NATIONAL WATER QUALITY MANAGEMENT STRATEGY

PAPER No. 4

Australian and New Zealand Guidelines for

Fresh and Marine Water Quality

Volume 1

The Guidelines

(Chapters 1–7)

October 2000

Australian and New Zealand

Environment and Conservation

Council

Agriculture and Resource

Management Council of Australia

and New Zealand
Copies of this publication may be obtained from:

Australian Water Association

PO Box 388

ARTARMON NSW 2064

Tel: (02) 9413 1288

Fax: (02) 9413 1047

OR

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and Townsville. For locations, contact and ordering

details go to: http://www.dofa.gov.au/infoaccess/

general/purchase_info_products.html

or phone132 447 (toll free in Australia, 24 hr service)

OR

NZ Water & Wastes Association

PO Box 13880

Onehunga, Auckland 1006

New Zealand

Tel: 64-9-636-3636

Fax: 64-9-636-1234

Email: water@nzwwa.org.nz

Material included in this document may be

freely reproduced provided that due

acknowledgment is given to the Australian and

New Zealand Environment and Conservation

Council and the Agriculture and Resource

Management Council of Australia and New

Zealand.

For further information on

acknowledgment, contact:

The Secretary

Australian and New Zealand Environment

and Conservation Council

GPO Box 787

CANBERRA ACT 2601

Tel: (02) 6274 1428

Fax: (02) 6274 1858

OR

The Secretary

Agriculture and Resource Management

Council of Australia and New Zealand

GPO Box 858

CANBERRA ACT 2601

Tel: (02) 6272 5216

Fax: (02) 6272 4772

Environment Australia

Cataloguing-in-Publication Data:

Australian and New Zealand guidelines for fresh and

marine water quality. Volume 1, The guidelines /

Australian and New Zealand Environment and

Conservation Council, Agriculture and Resource

Management Council of Australia and New Zealand.

Bibliography.

Includes index.

(National water quality management strategy; no.4)

ISBN 09578245 0 5 (set)

ISSN 1038 7072

1. Water quality – Australia – Measurement.

2. Water quality – New Zealand – Measurement.

3. Water – Pollution – Environmental aspects –

Australia. 4. Water – Pollution – Environmental

aspects – New Zealand. 5. Water quality

management – Australia. 6. Water quality

management – New Zealand. I. Australian and New

Zealand Environment and Conservation Council.

II. Agriculture and Resource Management Council of

Australia and New Zealand. III. Series

628.161’0994-dc21

Disclaimer

The contents of this document have been

compiled using a range of source materials and

while reasonable care has been taken in its

compilation, the member governments of

ANZECC and ARMCANZ and the

organisations and individuals involved with the

compilation of this document shall not be

liable for any consequences which may result

from using the contents of this document.

Printed in Australia on recycled paper for the

Australian and New Zealand Environment and

Conservation Council and the Agriculture and

Resource Management Council of Australia

and New Zealand.
Version — October 2000 page iii

Contents

Figures vii

Tables viii

Boxes xi

Case studies xi

Preamble xii

Acknowledgments xv

1 Introduction 1–1

1.1 Background 1–3

1.2 Guiding principles 1–5

1.3 Objectives 1–6

2 A framework for applying the guidelines 2–1

2.1 Water quality management framework 2–1

2.1.1 The broad strategy 2–1

2.1.2 Stakeholder involvement 2–4

2.1.3 Environmental values 2–6

2.1.4 Water quality guidelines 2–9

2.1.5 Water quality objectives 2–11

2.2 Application of the guidelines for water quality management 2–11

2.2.1 Philosophical approach to applying the guidelines 2–12

2.2.2 Mixing zones 2–17

2.2.3 Application of water quality prediction models 2–18

2.2.4 Deriving guidelines for compounds where no guidelines currently

exist 2–19

3 Aquatic ecosystems 3.1–1

3.1 Issues for all indicator types 3.1–1

3.1.1 Philosophy and steps to applying the guidelines 3.1–1

3.1.2 Features and classification of aquatic ecosystems in Australia and

New Zealand 3.1–7

3.1.3 Assigning a level of protection 3.1–9

3.1.4 Defining a reference condition 3.1–14

3.1.5 Decision frameworks for assessing test site data and deriving sitespecific water quality guidelines 3.1–17
Contents

