions, the Fitzroy plume may enter The Narrows region and potentially

supply dissolved metals to Port Curtis.

4. There were no conspicuous sources of trace metal within Port Curtis. Metals

in suspended and benthic sediments are low and are not a likely source of

trace metals to the water column. The trace metal distributions in Port Curtis

are likely to reflect a mixture of metal inputs including industrial discharges,

mobilisation of metals from mangrove regions in The Narrows and the Fitzroy

River plume. Deconvoluting these multiple sources is difficult. Modelling of

contaminant inputs using a version of the Port Curtis hydrodynamic model

modified to include inputs from the Fitzroy and The Narrows is probably the

most effective way of understanding this complex issue.

5. Salinity and pH gradients were observed in Port Curtis. Salinities tend to be

higher in the north of Port Curtis than in the surrounding coastal waters. This

could reflect evaporation losses in these more sheltered areas where water

circulation is restricted. Water column pH was lowest in The Narrows regions

and is most likely related to acid inputs from the adjacent mangrove regions.

6. Particulate metals data indicated that desorption of metals from suspended

sediments is unlikely to be a major source of dissolved trace metals. The

concentrations of copper and zinc in suspended sediments were, in most

parts, typical of the benthic sediments and did not indicate enrichment of

these metals. Trace metal inputs to Port Curtis which contribute to the

observed dissolved metal concentrations are most likely to be delivered in

solution form.


20

2.5 References Apte, S.C. and Day, G.M. (1998) Dissolved metal concentrations in the Torres Strait

and Gulf of Papua. Marine Pollution Bulletin, 30, 298�304.

Apte, S.C., Batley, G.E. and Maher, W.A. (2002) Monitoring of trace metals and

metalloids in natural waters. Chapter 6 in Handbook of Environmental

Monitoring, Eds F. Burden, U. Forstner, A. Guenther and I. McKelvie, McGraw

Hill, New York.

Apte, S.C., Batley, G.E., Szymczak, R., Rendell, P.S., Lee, R. and Waite, T.D.

(1998) Baseline trace metal concentrations in New South Wales coastal waters.

Marine and Freshwater Research, 49, 203�214.





Clark, M.W., McConchie, D., Lewis, D.W. and Saenger, P. (1998) Redox

stratification and heavy metal partitioning in Avicennia-dominated mangrove

sediments: a geochemical model. Chemical Geology, 149, 147�171.

Comber, S.D.W., Gunn, A.M. and Whalley, C. (1995) Comparison of the partitioning

of trace metals in the Humber and Mersey estuaries. Marine Pollution Bulletin,

30, 851�860.

Hatje, V., Apte, S.C., Hales, L.T. and Birch, G.F. (2003) Dissolved trace metal

distributions in Port Jackson estuary (Sydney Harbour) Australia. Marine

Pollution Bulletin, 46, 719�730.

Hatje, V., Birch, G.F. and Hill, D.M. (2001) Spatial and temporal variability of

particulate trace metals in Port Jackson Estuary, Australia. Estuarine Coastal

and Shelf Science, 53, 63�77.

Kremling, K. and Pohl, C. (1989) Studies on the spatial and seasonal variability of

dissolved cadmium, copper and nickel in north-east Atlantic surface waters.

Marine Chemistry, 27, 43�60.

Kristensen, E. (2000) Organic matter diagensis at the oxic/anoxic interface in

coastal marine sediments, with emphasis on the role of burrowing animals.

Hydrobiologia, 426, 1�24.

Mackey, D.J., O�Sullivan, R.J., Watson, G. and Pont, D. (2002) Trace metals in the

Western Pacific: temporal and spatial variability in the concentrations of Cd, Cu,

Mn and Ni. Deep Sea Research, 49, 2241�2259.


21

Powell, B. and Martens, M. (2005) A review of acid sulfate soil impacts, actions

and policies that impact on water quality in Great Barrier Reef catchments,

including a case study on remediation at East Trinity. Marine Pollution Bulletin,

51, 149�164.

Sanudo-Wilhemy, S.A. and Flegal, A.R. (1996) Trace metal concentrations in the

surf zone of and in coastal waters off Baja California, Mexico. Environmental

Science and Technology, 30, 1575�1580.

Van Cappellen, P. and Wang, Y. (1996) Cycling of iron and manganese in surface

sediments: A general theory for the coupled transport and reaction of carbon,

oxygen, nitrogen, sulfur, iron and manganese. American Journal of Science,

296, 197�243.

Wang, Y. and Van Cappellen, P. (1996) A multicomponent reactive model of early

diagensis: Application to redox cycling in coastal marine sediments.

Geochimica et Cosmochimica Acta, 60, 2993�3014.

Contaminant pathways in Port Curtis: Final report 3: Sediment contaminants

22

Chapter 3 Metal and polycyclic aromatic hydrocarbon contaminants in benthic sediments of Port Curtis

3.1 Introduction The screening level risk assessment (SLRA) of contaminants in Port Curtis (Apte

et al. 2005) found elevated arsenic, nickel and chromium in benthic sediments.

It was suspected that the observed metal enrichment was a natural phenomenon

reflecting the local geology of the region; however, further studies were

recommended to substantiate this possibility. The SLRA also identified the

polyaromatic hydrocarbon (PAH), naphthalene, in sediment as a contaminant of

potential concern. This finding was based on the results of a study conducted in

2000 where elevated concentrations of naphthalene (200�501 µg/kg) were found in

five out of 20 sediment samples from Port Curtis (WBM Oceanics Australia 2000).

The samples exceeded the ANZECC and ARMCANZ (2000) sediment guideline for

naphthalene of 160 µg/kg. Other PAHs were detected only at <10�20 µg/kg.

This study sought to determine:

• the sources of arsenic, nickel and chromium in benthic sediments

• the concentrations of PAH contaminants, particularly naphthalene in

benthic sediments

• the main contaminant deposition zones in Port Curtis and the deposition

rates of particulate contaminants.

The SLRA (Apte et al. 2005) investigated metal concentrations in estuarine

surficial sediments alone. In this study, sediment cores, especially from intertidal

mangrove sites, were collected and analysed. Mangroves or intertidal sites

generally trap fine sediments and hence may exhibit higher metal concentrations

than estuarine benthic sediments. Sediment cores can provide evidence and

history of contaminant accumulation provided they are not substantially disturbed

by natural forces (e.g. mixing or bioturbation) or human activities (e.g. dredging,

infrastructure or development works).

To assist in the determination of sediment sources, stable lead isotope ratio

measurements were performed on sediment samples from Port Curtis and some

sediments from Fitzroy catchment. Lead isotope ratios have been successfully

used to provide evidence of lead sources in sediments (Munksgaard et al. 2003).

Sediment geochronology using Pb-210 and Cs-137 gamma ray spectrometry was

utilised to estimate the age of intertidal sediments and their deposition rates.


23

3.2 Experimental Full details of the methods employed in this study may be found in the report by

Vicente-Beckett et al. (2006). The samples comprised 28 sediment grab samples

and 31 push cores (core lengths varying between 14 and 45 cm). Sample

locations are summarised in Appendix 3.1 of this chapter and are shown in

Figure 3.1. Grab samples were collected using a Van Veen sampler which is

designed for the collection of fine sediments from soft-bottomed locations

(Mudroch & Azcue 1995). Sediment cores were collected using 50 mm i.d. x 1 m

PVC pipes. They were pushed manually through intertidal or subtidal sediment

beds. The sediment cores were extruded from the pipes and sliced into two, 1cm

slices from the top of the core, followed by 2 cm slices for the next 8 cm, then

4 cm slices until the end of the core. All samples were kept frozen until analysis.

Frozen sediment samples were thawed at room temperature shortly before

analysis. Particle size distribution was determined gravimetrically on vacuum-dried

(at 40oC) fractions following wet-sieving through a 1 mm or 60 µm nylon sieve.

Dried sediments (≤1 mm particle size) were subjected to hot multi-acid digestion

and analysed for total metals using inductively coupled argon plasma emission

spectrometry (ICP-AES) (copper, nickel, zinc, aluminium, calcium and sulfur) or

inductively coupled argon plasma mass spectrometry (ICP-MS) (silver, cadmium

and lead), cold vapour atomic absorption spectrometry (CV-AAS) (mercury) and

neutron activation analysis (NAA) (arsenic, chromium, iron and antimony). Marine

sediment reference materials PACS-2 and BCSS-1 (NRC Canada) were also

analysed as a check on analytical accuracy. Spiked recoveries were 87�107% for

most metals. Organic carbon analyses were performed by a NATA-certified

analytical laboratory (Queensland Health Pathology and Scientific Services,

Brisbane). Sediments were prepared according to the Standards Australia method

AS4479. Total organic carbon was determined using a Leco C200 carbon analyser.

For the analysis of stable lead isotope ratios (PbIRs), sediments < 60 µm size

were digested in 1 mL concentrated nitric acid + 4 mL concentrated perchloric acid

in an open tube block digester at 200oC for 6 h. PbIRs were analysed as 208Pb/206Pb and 207Pb/206Pb calibrated to NIST standard reference material 981

(common lead). Experimental procedures used were similar to those given in

Munksgaard et al. (2003). Archived surficial sediments from the Phase I surveys

were also included in measurements of stable lead isotope ratios. Pb-210 and Cs-

137 activities were determined by gamma ray spectrometry on selected core

slices at the laboratories of CSIRO Land and Water, Canberra, ACT.


24

Figure 3.1. Distribution of sediments with particle size fraction < 60 µm (% mud)

PAH analyses were performed by a NATA-certified analytical laboratory

(Queensland Health and Pathology Scientific Services, Brisbane). About 50 g of

wet sediment was mixed with hydromatrix (diatomaceous earth) to form a free-

flowing powder which was then extracted using Dionex ASE100 or ASE300

(accelerated solvent extraction with 1:1 dichloromethane-acetone). The sample

was heated to 125oC with a static cycle of 5 min. Following extraction the solvent

extract was cleaned up using gel permeation chromatography (Waters Envirogel).

The extract was then concentrated and analysed by GC-MS (Shimadzu GC17a)

for PAHs. Following GC-MS the extracts were split, with one half undergoing

LC-MS/MS and the other half cleaned up using a Florisil column prior to analysis

by GC-ECD. Each batch of samples included a solvent blank and a sample spiked

with a mixture of PAHs. The limit of detection (LOD) for each PAH analysed was

2 µg/kg sediment dry weight.

Rating % Mud 1 0�20 2 20�40 3 40�55 4 55�70 5 70�85 6 85�100

15 Kilometres


25

3.3 Results and discussion

3.3.1 Metals in surficial sediments and sediment cores

The study obtained several push cores from mangrove sites around the Port

Curtis area, with the majority sited on the Gladstone side of the Port. The analyses

of the core slices are summarised in Appendix 3.2, together with the mean data for

surficial sediments from the main estuary obtained during the SLRA (Apte et al.

2005). Zones 1�7 represent sections of the estuary, as designated in the SLRA.

New zone designations in the present study are zones 8�13 which include

additional sites largely at mangrove sites, plus a few subtidal/estuarine sites from

Boyne River and one site at Awoonga Dam (upstream Boyne River). Figure 3.1

indicates that the fine sediment depositional zones in the estuary were largely

found in the intertidal and subtidal sites, particularly along The Narrows, Calliope

River and the South Trees Inlet�Boyne River, with some accumulation also

occurring at the north entrance

The sediment cores (depths 14�45 cm) taken from mangrove sites generally

showed constant metal concentrations with depth. Examples of core profiles for

key trace metals are shown in Figures 3.2 and 3.3. The relatively constant

concentration-depth profile suggests that sediments from the mangrove sites

received low inputs of anthropogenically derived trace metals and were probably

subjected to strong mixing by tidal and natural wave action.

-35

-30

-25

-20

-15

-10

-5

0

0 5 10 15 20 25 30 35 40

Metal (mg/kg dry weight)

Dep

th (c

m)

ArsenicChromiumNickelLeadCopper

Figure 3.2. Variation of metal concentrations with depth in a core from the

Calliope River mouth


26

-14

-12

-10

-8

-6

-4

-2

0

0 5 10 15 20 25 30 35 40 45

Metal (mg/kg dry weight)

Dep

th (c

m)

ArsenicChromiumNickelLeadCopper

Figure 3.3. Variation of metal concentrations with depth in a core from

Targinnie Creek

3.3.2 Estimates of background metal concentrations in sediments

Determining whether sediments have been enhanced or enriched in metals and

other contaminants by anthropogenic influences, and identifying contaminant

depositional zones, both require a knowledge of background metal concentrations

in sediments. As this information is often not known, the elemental composition of

the earth�s upper continental crust (UCC) or shale composition is often used as a

background reference level (Liaghati et al. 2003; Reimann and de Caritat 2005;

Selveraj et al. 2004).

A simple approach was used to estimate the background metal concentrations in

sediments. This was based on several methods reported in the literature which

involve the development of statistical models (e.g. Roussiez et al. 2005; Doherty

et al. 2000a,b; Liu et al. 2003). The mean metal concentrations were calculated

using only surficial sediments from the estuary, excluding the mangrove sites.

The original dataset of 100 measurements was reduced to a final dataset using

an outlier-based data elimination approach. The mean values of the final dataset

(n=11) are given in Table 3.1. Also included in Table 3.1 for comparison are

historical data for the Calliope River and the most recent estimate of the

composition of the upper continental crust of Queensland (UCC-MUQ) based on

25 river and 30 alluvial sediments around Queensland (Kamber et al. 2005). The

mean metal concentrations obtained from the reduced data set for sediments from

Port Curtis are consistent with the range of reported values for UCC-MUQ. Indeed,

the nickel concentration for Port Curtis is considerably lower than the UCC-MUQ


27

value. The estimation procedure used in this study has also provided background

concentrations for arsenic, cadmium and mercury which are lacking in the UCC-

MUQ dataset. It should also be noted that arsenic, chromium and nickel

concentrations did not change consistently with depth in any of the core samples

collected. This uniform distribution of metals is not consistent with anthropogenic

inputs of metals (surface enrichment expected). Based on three lines of evidence,

it is therefore concluded that the concentrations of particulate arsenic, chromium

and nickel in the benthic sediments of Port Curtis are related to the local geology

and do not reflect metal contamination by anthropogenic sources. This important

factor needs to be taken into account when applying the ANZECC/ARMCANZ

(2000) sediment quality assessment framework to this region.

Table 3.1. Metal concentrations in benthic sediments from various locations

Location Ag As Cd Cr Cu Hg Ni Pb Zn

mg/kg (dry wt) Port Curtis Estuary � mean of reduced dataset, n=11

0.054 20.4 0.05 65 29 0.0135 20 14 70

Upper continental crust of Queensland <150 µm fraction (Kamber et al. 2005)

nd* nd nd 65 32 nd 32 20 74

Calliope River <150 µm fraction, (Kamber et al. 2005)

nd nd nd 37 59 nd 19 8.3 74

ANZECC ISQG-low 1 20 1.5 80 65 0.15 21 50 200

*nd = no data

3.3.3 Sediment geochronology

Studies on chronology and/or sedimentation rates are often based on excess or

unsupported 210Pb activity. The technique makes use of the natural fallout

radionuclide 210Pb, a member of the uranium decay series, found on all surfaces

exposed to the atmosphere. This atmospherically derived excess 210Pb is

scavenged from the atmosphere by both wet and dry processes, subsequently

being incorporated in sedimentary deposits and decaying with a half-life of about

22 years. The unsupported 210Pb activity or excess 210Pb is the measured activity

of 210Pb which exceeds the activity in equilibrium with 226Ra in the sediment. 137Cs

fallout resulted from nuclear bomb detonations between 1945 and 1980 and is

globally distributed. Using several cores from northern Queensland, Pfitzner et al

(2004) demonstrated that 137Cs is also a useful independent tracer for sediment

dating purposes.

Four sediment cores from intertidal mangrove sites spread across the harbour

were taken for dating. No pronounced 210Pb excess or 137Cs activity was observed,

particularly at the top sections (0�10 cm) of all cores (Figures 3.4 and 3.5).


28

Figure 3.4. 137Cs activity in sediment cores from Port Curtis. Indicative error bars are

shown (± standard error)

Figure 3.5. 210Pb activity in sediment cores from Port Curtis. Indicative error bars are

shown (± standard error)

-35

-30

-25

-20

-15

-10

-5

0

0 0.5 1 1.5 2 2.5 3 3.5 Activity Bq/Kg (dry weight)

Narrows Rodds HarbourCalliope RiverSouth Trees

-35

-30

-25

-20

-15

-10

-5

0

0 5 10 15 20 25 30 Activity Bq/Kg (dry weight)

App

rox.

dep

th (c

m)

Narrows Rodds HarbourCalliope RiverSouth Trees

App

rox.

dep

th (c

m)


29

The lack of 210Pb excess count in the upper layers of the cores may be attributed

to sediment mixing, a phenomenon which has been observed in estuarine

sediment cores (Pfitzner et al. 2004). Longer cores would probably have provided

better estimates of the sedimentation history.

Taking the year 1958 as the time when 137Cs activity above background

concentrations was detected in the southern hemisphere (Pfitzner et al. 2004),

it appears that at least 28 cm of sediment was deposited over the past 47 years

(core sampled in 2005). This gives an estimated sediment accumulation rate of

at least 0.60 cm/y. This is roughly one-third of the sediment accumulation rate of

about 1.9 cm/y (also based on 137Cs activity) for a 1.2 m sediment core from

offshore Keppel Bay, Central Queensland sampled in 2000 (V. Vicente-Beckett,

unpublished data). This difference is not surprising considering the difference in

the hydrodynamics and nature of human activities between the two locations.

It is quite certain, however, that the sediments were post-1958 because of the

presence of 137Cs activity. This would be around the start of industrialisation of

the Gladstone area.

3.3.4 Stable lead isotope ratios (PbIRs)

The variety of lead ores used in various industrial applications has led to the

introduction of lead in the environment with distinct relative isotopic abundances.

The relative ratios of the four stable lead isotopes 206Pb (from radioactive decay

of 238U), 207Pb (from 235U decay), 208Pb (from 232Th decay) and 204Pb (no known

radioactive parent) depends upon the age and U/Pb and Th/Pb ratios of the ore

from which the lead was derived. Very old ores such as those from Broken Hill,

Australia contain small amounts of radiogenic lead isotopes; younger ores derived

from high U/Pb sources such as that mined in Missouri have much higher

proportions of 206Pb, 207Pb and 208Pb relative to 204Pb (Chillrud et al. 2003).

PbIRs of sediments from Port Curtis estuary (plus a grab from Awoonga Dam)

and selected soils/sediments from the lower Fitzroy catchment are plotted in

Figure 3.6. The mean PbIRs for all Port Curtis sediments were 208Pb/206Pb =

2.0758 ± 0.0111 and 207Pb/206Pb = 0.83676 ± 0.0068. The observed mean 208Pb/206Pb ratio is comparable to the value of 2.0635 measured for near-pristine

estuarine and marine tropical northern Australia (Munksgaard et al. 2003). The

modelled present-day average crustal values of these ratios was reported by

Stacey and Kramers (1975) as 208Pb/206Pb = 2.06058 and 207Pb/206Pb = 0.83572.

This point was included in Figure 3.6, as well as the lead isotope ratios for Mount

Isa (Queensland) lead deposits and those for oceanic sediments (Atlantic and

Pacific) (Stacey & Kramers 1975).


30

0.80

0.82

0.84

0.86

0.88

0.90

0.92

0.94

0.96

0.98

2.00 2.05 2.10 2.15 2.20 2.25208Pb/206Pb

207 Pb

/206 Pb

Port Curtis

Ave modern Pb

Mt Isa Pb

Awoonga Dam

Boyne River

Fitzroy catchment

Atlantic/PacificOceans sediments

Figure 3.6. Lead isotope ratios in Port Curtis sediments and other samples

The measured ratios for lead-contaminated soils/sludges from the Fitzroy

catchment are also included in the figure, with the sludge sample showing the

highest PbIRs. The plot shows that the Port Curtis sediments fall within a linear

trend (r2 = 0.65) starting from average modern or present-day lead and ending at

the most radiogenic Mount Isa lead (Munksgaard et al. 2003), with the Port Curtis

sediments being closer to the PbIRs for present-day lead. Higher PbIRs indicate

more anthropogenic lead inputs, probably via atmospheric lead (e.g. leaded petrol

emissions) and industrial sources (e.g. coal-fired operations), such as that found

for the contaminated Fitzroy soils and sludge samples. Duzgoren-Aydin et al.

(2004) reported (in converted ratios) 208Pb/206Pb = 2.2190 for alkyl lead sources

from Australian ores; the range of this ratio for Australia and New Zealand

atmospheric lead in 1997 was 2.1565-2.1847 (Bollhofer & Rosman 2000).

It is clear that the PbIRs measured for Port Curtis sediments do not show any

signs of anthropogenic lead contamination. This is consistent with the mean

particulate lead concentrations of the sediments (12.4 ± 3.9 mg/kg dry wt) which

are comparatively low and can be considered as baseline concentrations.


31

3.3.5 PAHs in sediments

PAHs are organic compounds with at least two fused benzene rings which are

toxic to many aquatic organisms. Higher molecular weight PAHs are particularly of

concern because they are potential carcinogens, mutagens and/or teratogens in

humans. They may be of fossil, biogenic or diagenic origin. PAHs generally enter

the air and are produced through incomplete burning of organic substances such

as coal, oil and gas and garbage. They can be produced by forest fires caused by

humans, coal-fired electricity power plants, petrol and diesel combustion engines,

incineration and burning of wood and coal. They are natural constituents of crude

oils, accounting for about 20% of total hydrocarbons (Kennish 1997). Their low

solubility in water makes them attractive to hydrophobic organic matter,

suspended particulates and sediments where they may remain for extended

periods. They appear to degrade only very slowly, mainly by microbial action and

photodegradation (Kennish 1997).

Seventeen PAHs (Table 3.2), representing mainly the priority pollutants identified

in the ANZECC/ARMCANZ (2000) interim sediment quality guidelines, were tested

in 25 sediment grabs and in one shallow core. Appendix 3.3 shows the individual

PAHs detected for the sediment grabs and Figure 3.7 maps the total PAHs

detected in the estuary. The distribution of selected PAHs is shown in Figures 3.8

to 3.10. The ANZECC trigger values were not exceeded in any of the samples.

Naphthalene concentrations were ≤ 5 µg/kg, in contrast to 200�501 µg/kg reported

earlier (WBM Oceanics Australia, 2000). Naphthalene constitutes a significant

fraction of crude oils and petroleum products with lighter fractions. The high

concentrations reported in 2000 could be indicative of a transient petroleum-

source PAH contamination (Tam et al. 2001).

The highest concentrations of the different PAHs were clearly found near the

industrial centre of Gladstone, that is, along the Calliope River and its mouth, and

at the South Trees Inlet (Figures 3.7 to 3.10). Sediments from the Clinton Coal

Facility (CCF) contained the greatest amount of PAHs, followed by sediments from

Red Mud Dam Outlet (RMDO), Auckland Creek (AC), downstream of the NRG

Power Station (CR-NRG) and the Marina (M). Sediments from the northern and

southern ends of the estuary contained only a few types of PAHs; no detectable

concentrations of PAHs were found in the Boyne River and Rodds Harbour.

Perylene was detected in most samples. Figure 3.11 shows a depth profile of

PAHs in a short core from Munduran Creek, which is seen to predominantly

contain perylene. Perylene is not included in the ANZECC sediment quality

guidelines and appears to be largely from natural sources (Venkatesan 1988).

It is not yet fully understood how it is produced naturally. It has also been found at

higher concentrations in deeper sediment core sections from sediment cores from


32

various locations (Jiang et al. 2000) in the world as well as in Keppel Bay (Vicente-

Beckett et al. 2005).

PAH ratios have been used to infer sources of PAHs in sediments. For example,

for PAHs of MW= 178, a concentration ratio of anthracene to the sum of

anthracene and phenanthrene <0.10 is taken as an indication of petrogenic

sources (e.g. fossil fuels), while a ratio >0.10 indicates a dominance of pyrogenic

or pyrolytic sources (from high-temperature and incomplete combustion of

biomass or fossil fuels) (Yunker et al. 2002). A third source category is diagenic

(e.g. perylene). It has been suggested that concentrations of perylene which are

greater than 10% of the total penta-aromatic isomers indicate a probable diagenic

input whereas those in which perylene is less than 10% indicate a probable

pyrolytic origin of the compound (Readman et al. 2002). Table 3.3 lists the ratios

for some PAH pairs observed in Port Curtis sediment grabs, and some reported

ratios for PAHs depending on the source. Combustion-derived PAHs were

apparently predominant in most samples.

Table 3.2. List of PAHs studied and their abbreviations

PAH Acronym MW Number of rings

Naphthalene NA 128 2

Acenaphthylene AYL 152 3

Acenaphthene AEN 154 3

Fluorene F 166 3

Anthracene AN 178 3

Phenanthrene PH 178 3

Fluoranthene FL 202 4

Pyrene PY 202 4

Benz[a]anthracene BaA 228 4

Chrysene CH 228 4

Benz[a]pyrene BaP 252 5

Benz[e]pyrene BeP 252 5

Benzo[b+k]fluoranthene BbkF 252 5

Perylene PER 252 5

Benzo[ghi]perylene Bghi 276 6

Indeno[123cd]pyrene IP 276 6

Dibenz[ah]anthracene DbA 278 5


33

Figure 3.7. Total PAHs in Port Curtis

Figure 3.8. Naphthalene in benthic sediments

Kilometres

Kilometres


34

Figure 3.9. Benzo[b+k]fluoranthene in benthic sediments

Figure 3.10. Perylene in benthic sediments

Kilometres

Kilometres


35

Figure 3.11. Depth profile of PAHs in Munduran Creek

0 10 20 30 40 50 60 70 80 90

0

3

7

11

15

Top

dept

h of

cor

e sl

ice,

cm

[PAH], ug/kg normalised to 1%TOC

NA AEN AYN F AN PN FL PY BaA CH BbkF BaP BeP PER IP Bghi DbA


36

Table 3.3. PAH ratios and origins

Sample AN/(AN+PH) FL/(FL+PY) BaA/(BaA+CH) IP/(IP+Bghi) PN/AN FL/PY BaA/CH PER/(sum 5-rings)

Narrows 1 1.00 Narrows 2 0.600 1.50 0.81 Munduran Creek 0.571 1.00 1.33 0.94 BC 0.500 1.00 0.64 BS 0.500 0.500 1.00 1.00 0.29 TC 0.545 0.375 1.18 0.60 0.30 GC 0.533 0.333 1.20 0.56 0.30 FP 0.556 0.300 1.20 0.50 0.24 NPI 0.533 0.375 1.18 0.63 0.28 SB 0.500 0.400 1.08 0.60 0.25 FL 0.059 0.567 0.368 12 1.25 0.56 0.19 WI 0.567 – 1.33 0.22 CCF 0.063 0.568 0.366 13 1.32 0.59 0.12 M 0.111 0.529 0.364 8 1.13 0.57 0.21 CR-U/S 0.077 0.500 0.350 9 1.05 0.77 0.28 CR-NRG 0.528 0.333 16 1.12 0.52 0.17 CR-STP 0.500 0.368 1.06 0.56 0.15 AC 0.100 0.537 0.370 12 1.16 0.59 0.17 QAL-RMDO 0.083 0.516 0.360 9.5 1.04 0.59 0.08 SPWC 0.536 0.381 1.15 0.62 0.17 Colosseum Inlet 0.538 0.400 0.40 1.17 0.67 0.36 Awoonga Dam 1.00 Pyrolytic sources (high temperature combustion of fossil fuels and biomass)

>0.10 >0.5 >0.35 >0.5 <10 >1 >0.9 <0.1

Petrogenic sources (e.g. fossil fuels, petroleum and shale oil)

<0.10 <0.5 <0.2 <0.2 >15 <1 < 0.4

Diagenic sources (formed from plant or biogenic precursors)

>0.1


37

3.4 Conclusions The study confirmed that intertidal (mangrove) sediments in Port Curtis tend to

collect fine sediments, which contained higher concentrations of metals and PAHs

than estuarine sediments. Using radiochemical dating methods, the top 28 cm of

subsurface sediments at intertidal/subtidal sites were estimated to have been

deposited since 1958 in Port Curtis, which is roughly the start of the

industrialisation of Gladstone. The rate of sediment deposition was at least

0.6 cm/y. Much deeper sediment cores (at least 1�2 m) are necessary to

determine sediment chronology more accurately. The sediment depositional zones

identified were the northern Narrows, lower Calliope River and South Trees Inlet�

Boyne River areas. Stable lead isotope ratios in Port Curtis sediments were

consistent with those reported for other sediments from northern Queensland.

Using three lines of evidence, it was shown that the concentrations of particulate

arsenic, chromium and nickel in the benthic sediments of Port Curtis are related

to the local geology and do not reflect metal contamination by anthropogenic

sources. This important factor needs to be taken into account when applying

the ANZECC/ARMCANZ (2000) sediment quality assessment framework to

this region.

PAH contaminants in sediments were highest around the industrial area of

Gladstone. Several types of PAHs characteristic of combustion sources were

detected at the middle harbour largely at the Clinton Coal Facility, along Calliope

River and at South Trees Inlet/Boyne River. Relatively high proportions of the

naturally-occurring PAH perylene were found in sediments from The Narrows

and Munduran Creek.


38


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41

Appendix 3.1. Sediment samples (2003–2005)

Date Sample ID Latitude Longitude Location No. of grabs

No. of cores

12/03 N S 23o35.925 E 151o02.343 Narrows 4 12/03 RH S 24o05.088 E 151o31.509 Rodds Harbour 4 7/12/04 N1 S 23o33.898 E 151o00.940 Narrows, near Fitzroy

mouth 1

7/12/04 N2 S 23o37.136 E 151o02.482 Narrows, near Fitzroy mouth

1

28/02/05 SPWC S 23o52.948 E 151o19.069 Spillway Creek 1 2 28/02/05 RMDO S 23o51.629 E 151o18.006 QAL red mud dam outlet 1 2 28/02/05 PB S 23o45.825 E 151o17.506 Pelican Bank 1 2 28/02/05 CR-U/S S 23o51.863 E 151o11.249 Upstream Calliope River 2 2 28/02/05 CR-NRG S 23o50.562 E 151o12.639 Calliope River, few km

downstream of NRG outlet

2 2

28/02/05 CR-STP S 23o50.142 E 151o13.379 Calliope River 2 2 28/02/05 FL S 23o47.355 E 151o10.486 Fisherman�s Landing 1 28/02/05 BC S 23o39.539 E 151o05.837 Northside Boatway Creek 1 28/02/05 BS S 23o40.726 E 151o07.396 Black Swan 1 28/02/05 TC S 23o43.376 E 151o08.216 Targinnie Creek 1 2 28/02/05 GC S 23o44.135 E 151o10.178 Grahams Creek 1 2 28/02/05 FP S 23o45.280 E 151o09.523 North Friend Point Flat 1 28/02/05 NPI S 23o45.682 E 151o10.604 North Passage Island 1 28/02/05 SB S 23o46.919 E 151o12.509 Stockyard Bay 1 28/02/05 WI S 23o48.936 E 151o12.865 Wiggins Island 1 28/02/05 CCF S 23o49.319 E 151o14.340 Clinton Coal Facility 1 28/02/05 M S 23o49.968 E 151o14.962 Marina 1 28/02/05 AC S 23o50.286 E 151o15.000 Auckland Creek 1 4/05 MC1 and 2 S 23 o39.465 E 151o02.933 Munduran Creek 2 2 8/5 N S 23 o61.954 E 151o04.142 Narrows (for dating) 1 8/05 CRM S 23 o82.677 E 151o22.049 Calliope River mouth

(for dating) 1

8/05 CRM S 23 o82.244 E 151o21.905 Calliope River mouth 1 8/05 STI S 23 o85.932 E 151o30.191 South Trees Inlet

(for dating) 1

8/05 RH S 24 o04.133 E 151o31.509 Rodds Harbour (for dating and PAHs analyses)

1 1

8/05 CI S 24 o01.597 E 151o44.200 Colosseum Inlet (for PAHs)

1

8/05 BR S 23 o94.191 E 151o35.429 Boyne River (for PAHs) 1 8/05 AD S 24 o09.510 E 151o30.620 Awoonga Dam 1

Contaminant pathways in Port Curtis: Final report 4: Hydrodynamic model evaluation

42

Appendix 3.2. Particulate metal concentrations (µg/kg, dry weight: Ag, Cd, Hg; mg/kg dry weight for all others) and other parameters Sample description Zone N Mean/

S.E.* Ag As Cd Cr Cu Hg Ni Pb Sb Zn Fe

Grabs; subtidal 1 10 mean 69 20.3 52.7 59.7 24.6 16.7 20.2 11.7 0.672 63.0 39 S.E. 8.0 2.8 0.5 7.4 3.3 3.6 2.9 1.1 0.060 8.3 45Grabs; subtidal 2 15 mean 98 14.7 52.2 54.2 21.5 19.4 17.3 11.7 0.579 62.6 35 S.E. 16 0.9 0.3 4.1 2.7 3.4 2.3 0.66 0.025 6.4 26Grabs; subtidal 3 15 mean 91 15.7 51.3 42.3 14.7 9.1 12.1 9.2 0.484 46.5 29 S.E. 29 2.0 0.2 4.3 1.9 1.7 1.3 0.68 0.040 4.2 28Grabs; subtidal 4 27 mean 53 20.9 51.1 40.3 12.7 5.4 10.6 10.0 0.479 39.7 27 S.E. 1.9 4.1 0.2 4.1 1.9 0.66 1.0 0.64 0.033 4.4 23Grabs; subtidal 5 14 mean 92 16.5 52.9 58.6 27.7 18.3 19.3 13.0 0.603 71.3 37 S.E. 31 1.6 0.4 5.6 3.6 3.2 2.3 0.90 0.038 7.7 28Grabs; subtidal 6 11 mean 66 17.6 73.7 45.6 16.7 12.9 14.2 11.1 0.465 48.4 31 S.E. 7.4 2.3 17.1 6.3 3.6 3.0 2.7 1.2 0.050 7.5 40Grabs; subtidal 7 8 mean 51 18.6 51.3 63.5 13.1 1.7 10.6 10.0 0.489 41.1 27 S.E. 0.73 2.6 0.7 28.5 5.2 0.24 3.0 1.9 0.088 13.5 78Narrows; intertidal 8 50 mean 82.9 21.3 59.3 79.7 30.3 40.6 27.1 17.7 0.631 86.7 37 S.E. 15.1 0.6 3.9 1.9 0.7 1.2 1.2 0.3 0.012 9.9 74Targinnie Ck, Graham Creek, BS, SB, BC

9 14 mean 226.6 19.5 61.0 38.6 31.2 nd 27.1 13.7 nd 83.5 44

Intertidal S.E. 44.7 1.5 13.8 2.8 0.7 1.2 0.3 2.3 91Calliope River-CCF-Marina-AC; intertidal or subtidal

10 31 mean 226.2 11.9 43.0 30.3 35.7 nd 21.2 11.6 nd 80.3 39

S.E. 25.8 0.5 3.3 1.7 1.2 0.6 0.3 2.1 99QAL/RMDO - Spillway Creek; intertidal

11 9 mean 164.7 12.1 31.1 36.8 17.5 nd 14.0 9.2 nd 57.7 26

S.E. 41.0 1.3 5.2 2.3 0.8 1.2 0.5 5.0 19Boyne River (subtidal); Awoonga Dam � fresh water

12 6 mean 165.7 11.9 31.9 43.1 14.9 nd 11.8 9.9 nd 43.4 21

S.E. 115.7 1.2 6.9 7.3 2.8 1.2 1.3 5.2 23Rodds Harbour; intertidal

13 35 mean 51.0 13.2 52.5 50.4 13.9 26.0 20.7 10.7 0.463 46.1 20

S.E. 0.1 0.5 1.5 1.8 0.6 1.1 5.4 0.3 0.015 8.2 95 Total 245 mean 106.0 16.9 52.6 52.2 22.8 21.8 22.6 12.4 0.543 63.6 32 S.E. 7.6 0.6 1.5 1.7 0.73 1.1 1.3 0.25 0.011 2.8 73ANZECC ISQG-low 1000 20 1500 80 65 21 20 2 200

*S.E. = Standard error


43

Appendix 3.3. Polycyclic aromatic hydrocarbons in Port Curtis surface sediments; µg/kg dry weight normalised to 1% TOC Sample ID AEN AYL AN F NA PN IP PER BaA BaP BeP Bb

Narrows 1 (N1-IC) <2 <2 <2 <2 <2 <2 <2 28 <2 <2 <2 <2Narrows 5 (N2-IC) <2 <2 <2 <2 <2 2 <2 15 <2 <2 <2 4Munduran Creek (MC2a) <2 <2 <2 <2 2.5 2.5 0.8 18 <2 <2 <2 1.3BC <2 <2 <2 <2 <2 <2 <2 2 <2 <2 <2 1BS <2 <2 <2 <2 <2 <2 <2 6 3 3 3 6TC <2 <2 <2 1 2 5 <2 7 3 2 4 9FP <2 <2 <2 <2 <2 7 <2 8 3 5 5 14GC <2 <2 <2 1 2 7 <2 9 3 3 5 11FL <2 <2 1 3 4 16 <2 10 7 7 9 19NPI <2 <2 <2 <2 3 6 <2 8 3 3 4 10CR-U/S <2 <2 1 3 3 12 <2 19 7 7 11 25SB <2 <2 <2 <2 <2 3 <2 4 2 2 3 6CR-NRG <2 <2 <2 4 4 21 <2 12 9 9 14 31WI <2 <2 <2 <2 <2 17 <2 9 <2 <2 9 22CR-STP <2 <2 <2 <2 <2 12 <2 8 7 7 10 22CCF 5 <2 2 8 5 30 <2 12 15 15 19 42M <2 <2 2 4 3 16 <2 13 8 7 11 23AC <2 <2 2 4 3 18 <2 13 10 11 13 28RMDO <2 <2 1 2 4 11 <2 8 9 14 17 51SPWC <2 <2 <2 <2 4 8.0 <2 12 8 11 13 28PB <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2Colosseum Inlet <2 <2 <2 <2 4 4 2 5 2 2 <2 6Awoonga Dam <2 <2 <2 <2 3 <2 <2 11 2 2 <2 6Boyne River <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2Rodds Harbour <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2 <2

ANZECC, ISQG-low, low MW PAHs

16

44

85

19

160

240

nd

nd

ANZECC (ISQG-low), high MW PAHs

261 430

ANZECC (ISQG-low), total PAHs


44

Chapter 4 Port Curtis hydrodynamic model evaluation

4.1 Background In the first phase of Coastal CRC activities, a pilot 3D hydrodynamic model was

developed for Port Curtis and the local region (Herzfeld et al. 2003). The pilot model

was subjected to limited calibration only but was adjudged to provide realistic, but

not necessarily accurate predictions of water column mixing. The pilot model

showed that the water circulation within Port Curtis estuary allows dissolved material

to be dispersed evenly throughout the estuary; however, material has difficulty

leaving the estuary into the offshore environment. The e-folding flushing time

(i.e. the time for total mass of material to decrease to ~1/3 of its original mass) for

the estuary was of the order of 19 days. Typical model output is shown in Figure 4.1.

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

0 0.0025 0.005

passive2 050

0000 01 Jan 1999 +100000 01 Jan 1999 +10

Figure 4.1. Typical output from the MECO model showing dispersion of a conservative tracer released from Fisherman�s Landing. See Herzfeld et al. (2003) for further details

The resources allocated in the Contaminant Pathways project were not sufficient

to carry out further model development and full model calibration, so, in its place,

a campaign-based field program followed by a less rigorous model evaluation was

carried out. This evaluation assessed the existing model�s performance, and made

recommendations as to what further work might be necessary to complete model

development.

The pilot model was designed to address environmental impacts, and for these

purposes, the dispersion of a tracer over multiple tidal cycles is one of the key

outputs. Evaluation of transport prediction is best addressed by comparing

predicted and observed distributions of a conservative tracer. The most readily

available tracer of this nature is salinity. An evaluation was therefore based around


45

a 2004�2005 wet season flood event, where freshwater inputs resulted in

measurable salinity fluctuations within Port Curtis. This chapter provides a

summary of the field versus model comparison. Further details may be found in

the report by Andrewartha and Herzfeld (2005).

4.2 Field program The output from the existing pilot model (Herzfeld et al. 2003) was used to

determine the best salinity sampling positions for the field program. These results

are summarised in Figures 4.2 and 4.3 which display:

• 12 sites sampled prior to an anticipated flood event, for the purpose of

initialising the model salinity field

• 3 sites to obtain continuous salinity measurements throughout a flood

event from moored loggers

• 2 transects for salinity profiling at regular intervals throughout and after

the flood event.

The full field program design may be found in the report by Herzfeld et al. (2004).

Three salinity loggers were fixed to channel pylons one metre below low-water at

sites T3, C5 and A4 in early November 2004 (see Figure 4.2). A significant rain

event began on 22 January 2005 and continued until 27 January, with a peak

rainfall in Gladstone on 23 January of 70 mm. Fieldwork was subsequently

conducted between 27 January and 24 February 2005. The January flood event

was only a 30 cumec flow event which is relatively small for the Calliope River in

summer. For instance, the March 1999 flood event which was used in designing

the evaluation program was 140 cumecs. In February 2003, a flood event of

1600 cumecs was recorded. Ideally, a flow event of over 150 cumecs is required

for the evaluation.


46

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

23o 40 / S

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

23o 40 / S

0 km 5

T3C5

A4

S1

S2

S3

S4S5

S6

S7

S8

S9

S10Curtis

Island

Gladstone

Facing

Island

Figure 4.2. Field program sampling sites

Red dots represent the fixed logger locations, while blue dots represent pre-flood sampling locations. Two further sites, S11 and S12, are not shown but lie further offshore east of Facing Island.

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

151o 10 / E 151o 15 / E 151o 20 / E

23o 50 / S

23o 45 / S

T1

T5

T8

T12

T15

T18

T22

T24T25

T26

T28T29

0 km 5

Figure 4.3. Field program transect sampling sites

Transect #1 is represented by sites T1 to T22, while transect #2 is represented by sites T24 to T29. There is no site T23, and T27 is co-located with T8.


47

4.3 Model description and development The hydrodynamic model used in the earlier pilot study was termed MECO

(Walker and Waring 1998). MECO is a general purpose model developed by

CSIRO Marine Research which is applicable to spatial scales ranging from

estuarine to regional ocean domains. MECO has been successfully applied to a

variety of applications encompassing these scales. Further technical details of the

model may be found in the report by Herzfeld et al. (2003). MECO has been

upgraded and extensively modified and is now called SHOC (Sparse

Hydrodynamic Ocean Code). This model utilizes the same computational physics

as MECO, but is cast in an alternate coordinate system to allow distributed

processing on super-computer platforms. It also has an enhanced suite of

diagnostics and contains some additional features.

There were four major differences between the model domain used for the pilot

study and that used for the current evaluation:

(i) In the earlier study, the Calliope River was only represented for several

hundred metres and in the absence of available data, the boundary

condition for salinity was set at 20 psu. For modelling the salinity

distribution resulting from a flood event, a more accurate representation

of the river was required. Without measured data, the only way to

accurately represent the river was to extend its length to the freshwater

boundary. The river length was therefore initially set at 20 km.

(ii) It is a requirement of this type of model that water depth exceeds a

minimum value, otherwise the models become unstable. For the pilot

model, a minimum water depth of 4 m was applied which meant that

large areas of water of <4 m depth e.g. the region between the Calliope

mouth and Fisherman�s Landing, were slightly misrepresented. The new

model is capable of handling shallower water and the minimum water

depth was set at 0.5 m.

(iii) In order to keep the model stable throughout all scenarios, with the 0.5 m

minimum water depth, the bathymetry was mathematically smoothed.

The overall bathymetry contains the same features, but bathymetric

changes are not as abrupt.