page iv Version — October 2000

3.1.6 Using management goals to integrate water quality assessment 3.1–19

3.1.7 Decision criteria and trigger values 3.1–20

3.1.8 Guidelines for highly disturbed ecosystems 3.1–22

3.2 Biological assessment 3.2–1

3.2.1 Introduction and outline 3.2–1

3.2.2 Matching indicators to problems 3.2–6

3.2.3 Recommended experimental design and analysis procedures for

generic protocols 3.2–10

3.2.4 Guidelines for determining an unacceptable level of change 3.2–13

3.2.5 Assessing the success of remedial actions 3.2–20

3.3 Physical and chemical stressors 3.3–1

3.3.1 Introduction 3.3–1

3.3.2 Philosophy used in developing guidelines for physical and chemical

stressors 3.3–3

3.3.3 Guideline packages for applying the guideline trigger values to sites 3.3–21

3.4 Water quality guidelines for toxicants 3.4–1

3.4.1 Introduction 3.4–1

3.4.2 How guidelines are developed for toxicants 3.4–1

3.4.3 Applying guideline trigger values to sites 3.4–11

3.5 Sediment quality guidelines 3.5–1

3.5.1 Introduction 3.5–1

3.5.2 Underlying philosophy of sediment guidelines 3.5–1

3.5.3 Approach and methodology used in trigger value derivation 3.5–2

3.5.4 Recommended guideline values 3.5–3

3.5.5 Applying the sediment quality guidelines 3.5–5

4 Primary industries 4.1–1

4.1 Introduction 4.1–1

4.2 Water quality for irrigation and general water use 4.2–1

4.2.1 Philosophy 4.2–1

4.2.2 Scope 4.2–1

4.2.3 Biological parameters 4.2–2

4.2.4 Irrigation salinity and sodicity 4.2–4

4.2.5 Major ions of concern for irrigation water quality 4.2–9

4.2.6 Heavy metals and metalloids 4.2–11

4.2.7 Nitrogen and phosphorus 4.2–12
Contents

Version — October 2000 page v

4.2.8 Pesticides 4.2–13

4.2.9 Radiological quality of irrigation water 4.2–13

4.2.10 General water uses 4.2–15

4.3 Livestock drinking water quality 4.3–1

4.3.1 Derivation and use of guidelines 4.3–1

4.3.2 Biological parameters 4.3–1

4.3.3 Major ions of concern for livestock drinking water quality 4.3–2

4.3.4 Heavy metals and metalloids 4.3–4

4.3.5 Pesticides and other organic contaminants 4.3–5

4.3.6 Radiological quality of livestock drinking water 4.3–6

4.4 Aquaculture and human consumption of aquatic foods 4.4–1

4.4.1 Background 4.4–1

4.4.2 Philosophy 4.4–2

4.4.3 Scope 4.4–2

4.4.4 Water quality guidelines for the protection of cultured fish, molluscs

and crustaceans 4.4–3

4.4.5 Water quality guidelines for the protection of human consumers of

aquatic foods 4.4–12

4.4.6 Some precautionary comments 4.4–18

4.4.7 Priorities for research and development 4.4–19

5 Guidelines for recreational water quality and aesthetics 5–1

5.1 Guidelines for users in New Zealand 5–1

5.2 Guidelines for users in Australia 5–1

5.2.1 Introduction 5–2

5.2.2 Recreational categories 5–3

5.2.3 Detailed water quality guidelines 5–4

6 Drinking water 6–1

6.1 Guidelines for users in New Zealand 6–1

6.2 Guidelines for users in Australia 6–1

6.2.1 Microbiological quality of drinking water 6–1

6.2.2 Chemical and radiological quality of drinking water 6–2

6.2.3 Small water supplies 6–2

6.2.4 Individual household supplies 6–3

6.2.5 Guideline values 6–3
Contents

page vi Version — October 2000

7 Monitoring and assessment 7.1–1

7.1 Introduction 7.1–1

7.1.1 Integrated monitoring strategies 7.1–1

7.1.2 Framework for a monitoring and assessment program 7.1–4

7.2 Choosing a study design 7.2–1

7.2.1 Recommendations for combinations of indicators for aquatic

ecosystems 7.2–1

7.2.2 Broad classes of monitoring design 7.2–3

7.2.3 Checklist of issues in refining program design 7.2–5

7.2.4 Sampling protocols and documentation 7.2–18

7.2.5 Sample processing and analysis 7.2–18

7.2.6 Data analysis, evaluation and reporting 7.2–19

7.3 Specific issues for biological indicators 7.3–1

7.3.1 Issues for univariate indicators 7.3–1

7.3.2 Issues for multivariate indicators 7.3–1

7.3.3 Use of AUSRIVAS 7.3–3

7.4 Specific issues for physical and chemical indicators (including

toxicants) of water and sediment 7.4–1

7.4.1 Hydrology and representative sampling 7.4–2

7.4.2 Chemical speciation in water samples 7.4–2

7.4.3 Quality Assurance and Quality Control (QA/QC) 7.4–3

7.4.4 Comparing test data with guideline trigger values 7.4–3

References

Chapter 1 Introduction R1–1

Chapter 2 A framework for applying the guidelines R2–1

Chapter 3 Aquatic ecosystems R3–1

Section 3.1 Introduction R3.1–1

Section 3.2 Biological assessment R3.2–1

Section 3.3 Physical and chemical stressors R3.3–1

Section 3.4 Water quality guidelines for toxicants R3.4–1

Section 3.5 Sediment quality guidelines R3.5–1

Chapter 4 Primary industries R4–1

Sections 4.1 Introduction R4.1–1

Sections 4.2 & 4.3 Agricultural water uses (irrigation and general water

use; livestock drinking water quality) R4.2/3–1
Contents

Version — October 2000 page vii

Section 4.4 Aquaculture and human consumers of aquatic foods R4.4–1

Chapter 5 Guidelines for recreational water quality and aesthetics R5–1

Chapter 6 Drinking water R6–1

Chapter 7 Monitoring and assessment R7–1

Appendices

Appendix 1 Acronyms and glossary of terms A–1

Appendix 2 The National Water Quality Management Strategy A–21

Appendix 3 Recent water quality documents of the NZ Ministry for the

Environment A–23

Appendix 4 Development of the revised guidelines A–25

Appendix 5 Basis of the proposed guidelines for recreational water

quality and aesthetics in Australia A–29

Index index 1

Figures

Figure 2.1.1 Management framework for applying the guidelines 2–2

Figure 3.1.1 Flow chart of the steps involved in applying the guidelines

for protection of aquatic ecosystems 3.1–4

Figure 3.1.2 Procedures for deriving and refining trigger values for

physical and chemical stressors and toxicants in water and

sediment 3.1–6

Figure 3.1.3 Classification of ecosystem type for each of the broad

categories of indicators 3.1–9

Figure 3.1.4 Graphical depiction of the relationship between indicator

response and strength of disturbance, and threshold for

management intervention 3.1–20

Figure 3.2.1 Decision tree for biological assessment of water quality 3.2–3

Figure 3.3.1 Decision tree framework (‘guideline packages’) for

assessing the physico-chemical stressors in ambient waters 3.3–2

Figure 3.3.2 Types of physical and chemical stressors 3.3–3

Figure 3.4.1 Simplified decision tree for assessing toxicants in ambient

waters 3.4–14

Figure 3.4.2 Decision tree for metal speciation guidelines 3.4–19

Figure 3.5.1 Decision tree for the assessment of contaminated

sediments 3.5–6

Figure 4.2.1 Flow diagram for evaluating salinity and sodicity impacts of

irrigation water quality 4.2–5
Contents

page viii Version — October 2000

Figure 4.2.2 Relationship between SAR and EC of irrigation water for

prediction of soil structural stability 4.2–9

Figure 4.4.1 Decision tree for determining water quality acceptable for

the protection of aquaculture species 4.4–6

Figure 7.1.1 Procedural framework for the monitoring and assessment

of water quality (the shaded area). 7.1–5

Figure 7.2.1 Flow chart depicting the broad categories of designs for

monitoring and assessment that apply in different contexts. 7.2–6

Figure 7.3.1 Schematic diagram of a river system with paired upstream

and downstream sites on each tributary. 7.3–2

Figure 7.4.1 Control chart showing physical and chemical data for test

and reference sites plotted against time and recommended actions 7.4–7

Figure 7.4.2 Control chart showing physical and chemical data for test

site plotted against default trigger value, time and recommended

actions 7.4–8

Figure A1 National Water Quality Management Strategy A–21

Tables

Table 3.1.1 Some features of Australian and New Zealand ecosystems

that have possible consequences for water quality assessment and

ecosystem protection 3.1–8

Table 3.1.2 Recommended levels of protection defined for each indicator

type 3.1–13

Table 3.2.1 Biological assessment objectives for different management

situations and the recommended methods and indicators 3.2–7

Table 3.2.2 Water quality issues and recommended biological indicators

for different ecosystem types 3.2–11

Table 3.2.3 Experimental design and analysis procedures to apply to

generic protocols 3.2–13

Table 3.2.4 Division of AUSRIVAS O/E indices into bands or categories

for reporting 3.2–16

Table 3.3.1 Summary of the condition indicators, performance indicators,

and location of default trigger value tables, for each issue 3.3–5

Table 3.3.2 Default trigger values for physical and chemical stressors for

south-east Australia for slightly disturbed ecosystems 3.3–10

Table 3.3.3 Ranges of default trigger values for conductivity, turbidity and

suspended particulate matter indicative of slightly disturbed

ecosystems in south-east Australia 3.3–11

Table 3.3.4 Default trigger values for physical and chemical stressors for

tropical Australia for slightly disturbed ecosystems 3.3–12
Contents

Version — October 2000 page ix

Table 3.3.5 Ranges of default trigger values for conductivity, turbidity and

suspended particulate matter indicative of slightly disturbed

ecosystems in tropical Australia 3.3–13

Table 3.3.6 Default trigger values for physical and chemical stressors for

south-west Australia for slightly disturbed ecosystems 3.3–14

Table 3.3.7 Range of default trigger values for conductivity, turbidity and

suspended particulate matter indicative of slightly disturbed

ecosystems in south-west Australia 3.3–15

Table 3.3.8 Default trigger values for physical and chemical stressors for

south central Australia — low rainfall areas — for slightly disturbed

ecosystems 3.3–16

Table 3.3.9 Ranges of default trigger values for conductivity, turbidity and

suspended particulate matter indicative of slightly disturbed

ecosystems in south central Australia — low rainfall areas 3.3–16

Table 3.3.10 Default trigger values for physical and chemical stressors in

New Zealand for slightly disturbed ecosystems 3.3–17

Table 3.3.11 Default trigger values for water clarity and turbidity indicative

of unmodified or slightly disturbed ecosystems in New Zealand 3.3–18

Table 3.4.1 Trigger values for toxicants at alternative levels of protection 3.4–5