(iv) The stability of the model was found to improve with the creation of a

slightly deepened channel from the mouth of the Calliope out to the main

shipping channel. This channel probably exists, but bathymetry data

used in the pilot model was not sufficient to resolve this feature.


48

4.4 Model forcing The model was forced with sea-surface elevation, wind, river flow, rainfall,

temperature and salinity data. Rainfall records for Gladstone were used as input

data. Riverflow records were obtained for the Calliope River at Castlehope and

Fitzroy River at the Gap. The Boyne River was omitted due to the presence of the

Awoonga Dam reducing flows to negligible levels. There are numerous other small

rivers and creeks that flow into Port Curtis, especially at times of heavy rain, but

no data was available for these sources. The model simulation period was

December 2004 to February 2005 inclusive. The final conditions adopted are

described in detail in the report by Andrewartha and Herzfeld (2005).

4.5 Model trials Some preliminary model runs were performed to test sea level prediction.

Modelled sea-levels are compared to those measured at South Trees Inlet in

Figure 4.4, from which it is observed that agreement was good.

−2

−1.5

−1

−0.5

0

0.5

1

1.5

2

2.5

Sea

−Le

vel

(m)

3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18JAN 05

−2

−1.5

−1

−0.5

0

0.5

1

1.5

2

2.5

Sea

−Le

vel

(m)

3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18JAN 05

Figure 4.4. Measured sea level (blue line) compared with modelled sea level (red dots) for

a site near South Trees Inlet during one spring neap tidal cycle

Initial modelling with SHOC indicated that salinity distribution in the estuary was

sensitive to the volume of water contained in the Calliope River (i.e. river length,

width and depth). While the actual length of 20 km from the mouth to Castlehope

may seem the most appropriate, this length had the disadvantage of containing

too great a volume of water, because the model resolution does not allow the

width to be realistically represented. Without data to prescribe the river cross-

sectional area, the volume can only be approximated. Three river lengths were

tested: the full length (~20 km), a half-length and a short (~3.3 km) length. In each

case, the river depth was maintained at 4 m. The results indicated that the short

river produced markedly lower salinities in the estuary, more in keeping with the


49

field measurements. In future studies, it is advisable to place a salinity logger

closer to the mouth and prescribe the salinity boundary condition from measured

data, thus circumventing the need to resolve the entire river length.

4.6 Evaluation results Model runs using the �short� Calliope River and the �half-length� Calliope River

scenarios were implemented using the best configuration and forcing methods as

determined by the trials outlined in the previous section. Further details may be

found in the report by Andrewartha and Herzfeld (2005). Owing to equipment

failure, no useable data was obtained from the data loggers deployed at sites in

Port Curtis.

Time series comparisons of modelled and measured salinity are shown in

Figure 4.5 for the half-length river and in Figure 4.6 for the short river. The

model predicts the two salinity depressions, shown about 2 weeks apart in the

measurements, fairly well. This is best seen in the plots for station C5. The model

and field measurements were in agreement with respect to the vertical structure

which was well mixed.

The comparisons present two definite discrepancies between model predictions

and field measurements. Figures 4.5 and 4.6 indicate that the modelled salinity

drop due to the flood occurs before the measured data and recovers more rapidly.

The model flood peak occurs too early, and for the half-length river, the freshwater

input does not reduce salinity by as much as the measured data suggests. This

phenomenon is no doubt due to the treatment of the river in the model, particularly

the river length and velocity profile established in the channel in response to the

depth. Although a definitive judgment is difficult, the short river is probably

performing the better of the two simulations.

The second discrepancy concerns rapid salinity recovery in the days following the

flood event. While the field transect measurements showed that salinity was

depressed throughout most of February at levels of about 32 to 33 �, the model

predicts salinity to gradually recover to pre-flood levels during that time. Either the

model overestimated flushing rates or, as is more likely, there were additional

inputs of freshwater from sources unaccounted for by the model. Simple mass

balance calculations indicated that the flood event alone was not capable of

lowering salinity to that observed, and freshwater input�from other creeks and

streams entering the estuary, or input through The Narrows and offshore regions,

and probably combined with direct rainfall over the estuary surface�must also

have contributed to lowering salinity. This is a result of the nature of the rainfall


50

event responsible for the flood; rainfall was spread over the lower catchment and

estuary itself. All of these other sources of fresh water into the estuary were

required to be well quantified as inputs if the model was to respond realistically.

.

30

32

34

36S

alin

ity

(PS

U)

Station T3

30

32

34

36S

alin

ity

(PS

U)

Station T3

30

32

34

36

Sal

inity

(PS

U)

Station C5

30

32

34

36

Sal

inity

(PS

U)

Station C5

30

32

34

36

Sal

inity

(PS

U)

5 10 15 20 25 30DEC 04

5 10 15 20 25 30JAN 05

5 10 15 20 25FEB 05

Station A4

30

32

34

36

Sal

inity

(PS

U)

5 10 15 20 25 30DEC 04

5 10 15 20 25 30JAN 05

5 10 15 20 25FEB 05

Station A4

Figure 4.5. Time series comparisons of salinity from the transect measurements (blue) and the model (red) for the half-length river


51

30

32

34

36

Sal

inity

(PS

U)

Station T3

30

32

34

36

Sal

inity

(PS

U)

Station T3

30

32

34

36

Sal

inity

(PS

U)

Station C5

30

32

34

36

Sal

inity

(PS

U)

Station C5

30

32

34

36

Sal

inity

(PS

U)

5 10 15 20 25 30DEC 04

5 10 15 20 25 30JAN 05

5 10 15 20 25FEB 05

Station A4

30

32

34

36

Sal

inity

(PS

U)

5 10 15 20 25 30DEC 04

5 10 15 20 25 30JAN 05

5 10 15 20 25FEB 05

Station A4

Figure 4.6. Time series comparisons of salinity from the transect

measurements (blue) and the model (red) for the short river


52

4.7 Conclusions The evaluation exercise described here has not allowed a rigorous assessment of

the performance of the model with respect to transport of salinity. The key problem

was an inability to quantify the inputs of fresh water with sufficient accuracy during

the chosen flood event. In particular, it appears highly likely that fresh water from a

major flow event in the Fitzroy has entered via The Narrows and dominated local

inputs from the Calliope. It was not possible to capture this effect without a

continuous salinity record from The Narrows.

The flood event was not ideal. Ideally, a flood event was required resulting from

rainfall confined to the upper catchment, so as to create an isolated singular pulse

of fresh water propagating down the river which could be easily quantified with no

confounding influences from other creek systems, rainfall or fresh water entering

through open boundaries.

Despite the identified flaws, there are grounds to be optimistic that the model

represents tracer transport reasonably well. The transport regime in the estuary is

predominantly tidally driven, and the distribution of passive tracer will reflect this

dominant forcing. The model reproduces tidal elevation well. A full calibration

exercise incorporating a comprehensive field program, carried out for a similar

model implemented for the Fitzroy River�Keppel Bay region, showed that model to

reproduce salinity distributions well (Atkinson 2004).

The Port Curtis estuary and surrounds is clearly a complex region in terms of its

shallow topography, high tidal regime, and many rivers and creeks which impact

during the wet season. Ideally, a hydrodynamic model of such a region should be

fully calibrated. This requires an improved physical characterisation of the estuary

with quantification of all freshwater sources and sinks. The Calliope River and The

Narrows boundaries could be dealt with more effectively with appropriate field

sampling. Dedicated field data collection programs using moored instruments,

sampling transects and profiles in both the wet and dry seasons are also

recommended.


53

4.8 References Atkinson I. (2004) Field report on the second survey of dry season water column

and sediment properties in the Fitzroy Estuary and Keppel Bay, Rockhampton

Queensland, August 15 – September 1, 2004. Report for the CRC for Coastal

Zone, Estuary and Waterway Management, Brisbane.

Andrewartha, J.R. and Herzfeld, M. (2005) Port Curtis hydrodynamic model

evaluation. Report for the CRC for Coastal Zone, Estuary and Waterway

Management, Brisbane.

Herzfeld M., Parslow J., Andrewartha J.R., Sakov P. and Webster I.T. (2003)

Numerical modelling of the Port Curtis Region. Technical Report No. 7,

CRC for Coastal Zone, Estuary and Waterway Management, Brisbane.

Herzfeld M., Parslow J. and Andrewartha J.R. (2004) Model evaluation field

program design, Report for the CRC for Coastal Zone, Estuary and Waterway

Management, Brisbane.

Walker S.J. and Waring J.R. (1998) MECO – Model for estuaries and coastal

oceans. CSIRO Marine Research internal Report OMR 118/120, June 1998.

Contaminant pathways in Port Curtis: Final report 5: Metal bioaccumulation in food web

54

Chapter 5 Metal bioaccumulation through foodweb pathways in Port Curtis

5.1 Introduction The Port Curtis screening level risk assessment (Apte et al. 2005) showed that

despite relatively low metal concentrations in the water column, there appears

to be enhanced bioaccumulation of some metals in the marine organisms that

inhabit Port Curtis relative to control sites outside of the harbour. This finding

was supported by earlier studies that flagged concentrations of some metals; in

particular copper and zinc in mud crabs (Andersen & Norton 2001) and copper in

seagrass (Prange 1999) and fiddler crabs (Andersen et al. 2002) as being

potentially anomalous.

Marine organisms can accumulate trace metals from both the dissolved phase and

from ingested food (Fisher & Reinfelder 1995). The relative importance of each

pathway is dependent on both the metal and the organism in question. In order to

better understand dietary routes of metal bioaccumulation, it is first necessary to

elucidate foodweb structure (i.e. who eats what). There are several techniques

available for understanding the diet of organisms, including gut content analysis

and direct observation of feeding behaviour. The use of stable isotopes as an

alternative to gut content analyses has been successfully applied to define aquatic

foodweb interactions (France 1998; Fantle et al. 1999; Kang et al. 1999). The

stable isotope ratios of carbon and nitrogen in the tissues of plants and animals

can give an indication of the energy source, and the position in the food chain of

those organisms, respectively. The isotope approach has the advantage of

measuring assimilated carbon as opposed to carbon merely resident in the gut of

the organism which may or may not be digested.

It was hypothesised that the dietary route of metal accumulation could account for

the observed elevated levels of metals in biota, and therefore this pathway/

mechanism was investigated in this study. Particular emphasis was placed on the

mud crab as an example of a higher trophic consumer. Animals and plants

believed to be within the mud crab food web from a number of sites within (Boat

Creek, Graham Creek and Black Swan) and outside of Port Curtis (Yellow Patch)

were examined for metal accumulation in order to identify possible site-related

differences in metal bioaccumulation. Stable isotope carbon and nitrogen ratios

were also measured on selected samples in order to elucidate the food web and

aid the interpretation of metal bioaccumulation data.


55

5.2 Experimental Full sample collection details may be found in the report by Andersen et al. (2005).

Three sites: Boat Creek (Site 1), Graham Creek (Site 2) and Black Swan (Site 3)

were selected along a north-west transect towards The Narrows, representing

increasing distances from likely sources of anthropogenic inputs (Figure 5.1).

Yellow Patch (Site 4, see Figure 5.1), an unimpacted oceanic reference site on the

eastern side of Curtis Island, was selected for comparison. A list of organisms

collected is given in Table 5.1 and photographs of selected specimens collected

shown in Figure 5.2. The majority of samples were collected from April to June

2001 except for seston and mangrove snails at Site 4 and mullet at Site 2, which

were collected in October 2002. Samples of epiphytic and filamentous algae and

prawns and repeated samples of macroalgae and particulate organic matter

(POM) were collected in June 2004. Typically, five replicates of each sample type

were taken at each site. Mud crab samples for isotopic analysis were taken from

all four sites. For the remaining organism types, isotopic analyses were carried out

on samples predominantly from Site 2. Metals analysis was carried out on

samples collected at all four sites.

0-10 Kilometres

Site 1 - Boat Creek

Site 2 � Graham Creek

Site 3 � Black Swan

Gladstone

Curtis Island

Site 4 � Yellow Patch

Figure 5.1. Location of organism sampling sites in Port Curtis: Site 1 � Boat Creek,

Site 2 � Graham Creek and Site 3 � Black Swan, and reference site outside of Port Curtis: Site 4 � Yellow Patch. Shaded areas indicate mangrove zones


56

Samples were analysed for eight different metals (aluminium, arsenic, cadmium,

chromium, copper, nickel, selenium and zinc) at NMI Sydney and reported on a

wet weight basis. Stable isotope determinations (carbon and nitrogen) were

carried out by CSIRO Marine Laboratories, Hobart. The combined datasets were

then interpreted using a variety of statistical tests. Further details may be found in

the report by Andersen et al. (2005).

Table 5.1. Organisms collected for the foodweb study

Organism Mud crab (Scylla serrata) Fiddler crab (Uca coarctata) Metopograpsus (Metopograpsus frontalis) Banana prawn (Penaeus merguiensis) Mullet (Mugil cephalus) Mud whelk (Telescopium telescopium) Mangrove snail (Nerita balteata) Oyster (Saccostrea glomerata) Mangrove leaves (Rhizophera stylosa) Seagrass (Zostera capricorni) Macroalgae (Catenella nipae) Filamentous algae (predominantly Lyngbya majuscula) Seston (zooplankton and phytoplankton) Epiphytic algae Particulate organic matter (POM)


57

Mud crab (Scylla serrata) Fiddler crab (Uca coarctata)

Grapsid crab (Metopograpsus frontalis) Banana prawn (Penaeus merguiensis) Oyster (Saccostrea glomerata)

Mullet (Mugil cephalus)

Mud whelks (Telescopium telescopium) Mangrove snails (Nerita balteata)

(All photographs courtesy of Leonie Andersen)

Figure 5.2. Examples of organisms collected as part of the foodweb study


58


5.3.1 Foodweb elucidation

Carbon and nitrogen have more than one isotope, and the isotopic composition of

natural materials such as animal and plant tissue change, in predictable ways, as

these elements cycle upward through the food web (Peterson & Fry 1987). Carbon

isotopes allow organic sources to be traced through consumers (Rodelli et al.

1984), whereas nitrogen isotopes provide information on the trophic level of an

organism (Peterson 1999). Biochemical reactions cause fractionation of stable

isotopes which change the isotopic ratio. For example, in the metabolism of

nitrogen, the light isotope is concentrated in nitrogenous excretion products while

the heavy isotope is retained in the body tissues. As a result, the 15N:14N ratio

increases with trophic level. Isotopic ratios of isotopes 13C/12C and 15N/14N are

expressed as delta (δ) values, which are the relative difference (�, parts per

thousand) between the sample and conventional standard reference materials

(Peterson 1999). Increases in the δ value denote an increase in the amount of

heavy isotope component (the sample is therefore enriched in 13C or 15N) and

therefore will have a heavier δ value. Conversely, a sample depleted in 13C or 15N will have a lighter or more negative δ value.

The mean isotopic values of the collected biological specimens are presented in

Figure 5.3. Mullet, mud crabs and prawns tended to share a similar trophic

position and carbon signature in the food web, relying predominantly on

filamentous algae and to a lesser extent epiphytes, seston and seagrass for their

primary carbon sources. There was a large range in carbon signatures (-13.6 to

-31.5�) among the major primary producers [epiphytic algae, filamentous algae,

seston, POM (sediment organic matter), seagrass, macroalgae and mangrove

leaves]. In some cases the signatures overlapped, for example, POM and seston.

Mangrove carbon contributed to the diet of very few organisms indicating that

very few species rely on mangroves as a predominant food source but are more

likely to be dependent on benthic organic matter and algae. Generally the food

web established in this study for Port Curtis was similar in structure to estuarine

food webs of other authors (Rodelli et al. 1984; Primavera 1996; Thimdee et al.

2001, 2004).

The low contribution of mangroves or mangrove detritus to the diet of prawns in

this study supports the findings of other researchers (Primavera 1996; Thimdee

et al. 2004) that prawns are not predominantly detritivores. Both earlier findings

suggested macro- and microalgae, seagrass, epiphytes or seston, or

combinations of these producers, are carbon sources for both juvenile and


59

adult prawns. The prawns in Port Curtis appear to be consumers of meiofauna

(animals of microscopic size living in marine sediments), which in turn feed on

benthic algae.

-6

-4

-2

0

2

4

6

8

10

12

-35 -30 -25 -20 -15 -10 -5

Delta 13C (�)

Delta

15N

(�)

Macroalgae

Mangrove leaves

Mangrove snails Oysters

Seston

Grapsid crabs

Fiddler crabs

Mullet Mud crabs

Mud whelks

Seagrass

Epiphytes

Filamentous algae

Sediment POM

Prawns

-6

-4

-2

0

2

4

6

8

10

12

-35 -30 -25 -20 -15 -10 -5

Delta 13C (�)

Delta

15N

(�)

Macroalgae

Mangrove leaves


Seston

Grapsid crabs

Fiddler crabs

Mullet Mud crabs

Mud whelks

Seagrass

Epiphytes

Filamentous algae

Sediment POM

Prawns

Macroalgae

Mangrove leaves


Seston

Grapsid crabs

Fiddler crabs

Mullet Mud crabs

Mud whelks

Seagrass

Epiphytes

Filamentous algae

Sediment POM

Prawns

Figure 5.3. Relationship of δ 13C and δ15N (mean ± 1 S.D.) of selected primary producers

and consumers in a Port Curtis food web

There were no ecologically significant site or gender differences among male and

female mud crabs in terms of their carbon and nitrogen signatures, indicating

crabs from all four sites have similar diets and trophic positions. A previous study

(Andersen & Norton 2001) determined there was a significant difference in the

carbon signatures of mud crabs from Port Curtis compared to those from the

reference site in Ayr (North Queensland). The difference could have been due to a

natural spatial variability in the same carbon source such as algae or alternatively

that the mud crabs from the two locations were consuming different foods. The

study also determined that there was a trend for a correlation between

hepatopancreas copper concentrations and muscle δ 13C, suggesting the diets

of mud crabs may be a major source of copper for mud crabs in Port Curtis.

The δ13C of prawns (-19.3 ± 0.1�) was similar to that of the filamentous algae

(-17.8 ± 0.3�), which were identified predominantly as Lyngbya majuscula

(a blue-green algae) with some green algal filaments interspersed within. Visual

confirmation that the prawns were feeding on the algae directly or on smaller

invertebrates that had fed on the algae, were obtained when after 24 h,

breakdown of the algae caused the release of purple pigments (most likely

phycoerythrin and phycocyanin). The same colouring was noted in the digestive

tracts of the prawns (Figure 5.4). The higher δ15N of prawns (8.0 ± 0.3�)


60

compared to the algae (1.1 ± 0.2�) indicates a higher position in the food chain

of approximately two trophic levels, considering fractionation of nitrogen of ~3�

per trophic level. This suggests the prawns were mainly feeding on primary

consumers that may have been feeding directly on the algae. If the prawns are

accumulating toxins contained in the algae, then potentially these toxins may be

transferred up the food chain. Should the toxin accumulate in the prawn muscle

tissue, this could also have human health ramifications for consumers of banana

prawns in Port Curtis.

(Photograph courtesy of Leonie Andersen)

Figure 5.4. Blue green algae (Lyngbya majuscula) demonstrating released pigment (arrowed) and the same pigment observed in the hepatopancreas (liver) of a banana

prawn from the same site (also arrowed)

5.3.2 Metal distributions and relation to food web structure

A summary of mean concentration data for each organism/metal is given in

Table 5.2. Detailed statistical analyses of the data may be found in the report by

Andersen et al. (2005). With the exception of aluminium and arsenic, there were

no noticeable between-site variations in tissue metal concentrations indicative of

�hotspots� or gradients of metal bioaccumulation. On the whole, organisms from

the inner harbour sites�Boat Creek (Site 1), Grahams Creek (Site 2) and Black

Swan (Site 3)�tended to have more elevated metal concentrations than those

from Yellow Patch (Site 4). Aluminium and arsenic data are summarised in

Figures 5.5 and 5.6 respectively. Yellow Patch (reference site) organisms

tended to have the lowest aluminium accumulations compared to all the other

sites but the highest arsenic concentrations. This does not necessarily imply

contamination of arsenic at the reference site, but may reflect the complex

interactions (e.g. antagonistic effects of other metals/chemical species) affecting

metal bioaccumulation.


61

Table 5.2. Mean trace metal concentrations for the organisms collected

Organism Al As Cd Cr Cu Ni Se Zn mg/kg wet weight (± S.D.)

Mud crab hepatopancreas

0.6±0.3 7.8±2.6 0.37±0.32 0.03±0.03 210±134 1.4±1.0 3.6±2.0 36.0±8.3

Oysters 10.0±6.0 5.4±4.7 0.18±0.07 0.08±0.02 118±32 0.8±2.0 1.1±0.2 315±95

Algae 274±194 2.6±0.5 <0.01 0.4±0.3 1.8±0.4 0.8±0.5 0.2±0.1 3.2±0.8

POM 1340±510 7.5±2.4 0.01±0.01 3.7±1.2 9.3±4.7 2.0±0.8 0.1±0.1 8.9±4.2

Seston 2080±1380 29±45 0.03±0.02 4.3±2.7 5.5±3.5 4.1±2.3 0.2±0.1 36±26

Mangrove leaves 47±38 0.1±0.1 <0.01 0.14±0.05 0.4±0.2 0.1±0.1 <0.05 1.5±0.2

Metapograspus 264±98 2.9±0.7 0.01±0.01 0.55±0.17 29.5±5.7 0.2±0.1 0.2±0.2 19.4±2.0

Mullet 2.1±4.1 1.2±0.4 <0.01 0.06±0.05 0.4±0.1 <0.1 0.2±0.1 6.7±2.3

Mud whelks 0.8±0.8 1.9±0.5 0.01±0.01 0.03±0.01 21.4±7.1 0.1±0.1 0.5±0.1 12.7±2.1

Snails 10.3±5.6 2.6±0.9 0.02±0.01 0.13±0.07 2.1±0.5 0.5±0.2 0.8±0.1 16.8±1.5

Seagrass 263±122 0.9±0.6 0.04±0.01 0.39±0.13 2.5±0.7 0.7±0.2 <0.05 4.9±1.7

Correlation of stable isotope data with metal concentration data yielded few

relationships of any significance. An attempt was made to explore the possibility

that different dietary sources of carbon could explain some of the differences in

metal accumulations in mud crabs. The carbon signatures of individual male mud

crabs from all sites were plotted against their individual accumulated metal

concentrations to determine if there was commonality between metal

accumulations and carbons signatures at each site. There was clustering of

some crabs from each site in respect to δ13C and copper, arsenic and zinc

concentrations, respectively (Andersen et al. 2005), suggesting a site relationship

between metal accumulations and carbon signatures. Aside from this, no clear

trends were observed.