Table 3.4.2 General framework for applying levels of protection for

toxicants to different ecosystem conditions 3.4–11

Table 3.4.3 General form of the hardness-dependent algorithms

describing guideline values for selected metals in freshwaters 3.4–21

Table 3.4.4 Approximate factors to apply to soft water trigger values for

selected metals in freshwaters of varying water hardness 3.4–21

Table 3.5.1 Recommended sediment quality guidelines 3.5–4

Table 4.2.1 Key issues concerning irrigation water quality effects on soil,

plants and water resources 4.2–2

Table 4.2.2 Trigger values for thermotolerant coliforms in irrigation waters

used for food and non-food crops 4.2–3

Table 4.2.3 Soil type and average root zone leaching fraction 4.2–6

Table 4.2.4 Soil and water salinity criteria based on plant salt tolerance

groupings 4.2–7

Table 4.2.5 Tolerance of plants to salinity in irrigation water 4.2–8

Table 4.2.6 Chloride concentrations causing foliar injury in crops of

varying sensitivity 4.2–10

Table 4.2.7 Risks of increasing cadmium concentrations in crops due to

chloride in irrigation waters 4.2–10

Table 4.2.8 Sodium concentration causing foliar injury in crops of varying

sensitivity 4.2–10
Contents

page x Version — October 2000

Table 4.2.9 Effect of sodium expressed as sodium adsorption ratio on

crop yield and quality under non-saline conditions 4.2–11

Table 4.2.10 Agricultural irrigation water long-term trigger value, shortterm trigger value and soil cumulative contaminant loading limit