Biomagnification of metals describes the increasing accumulation of metals with

increasing trophic level (Reinfelder et al. 1998). Biomagnification has been

identified in some food webs for some contaminants, most notably mercury.

Biomagnification per se was not demonstrated for the eight metals examined in

this study.


62

Aluminium

Figure 5.5. Mean aluminium concentrations (mg/kg wet wt)(± 1S.E.) in biota at three inner harbour sites (1�3) in Port Curtis and an outer harbour reference site (4)

Epiphytes

0100020003000400050006000

1 2 3 4

Site

mg/

kg w

et w

t.

POM

01000200030004000500060007000

1 2 3 4

Site

mg/

kg w

et w

t.

Seston

0

1000

2000

3000

4000

5000

1 2 3 4

Site

mg/

kg w

et w

t.

Oysters

02468

10121416

1 2 3 4

Site

mg/

kg w

et w

t.Mangrove snails

0

5

10

15

20

1 2 3 4

Site

mg/

kg w

et w

t.Mud whelks

0

0.5

1

1.5

2

2.5

1 2 3 4

Site

mg/

kg w

et w

t.

Mullet

012345678

1 2 3 4

Site

mg/

kg w

et w

t.

Metopograspus

0

100

200

300

400

500

1 2 3 4

Site

mg/

kg w

et w

t.

F iddler crab

0

500

10001500

2000

2500

3000

1 2 3 4

Site

mg/

kg w

et w

t.

Mud crab

0

0.2

0.4

0.6

0.8

1

1 2 3 4

Site

mg/

kg w

et w

t.

Mangrove leaves

0

2040

60

80100

120

1 2 3 4

Site

mg/

kg w

et w

t.


63

Figure 5.6. Mean arsenic concentrations (mg/kg wet wt)(± 1 S.E.) in biota at three inner harbour

sites (1�3) in Port Curtis and an outer harbour reference site (4)

Arsenic

E p ip h y te s

0

2

4

6

8

1 2 3 4S ite

mg/

kg w

et w

t.POM

012345678

1 2 3 4

Site

mg/

kg w

et w

t.

Seston

020406080

100120

1 2 3 4

Site

mg/

kg w

et w

t.

Oysters

02468

10121416

1 2 3 4

Site

mg/

kg w

et w

t.

Mangrove snails

0

1

2

3

4

5

1 2 3 4

Site

mg/

kg w

et w

t.

Mud whelks

0

0.5

1

1.5

2

2.5

3

1 2 3 4

Site

mg/

kg w

et w

t.

Mullet

0

0.5

1

1.5

2

1 2 3 4

Site

mg/

kg w

et w

t.

Metopograspus

0

1

2

3

4

5

1 2 3 4

Site

mg/

kg w

et w

t.

F iddler crab

00.5

11.5

22.5

33.5

1 2 3 4

Site

mg/

kg w

et w

t.

Mud crab

02468

101214

1 2 3 4

Site

mg/

kg w

et w

t.

Mangrove leaves

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

1 2 3 4

Site

mg/

kg w

et w

t.

Mangrove leaves

0 0.02 0.04 0.06 0.08

0.1 0.12 0.14

1 2 3 4 Site

mg/

kg w

et w

t.


64

5.4 Conclusions A food web including mud crabs, other crustaceans, fish, molluscs and a variety of

plants was identified in Port Curtis. In general, the food web was not unlike those

established for other estuarine embayments. Mangrove carbon contributed to the

diet of very few organisms. It appears that very few species rely on mangroves as a

predominant food source but are more likely to be dependent on benthic organic

matter and algae. Mud crabs were identified as one of the dominant predators in the

food chain. There were no ecologically significant site or gender differences among

male and female mud crabs in terms of their carbon and nitrogen signatures,

indicating crabs from all four sites have similar diets and trophic positions. Carbon

isotopes suggested that prawns were feeding either directly or indirectly on blue-

green algae (Lyngbya majuscula) and this was supported by observations of

pigment from the algae being visually evident in the prawns. The finding may have

consequences for consumers should the toxin produced by the algae follow similar

uptake pathways to the pigment and accumulate in the prawn muscle tissue.

Although there were very few significant between-site differences in metal

bioaccumulation, organisms from inner harbour sites tended to be more enriched in

metals than those from the unimpacted reference site outside the harbour. More

recent work by the CRC Contaminant Pathways team (reported in this volume) has

determined that dissolved metal concentrations in the water column may actually

increase through The Narrows. Recent findings suggest that The Narrows could be

a sink (or source) for dissolved metals in Port Curtis and that metal concentrations

may not decrease appreciably until outside Port Curtis. In addition, the Port Curtis

hydrodynamic model (Herzfeld et al. 2004) predicted a reduced flushing of the

harbour and a greater retention time of the water body than reported in previous

models. The two factors�elevated dissolved metals and reduced flushing�could

contribute to the anomalous bioaccumulation of metals in biota in inner harbour

sites compared to outer harbour sites recorded in this and previous studies.

In summary, this study highlights the complexity of interactions that are likely to

occur in metal pathways in estuarine food webs. Although uptake of metals from

the dissolved phase is still important, many studies are highlighting the

significance of trophic transfer in metal accumulation by aquatic invertebrates.

The findings of this study indicate that for the majority of organisms, the uptake of

metals through food pathways is likely to be complex and integrated, particularly

for those in higher trophic positions and those that have the ability to regulate

metal accumulations. The adage �you are what you eat� may hold true for carbon

sources, but not necessarily for metals accumulated by consumers in complex

mangrove ecosystems.


65

5.5 References Andersen, L.E. and Norton, J.H. (2001) Port Curtis mud crab shell disease –

nature, distribution and management. FRDC Project No. 98/210, Central

Queensland University, Gladstone, 115 pp.

Andersen, L.E., Revill, A. and Storey, A. (2005) Metal bioaccumulation through

food web pathways in Port Curtis. Technical Report No. 31,CRC for Coastal

Zone, Estuary and Waterway Management, Brisbane, 49 pp.

Andersen, L.E., Boundy, K. and Melzer, A. (2002) Intertidal crabs as potential

biomonitors in Port Curtis. Centre for Environmental Management, Central

Queensland University and Cooperative Research Centre for Coastal Zone,

Estuary and waterway Management, Gladstone, 23 pp.




Coastal Zone, Estuary and Waterway Management, 146 pp.

Fantle, M.S., Dittel, A.I., Schwalm, S.M., Epifanio, C.E. and Fogel, M.L. (1999)

A foodweb analysis of the juvenile blue crab, Callinectes sapidus, using

stable isotopes in whole animals and individual amino acids. Oecologia,

120, 416�426.

Fisher, N.S. and Reinfelder, J.R. (1995) The trophic transfer of metals in marine

systems In: A. Tessier and D.R. Turner (Eds), Metal speciation and

bioavailability in aquatic systems. John Wiley and Sons Ltd, New York,

pp. 363�406.

France, R. (1998) Estimating the assimilation of mangrove detritis by fiddler crabs

in Laguna Joyuda, Puerto Rico, using dual stable isotopes. Journal of Tropical

Ecology, 14, 413�425.

Herzfeld, M., Parslow, J., Andrewartha, J., Sakov, P. and Webster, I.T. (2004)

Hydrodynamic modelling of the Port Curtis region. Technical Report No. 7,

CRC for Coastal Zone, Estuary and Waterway Management, Brisbane, 51 pp.

Kang, C.K., Sauriau, P.G., Richard, P. and Blanchard, G.F. (1999) Food sources

of the infaunal suspension-feeding bivalve Cerastoderma edule in a muddy

sandflat of Marennes-Oleron Bay, as determined by analyses of carbon and

nitrogen stable isotopes. Marine Ecology Progress Series, 187, 147�158.

Peterson, B.J. (1999) Stable isotopes as tracers of organic matter input and

transfer in benthic food webs: a review. Acta Oecologia, 20, 479�487.

Peterson, B.J. and Fry, B. (1987) Stable isotopes in ecosystem studies. Annual

Review of Ecology and Systematics, 18, 293�320.


66

Prange, J.A. (1999) Physiological responses of five seagrass species to trace

metals. B.Sc. honours thesis. Botany Department, University of Queensland,

Brisbane, 52 pp.

Primavera, J.H. (1996) Stable carbon and nitrogen isotope ratios of penaeid

juveniles and primary producers in a riverine mangrove in Guimaras,

Philippines. Bulletin of Marine Science, 58, 675�683.

Reinfelder, J.R., Fisher, N.S., Luoma, S.N., Nichols, J.W. and Wang, W-X. (1998)

Trace element trophic transfer in aquatic organisms: a critique of the kinetic

model approach. Science of the Total Environment, 219, 117�135.

Rodelli, M.R., Gearing, J.N., Gearing, P.J., Marshall, N. and Sasekumar, A. (1984)

Stable isotope ratio as a tracer of mangrove carbon in Malaysian ecosystems.

Oecologia, 61, 326�333.

Thimdee, W., Deein, G., Sangrungruang, C. and Matsunaga, K. (2004) Analysis of

primary food sources and trophic relationships of aquatic animals in a

mangrove-fringed estuary, Khung Krabaen Bay (Thailand) using dual stable

isotope techniques. Wetlands Ecology and Management, 12, 135�144.

Thimdee, W., Deein, G., Sangrungruang, C. and Matsunaga, K. (2001) Stable

carbon and nitrogen isotopes of mangrove crabs and their food sources in a

mangrove-fringed estuary in Thailand. Benthos Research, 56, 73�80.


67

Contaminant pathways in Port Curtis: Final report 6: Occurrence of imposex

68

Chapter 6 Occurrence of imposex in Port Curtis

6.1 Introduction Tributyltin (TBT) is a broad spectrum biocide used to coat the bottom of ships to

prevent attachment of marine organisms. Introduced in the 1960s, it is one of the

most effective antifouling agents ever developed. However, it became apparent in

the 1970s that leachates of organotin compounds were having deleterious effects

on non-target organisms such as gastropods. The extensive use of TBT on all

types of boats led to the virtual collapse of oyster farming industries in several

countries during the 1980s. The harmful effects of organotin compounds were

recognised in 1989 by the International Maritime Organisation (IMO), of which

Australia is a member. In 1990 the IMO recommended a ban on the use of TBT

on vessels less than 25 m in length, as well as the elimination of all paints with a

leaching rate of more than four µg of TBT per day. In 1999, the IMO adopted a

resolution calling for a global ban on the application of all organotin compounds by

2003 and a complete prohibition by 2008 (IMO 2002).

The issue of elevated butyltin concentrations in Port Curtis was first identified

during the early stages of the Port Curtis contaminant risk assessment (Apte et al.

2005). Water column TBT concentrations were above the trigger value of 0.006 µg

Sn/L, although still much lower than concentrations in many world harbours. TBT

concentrations were elevated in 13% of the 56 sediment samples analysed, but

again were low compared to severely polluted harbours in other parts of the world.

TBT was also found to have bioaccumulated in resident oysters, mud whelks and

mud crabs from Port Curtis, indicating exposure of these organisms.

A follow-up study was therefore undertaken to characterise the effects of TBT

exposure on marine organisms resident in Port Curtis. An imposex survey of the

gastropod Morula marginalba (mulberry whelk) was conducted. Full details of this

study may be found in the publications by Andersen and co-workers (Andersen et

al. 2004a, b). A summary is given below.

6.2 Imposex in marine gastropods There have been a number of deleterious impacts of TBT on non-target

organisms, most notably, the imposex phenomenon in marine gastropods.

Imposex is the imposition of male sexual characteristics (notably a penis) on

female marine snails. Reproductive failure and death of affected females can


69

occur, with the eventual decimation of entire populations of severely affected

snails. The term �imposex� was first coined by Smith (1971) to describe the

imposition of male characteristics on the female intertidal mud snail, Nassarius

obsoletus and was subsequently linked to the presence of TBT (Smith 1981).

Bryan et al. (1986, 1987) later confirmed the association of imposex with TBT

through a series of laboratory and field transplant studies with the gastropod

Nucella lapillus. Alzieu (1991) noted that imposex had been described in over

72 species belonging to 49 genera, with more species being identified every year.

Imposex occurs at extremely low (parts per trillion) concentrations of ambient TBT

contamination and is considered a sensitive bioindicator of the effects of TBT

exposure (Gibbs et al. 1987).

Gibbs and Bryan (1986) described three stages of imposex development: an

�early� stage involving the formation of a vas deferens and small penis; an

�intermediate� stage characterised by an enlarged female penis approaching the

size of a male penis; and a �late� stage where there is blockage of the female

opening preventing the release of egg capsules. In the latter stages, reproductive

failure and most likely premature death of the female occur. As the development of

the vas deferens precedes that of the penis in Nucella lapillus, Gibbs et al. (1987)

developed the vas deferens sequence (VDS) index, which categorises six stages

of imposex development. The above described indices or modified versions of

them have since been accepted and used to measure imposex in other species

worldwide (Liu et al. 1997; Tan 1999; Ramon and Amor 2001; Terlizzi et al. 2004).

6.3 Experimental Morula marginalba Blainville (1882) (Order Neogastropoda, Family Muricidae)

commonly known as the �mulberry whelk� (Figure 6.1) is the major carnivorous

predator of macro-invertebrates in the mid-intertidal zone of rocky shores on the

Australian east coast. (Moran 1985). M. marginalba and its close relative Morula

granulata have been previously used as bioindicators of TBT contamination in

Australia (Wilson et al. 1993; Reitsema and Spickett 1999). As M. marginalba is

extremely abundant in Port Curtis, it appeared to be a suitable species to

determine the distribution and severity of imposex in Port Curtis.

Whelks were collected from ten selected sites in Port Curtis (Table 6.1) over a

two-day period in October 2003. Where possible, at least 100 whelks were

collected from each site. Sites were distributed in an array fashion with increasing

distance from major shipping activity (Table 6.1, Figure 6.2). Sites were located in

the inner harbour (wharf sites), middle harbour (adjacent to shipping channels)


70

and outer harbour (reference sites) to establish if there were differences in the

frequency and severity of imposex that could be associated with differences in

shipping intensity.

(Photo courtesy of Leonie Andersen)

Figure 6.1. The mulberry whelk, Morula marginalba

Table 6.1. Site locations and shipping intensity in Port Curtis

Site Location Array Shipping intensity/type

1 Clinton Coal wharf (CCW) Inner Major, large vessels 2 Tide Island Middle Minor, all vessels 3 Tug berth Inner Major, large vessels 4 QCL berth Inner Major, large vessels 5 Quoin Island Middle Minor, all vessels 6 BSL berth Inner Major, large vessels 7 Worthington Island Reference Occasional, small vessels 8 Rat Island Middle Minor, small/medium vessels 9 Gatcombe Head Middle Minor, all vessels 10 Blackhead Reference Occasional, small vessels

The percentage of females affected at each site, the length of the female

pseudopenis, shell length and length of male penes were recorded. The identity

of a subsample of whelks was verified as Morula marginalba by Queensland

Museum. Female pseudopenis length was difficult to measure in some females

due to the small size and non-uniform shape of the female penis. Additional

observations on penis length and vas deferens development were therefore made

and a grading system developed which reflected the extent of penis development

(Table 6.2).

3.5 mm


71

Figure 6.2. Collection sites for M. marginalba at ten sites in Port Curtis in relation to the shipping channel

Table 6.2. Imposex grading system for M. marginalba

Grade Female penis description

1 <0.5 mm raised discoloured area up to a bump (with or without vas deferens) but unable to be measured

2 0.5 mm � <1.0 mm 3 1.0 mm � <2.0 mm 4 2.0 mm and greater

6.4 Results and discussion A summary of the field collection data is presented in Table 6.3. A number of

female M. marginalba were found to exhibit the imposex phenomenon

(Figure 6.3). The highest incidence of imposex was found at the sites having the

highest shipping intensity (BSL berth, Clinton Coal Wharf, Tug berth and QCL

berth. The highest imposex frequency (43%) was found at the BSL berth (Site 6),

which averaged 39 vessels totalling 473 000 tonnes per year for the last ten years

(Figures 6.2 and 6.4). Imposex was absent at the reference sites. There was also

a significant positive relationship between imposex frequency and the array of

shipping intensity (r= -0.705, p=0.023), but not between imposex frequency and

distance to the major shipping channel.

PORT CURTIS

10 Kilometres


72

The incidence reported here (0�43 %) is low in comparison to previous studies

carried out in Australian waters of this and other species in which up to 100% of

females were affected (Foale 1993; Wilson et al. 1993; Reitsema & Spickett 1999;

Gibson and Wilson 2003). The severity of imposex (based on grade) was not

severe in comparison to other studies and indicates that the degree of imposex in

this population was in the early stages. The majority of penes did not form a

measurable bud, with the largest pseudopenis of 2.5 mm being significantly smaller

then the average male penis of 8.9 mm. A higher frequency would most likely have

occurred in the Gladstone marina where some of the highest water and sediment

TBT concentrations have been previously found (Apte et al. 2005). Unfortunately no

whelks were available to sample in this location. Reitsema and Spickett (1999)

found a similar frequency of imposex (0�57%), for M. granulata in a survey of the

Dampier Archipelago, Western Australia, the largest tonnage port in Australia.

Wilson et al. (1993) found a correlation on the east coast of NSW, between the

amount of boating activity (high, medium and low) and the degree to which the

M. marginalba population were affected. Reitsema and Spickett (1999) also noted a

correlation between distance to nearest vessel activity and imposex in M. granulata,

a very close relative of M. marginalba. The relationship has been supported by the

majority of imposex surveys conducted worldwide, especially those undertaken

prior to or shortly after 1990, when TBT use was banned on vessels <25 m (Bryan

et al. 1987; Foale 1993; Gibbs & Bryan 1996; Reitsema & Spickett 1999).

Although leaching of antifouling paint on vessels at wharves is likely to be a major

source of organotin pollution, shipyard activities such as hull painting, slipways

and paint removal offer an alternate source. A large number of non-merchant

vessels also access other parts of the harbour including the northern harbour

entrance. It is not surprising, therefore, to find imposex at sites other than

commercial moorings, indicating the widespread contamination of TBT. This could

explain the 4% imposex frequency at Rat Island (Site 8) adjacent to the northern

entrance, which serves as a passage to trawlers, island ferry services and supply

barges >25 m in length that are still legally able to use TBT.

Species differences in the sensitivity to TBT (Wilson et al. 1993; Tan 1999) and

the bioaccumulation of TBT (Liu et al. 1997) have been demonstrated in other

surveys. Liu et al. (1997) found that imposex was much more severe in Thais

species than Morula despite similar organotin burdens, and suggested a genus-

specific susceptibility to organotin pollution with the ranking order of Nucella, Thais

and Morula. Differences in habitat (e.g. high-shore versus low-shore species), diet

and physiology have been suggested for interspecific differences in imposex

(Tan 1999).


73

The life cycle of the gastropod also has a bearing on the impact of TBT at the

population level. The decline of N. lapillus in populations severely affected by

imposex has been attributed to a reduction in recruitment caused by a decrease

in reproductive capacity (Bryan et al. 1986). This gastropod does not have a

planktonic larval stage and apart from the small number of juveniles that may reach

other sites via pieces of floating debris, maintenance of a population relies solely on

its ability to reproduce. M. marginalba, however, has a long-term planktonic larval

stage (Underwood 1974), allowing recruitment of individuals from other locations.

Therefore complete decimation of this species at severely affected sites is unlikely.

In conclusion imposex was found to be present in M. marginalba collected from

Port Curtis confirming a sublethal, biological response to TBT exposure. Although

related to shipping intensity, the frequency and grade of the imposex condition

were not severe in comparison to other port surveys. Due to the mulberry whelks�

ability to recruit juveniles from unaffected locations, conservation of the species is

highly likely despite the effects of imposex. Other more TBT-sensitive species

such as Thais which have non-planktonic larval stages may be more affected.

Some subsequent re-surveying studies (post TBT ban) in Australia have noted an

overall trend for decline in either imposex frequency and/or severity in major ports

and coast sites since the introduction of the ban (Gibson & Wilson 2003; Reitsema

et al. 2003). Globally, the condition is likely to slowly improve with the introduction

of further restrictions on the use of TBT in 2008.


Figure 6.3. Imposex in M. marginalba with penis bud (arrowed)

RT = right tentacle, VD = vas deferens

0.75 mm

RT

VD


74

Table 6.3. Field data for M. marginalba in Port Curtis

Site

Location

Number examined

% Females

% Imposex

Mean grade imposex

Mean shell length (mm ± S.D.)

Mean male penis length (mm ± S.D.)