triggers for heavy metals and metalloids 4.2–11

Table 4.2.11 Agricultural irrigation water long-term trigger value and

short-term trigger value guidelines for nitrogen and phosphorus 4.2–13

Table 4.2.12 Interim trigger value concentrations for a range of herbicides

registered in Australia for use in or near waters 4.2–14

Table 4.2.13 Trigger values for radiological contaminants for irrigation

water 4.2–15

Table 4.2.14 Corrosion potential of waters on metal surfaces as indicated

by pH, hardness, Langelier index, Ryznar index and the log of

chloride:carbonate ratio 4.2–15

Table 4.2.15 Fouling potential of waters as indicated by pH, hardness,

Langelier index, Ryznar index and the log of chloride:carbonate ratio 4.2–16

Table 4.3.1 Tolerances of livestock to total dissolved solids in drinking

water 4.3–4

Table 4.3.2 Recommended water quality trigger values for heavy metals

and metalloids in livestock drinking water 4.3–5

Table 4.3.3 Trigger values for radiological contaminants in livestock

drinking water 4.3–6

Table 4.4.1 Representative aquaculture species, occurrence and culture

status 4.4–4

Table 4.4.2 Physico-chemical stressor guidelines for the protection of

aquaculture species 4.4–7

Table 4.4.3 Toxicant guidelines for the protection of aquaculture species 4.4–8

Table 4.4.4 Guidelines for the protection of human consumers of fish and

other aquatic organisms from bacterial infection 4.4–14

Table 4.4.5 Guidelines for chemical compounds in water found to cause

tainting of fish flesh and other aquatic organisms 4.4–16

Table 5.2.1 Water quality characteristics relevant to recreational use 5–2

Table 5.2.2 Summary of water quality guidelines for recreational waters 5–3

Table 5.2.3 Summary of water quality guidelines for recreational

purposes: general chemicals 5–9

Table 5.2.4 Summary of water quality guidelines for recreational

purposes: pesticides 5–10

Table 7.2.1 Broad categories of design relevant to the Guidelines listed

together with the assessment objectives that could be fulfilled by

each category 7.2–7
Contents

Version — October 2000 page xi

Table 7.4.1 Checklist for sampling and analysis of physical and chemical

indicators with cross-reference to details provided in the Monitoring

Guidelines 7.4–1

Boxes

Box 1.1 Water quality guidelines may be used to trigger action 1–2

Box 1.2 Application of the guidelines to groundwater 1–2

Box 2.1 Examples of stakeholder involvement in Australia and New

Zealand 2–5

Box 2.2 Cultural importance of water 2–7

Box 2.3 An alternative approach to statistical decision making 2–14

Box 2.4 Mixing zones adjacent to effluent outfalls 2–17

Box 3.1.1 Human activities affecting aquatic ecosystems 3.1–2

Box 3.1.2 Protecting biodiversity

Box 3.1.3 How to apply the guidelines 3.1–3

Box 3.1.4 Examples of water quality assessment programs conducted

in major urban regions of Australia 3.1–24

Box 3.2.1 A cautionary note on the use of the AUSRIVAS RBA

approach for site-specific assessments 3.2–9

Box 3.3.1. Sources of information for use when deriving low-risk trigger

values 3.3–6

Box 7.1.1 Enhancing inferences and defraying costs in environmental

monitoring programs 7.1–3

Box 7.2.1 Issues for restoration and rehabilitation 7.2–10

Box 7.2.2 Hypothesis testing in environmental monitoring and

assessment 7.2–12

Box 7.2.3 Application of ‘bioequivalence testing’ for environmental

restoration 7.2–13

Box 7.2.4 Effect sizes are implicit in some procedures 7.2–15

Box 7.2.5 Some suggestions for setting an effect size 7.2–16

Case studies

Case Study 1. Assessing the risk of cyanobacterial blooms in a lowland

river 3.3–28

Case Study 2. Establishing sustainable organic matter loads for

standing waterbodies 3.3–29

3.1–3
page xii Version — October 2000

Preamble

The Australian National Water Quality Management Strategy (NWQMS) aims to

achieve the sustainable use of Australia’s and New Zealand’s water resources by