1 Clinton Coal wharf 100 55 9 1 22 ± 1.3 10.4 ± 1.4 2 Tide Island 100 60 2 1 26 ± 2.4 10.4 ± 1.0 3 Tug berth 100 69 17 1.2 20 ± 1.1 7.8 ± 1.2 4 QCL berth 100 65 17 1.1 29 ± 2.4 11.0 ± 0.7 5 Quoin Island 100 63 3 1 21 ± 1.7 8.7 ± 1.3 6 BSL berth 91 86 43 2.4 19 ± 3.4 6.0 ± 0.9 7 Worthington Island 50 62 0 0 30 ± 2.7 10.6 ± 1.1 8 Rat Island 100 51 4 2.5 18 ± 1.8 7.4 ± 1.3 9 Gatcombe Head 100 57 5 2 16 ± 1.4 7.9 ± 1.4 10 Blackhead 100 66 0 0 17 ± 1.4 8.7 ± 1.2

0

5

10

15

20

25

30

35

40

45

50

1 -Clint

on co

al wha

rf IA

3 -Tug

berth

IA

4 -QCL b

erth I

A

6 -BSL b

erth I

A

2 -Tide

Islan

d MA

5 -Quo

in Isla

nd M

A

8 -Rat

Island

MA

9 -Gatc

ombe

Head M

A

7 -Wort

hingto

n Isla

nd REF

10 -B

lack H

ead R

EF

Site

Impo

sex

freq

uenc

y (%

)

Figure 6.4. Imposex frequency in female M. marginalba at ten sites (1�10) in Port Curtis in 2003


75

Figure 6.5. Frequency of imposex in M. marginalba at ten sites in Port Curtis in 2003 in relation to the major shipping channel

Site 10

Site 6

Site 5 Site 3

Site 7

Site 4

Site 2

Site 1

Site 8

Site 9

IMPOSEX FREQUENCY %

0 (Reference) 1 �5 (Middle array) 6 � 20 (Inner array)

21 � 45 (Inner array)

Shipping channel Dredge spoil ground

GLADSTONE

FACING ISLAND

CURTIS ISLAND

Kilometres

Site legendSite 1 � Clinton Coal wharf

Site 2 � Tide Island

Site 3 � Tug berth

Site 4 � QCL berth

Site 5 � Quoin Island

Site 6 � BSL berth

Site 7 � Worthington Island

Site 8 � Rat Island

Site 9 � Gatcombe Head

Site 10 � Black Head


76

6.5 References Andersen, L. (2004a) Imposex: A biological effect of TBT contamination in Port

Curtis, Queensland. Australasian Journal of Ecotoxicology, 13, 5�-61.

Andersen, L. (2004b). Imposex in the city: a survey to monitor the effects of TBT

contamination in Port Curtis, Queensland. Technical Report CP20, CRC for

Coastal Zone, Estuary and Waterway Management, Brisbane, 25 pages.

Alzieu, C. (1991) Environmental problems caused by TBT in France: assessment,

regulations, prospects. Marine Environmental Research, 32, 7�17.





Bryan, G.W., Gibbs, P.E., Burt, G.R. and Hummerstone, L.G. (1987) The effects of

tributyltin (TBT) accumulation on adult dog-whelks, Nucella lapillus: long-term

field and laboratory experiments. Journal of the Marine Biological Association

of the UK, 67, 525�544.

Bryan, G.W., Gibbs, P.E., Hummerstone, L.G. and Burt, G.R. (1986) The decline

of the gastropod Nucella lapillus around south-west England: evidence for the

effect of tributyltin from antifouling paints. Journal of the Marine Biological

Association of the UK, 66, 611�640.

Foale, S. (1993) An evaluation of the potential of gastropod imposex as a

bioindicator of tributyltin pollution in Port Phillip Bay, Victoria. Marine Pollution

Bulletin, 26, 546�552.

Gibbs, P.E. and Bryan, G.W. (1986) Reproductive failure in populations of the

dog-whelk, Nucella lapillus, caused by imposex induced by tributyltin from

antifouling paints. Journal of the Marine Biological Association of the UK, 66,

767�777.

Gibbs, P.E. and Bryan, G.W. (1996) TBT-induced imposex in neogastropod snails:

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79

Contaminant pathways in Port Curtis: Final report 7: Antioxidant enzymes as biomarkers

80

Chapter 7 Antioxidant enzymes as biomarkers of environmental stress in oysters in Port Curtis

7.1 Introduction The screening level risk assessment of contaminants in Port Curtis (Apte et al.

2005) found that concentrations of metals in sediments and dissolved metals in

the waters were generally below levels of regulatory concern. However,

concentrations of a variety of metals were significantly enriched in marine biota in

comparison with organisms sampled at reference sites. Studies prior to this had

also flagged concentrations of some metals, in particular, copper and zinc in mud

crabs (Andersen & Norton 2001) and copper in seagrass (Prange 1999) and

fiddler crabs (Andersen et al. 2002), as potentially anomalous in Port Curtis

relative to background levels. The demonstration of bioaccumulation of a

contaminant, however, does not necessarily mean that organisms will display

adverse effects. There is a need to demonstrate a link between exposure and an

adverse biological response.

Biomarkers are biochemical, physiological or histological changes that measure

sublethal effects of, or exposure to, toxic chemicals (Weeks 1995; Luebke et al.

1997), and generally but not exclusively pertain to a response at a specific organ,

cellular or subcellular level of organisation (O'Halloran et al. 1998). These cellular

and molecular responses can be used as early warning signals of environmental

stress, before whole organism effects become apparent (Regoli et al. 1998).

Environmental pollutants generally cause an increase in peroxidative processes

within cells, causing oxidative stress (Winston & Giulio 1991; Cheung et al. 2001;

Nusetti et al. 2001). Lipid peroxidation (LPO) has often been used as a biomarker

of environmental stress, reflecting damage to cell membranes from free radicals

(Ringwood et al. 1999). The extent of damage caused by oxyradical production is

dependent on antioxidant defences, which include antioxidant enzymes and free

radical scavengers, such as glutathione (Doyotte et al. 1997). Therefore,

antioxidant enzymes are some of the most common biomarkers used in

environmental monitoring (Regoli et al. 1998).

The enzymes usually respond rapidly and sensitively to biologically active

pollutants (Fitzpatrick et al. 1997). Some of the most commonly used antioxidant

enzyme biomarkers include catalase (CAT) and glutathione-s-transferase (GST)

(Winston & Giulio 1991; Regoli & Principato 1995; Regoli et al. 1998). Glutathione

(GSH) is often used in biomarker studies, as it is an overall modulator of cellular


81

homeostasis (Ringwood et al. 1999). Glutathione (GSH) is a low molecular weight

scavenger of oxygen radicals which is often found to be depleted in contaminant-

exposed organisms (Regoli et al. 1998).

The major objective of this study was to determine whether selected biomarkers

can be used as bioindicators of metal-induced stress in oysters, both in the field

and in the laboratory. The laboratory experiment was used to establish clear

cause�effect relationships without any confounding variables often found in the

field environment, and without the presence of unknown mixtures of contaminants.

Copper was selected for the laboratory exposures because, in addition to being

identified as a contaminant of concern in Port Curtis (Andersen & Norton 2001;

Andersen et al. 2005a), the metal had been shown to induce strong biomarker

responses in other studies (Regoli & Principato 1995; Doyotte et al. 1997; Regoli

et al. 1998; Brown et al. 2004). Bivalves have also been successfully used in

biomarkers studies, showing significant variation in a range of biochemical

markers, in both gill and digestive gland tissues (Cheung et al. 2001; Cheung

et al. 2002; Irato et al. 2003).

In this study, the Sydney rock oyster (Saccostrea glomerata) was chosen as a

suitable biomonitor. Oysters are suspension feeders and take up metals both

directly from sea water and from suspended particles collected during feeding

(Rainbow 1995). Due to their ability to accumulate contaminants, oysters have

been successfully used as biomonitors in many pollution assessment studies in

Port Curtis (Andersen et al. 2003; Andersen et al. 2004; Andersen et al. 2005b),

and elsewhere (Odzak et al. 2001). The use of transplanted oysters has several

advantages and has been used successfully in several previous studies (Curran

et al. 1986; Chan et al. 1999) including those in Port Curtis mentioned previously.

7.2 Experimental The field component of the research involved the measurement of biomarkers

(CAT, LPO, GSH and GST) and metal concentrations in oysters deployed at two

sites; one in the inner harbour area and the other outside of Port Curtis. Both

sites have been monitored previously for other research in Port Curtis. Site 1 is

considered an impacted site located adjacent to the Fisherman�s Landing trade

waste effluent outfall (Figure 7.1) where metal bioaccumulation has been

demonstrated (Andersen et al. 2005b). Site 2 is relatively pristine, located on the

oceanic side of Curtis Island. Previous studies indicate metal bioaccumulation in

this area to be low (Andersen et al. 2005a).


82

The oysters used in the experiments (Saccostrea glomerata) were obtained from a

commercial lease located in Moreton Bay, Queensland. Oysters were deployed in

a series of mesh bags (18 oysters per bag), with seven bags deployed per site.

The bags were attached approximately 0.5 m below the water surface, to

anchored buoys (Figure 7.2). One bag was collected from each site on days 3, 5,

8, 12, 15, 22 and 29 following deployment. Ten of the retrieved oysters were used

for biomarker analysis and six oysters for tissue metal analysis (two oysters

pooled to form one replicate).

On one occasion, the same number of resident oysters from both sites were

collected from adjacent rocks for both biomarker and metal concentration

analyses. Resident oysters from Site 1 were identified as the same species of

oysters as those from the lease (Saccostrea glomerata); however the dominant

oyster sampled at Site 2 was a different, but closely related species (Saccostrea

cucullata).

Figure 7.1. Location of Sites 1 and 2 for oyster field experiments in Port Curtis Harbour

Kilometres


83


Figure 7.2. Individual bags of oysters attached to buoys ready for deployment

A laboratory bioassay experiment (Figure 7.3) was undertaken in order to

determine the effect of dissolved copper exposure on biomarker response

(CAT, LPO, GSH and GST) in oysters. Full details of these experiments may be

found in the report by Andersen et al. (2006). Briefly, the experiment involved a

copper exposure phase (21 days) and a depuration phase (7 days). Oysters

were maintained in aerated tanks containing 10 L filtered sea water at 25°C,

with a 12:12 h light:dark cycle. Each tank contained typically 26 oysters. The

oysters were fed three times a week with 200 mL of cultured marine algae,

Nanochloropsis occulata.


Figure 7.3. Oysters in treatment tanks in copper bioassay


84

Oysters were acclimated for 7 days in clean sea water in the laboratory prior to

the start of the experiment. The seawater treatments were prepared by spiking

filtered seawater (background copper concentrations: ~3 µg/L) with inorganic

copper to attain final nominal added copper concentrations of 0, 3.75, 7.5, 15

and 30 µg/L in addition to the background concentrations. Each treatment was

replicated five times. Treatment water was renewed three times weekly. Samples

for tissue metals and biomarker analysis were taken on days: 2, 5, 8, 12, 15, 23

and 28. Ten oysters were used for biomarker analysis and six oysters (two

oysters pooled to form one composite) were sampled for metal analysis.

After removal from the field or from the laboratory treatment tanks, oysters were

dissected and gills and hepatopancreas removed then placed into centrifuge tubes

and immediately frozen on dry ice. Samples were then stored frozen in liquid

nitrogen (-80 ºC) before transportation on dry ice to City University, Hong Kong,

for biomarker analysis. Biomarker analysis in both the dissected gill and

hepatopancreas samples was carried out using a method based on the

procedures developed by Cheung et al. (2001).

Oysters collected for metal analysis were frozen whole until processing. The

samples were thawed overnight in a refrigerator, then the soft tissue extracted

from the shell and blotted dry. The tissues of the two replicate oysters from each

treatment tank were pooled to form one composite sample, placed in polyethylene

jars, and frozen until analysis at Griffith University, Queensland. The samples

were analysed using inductively coupled plasma mass spectrometry (ICP-MS).

Preparation and chemical digestion of oyster tissues followed a method similar to

that used by Andersen et al. (1996).


7.3.1 Oyster metal concentrations

Tissue metal concentrations are summarised in Table 7.1. Metal concentrations

displayed very few convincing trends with time. The tissue concentrations of

aluminium, copper, zinc and chromium after 29 days of deployment were

significantly greater than those at Site 2 (Table 7.1). Conversely, concentrations

of arsenic and nickel were significantly more elevated in oysters from Site 2.


85

Table 7.1. Mean ±1 S.E. concentration of metals in oysters at Sites 1 and 2 throughout the 29-day deployment period, including one collection of resident oysters from each site

Site Day Cu Zn As Cd Pb Al Cr Ni (µg/g dry wt) 1 0 53 ± 7 623 ± 52 11 ± 0 3 ± 0 0.4 ± 0.0 131 ± 44 0.8 ± 0.1 0.8 ± 0.0 3 51 ± 12 520 ± 151 10 ± 1 3 ± 1 0.2 ± 0.0 108 ± 35 0.8 ± 0.1 0.8 ± 0.1 5 95 ± 19 881 ± 190 11 ± 1 4 ± 1 0.2 ± 0.0 73 ± 48 0.8 ± 0.1 0.9 ± 0.1 8 60 ± 4 568 ± 77 11 ± 1 2 ± 0 0.2 ± 0.0 93 ± 5 0.7 ± 0.1 0.8 ± 0.1 12 58 ± 2 665 ± 70 10 ± 0 2 ± 0 0.2 ± 0.0 88 ± 3 0.6 ± 0.1 0.9 ± 0.0 15 68 ± 10 706 ± 124 10 ± 2 2 ± 0 0.2 ± 0.0 86 ± 3 0.6 ± 0.0 0.9 ± 0.1 22 74 ± 12 602 ± 37 10 ± 1 2 ± 1 0.1 ± 0.0 51 ± 14 0.5 ± 0.1 0.8 ± 0.1 29 138 ± 48 967 ± 162 9 ± 1 4 ± 0 0.2 ± 0.1 61 ± 2 0.8 ± 0.2 1.2 ± 0.1

Resident 583 ± 91 2563 ± 182 8 ± 0 1 ± 0 0.1 ± 0.0 28 ± 5 0.6 ± 0 1 ± 0 2 0 53 ± 7 623 ± 52 11 ± 0 3 ± 0 0.4 ± 0.0 131 ± 44 0.8 ± 0.1 0.8 ± 0.0 3 51 ± 4 485 ± 70 15 ± 1 4 ± 1 0.2 ± 0.1 91 ± 17 0.8 ± 0.0 1.1 ± 0.1 5 51 ± 2 386 ± 37 15 ± 2 3 ± 1 0.2 ± 0.0 69 ± 20 0.7 ± 0.0 0.8 ± 0.0 8 49 ± 8 373 ± 43 16 ± 1 3 ± 0 0.2 ± 0.0 30 ± 7 0.6 ± 0.1 1.2 ± 0.2 12 58 ± 30 549 ± 311 11 ± 2 3 ± 1 0.2 ± 0.0 25 ± 4 0.5 ± 0.1 1.3 ± 0.2 15 49 ± 4 374 ± 77 14 ± 2 3 ± 1 0.2 ± 0.0 43 ± 13 0.6 ± 0.1 1.2 ± 0.0 22 53 ± 18 429 ± 231 16 ± 5 4 ± 1 0.2 ± 0.0 50 ± 25 0.6 ± 0.0 1.5 ± 0.1 29 40 ± 7 339 ± 78 14 ± 0 2 ± 1 0.2 ± 0.1 23 ± 6 0.5 ± 0.1 1.2 ± 0.2

Resident 256 ± 16 490 ± 71 31 ± 2 1 ± 0 0.1 ± 0.1 38 ± 21 0.7 ± 0.1 1.7 ± 0.2

The deployed oysters did not attain the same metal concentrations as the resident

oysters (Table 7.1). At Site 1, deployed oysters on day 29 contained only one-

quarter of the copper, and approximately one-third of the zinc of the resident

oysters. Previous studies (Andersen et al. 2003; Andersen et al. 2004; Andersen

et al. 2005b) have used deployment periods of eight to ten weeks and have found

this time period to be sufficient to allow significant separation of sites in terms of

metal accumulation, representative of the environmental conditions in Port Curtis.

This is substantially longer than the 28-day deployment in the current study.

However, it is unlikely that deployed oysters could accumulate to the same degree

as resident oysters due to the differences in environmental exposure history.


86

7.3.2 Oyster biomarker concentrations

Since the gills are the first point of contact for metal exposure and the digestive

gland (i.e. hepatopancreas) is an important organ to which metals are known to

sequester, these tissues were chosen to measure biomarker responses. The

�ideal response� was deemed to be an easily measurable biomarker increase

(or decrease) with time, associated with the bioaccumulation of metals.

Concentrations of biomarkers at both sites were variable over the deployment

period with responses in the gill and hepatopancreas not necessarily following

the same patterns (Table 7.2, Figure 7.4). The only enzyme to demonstrate a

significant relationship with time was CAT in hepatopancreas at Site 1 and CAT

in both tissues at Site 2. In hepatopancreas tissue, CAT tended to follow the

same pattern with similar concentrations at both sites and with a substantial initial

increase in concentrations from baseline to three days, which continued to be

maintained.

Handling stress in oysters may also have an effect on some enzyme responses.

A comparison of biomarker concentrations of oysters a) within hours of collection

from the lease and b) baseline oysters prior to deployment or allocation to

acclimation facilities (Andersen et al. 2006), determined that there was a large

decline in CAT in hepatopancreas from when oysters were sampled at the lease to

their arrival two days later and prior to deployment in the field or allocation to the

bioassay. The low initial concentration of CAT in hepatopancreas may therefore

be due to transportation stress and may be considered an anomaly rather than a

true baseline reference point. It is interesting to note that at both sites the

concentration of biomarkers in the resident organisms was generally higher than in

the transplanted oysters (Table 7.2). This may indicate increased stress owing to

higher metal burdens or historical exposure to other environmental stressors.

The concentration of biomarkers in field samples varied consistently across the

tissue types (Table 7.2, Figure 7.4). Both CAT and GST were generally more

elevated in the hepatopancreas, whereas GSH was at slightly lower concentrations

in the hepatopancreas than in the gills. LPO was found at similar concentrations in

both the gills and hepatopancreas.

Several significant correlations were found between enzyme concentrations and

metal concentrations in deployed oysters (Pearson Product Moment correlations,

see Table 7.3). CAT and LPO exhibited significant linear responses to increased

concentrations of certain metals, namely aluminium, cadmium, chromium, copper

and nickel with the majority of responses at Site 1, the more impacted site. This

indicates that some biomarker responses could be associated with accumulated


87

metal concentrations. In some cases there was an initial increase in biomarker

concentration followed by a decline, which could indicate adaptation or acclimation

of the oyster to the new ambient environment.

Table 7.2. Concentrations (µmol/g) of antioxidant enzymes in oysters including residents at Sites 1 and 2 throughout the deployment period. N=10 except where 8 (n=9) due to insufficient

protein in the sample for analyses

Catalase Lipid peroxidase

Glutathione-S-transferase

Glutathione Site Day

Gills Hepato Gills Hepato Gills Hepato Gills Hepato 1 0 *1475 ± 138 739 ± 103 54 ± 4 81 ± 5 *48 ± 4 125 ± 10 *16 ± 3 10 ± 3 3 1455 ± 187 4784 ± 392 87 ± 2 75 ± 6 48 ± 2 121 ± 15 14 ± 2 6 ± 0 5 1541 ± 78 5347 ± 672 85 ± 5 74 ± 9 23 ± 5 123 ± 8 15 ± 1 4 ± 1 8 *1510 ± 116 *6095 ± 933 *74 ± 3 69 ± 14 *47 ± 3 *129 ± 13 14 ± 2 4 ± 1 12 1423 ± 162 6210 ± 919 62 ± 3 65 ± 4 41 ± 3 95 ± 7 14 ± 2 3 ± 1 15 1419 ± 134 4769 ± 363 65 ± 5 69 ± 5 49 ± 5 133 ± 7 10 ± 1 9 ± 1 22 1420 ± 149 5895 ± 756 80 ± 4 63 ± 7 50 ± 4 135 ± 12 14 ± 2 10 ± 1 29 1573 ± 139 3130 ± 379 70 ± 4 78 ± 4 49 ± 4 109 ± 11 16 ± 2 5 ± 1 Resident 1515 ± 178 4347 ± 959 183 ± 31 121 ± 11 31 ± 4 200 ± 33 21 ± 6 12 ± 2 2 0 *1475 ± 138 739 ± 103 54 ± 5 81 ± 5 *48 ± 4 125 ± 10 *16 ± 3 10 ± 3 3 1277 ± 157 5348 ± 498 56 ± 8 95 ± 5 24 ± 2 112 ± 12 14 ± 1 8 ± 1 5 1496 ± 175 4636 ± 697 72 ± 7 90 ± 9 81 ± 9 138 ± 9 23 ± 6 10 ± 2 8 1705 ± 170 6356 ± 968 50 ± 4 64 ± 4 52 ± 4 112 ± 10 14 ± 2 7 ± 1 12 1902 ± 108 6074 ± 568 63 ± 9 66 ± 4 61 ± 5 101 ± 7 16 ± 2 10 ± 1 15 1857 ± 190 *4879 ± 667 63 ± 10 78 ± 10 70 ± 4 117 ± 7 17 ± 2 10 ± 2 22 1704 ± 264 *5549 ± 744 68 ± 10 61 ± 6 76 ± 5 138 ± 13 16 ± 2 10 ± 1 29 1438 ± 169 3361 ± 231 53 ± 4 70 ± 5 34 ± 2 93 ± 7 19 ± 2 8 ± 1 Resident 2922 ± 413 *15468 ± 3372 163 ± 20 142 ± 21 50 ± 13 234 ± 27 25 ± 4 14 ± 2


88

a) CAT Gill b) CAT Hepatopancreas

Time (day)

0 5 10 15 20 25

Gill

CAT

con

cent

ratio

ns (u

mol

/g)

1000

1200

1400

1600

1800

2000

2200Site 1Site 2

Time (day)

0 5 10 15 20 25

Hep

atop

ancr

ease

CAT

con

cent

ratio

ns (u

mol

/g)

0

1000

2000

3000

4000

5000

6000

7000

8000Site 1Site 2

c) LPO Gill d) LPO Hepatopancreas

Time (day)

0 5 10 15 20 25

Gill

LPO

con

cent

ratio

n (u

mol

/g)

40

50

60

70

80

90

100Site 1Site 2

Time (day)

0 5 10 15 20 25

Hep

atop

ancr

eas

LPO

con

cent

ratio

n (u

mol

/g)

50

60

70

80

90

100

110

Site 1Site 2

e) GST Gill f) GST Hepatopancreas

Time (day)

0 5 10 15 20 25

Gill

GST

con

cent

ratio

n (u

mol

/g)

0

20

40

60

80

100Site 1Site 2

Time (day)

0 5 10 15 20 25

Hep

atop

ancr

eas

GS

T co

ncen

tratio

n (u

mol

/g)

80

90

100

110

120

130

140

150

160Site 1Site 2

g) GSH Gill h) GSH Hepatopancreas

Time (day)

0 5 10 15 20 25

Gill

GSH

con

cent

ratio

n (u

mol

/g)

5

10

15

20

25

30Site 1Site 2

Time (day)

0 5 10 15 20 25

Hep

atop

ancr

eas

GS

H c

once

ntra

tion

(um

ol/g

)

2

4

6

8

10

12

14

16Site 1Site 2

Figure 7.4. Mean ±1 S.E. concentration (µmol/g) of biomarkers in oysters from

Site 1 and Site 2 over time


89

Table 7.3. Correlations between metal concentrations and enzyme concentrations in gills and hepatopancreas of oysters in Sites 1 and 2. Only significant correlations shown (α = 0.05)

Site Metal Tissue Enzyme R value P value

1 Copper Gill CAT 0.722 0.043 Aluminium Hepatopancreas CAT -0.736 0.038 Aluminium Hepatopancreas LPO 0.788 0.020 Chromium Hepatopancreas LPO 0.895 0.003 Cadmium Gill CAT 0.860 0.006 Chromium Gill CAT 0.735 0.038

2 Chromium Hepatopancreas LPO 0.766 0.016 Nickel Hepatopancreas LPO -0.744 0.036

7.3.3 Laboratory bioasssay

Tissue copper concentrations in oysters generally increased with exposure to

dissolved copper (Figure 7.5). The concentration of biomarkers (Table 7.4)

showed a similar pattern to the field-deployed oysters. CAT and GST

concentrations were higher in the hepatopancreas than the gills, while GSH

concentrations were slightly higher in the gills than the hepatopancreas and

LPO concentrations were similar across the two tissues (Table 7.4). Patterns of

response for all biomarkers in both tissues were similar in the control group and

the treatment group.

GST and GSH were the only biomarkers significantly correlated with oyster copper

concentration in the laboratory experiment, and only in the highest treatment

(30 µg/L). Hepatopancreas GST increased as oyster copper concentrations

increased, while after an initial stimulation gill GSH concentrations decreased over

the exposure phase and combined exposure/depuration phases (Table 7.4). The

relationships over time were significant in only the highest three treatment groups

and became stronger and more significant as copper treatment concentrations

increased. This indicates adaptation or acclimation of biomarker responses to

changed exposure conditions (Figure 7.6).


90

Time (day)

0 5 10 15 20 25

Oys

ter c

oppe

r con

cent

ratio

n (u

g/g)

0

50

100

150

200

250

300

0 ug/L3.75 ug/L7.5 ug/L15 ug/L30 ug/L

Figure 7.5. Accumulation in copper-exposed oysters from the five treatment concentrations. The depuration period started at 21 days

Time (day)

0 5 10 15 20 25

Gill

GSH

con

cent

ratio

n (u

mol

/g)

0

5

10

15

20

25

30

Figure 7.6. Regression of mean GSH concentration in gills against time a) 23 days and b) 28 days in each treatment (0, 3.75, 7.5, 15 and 30 µg/L) including baseline. Regressions

were almost significant in 15 µg/L at 23 days (r2 = 0.77, p = 0.053)

R2 = 0.72 peak, p=0.04

R2 = 0.85 peak, p=0.009

R2 = 0.92 peak, p=0.002


91

Table 7.4. Concentration of biomarkers in gill and hepatopancreas of copper-exposed oysters

Catalase Lipid peroxidase

Glutathione-S-transferase

Glutathione Copper (µg/L)

Day

Gills Hepato Gills Hepato Gills Hepato Gills Hepato0 0 1750 ± 112 3064 ± 378 87 ± 8 64 ± 6 55 ± 4 104 ± 9 15 ± 2 5 ± 2

3 1695 ± 168 4798 ± 1009 65 ± 8 101 ± 10 39 ± 3 195 ± 13 15 ± 1 6 ± 1

5 1818 ± 195 5475 ± 456 95 ± 12 106 ± 9 90 ± 9 230 ± 24 25 ± 5 7 ± 1

8 2892 ± 406 2602 ± 341 102 ± 18 103 ± 5 73 ± 5 184 ± 18 17 ± 5 4 ± 1

12 1338 ± 234 1601 ± 154 124 ± 12 123 ± 9 84 ± 9 198 ± 14 15 ± 2 4 ± 1

15 1278 ± 86 1924 ± 223 60 ± 7 50 ± 4 17 ± 1 136 ± 13 5 ± 1 4 ± 1

23 1953 ± 156 2518 ± 187 41± 4 57 ± 5 49 ± 2 286 ± 12 11 ± 1 5 ± 1

28 1482 ± 148 3316 ± 401 26 ± 2 43 ± 6 69 ± 7 331 ± 34 3 ± 1 4 ± 1

3.75 0 1750 ± 112 3064 ± 378 87 ± 8 64 ± 6 55 ± 4 104 ± 9 15 ± 2 5 ± 2

3 1905 ± 80 4137 ± 729 70 ± 6 91 ± 6 51 ± 6 177 ± 19 13 ± 1 8 ± 2

5 3798 ± 683 3544 ± 590 101 ± 9 105 ± 4 81 ± 8 170 ± 17 21 ± 3 6 ± 2

8 1925 ± 338 3238 ± 497 81 ± 8 117 ± 11 62 ± 6 168 ± 11 17 ± 1 5 ± 1

12 1540 ± 186 1989 ± 211 113 ± 13 109 ± 5 80 ± 8 149 ± 17 18 ± 2 6 ± 1

15 1048 ± 55 2355 ± 208 48 ± 4 44 ± 4 20 ± 3 178 ± 13 5 ± 1 4 ± 1

23 1751 ± 223 2222 ± 336 36 ± 5 61 ± 8 56 ± 4 263 ± 22 9 ± 1 3 ± 1

28 2371 ± 224 4359 ± 394 62 ± 12 71 ± 5 60 ± 4 289 ± 36 3 ± 0 4 ± 1

7.5 0 1750 ± 112 3064 ± 378 87 ± 8 64 ± 6 55 ± 4 104 ± 9 15 ± 2 5 ± 2

3 1524 ± 114 3994 ± 476 75 ± 6 92 ± 3 44 ± 4 182 ± 15 16 ± 1 7 ± 2

5 1843 ± 119 5035 ± 989 95 ± 8 99 ± 9 64 ± 6 126 ± 16 22 ± 3 5 ± 1

8 2048 ± 204 2674 ± 256 92 ± 10 123 ± 10 75 ± 5 202 ± 28 18 ± 1 5 ± 1

12 1626 ± 188 1491 ± 145 89 ± 10 88 ± 7 59 ± 6 125 ± 13 17 ± 3 4 ± 1

15 920 ± 60 1616 ± 116 39 ± 3 48 ± 4 21 ± 1 179 ± 14 7 ± 1 3 ± 0

23 1869 ± 184 2858 ± 196 48 ± 6 55 ± 6 50 ± 2 269 ± 20 11 ± 1 5 ± 1

28 1859 ± 281 2578 ± 259 55 ± 7 53 ± 4 53 ± 4 303 ± 40 2 ± 0 4 ± 1

15 0 1750 ± 112 3064 ± 378 87 ± 8 64 ± 6 55 ± 4 104 ± 9 15 ± 2 5 ± 2

3 1411 ± 165 3257 ± 527 79 ± 7 92 ± 8 53 ± 5 166 ± 16 15 ± 2 6 ± 1

5 2355 ± 335 3676 ± 486 92 ± 10 85 ± 5 61 ± 4 142 ± 6 18 ± 2 5 ± 1

8 1501 ± 153 2086 ± 320 110 ± 8 126 ± 10 82 ± 11 205 ± 24 18 ± 2 6 ± 2

12 1586 ± 218 1579 ± 177 105 ± 15 112 ± 6 63 ± 4 168 ± 15 19 ± 2 5 ± 1

15 812 ± 56 1831 ± 165 40 ± 4 41 ± 4 17 ± 2 122 ± 15 6 ± 1 4 ± 0

23 1810 ± 252 3067 ± 250 41 ± 7 48 ± 4 71 ± 13 323 ± 30 4 ± 1 3 ± 1

28 1698 ± 233 2870 ± 264 76 ± 23 73 ± 5 52 ± 5 297 ± 23 ^2 ± 0 3 ± 1

30 0 1750 ± 112 3064 ± 378 87 ± 8 64 ± 6 55 ± 4 104 ± 9 15 ± 2 5 ± 2

3 1374 ± 108 5401 ± 839 74 ± 9 110 ± 8 67 ± 3 186 ± 13 17 ± 2 7 ± 1

5 2097 ± 264 4535 ± 414 103 ± 9 100 ± 6 73 ± 5 134 ± 17 19 ± 3 6 ± 1

8 2403 ± 364 2025 ± 302 101 ± 11 131 ± 13 80 ± 6 213 ± 23 18 ± 2 8 ± 2

12 1408 ± 165 2039 ± 511 105 ± 11 139 ± 13 52 ± 4 258 ± 36 17 ± 3 5 ± 1

15 1289 ± 75 1505 ± 229 63 ± 9 51 ± 4 27 ± 4 133 ± 15 5 ± 1 4 ± 1

23 2044 ± 155 2695 ± 329 61 ± 9 59 ± 7 52 ± 3 294 ± 17 ^2 ± 1 3 ± 0

28 2881 ± 599 2693 ± 198 72 ± 10 74 ± 9 51 ± 4 289 ± 22 ^2 ± 1 3 ± 1


92

7.4 Conclusions 1. Metal concentrations in transplanted oysters were lower than those in

resident oysters. It is therefore likely that deployed organisms had not

attained equilibrium and may not achieve the same levels of

accumulation as resident oysters. The concentrations of copper and

zinc increased with time in the transplanted oysters whereas the

concentration of tissue aluminium decreased.

2. Field biomarker concentrations were quite variable and few consistent

trends were observed both with exposure time and between site.

Although some statistically significant correlations between biomarker

concentrations and tissue metal concentrations were observed, no firm

conclusions could be drawn regarding the suitability of these biomarkers

for biomonitoring in Port Curtis.

3. Under controlled laboratory conditions glutathione (GSH) and

glutathione-s-transferase (GST) exhibited consistent responses to

dissolved copper exposure at elevated copper concentrations (30 µg/L).

After initial stimulation there may also be adaptation or acclimation of

biomarker responses to new exposure conditions. The other biomarkers

(CAT and LPO) did not respond in a consistent manner to dissolved

copper exposure.

4. The use of biomarker responses as a suitable measure of �stress� in

oysters in Port Curtis could not be determined from this study alone. The

causes of biomarker variability and the use of other biomarkers directly

linked to metal metabolism should be further investigated.

7.5 References Andersen, L.E., Boundy, K. and Melzer, A. (2002) Intertidal crabs as potential

biomonitors in Port Curtis. Centre for Environmental Management, Central

Queensland University and Cooperative Research Centre for Coastal Zone,

Estuary and Waterway Management, Gladstone, 23 pp.

Andersen, V., Maage, A. and Johannessen, P.J. (1996) Heavy metals in blue

mussels (Mytilus edulis) in the Bergen harbour area, western Norway. Bulletin

of Environmental Contamination and Toxicology, 57, 589�596.


93

Andersen, L.E. and Norton, J.H. (2001) Port Curtis mud crab shell disease –

nature, distribution and management. FRDC Project No. 98/210, Central

Queensland University, Gladstone, 115 pp.

Andersen, L.E., Revill, A.T. and Storey, A.W. (2005a) Metal bioaccumulation

through food web pathways in Port Curtis. Technical Report No. 31, CRC for


Andersen, L.E., Siu, W.H.L., Ching, E.W.K., Kwok, C.T., Melville, F., Plummer, C.,

Storey, A.W. and Lam, P.K.S. (2006) Antioxidant enzymes as biomarkers of

environmental stress in oysters in Port Curtis. Technical Report, CRC for


Andersen, L.E., Storey, A.W. and Fox, S. (2004) Assessing the effects of harbour

dredging using transplanted oysters as biomonitors. Centre for Environmental

Management, Central Queensland University, Gladstone, 214 pp.

Andersen, L.E., Storey, A.W., Sinkinson, A.W. and Dytlewski, N. (2003)

Transplanted oysters and resident mud crabs as biomonitors in Spillway

Creek. Centre for Environmental Management, Central Queensland

University, Gladstone, 30 pp.

Andersen, L.E., Teasdale, P., Jordan, M. and Storey, A.W. (2005b) Transplanted

oysters and DGT devices to measure bioavailable metals: comparison of

techniques. Reports to Comalco Alumina Refinery and Institute of Sustainable

Regional Development. Centre for Environmental Management, Central

Queensland University, Gladstone.





Brown, R.J., Galloway, T.S., Lowe, D.M., Browne, M.A., Dissanayake, A., Jones,

M.B. and Depledge, M.H. (2004) Differential sensitivity of three marine

invertebrates to copper assessed using multiple biomarkers. Aquatic


Chan, K.W., Cheung, R.Y.H., Leung, S.F. and Wong, M.H. (1999) Depuration of

metals from soft tissues of oysters (Crassostrea gigas) transplanted from a

contaminated site to clean sites. Environmental Pollution, 105, 299�310.

Cheung, C.C.C., Zheng, G.J., Lam, P.K.S. and Richardson, B.J. (2002)

Relationships between tissue concentrations of chlorinated hydrocarbons

(polychlorinated biphenyls and chlorinated pesticides) and antioxidative


94

responses of marine mussels, Perna viridis. Marine Pollution Bulletin,

45, 181�191.

Cheung, C.C.C., Zheng, G.J., Li, A.M.Y., Richardson, B.J. and Lam, P.K.S. (2001)

Relationships between tissue concentrations of polycyclic aromatic

hydrocarbons and antioxidative responses of marine mussels, Perna viridis.

Aquatic Toxicology, 52, 189�203.

Curran, J.C., Holmes, P.J. and Yersin, J.E. (1986) Moored shellfish cages for

pollution monitoring. Marine Pollution Bulletin, 17, 464�465.

Doyotte, A., Cossu, C., Jacquin, M., Babut, M. and Vasseur, P. (1997) Antioxidant

enzymes, glutathione and lipid peroxidation as relevant biomarkers of

experimental or field exposure in the gills and the digestive gland of the

freshwater bivalve Unio tumidis. Aquatic Toxicology, 39, 93�110.

Fitzpatrick, P.J., O'Halloran, J., Sheehan, D. and Walsh, A.R. (1997) Assessment

of glutathione-S-transferase and related proteins in the gills and digestive

gland of Mytilus edulis (L.), as potential organic pollution biomarkers.

Biomarkers, 2, 51�56.

Irato, P., Santovito, G., Cassini, A., Piccinni, E. and Albergoni, V. (2003) Metal

accumulation and binding protein induction in Mytilus galloprovincialis,

Scapharca inaequivalvis, and Tapes philippinarum from the Lagoon of Venice.

Archives of Environmental Contamination and Toxicology, 44, 476�484.

Luebke, R.W., Hodson, P.V., Faisal, M., Ross, P. S., Grasman, K.A. and Zelikoff,

J.T. (1997) Aquatic pollution-induced immunotoxicity in wildlife species.

Fundamental and Applied Toxicology, 37, 1�15.

Nusetti, O., Esclapes, M., Salazar, G., Nusetti, S. and Pulido, S. (2001)

Biomarkers of oxidative stress in the polychaete Eurythoe complanata

(Amphinomidae) under short-term copper exposure. Bulletin of Environmental

Contamination and Toxicology, 66, 576�581.

Odzak, N., Zvonaric, Z., Kljakovic, G. and Barie, A. (2001) Biomonitoring of

copper, cadmium, lead, zinc and chromium in the Kastela Bay using

transplanted mussels. Environmental Bulletin, 10, 37�41.

O'Halloran, K., Ahokas, J. and Wright, P. (1998) The adverse effects of aquatic

contaminants on fish immune responses. Australasian Journal of

Ecotoxicology, 4, 9�28.

Prange, J.A. (1999) Physiological responses of five seagrass species to trace

metals. B.Sc. honours thesis. Botany Department, University of Queensland,

Brisbane, 52 pp.


95

Rainbow, P.S. (1995) Biomonitoring of heavy metal availability in the marine

environment. Marine Pollution Bulletin, 31, 183�192.

Regoli, F., Nigro, M. and Orlando, E. (1998) Lysosomal and antioxidant responses

to metals in the Antarctic scallop Adamussium colbecki. Aquatic Toxicology,

40, 375�392.

Regoli, F. and Principato, G. (1995) Glutathione, glutathione-dependent and

antioxidant enzymes in mussel, Mytilus galloprovincialis, exposed to metals

under field and laboratory conditions: implications for the use of biochemical

biomarkers. Aquatic Toxicology, 31, 143�164.

Ringwood, A.H., Conners, D.E., Keppler, C.J. and Dinovo, A. (1999) Biomarker

studies with juvenile oysters (Crassostrea virginica) deployed in situ.

Biomarkers, 4, 400�414.

Weeks, J.M. (1995) The value of biomarkers for ecological risk assessment:

academic toys or legislative tools? Applied Soil Ecology, 2, 215�216.

Winston, G.W. and Giulio, R.T. (1991) Prooxidant and antioxidant mechanisms in

aquatic organisms. Aquatic Toxicology, 19, 137�161.

Contaminant pathways in Port Curtis: Final report 8: Effect of pulse events

96

Chapter 8 Effect of pulse events on biological responses to contaminants

8.1 Background Coastal waters adjacent to industrialised regions such as Port Curtis are likely to

receive contaminant inputs from urban runoff and stormwater drains and from

industrial effluent discharges. While routine monitoring may indicate that

contaminant concentrations are below levels of regulatory concern, there is

generally a poor understanding of the rate, concentration and consistency of

contaminant inputs and the associated biological effects. In particular, the

intermittent discharge of industrial effluent is expected to result in significant

temporal fluctuations in contaminant concentrations, especially within discharge

mixing zones. These fluctuating contaminant inputs (pulses), coupled with the

dynamic natural processes of tides, seasonal rainfall events, and sediment

resuspension (e.g. during dredging), may result in routine monitoring failing to

measure many major contaminant inputs that cause short-term ecological effects.

In ecological risk assessments, toxicity tests play a critical role in quantifying the

biological effects of contaminants within effluents and receiving waters. Standard

toxicity tests involve exposing test organisms to waters for a predetermined time

period, during which the contaminant concentrations and exposure conditions are

generally considered to be constant. While these methods may correctly quantify

effects occurring during laboratory-based toxicity tests, they may poorly represent

how organisms respond in the field to fluctuating contaminant concentrations

(Burton et al. 2000).

To date, most studies of fluctuating contaminant exposures have considered

organic contaminants such as insecticides and pesticides, with far fewer studies

of metal contaminants. Comparisons of continuous and pulsed exposures with

equivalent contaminant doses have reported varied results. Some studies have

shown that fluctuating, pulsed exposures cause greater uptake of contaminants

and greater toxicity than continuous contaminant exposures (Curtis et al. 1985;

Holdway et al. 1994; Ingersoll & Winner 1982; Parsons & Surgeoner 1991; Schulz

& Liess 2000; Siddens et al. 1986; Thurston et al. 1981). In these cases it appears

that the pulse concentration has a greater influence on the toxic response of the

test organism than the duration of exposure (Reinert et al. 2002), with high

contaminant concentrations appearing to overwhelm the test organism. In other

studies, continuous exposure has been reported as being more toxic to test

organisms than pulsed exposures to the equivalent contaminant dose (Hosmer


97

et al. 1998; Jarvinen et al. 1988; Kallander et al. 1997; Mancini 1983; Marr et al.

1995; Pascoe & Shazili 1986). In these cases it appears that the exposure

duration has a greater influence on the toxic response of the organism than the

contaminant concentration (Reinert et al. 2002), allowing adequate time for

uptake during longer exposures.

The aim of this work was to compare the effects of pulse exposures to dissolved

copper with continuous exposures, using a metal-sensitive microalga. The

pennate marine diatom Phaeodactylum tricornutum was selected for study as it is

widely distributed in temperate and tropical environments and is an important food

source for invertebrates and fish. Copper was an obvious choice as a model

contaminant given its elevated concentration in waters and biota in Port Curtis

(Apte et al. 2005).

8.2 Experimental

8.2.1 Chemical analysis

General trace metal sample treatment and analysis procedures described by the

authors in previous publications (Apte et al. 2005; Simpson et al. 2003) were used

in this study. Dissolved copper concentrations in saline solutions were determined

by inductively coupled argon plasma atomic emission spectroscopy (ICP-AES,

Spectroflame EOP) calibrated using matrix-matched standard solutions. The

detection limit for copper was 2 µg/L. Intracellular copper analyses were made by

graphite furnace atomic absorption spectroscopy (GFAAS, Perkin Elmer 4100ZL)

using Zeeman effect background correction and operating conditions

recommended by the manufacturer. Extracellular copper analyses were made by

square-wave anodic stripping voltammetry (SW-ASV) (Metrohm 646 Voltammetric

Analyser) with a hanging mercury drop electrode. Samples were stirred and

de-aerated with nitrogen for 300 s before deposition for 300 s at -0.6V vs SCE.

A potential scan was initiated (scan rate 3.3 mV/s, pulse height 50mV, pulse step

2 mV) and the copper oxidation peak areas recorded between -0.2 and 0.2V.

A calibration curve was constructed using matrix-matched standards.

8.2.2 Algal bioassay procedure

The algal bioassay measured the decrease in growth rate and cell yield of the

marine unicellular alga Phaeodactylum tricornutum. P. tricornutum was chosen for

the study because it has previously been shown to be sensitive to copper (growth

rate 72-h IC50 = 10 ± 4 µg/L, Franklin et al. 2001b), it is easy to count and does


98

not clump or adsorb to the walls of the test containers. The bioassay protocol was

based on the OECD Guideline 201 (2005) and the protocol of Stauber et al.