protecting and enhancing their quality while maintaining economic and social

development. The NWQMS is a joint strategy developed by two Ministerial

Councils — the Agriculture and Resources Management Council of Australia and

New Zealand (ARMCANZ) and the Australian and New Zealand Environment and

Conservation Council (ANZECC). The Australian National Health and Medical

Research Council (NHMRC) is involved in aspects of the strategy that affect public

health. The NWQMS aims to meet future needs by providing policies, a process

and national guidelines for water quality management.

Further information on the National Water Quality Management Strategy is

provided in Appendix 2.

The Australian Water Quality Guidelines for Fresh and Marine Waters (ANZECC

1992) was one of a suite of 21 documents forming the NWQMS and was released

in 1992 as one of the first guideline documents. In 1993 the ANZECC Standing

Committee on Environmental Protection (SCEP) agreed to review the water quality

guidelines to incorporate current scientific, international and national information

in a clear and understandable document.

Since the ANZECC Guidelines were published in 1992 there have been a number

of important advances. First, there have been some major policy initiatives at

federal and state level that, combined with the National Water Quality

Management Strategy, have increased the focus of attention on ecologically

sustainable management of water resources in Australia and New Zealand (e.g.

Council of Australian Governments (COAG) reform framework, State of the

Environment reporting, and modification and implementation of the NZ Resource

Management Act). Second, there is a pleasing trend towards a more holistic

approach to the management of aquatic systems. Third, as initially recommended

in the 1992 ANZECC Guidelines, there has been an increased use of biological

indicators to assess and monitor the ‘health’ of aquatic ecosystems. Finally, a

number of major environmental studies (e.g. the Port Phillip Bay Study in Victoria,

the Southern Metropolitan Coastal Waters Study in Western Australia) have led to

significant advances in knowledge about estuarine and coastal ecosystems.

The scope of this revised version, the Australian and New Zealand Guidelines for

Fresh and Marine Water Quality, has also been extended to include a consideration

of both Australia’s and New Zealand’s water resources. The review program is

outlined in Appendix 4.

The Guidelines have been revised using data, relevant literature, and other

information available to at least 1996, specifically:

• Databases used to derive guideline values for toxicants and sediments

(Chapter 3) and aquaculture and human consumers of aquatic foods

(Chapter 4) have been updated to include information available to late 1996,

while default guidelines for physical and chemical stressors (Chapter 3) have

been derived from databases current to early 2000.

• The guidelines for biological indicators (Chapter 3), advice for monitoring and

assessment (Chapter 7) and support text for physical and chemical stressors
Preamble

Version — October 2000 page xiii

(Chapter 3) have been revised to include information available to late 1998.

However, all support text for aquatic ecosystems (Chapters 3 and 8) and

aquaculture and human consumers of aquatic foods (Chapters 4 and 9) capture

important developments and key references available to early 2000.

• The guidelines for agricultural water uses (irrigation and general water use and

livestock drinking water, Chapters 4 and 9) have been revised to include

information available to early 2000.

• The guidelines for recreational water quality and aesthetics (Chapter 5) are still

in revision in Australia, while New Zealand readers are referred to the relevant

1999 guidelines. For guidelines for drinking water (Chapter 6), Australian and

New Zealand readers are referred to the relevant 1996 and 1995 guidelines

respectively.