(1994). P. tricornutum cultures between 3�5 days were used for inoculation to

ensure that the algae were in the exponential phase of growth. Cultures were

washed three times in clean sea water to ensure the complete removal of any

culture medium and cellular exudates from the algal solution (Stauber & Florence

1987). Cells were inoculated into 250 mL borosilicate glass Erlenmeyer flasks

(pre-silanised with Coatasil to reduce metal adsorption to flask walls) containing

50 or 100 ml of clean filtered sea water. The initial cell density was typically

4�6 ×103 cells/mL. The flask contents were supplemented with 0.5 mL of 26 mM

sodium nitrate (15 mg NO3−/L) and 0.5 mL of 1.3 mM potassium dihydrogen

phosphate (1.5 mg PO43−/L) in order to maintain exponential growth over 72 h.

Copper was added to treatment flasks (Day 0) and cell densities determined daily

over the next 3 days. Flasks were shaken twice daily by hand to avoid CO2

limitation. Cell density measurements were made daily using a FACSCalibur flow

cytometer (BD Biosciences). All tests included a reference toxicant copper (tested

at five concentrations, 0�40 µg Cu/L) to determine an IC50 value (i.e. the inhibitory

concentration to cause a 50% decrease in growth rate or cell yield). Bioassay test

results were processed using standard statistical procedures described elsewhere

(Franklin et al. 2001a; ToxCalc 1984; Sprague & Fogels 1977).

8.2.3 Pulsed exposures to dissolved copper

To generate contaminant pulses, the algae needed to be repeatedly cycled

between seawater solutions containing different amounts of copper. A gentle

centrifugation method (3500 rpm for 4 min, ~1500 g) was developed for isolating

algal cells post-exposure which did not damage algal cells or affect growth. Three

types of copper exposure scenarios were investigated (Figure 8.1):

(i) exposures of equivalent copper �dose�, but varying pulse duration and

magnitude (four experiments with different copper dose, each comprising

continuous, 1-, 2-, 4-, and 8-h pulse exposures)

(ii) equivalent copper �dose� and concentration, but dose applied at varying

pulse frequency (two experiments with varying pulse frequency)

(iii) high copper concentration pulses of decreasing pulse duration. The

copper concentrations for each exposure were calculated so that the

algae would receive an equal dose of copper per day and this was

generally equivalent to the IC50 copper concentration for the continuous

exposure (7�10 µg/L).


99

0

5

10

15

20

25

30

35

0 12 24 36 48 60 72Time, h

Dis

solv

ed c

oppe

r, µg

/L (i)

0

10

20

30

40

50

60

0 12 24 36 48 60 72Time, h

Dis

solv

ed c

oppe

r, µg

/L (ii)

0

40

80

120

160

200

0 12 24 36 48 60 72Time, h

Dis

solv

ed c

oppe

r, µg

/L (iii)

Figure 8.1. Copper exposure scenarios tested: (i) equivalent copper �dose�, but varying duration and magnitude; (ii) equivalent copper �dose� and concentration, but varying pulse frequency;

and (iii) high copper concentration pulses of decreasing pulse duration

8.2.4 Intracellular and extracellular copper determinations

To investigate the uptake of copper during pulsed and continuous exposure, the

copper bound to the surface of the algae (extracellular) and the internalised

copper (intracellular) were determined at various exposure times. Extracellular

copper was isolated by suspending the algal pellet in 20 ml of 0.01 M EDTA in

NaCl solution (3.5% m/v) for 35 min. Following centrifugation, a portion of this

extract was then acidified with 1% HNO3 to pH 3 and analysed by SW-ASV. The

residual algal pellet was then washed with approximately 10 mL of clean sea

water, acidified (2 mL of concentrated HNO3), and was left for at least 30 min prior

to microwave digestion (90 W for 5 min, then cooling to room temperature). The

digests were then diluted to 20 mL with deionised water and analysed by GF-AAS.

This fraction was deemed the �intracellular Cu.� The environmentally realistic, low

algal cell densities used in these experiments meant that multiple flasks often had

to be combined so that the extracted intracellular and extracellular copper were

above the detection limit of the instruments used in their analysis.


100

8.2.5 Modelling bioassay response with fluctuating copper concentrations

In the absence of a toxicant, and under constant light, temperature and nutrient

conditions (assuming no lag phase), algal growth is exponential (Nyholm 1985)

and the growth rate (µt) is constant. A plot of log cell density (log N) versus time

is linear with slope equal to µt (equation 1):

Nt = N0·eµt (t-t0) (Eqn 1)

Algal growth rate, µt, decreases with increasing toxicant concentration and was

modelled using a four-parameter logistic model (Nyholm et al. 1992; Simpson et

al. 2003) according to:

µt = α � β/(1 + δ·e(-φ·Ct)) (Eqn 2)

where µt is the growth rate at time t, Ct is the toxicant concentration at time t, α, β,

δ and φ are constants.

In the model, Ct, and subsequently µt were calculated at hourly intervals over the

72-h test period and algal cell densities were calculated according to equation 3.

Nt = Ntd·eµt (t - td) (Eqn 3)

where Nt is the number of cells at any given time t, Ntd is the number of cells at

time td (td< t), µt is the growth rate at time t (varying due to changing

concentration, Ct).

Measurements of algal biomass at time periods t = 0, 24, 48, and 72 h during

exposure of the algae to copper concentrations of 0, 2, 4, 8, 16, 30 and 50 µg/L

copper was used to develop a relationship between the algal growth-rate

parameter, µt, and copper. The fit between the measured algal biomass and the

model predictions was optimised using the Solver tool application of Microsoft

Excel® which minimises the sum-of-the-squares of difference between the

measured and the model algal biomass data by changing the variables in the

model (equations 2 and 3). Only the effects-data for the continuous exposure

experiments were used for optimising the model fit.

Two exposure effect models were developed. In Model 1, the external copper

concentration in solution was treated as the �exposure� and the cause of the toxic

effects to algal growth rate. In Model 2, the intracellular copper concentration,

within the algae, was treated as the �exposure� and the cause of the toxic effects to

algal growth rate. In Model 2, the intracellular exposure was calculated by

multiplying the external copper concentration in solution by an uptake rate


101

constant, ku-I. In this model, the algal growth rate, µt, decreases with increasing

intracellular copper concentration. A copper efflux mechanism was not included in

Model 2.


8.3.1 Continuous exposure

The inhibition of growth (cell biomass) during continuous exposure to dissolved

copper, measured at time periods of 24, 48, and 72 h is shown in Figure 8.2. For

the 24-, 48-, and 72-h exposure periods, IC50 concentrations were calculated as

40 (18-88), 8.6 (6.3-12) and 5.4 (4.5-6.5) µg Cu/L, respectively. The data

illustrates that inhibition concentrations (e.g. IC50s) calculated using cell biomass

data decrease as the test duration increases.

0

20

40

60

80

100

0 10 20 30 40

Dissolved copper, µg/L

Alg

al c

ell b

iom

ass,

% c

ontro

l .

24 h 48 h 72 h

Figure 8.2. Measured (symbols) and predicted (three models) effect

of dissolved copper concentrations on algal cell biomass

The error bars represent standard deviations of triplicate measurements. The three models are for effect due to (a) exposure copper concentration (solid line) and (b) internalised copper concentration (short-dash line).

8.3.2 Pulsed copper exposures

The results for algal cell biomass and the respective growth inhibition following

72 h exposure to the different pulse scenarios are shown in Table 8.1. Pulsed

exposures to copper caused similar or less inhibition of algal growth than

continuous exposures. Of the pulse scenarios, short (1�2 h) copper pulses at

high copper concentrations (51 µg/L) caused greater inhibition of algal growth

(82%) than longer (4-8 h) copper pulses of intermediate copper concentrations

(18�28 µg/L , 65�70% inhibition), and had a similar inhibitory effect as continuous

exposure. This may be because between the short pulses (the recovery period)


102

the copper concentration was sufficiently high (5 µg/L) to cause continuous growth

inhibition, however for the lower pulsed exposure scenarios, between the pulses

the copper concentrations were lower (1.8-2.8 µg/L) and below the lowest

observable effects concentration (LOEC).

Table 8.1. Pulse exposure scenarios and the biomass inhibition observed at 72 h

Copper concentration, µg/L

Pulse time, h Pulse / recovery scenario

Measured mean copper (dose)

Inhibition, % control

Pulse Scenario 1 Control 1× 72 h 0 <1 0 2 h pulse 3× 2 h (1 per day) 38 / 7.5 5.1 81 4 h pulse 3× 4 h (1 per day) 30 / 6.0 6.3 76 8 h pulse 3× 8 h (1 per day) 21 / 4.3 6.3 80 Continuous 1× 72 h 10 6.2 84



Pulse Scenario 4 Control 1× 72 h 0 <1 0 1 h pulse 3× 1 h (1 per day) 180 / 1.8 18.8 85 2 h pulse 3× 2 h (1 per day) 97 / 0.9 15.2 82 4 h pulse 3× 4 h (1 per day) 51 / 0.5 10.7 66 8 h pulse 3× 8 h (1 per day) 27 / 0.3 11.6 77 Continuous 1× 72 h 10 7.6 86

8.3.3 Copper uptake

Measurements of intracellular copper in both continuous and pulsed exposure

scenarios (Table 8.2) showed that there was more intracellular copper in cells in

both the continuous and short, high-pulse exposure scenarios than in the other

pulse scenarios, corresponding to the greater growth inhibition observed.

Extracellular copper, that is, copper loosely bound to the cell wall, was also higher

in the continuous and short, high pulse exposure scenario. There was a significant

positive relationship (p<0.01) between the extracellular copper concentrations on

the algal cells and the copper concentrations of the exposure solutions at 72 h,


103

indicating that the extracellular copper may just reflect the copper concentrations

in the exposure solution. Although it is possible that higher extracellular copper

concentrations may contribute to growth-related effects by disrupting cell surface

transport processes, it is more likely that the copper has to be internalised to

cause growth inhibition (Franklin et al., 2002, de Schamphelaere et al. 2005).

Table 8.2. Extra- and intra-cellular copper determined following the different 72 h copper pulse exposure scenarios

Exposure Planned exposure [Extracellular Cu] [Intracellular Cu] Inhibition scenario (pulse/ nonexposure) (x10-8 ng Cu/µm2) (x10-8 ng Cu/µm3) (%)

Control 0, 0 µg/L 1.13 0.12 0% 1 h pulse 51.0, 5.0 µg/L 11.2 6.2 82% 4 h pulse 28.0, 3.0 µg/L 3.4 3.7 65% 8 h pulse 18.0, 2.0 µg/L 2.2 4.5 70%

Continuous 7.0, 7.0 µg/L 10.7 6.1 82%

The kinetics of copper uptake for different copper concentrations was investigated

to better understand why the algae responded differently to the various exposure

scenarios. Extracellular copper increased rapidly on the algal cells (within

minutes), and then seemed to plateau to a relatively constant concentration for

each of the exposure concentrations (Figure 8.3).

0.0

0.5

1.0

1.5

2.0

2.5

3.0

0 20 40 60 80Time (h)

Ext

race

llula

r Cu

(ng/

104 c

ells

) A

0.0

0.5

1.0

1.5

2.0

2.5

3.0

0 20 40 60 80Time (h)

Ext

race

llula

r Cu

(ng/

10 4 c

ells

)

B

0.0

0.5

1.0

1.5

2.0

2.5

3.0

0 20 40 60 80Time (h)

Ext

race

llula

r Cu

(ng/

10 4 c

ells

)

C

y = 0.0029x + 0.019R2 = 0.57

0.0

0.2

0.4

0.6

0.8

1.0

0 20 40 60 80Time (h)

Intra

cellu

lar C

u (n

g/10

4 cel

ls) A y = 0.0066x - 0.0046

R2 = 0.88

0.0

0.2

0.4

0.6

0.8

1.0

0 20 40 60 80Time (h)

Intra

cellu

lar C

u (n

g/10

4 cel

ls)

B y = 0.0074x - 0.0069R2 = 0.93

0.0

0.2

0.4

0.6

0.8

1.0

0 20 40 60 80Time (h)

Intra

cellu

lar C

u (n

g/10

4 cel

ls)

C

Figure 8.3. Relationships between the exposure time and extracellular (upper) and intracellular (lower) copper concentrations in P. tricornutum cells for exposures to

varying copper concentrations

Copper concentrations are: A (10 µg/L- #,$, and !); B (30 µg/L- #, and !); and C (50 µg/L µg/L- #). Error bars represent the standard deviation of the mean for triplicate treatments in each experiment, and separate experiments are denoted by different symbols (i.e. #,$, and !).


104

In general, intracellular copper increased very slowly in the initial 20 h, and then

began to accumulate in cells as exposure time increased (Figure 8.3). Once

copper internalisation commenced, the intracellular copper uptake was still

relatively linear over the duration of the bioassay for the 10, 30, and 50 µg/L

copper bioassays with rates of 29, 66, and 74 ng/×108 cells/h, respectively

(Figure 8.3). Higher exposure concentrations and longer exposure times resulted

in greater intracellular concentrations.

These results indicated that the internalisation of copper occurs more slowly than

copper binding at the cell surface (extracellular copper) and may be a rate-limiting

step for toxic effects. This is in agreement with other studies (Knauer et al. 1997;

Hassler et al. 2004; Slaveykova & Wilkinson 2002). Furthermore, the rate of

copper internalisation did not increase linearly with increasing copper

concentration in the external exposure solution. These observations have

important implications for how pulse copper exposures may cause toxic effects.

For short-duration copper pulses, there may be insufficient time for the toxic

effects to occur before the external exposure is removed.

8.3.4 Copper elimination

To investigate potential copper elimination by cells and algal growth recovery

following exposure to 10 µg Cu /L for 72 h, the exposure solution was removed

and the algal cells resuspended in clean sea water containing nutrients. When

algal cells were placed in clean sea water, the extracellular and intracellular

copper decreased with time and was below detection after 27 h (Figure 8.4A, B).

The results indicated that the elimination of extracellular copper from the cells

occurs through both desorption and dilution following cell division, while the

elimination of intracellular copper from the cells was due to dilution via cell

division, rather than due to efflux of copper from the cell.

Possible mechanisms for the elimination of copper from the algae cells include

(i) desorption of extracellular copper into clean sea water, (ii) efflux of intracellular

copper from cells, and (iii) dilution of extracellular and intracellular copper through

cell division and growth of the algae in the clean sea water. During the desorption

period, the intracellular copper concentration remained reasonably constant for the

initial 6 h, indicating negligible efflux of intracellular copper occurred during the this

period (Figure 8.4B). The intracellular copper concentration began to decrease

rapidly 6 h after the desorption experiments commenced and coincided with the

increase in cell division. P. tricornutum are also capable of incorporating

intracellular copper into inert bodies such as vacuoles or binding copper with

phytochelatins to reduce toxicity (Knauer et al. 1997).


105

0.00

0.05

0.10

0.15

0.20

0 10 20 30 40 50

Recovery time (h)

Intra

cellu

lar C

u (n

g/10

4 cel

ls)

0

15

30

45

60

Cel

l den

sity

(x10

4 cel

ls/m

l)

B

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0 10 20 30 40 50

Recovery time (h)

Ext

race

llula

r Cu

(ng/

104 c

ells

) A

Figure 8.4. Copper efflux after placing P. tricornutum cells in clean sea water

(A) Extra-cellular copper (!) on a cell-density basis, (B) intra-cellular copper on a cell-density basis (!) and on a per/cell basis (%). The accumulation stage involved a 72 h continuous copper exposure of 7 µg/L.

8.3.5 Modelling effects of pulsed copper exposures on algal growth

Rather than having to always measure toxicity responses of biota to contaminants,

various modelling approaches can be used to predict toxicity under a variety of

water quality conditions. Exposure-effect models are one type of modelling

approach that should enable prediction of toxic effects after careful validation with

experimental data.

The effect of the solution copper concentration on algal growth rate was described

in Model 1 by µt = 0.046 - 0.035/(1 + 4.6�e-5.5�Ct) (where Ct is the external copper

concentration at time t). For the continuous copper exposure experiments, the fit

between the observed data and Model 1 is shown in Figure 8.2. The fit becomes

increasingly worse for short exposure times (e.g. 24-h exposure data) and may be

related to the �lag� between the exposure (external copper concentration) and the

effects (presumably occurring due to internalised copper). The fit between the

observed data and Model 1 for the pulsed exposure experiments is shown in

Figure 8.5. Model 1 adequately predicted the effects of the continuous copper

exposures on P. tricornutum growth, but for the 1�8 h pulsed copper exposures

Model 1 generally underestimated the effects of the copper exposure on algal

growth (Figure 8.5a). It is likely that algal growth rate is not directly linked by the

external copper concentration (solution) or the extracellular copper, but instead to

the intracellular copper (Knauer et al. 1997; Hassler et al. 2004; Slaveykova &

Wilkinson 2002).

Using Model 2, the effect of the internalised copper concentration (toxicant) on

algal growth rate was described by µt = 0.043 + 0.064/(1 + 2.6�e-3.0�C(int)t) (where

C(int)t is the internalised copper concentration at time t). The fit between the


106

observed data and Model 2 (Figure 8.5b, r2 = 85) for the pulsed exposure

experiments was better than Model 1 (Figure 8.5a, r2 = 0.67). While neither model

provided a completely accurate description of effects the copper exposure had on

algal growth, in general the models reasonably predicted effects due to the various

pulsed copper exposure scenarios (Figure 8.5).

0%10%20%30%40%50%60%70%80%90%

100%

72 h

Yie

ld, %

Con

trol

Measured Model 1 (dissolved Cu exposure)

Exposures: C, 1, 2, 4, 8 = continuous, 1 h, 2 h, 4 h, 8 h pulses

C 2 4 8 C 1 4 8 C 1 4 8 C 1 2 4 8

Time averaged concentration, µg/L

9 7 8 8 5 4 4 5 6.5 7 7 7 7 11 12 16 21

(a)

0%10%20%30%40%50%60%70%80%90%

100%

72 h

Yie

ld, %

Con

trol

Measured Model 2 (intracellular Cu exposure)

Exposures: C, 1, 2, 4, 8 = continuous, 1 h, 2 h, 4 h, 8 h pulses

C 2 4 8 C 1 4 8 C 1 4 8 C 1 2 4 8

(b)

Figure 8.5a,b. Measured and predicted effect of copper exposure scenarios tested with equivalent copper �dose�, but varying duration and magnitude

The models treated the �exposure� causing the toxicity as (a) the external copper concentration (Model 1), and (b) the internalised copper concentration (Model 2).

The discrepancies (Figure 8.5) between the measured effects of copper on algal

growth and the predictions from Models 1 and 2 for the pulsed copper experiments


107

may be due, in part, to changes in cell morphology due to the copper exposure.

Optical and transmission electron microscopy showed that dissolved copper

caused the algal cells to swell and clump together, changing their surface area

and internal volumes in both continuous and pulsed copper exposures

(Figure 8.6). Greater copper concentrations and longer exposure times caused

greater cell swelling. Swelling following copper exposure has been observed for

the marine alga, Nitzchia closterium, a species of pennate diatom similar to

P. tricornutum (Stauber & Florence 1987).

(Images courtesy of University of Wollongong and CSIRO Land and Water)

Figure 8.6. Transmission electron microscopy of P. tricornutum: (a) control cells grown in clean sea water and (b) cells grown in 15 µg/L copper for 72 h

8.4 Conclusions Upon exposure to dissolved copper, P. tricornutum rapidly accumulated

extracellular copper and, after a delay of approximately 20 h, accumulation

of intracellular copper began. Copper efflux measurements indicated that

P. tricornutum did not have an effective mechanism for eliminating copper from

cells; rather the intracellular copper decreased as a result of dilution by cell

division. Exposure-effect models, based on intracellular and external solution

copper concentrations, were developed in an attempt to predict copper toxicity to

algae under different exposure scenarios. The model based on internalised copper

gave the best fit to the observed data; however further refinement of the model is

necessary to take into account physiological changes in algae as a result of

copper exposure. The ability to accurately model the toxic effects of copper to the

alga is complicated by varying copper exposure concentrations in solution during

tests and by rates of extracellular and intracellular copper uptake and elimination.

These studies suggest that, at least for microalgae over short exposure times (up

to 72 h), bioaccumulation of copper from exposure to dissolved copper pulses is

no greater than bioaccumulation from continuous exposures.

(a) (b)


108

Given the complexity of the experimental approaches needed to study pulse

exposures it is recommended that modelling of organism response is pursued

further. In this study, water column only exposure was evaluated. Further work on

pulsed exposures for organisms such as invertebrates which are also exposed to

metals via dietary uptake is required. If models were developed for key organisms

in Port Curtis, this would allow better assessment of pulse exposure. This is

currently the best practicable approach to solving this complex problem.


109

8.5 References Apte, S.A., Jones, M.-A., Simpson, S.L., Stauber, J.L., Vicente-Beckett, V.,




Burton, G.A. Jr., Pitt, R. and Clark, S. (2000) The role of traditional and novel

toxicity test methods in assessing stormwater and sediment contamination.

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Curtis, L.R., Seim, W.K., Chapman, G.A. (1985) Toxicity of fenvalerate to

developing steelhead trout following continuous or intermittent exposure.

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De Schamphelaere, K.A.C., Stauber, J.L., Wilde, K.L., Markich, S.J., Brown, P.L.,

Franklin, N.M., Creighton, N.M. and Janssen, C.R. (2005) Towards a biotic

ligand model for

algae: surface-bound and internal copper explain the effect of pH on copper

toxicity to Chlorella sp. and Pseudokirchneriella subcapitata. Environmental

Science and Technology, 39, 2067�2072.

Franklin, N.M., Stauber, J.L., Apte, S.C. and Lim, R. P. (2002) Effect of initial cell

density of copper in microalgae bioassays. Environmental Toxicology and

Chemistry, 21, 742�751.

Franklin, N.M., Adams, M.S., Stauber, J.L. and Lim, R.P. (2001a) Development of

a rapid enzyme inhibition bioassay with marine and freshwater microalgae

using flow cytometry. Archives of Environmental Contamination and


Franklin, N.M., Stauber, J.L. and Lim, R.P. (2001b) Development of flow

cytometry-based algal bioassays for assessing the toxicity of metals in natural

waters. Environmental Toxicology and Chemistry, 20, 160�170.

Hassler, C.S., Slaveykova, V.L. and Wilkinson, K.J. (2004) Discriminating

between intra- and extra-cellular metals using chemical extractions.

Limnology and Oceanography Methods, 2, 237�247.

Holdway, D.A., Barry, M.J., Logan, D.C., Robertson, D., Young, V. and Ahokas,

J.T. (1994) Toxicity of pulse-exposed fenvalerate and esfenvalerate to larval

Australian crimson-spotted fish (Melanotaenia fluviatilis). Aquatic Toxicology,

28, 169�187.


110

Hosmer, A.J., Warren, L.W. and Ward, T.J. (1998) Chronic toxicity of pulse-dosed

fenoxycarb to Daphnia magna exposed to environmentally realistic

concentrations. Environmental Toxicology and Chemistry, 17, 1860�1866.

Ingersoll, C.G. and R.W. Winner. (1982) Effect on Daphnia pulex (De Geer) of

daily pulse exposures to copper or cadmium. Environmental Toxicology and

Chemistry, 1, 321�327.

Jarvinen, A.W., Tanner, D.K. and Kline, E.R. (1988) Toxicity of chlorpyrifos,

endrin, or fenvalerate to fathead minnows following episodic or continuous

exposure. Ecotoxicology and Environmental Safety, 15, 78�95.

Kallander, D.B., Fisher, S.W. and Lydy, M.J. (1997) Recovery following pulsed

exposure to organophosphorus and carbamate insecticides in the midge,

Chironomus riparius. Archives of Environmental Contamination and


Knauer, K., Behra, R. and Sigg, L. (1997) Adsorption and uptake of copper by the

green alga Scenedesmus subspicatus (Chlorophyta). Journal of Phycology,

33, 596�601.

Mancini, J.L. (1983) A method for calculating effects, on aquatic organisms, of

time varying concentrations. Water Research, 17, 1355�1362.

Marr, J.C.A., Bergmenn, H.L., Parker, M., Lipton, J., Cacela, D., Erikson, W.,

Phillips, G.R. (1995) Relative sensitivity of brown and rainbow trout to pulsed

exposures of an acutely lethal mixture of metals typical of the Clark Fork

River, Montana. Canadian Journal of Fisheries and Aquatic Science, 52,

2005�2015.

Nyholm, N. (1985) Response variable in algal growth inhibition tests � biomass or

growth rate? Water Research, 19, 273�279.

Nyholm, N., Sørensen, P.S., Kusk, K.O. and Christensen, E.R. (1992) Statistical

treatment of data from microbial tests. Environmental Toxicology and

Chemistry, 11, 157�167.

Parsons, J.T. and Surgeoner, G.A. (1991) Acute toxicities of permethrin,

fenitrothion, carbaryl and carbofuran to mosquito larvae during single- or

multiple-pulse exposures. Environmental Toxicology and Chemistry, 10,

1229�1233.

Pascoe, D. and Shazili, N.A.M. (1986) Episodic pollution � a comparison of brief

and continuous exposure of rainbow trout to cadmium. Ecotoxicology and

Environmental Safety, 12, 189�198.


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Reinert, K.H., Giddings, J.M. and Judd, L. (2002) Effects analysis of time-varying

or repeated exposures in aquatic ecological risk assessment of

agrochemicals. Environmental Toxicology and Chemistry, 21, 1977�1992.

Schulz, R. and Liess, M. (2000) Toxicity of fenvalerate to caddisfly larvae: Chronic

effects of

1-h vs 10-h pulse exposure with constant doses. Chemosphere, 41, 1511�

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Siddens, L.K., Seim, W.K., Curtis, L.R. and Chapman, G.A. (1986) Comparison of

continuous and episodic exposure to acidic, aluminum-contaminated waters

of brook trout (Salvelinus fontinalis). Canadian Journal of Fisheries and

Aquatic Science, 43, 2036�2040.

Simpson S.L., Roland, M.G.E., Stauber, J.L. and Batley G.E. (2003) Effect of

declining toxicant concentrations on algal bioassay endpoints. Environmental

Toxicology and Chemistry,

22, 2073�2079.

Slaveykova, W. I. and Wilkinson, K. J. (2002). Physicochemical aspects of lead

bioaccumulation by Chlorella vulgaris. Environmental Science and

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36, 969-975.

Sprague, J.B. and Fogels, A. (1977) Watch the y in bioassay. Proceedings of the

third aquatic toxicity workshop, Halifax, N.S., Nov 2–3, 1976. Environmental

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Fogels, Halifax, NS, Canada,

pp 107�118.

Stauber, J.L. and Florence, T.M. (1987) Mechanism of toxicity of ionic copper and

copper-complexes to algae. Marine Biology, 94, 511�519.

Stauber, J.L., Tsai, J., Vaughan, G.T., Peterson, S.M. and Brockbank, C.I. (1994)

Algae as indicators of toxicity of the effluent from bleached eucalypt kraft pulp

mills. National pulp mills research program technical report No. 3, CSIRO,

Canberra, Australia.

Thurston, R.V., Chakoumakos, C. and Russo, R.C. (1981) Effect of fluctuating

exposures on the acute toxicity of ammonia to rainbow trout (Salmo gairdneri)

and cutthroat trout (S. Clarki). Water Research, 15, 911�917.

ToxCalc (1994) ToxCalc users’ guide. Comprehensive toxicity data analysis and

database software, version 5.0.23C, Tidepool Scientific Software,

McKinleyville, CA, USA.

Contaminant pathways in Port Curtis: Final report 9: Conclusions and future directions

112

Chapter 9 Conclusions and future directions

9.1 Conclusions 9.1.1 Water quality

The Contaminant Pathways project has produced the first accurate data on

dissolved trace metal concentrations in the coastal waters of Central Queensland

and in close proximity to the Great Barrier Reef. In the offshore coastal waters,

dissolved metal concentrations were extremely low and were comparable to those

measured at open Pacific Ocean and New South Wales coastal water locations.

Intensive surveying of Port Curtis has confirmed the presence of elevated metal

concentrations within the harbour. The Narrows region was found to have the

highest concentrations of dissolved copper and nickel and this could be attributed

to natural geological sources.

The Fitzroy River is a source of dissolved metals to the local coastal region. In

particular, the Fitzroy contains elevated dissolved nickel concentrations. Under

some flow conditions, the Fitzroy plume may enter The Narrows region and supply

dissolved metals to Port Curtis. There were no conspicuous sources of trace

metals within Port Curtis. The trace metal distributions in Port Curtis are likely to

reflect a subtle mixture of metal inputs including industrial and other anthropogenic

discharges, inputs from unidentified sources in The Narrows and the Fitzroy River

plume. Survey measurements showed that trace metal inputs to Port Curtis which

contribute to the observed dissolved metal concentrations are most likely to be

delivered in solution form and not by release of metals from particulates.

9.1.2 Sediment quality

Using multiple lines of evidence, it was shown that the concentrations of

particulate arsenic, chromium and nickel in the benthic sediments of Port Curtis

are elevated because of the local geology and not because of metal contamination

from anthropogenic sources. This important factor needs to be taken into account

when applying the ANZECC/ARMCANZ (2000) sediment quality assessment

framework to this region. PAH contaminants in sediments were highest around the

industrial area of Gladstone, although concentrations at all locations were below

ANZECC trigger values. Several types of PAHs characteristic of combustion

sources were detected in the middle harbour, largely at the Clinton Coal Facility,

along the Calliope River and at South Trees Inlet/Boyne River, but again

concentrations were considered relatively low. Relatively high proportions of the

naturally-occurring PAH, perylene, were found in sediments from The Narrows


113

and Munduran Creek. At least the top 28 cm of sediments at intertidal and subtidal

sites were estimated to have been deposited since 1958 in Port Curtis, which is

roughly the start of the industrialisation of Gladstone. The rate of sediment

deposition was at least 0.6 cm/y. The sediment depositional zones identified were

the northern Narrows, lower Calliope River and South Trees Inlet-Boyne River

areas and therefore these areas are potentially sinks for metal deposition.

9.1.3 Hydrodynamic modelling

Some field data problems were encountered which did not allow a full evaluation

of model performance. The comparison of modelled and field data for a modest

flood event did however show that the model under-predicted salinity. This was

most likely due to inputs of fresh water occurring during the flood event that were

not included in the model (e.g. freshwater flow from the Fitzroy via The Narrows).

Nevertheless, there are grounds to be optimistic that the model represents tracer

transport reasonably well. The transport regime in the estuary is predominantly

tidally driven, and the distribution of passive tracer will reflect this dominant

forcing. The model reproduced tidal elevation satisfactorily.

9.1.4 Sublethal indicators of contaminant exposure

Imposex was detected in mulberry whelk specimens collected from Port Curtis

confirming a sublethal, biological response to TBT exposure. Although related to

local shipping intensity, the frequency and grade of the imposex condition were

not severe in comparison to surveys of other Ports in Australia and overseas.

Globally, the condition is likely to slowly improve with the introduction of further

restrictions on the use of TBT in 2008.

The concentrations of stress biomarkers (glutathione, glutathione-s-transferase,

catalase and lipid peroxidase) in field-deployed oysters were quite variable and

few consistent trends were observed that could be related to contaminant

exposure. No firm conclusions could be drawn regarding the suitability of these

biomarkers for biomonitoring in Port Curtis.

9.1.5 Contaminant foodweb dynamics

A food web including mud crabs, other crustaceans, fish, molluscs and a variety of

plants was characterised in Port Curtis. In general, the food web was not unlike

those established for other estuarine embayments. It appears that very few

species rely on mangroves as a predominant food source but are more dependent

on benthic organic matter and algae. Mud crabs were identified as one of the

dominant predators in the food chain. Carbon isotopes suggested that prawns


114

were feeding either directly or indirectly on a blue-green algal bloom (Lyngbya

majuscula) and this was supported by observations of pigment from the algae

being visually evident in the prawns. The finding may have consequences for

consumers should the toxin, sometimes produced by the algae, follow similar

uptake pathways to the pigment and accumulate in the prawn muscle tissue.

Although there were very few significant between-site differences in metal

bioaccumulation, organisms from inner harbour sites tended to be more enriched

in metals than those from the reference site outside the harbour. The findings of

this study indicate that for the majority of organisms the uptake of metals through

food pathways is likely to be complex and integrated, particularly for those in

higher trophic positions and those that have the ability to regulate metal

accumulations.

9.1.6 Pulse exposure to contaminants

Contaminant pulse studies were conducted in the laboratory using the marine

algae Phaeodactylum tricornutum as the model organism and copper as the

model contaminant. These studies suggest that, at least for microalgae over short

exposure times (up to 72 h), bioaccumulation and toxicity of copper from pulse

exposure is no greater than bioaccumulation from continuous exposure. Copper

bioaccumulation measurements indicated that P. tricornutum did not have an

effective mechanism for eliminating copper from cells; rather the intracellular

copper decreased as a result of dilution by cell division. If predictive models were

developed for key organisms in Port Curtis, this would allow better assessment of

pulse exposure. This is currently the best practicable approach to solving this

complex problem.

9.2 Future directions

After six years of activity, the Coastal Zone CRC has left a lasting legacy in Port

Curtis. There is an increased awareness among stakeholders of contaminant

issues based on good quality data. The CRC study was the first to adopt a

whole-of-port approach to understanding contaminants in Port Curtis. With a

few exceptions, the majority of previous research had either not focussed on

contaminants or their effects or had been limited to studies of particular receiving

environments. Specific project outputs, including reports, press releases and

research papers, are listed in the appendix. A considerable database of accurate

contaminant distributions is now available for utilisation by local industry,

researchers and regulators alike. The �report card� for contaminants in Port Curtis

is generally quite good, although a recent oil spill event illustrates the sensitivity of


115

the ecosystem and the need for reliable baseline information and strong

environmental management.

In the future, is it envisaged that the Port Curtis Integrated Monitoring Program

(PCIMP) and the Centre for Environmental Management (based at the Gladstone

campus of Central Queensland University) will carry on legacy of the CRC. PCIMP

is a consortium of members from 14 bodies representing industry, government

(both local and state), research institutions and other stakeholders to develop a

cooperative, integrated monitoring program for monitoring the ecosystem health of

Port Curtis. A strong, long-term, annual monitoring program building on the initial

groundwork established by the CRC is being formulated by PCIMP members in

consultation with local stakeholders. Research will focus on water and sediment

quality�particularly bioavailable contaminants�and on mangrove ecosystems.

Results will be presented to the community in the form of an Ecosystem Health

Report Card for Port Curtis.

Based on the CRC studies of the last six years, we suggest some directions

for future contaminants management and research, as described in the following

sections (sections 9.2.1�9.2.4).

9.2.1 Risk-based management

The screening level risk assessment (SLRA) illustrated the utility of using a risk-

based approach to contaminant management. Owing to limited resources, the

CRC was not able to assess the risks posed by many organic contaminants. We

recommend that this issue now be covered, but with a stage-based approach. The

CRC research has shown the value of using contaminant bioaccumulation as an

indicator of ecosystem health. A first stage would therefore be the measurement of

organic contaminants in indicator organisms in Port Curtis. Further investigations

may be necessary if bioaccumulated organic contaminants prove to be significant.

Mercury in piscivorous fish such as barramundi can be elevated owing to food

chain biomagnification. This is a regional issue of importance and should not be

forgotten given the large recreational and commercial fishing industries present

in Port Curtis and surrounding regions. The characterisation of mercury

bioaccumulation and biomagnification through food webs over the coastal region

of the whole Central Queensland region is considered appropriate.


116

9.2.2 Improved monitoring

The Contaminant Pathways study and the SLRA have shown the value of �good

quality data�. It is recommended that future monitoring adopt and enforce rigorous

quality assurance protocols to ensure quality is maintained.

As noted earlier, contaminant concentrations may fluctuate over various time-

scales in Port Curtis. Such variations are not easily picked up with a discrete

sampling approach. Time-integrated monitoring such as biomonitoring using

deployed organisms (e.g. oysters) and chemical surrogates such as diffusive

gradients in thin films (DGT) for metals and solvent-filled dialysis cells for organics

(passive samplers) is therefore recommended. Seasonal monitoring may also be

considered to determine if there are any seasonal fluctuations to contaminant

loads in Port Curtis.

9.2.3 Ability to predict contaminants concentrations and effects

A key tool in sustainable management is the ability to predict impacts. We

recommend that the further development of the Port Curtis hydrodynamic model in

conjunction with contaminant data and a subsequent suite of predictive models is

pursued. The model should be expanded to include The Narrows region and some

of the major estuaries in Port Curtis to assist in understanding the contribution of

metal load from point sources, and flows from the Fitzroy to Port Curtis. Predictive

models of metal bioaccumulation and biological impact (or establishing if

environmental harm has occurred) are also worthy of consideration. The effect of

pulse versus continuous discharges on bioaccumulation should also be pursued.

Information could result in management changes for a more favourable controlled

release of contaminants at point source discharges.

9.2.4 Future concerns

It is not easy to predict the priorities for contaminant management over the next

decade. It is necessary to keep a watching brief on emerging issues such as the

introduction of new �contaminants of potential concern� (COPCs) from developing

industries. Unforseen events such as oil spills and resulting PAH contamination

will always be a threat in a busy commercial port and stringent protocols should be

in place to manage and subsequently assess the impacts of such events. Ross

(2002) recently reported a survey of acid sulfate soils (ASS) in the Central

Queensland coast and found high occurrence of these soils on the coastal plain

along the Curtis and Capricorn coasts and at Shoalwater Bay and Broadsound.

ASS are soils containing sulfides or acid-producing soil layers resulting from the

oxidation of sulfides. When exposed to air, sulfides are oxidised, producing sulfuric


117

acid, and they can also release iron, aluminium and other heavy metals. This is an

issue of potential concern in Port Curtis, especially in areas where development

results in the aerial exposure of sulfide-containing sediments.

The reduced flushing of the estuary highlighted by the hydrodynamic model also

brings into question the resilience of the estuary in terms of its ability to cope with

increased contaminant loads from new industries or current industry expansions.

A number of new industries are proposed in the near future, including a nickel

refinery and aluminium smelter, in addition to the expansion of already existing

industries. Bioaccumulation of metals in the inner harbour area has already been

demonstrated, indicating that the harbour has potentially a limited threshold for

contaminant loads. Ecosystem health should continue to be monitored over the

long term to ensure the threshold is not exceeded, resulting in a decline in the

current state of the harbour.


marine water quality, Volume 1: The guidelines. Australian and New Zealand

Environment and Conservation Council (ANZECC) and Agriculture and

Resource Management Council of Australia and New Zealand (ARMCANZ).

Ross, D.J. (2002). Acid sulfate soils – Tannum Sands to St Lawrence, Central

Queensland Coast. Queensland Department of Natural Resources and Mines,

Rockhampton.

Contaminant pathways in Port Curtis: Final report Appendix: Specific project outputs

118

Appendix Specific project outputs Technical reports

(Available from the CRC website)

Andersen, L.E. (2004) Imposex in the city: a survey to monitor the effects of TBT

contamination in Port Curtis, Queensland. Technical Report No. 16, CRC for


Andersen, L.E., Revill, A. and Storey, A. (2005) Metal bioaccumulation through

foodweb pathways in Port Curtis. Technical Report No. 31, CRC for Coastal

Zone, Estuary and Waterway Management, Brisbane.


Storey, A.W. and Lam, P.K.S. (2006) Antioxidant enzymes as biomarkers of

environmental stress in oysters in Port Curtis. Technical Report, CRC for


Research papers

Published Andersen, L.E. (2004) Imposex: A biological effect of TBT contamination in Port

Curtis, Queensland. Australasian Journal of Ecotoxicology, 13, 5�-61.

In preparation Angel, B.M., Apte, S.C., Simpson, S.L., and Jolley, D.F. (2006) Concentrations

and sources of trace metals in Port Curtis and surrounding coastal waters,

Queensland, Australia. For submission to Marine Environmental Research.

Angel, B.M., Simpson, S.L., Stauber, J.L., and Jolley, D.F. (2006) An exposure-

effect model describing the extra- and intra-cellular copper uptake, elimination

and toxicity to marine algae during pulsed copper exposures. For submission

to Environmental Toxicology and Chemistry.

Conference presentations

Angel, B.A., Simpson, S.L., Stauber, J.L., Jolley, D.F. (2004) Poster presentation

on pulse work at Interact II, Gold Coast, July 2004.

Andersen, L.E. (2005). �Metal accumulation through food pathways in Port Curtis�.

Platform presentation at the International conference on the biogeochemistry

of trace elements (ICOBTE), April 3�7, Adelaide, South Australia.


119

Angel, B.A., Simpson, S.L., Stauber, J.L., Jolley, D.F. (2005) �The effects of

continuous and fluctuating copper exposures on the marine alga

Phaeodactylum tricornutum’. Conference Proceeding, International conference

on the biogeochemistry of trace elements (ICOBTE), April 3�7, Adelaide,

South Australia.


Storey, A.W. and Lam, P.K.S. (2005) �Antioxidant enzymes as biomarkers of

environmental stress in oysters in Port Curtis�. Conference presentation,

Research for Coastal Management, CRC Coastal Zone, 14 September,

Coolangatta, Queensland.

Andersen, L.E. (2004)� Imposex in the City: A survey to monitor the effects of TBT

contamination in Port Curtis, Queensland�. Conference presentation, CRC

Coastal Zone, 16 September, Coolangatta, Queensland.

Fabbro L.F. and Andersen, L.E. (2004) �Toxins and contaminants causing

environmental harm? - Bioaccumulation and identification of biological effects�.

Faculty of Arts health and Sciences 2005 Lecture Series, Central Queensland

University, Rockhampton.

All team members presented talks at the CRC-organised one-day seminar on

�Contaminants in Port Curtis�, Gladstone, 2005.

Dissemination of information through local media

A phone interview was held with a reporter from ABC Science/News on

19 October 2004 and an article appeared on their website.

Radio interview ABC Capricornia fishing segment re: oysters as biomonitors.

A phone interview was held with a reporter from The Veterinarian magazine on

9 November 2004.

The following press release was issued following a one-day Contaminant

Pathways workshop in Gladstone in July 2005. It was published in CQU News

and local newspapers:

Harbour Gets Second Clean Bill of Health A second phase of research in Port Curtis recently conducted by the Coastal Cooperative Research Centre has determined that the harbour still remains relatively healthy. A recent presentation of results to industry, managers and stakeholders by the combined team from CSIRO and Central Queensland University indicated positive findings on the health of local waterways. Extensive surveys of the port included measuring contaminants in water, sediments and marine organisms and examining the health effects of these contaminants on marine life. Although concentrations of some dissolved metals were elevated in harbour water, they were not above levels of regulatory concern and returned to natural


120

background levels in coastal waters. Metal concentrations were higher in The Narrows and this may be associated with natural inputs from mangrove areas. A new hydrodynamic model being tested in the harbour determined that water flushing times are probably longer than previously thought; around 19–24 days. Metal concentrations were high in a range of plants and animals living within the port compared to those from a coastal reference area and this may be due to the high retention time of water (and therefore available contaminants) in the harbour. New tools were used to assess the health of organisms exposed to contaminants in the harbour. These included the use of biochemical markers such as stress enzymes in oysters and imposex (growth of male genitalia in females) in snails. In addition, laboratory studies explored the different responses of organisms to periodic or continuous exposure to discharged contaminants. Sediments were found to contain much lower levels of naphthalene (a potentially harmful polycyclic aromatic hydrocarbon or PAH) than previously thought. PAHs are derived from a number of sources including oil shale and coal, but levels across the harbour were well below guidelines. Sediment cores also indicated that there have been no major contaminant inputs in recent history. The pathway of contaminants up through the aquatic food chain was determined to be complex and scientists are still unravelling the story.

Two articles were published in the Gladstone Observer on: (i) the oyster

bioaccumulation/biomarker studies (Figure A.1) and (ii) collaborations with Griffith

and City (Hong Kong) Universities.

Figure A.1. Example of a contaminant pathways article published in the Gladstone Observer, 2005

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