To be continuously relevant to its users, the Australian and New Zealand

Guidelines for Fresh and Marine Water Quality, like other NWQMS benchmark

documents, will require ongoing review and revision. The present version was

current up to October 2000. Users are invited to comment on the Australian and

New Zealand Guidelines for Fresh and Marine Water Quality by contacting the

offices listed on the next page. These addresses can also receive comments on the

Australian Guidelines for Water Quality Monitoring and Reporting, so users

should name the document to which their comments apply.
page xiv Version – October 2000

Contacts for comments and information

For information and advice about the Water Quality Guidelines and to advise of possible errors, omissions and

changes required for future revisions, please contact the designated agency for your state or territory in

Australia or for New Zealand. The agency contacts are listed below.

Australian Capital Territory

Bob Neil

Environment ACT

PO Box 144, Lyneham, ACT 2602

Tel: (02) 6207 2581 Fax: (02) 6207 6084

e-mail: robert.neil@act.gov.au

New South Wales

Pollution Line

NSW Environment Protection Authority

PO Box A290, Sydney South NSW 1232

Tel: 131 555 Fax: (02) 9995 5911

e-mail: info@epa.nsw.gov.au

Northern Territory

Director Resource Management

Natural Resources Division

Department of Lands Planning & Environment

PO Box 30, Palmerston NT 0831

Tel: (08) 8999 4455 Fax: (08) 8999 4403

e-mail: michael.lawton@nt.gov.au

Queensland

All enquiries for Chapter 4 (volume 1) and Volume

3 should be addressed to:

Heather Hunter

Department of Natural Resources

Block B 80 Meiers Road

Indooroopilly, QLD 4068

Tel: (07) 3896 9637 Fax: (07) 3896 9591

e-mail: heather.hunter@dnr.qld.gov.au

All other enquiries to:

Andrew Moss

Environmental Protection Agency

PO Box 155

Brisbane Albert Street, QLD 4002

Tel: (07) 3896 9245 Fax: (07) 3896 9232

e-mail: andrew.moss@env.qld.gov.au

South Australia

Manager

Evaluation Branch

South Australian Environment Protection Agency

GPO Box 2607, Adelaide SA 5001

Tel: (08) 8204 2055 Fax: (08) 8204-2107

e-mail: Cugley.John@saugov.sa.gov.au

Tasmania

Director of Environmental Management

Environmental Planning and Scientific Services

Scientific and Technical Branch, Water Section

Department of Primary Industries, Water and

Environment

GPO Box 44A, Hobart, TAS 7001

Tel: (03) 6233 6518 Fax: (03) 6233 3800

e-mail: waterinfo@dpiwe.tas.gov.au

Victoria

Lisa Dixon

Manager Freshwater Sciences

Environment Protection Authority

GPO Box 4395QQ, Melbourne VIC 3001

Tel: (03) 9616 2361 Fax: (03) 9614 3575

e-mail: lisa.dixon@epa.vic.gov.au

Western Australia

Victor Talbot

Department of Environmental Protection

PO Box K822, Perth WA 6842

Tel: (08) 9222 8655 Fax: (08) 9322 1598

e-mail: victor_talbot@environ.wa.gov.au

NEW ZEALAND

Nigel Bradly

Land and Water Group

Ministry for the Environment

PO Box 10362

Wellington NEW ZEALAND

Tel: NZ (04) 917 7489 Fax: NZ (04) 917 7523

Mobile: NZ 025 379 391

e-mail: nigel.bradly@mfe.govt.nz
page xv Version — October 2000

Acknowledgments

The revision of the Australian and New Zealand Guidelines for Fresh and Marine

Water Quality with its extended scope was a challenging and difficult task given

the complexity of the issues and the many advances made recently in water

resource management worldwide. The task could not have been possible without

the substantial assistance received from a large number of people and organisations

across Australia and New Zealand over the last four years.

Both the Project Committee and ANZECC and ARMCANZ Contact Group and its

working parties (Appendix 4) co-ordinated input from all the relevant government

jurisdictions, water quality experts, industry and conservation groups, and kept the

revision on track.

ANZECC and ARMCANZ would like to acknowledge the efforts of the following

organisations which coordinated the drafting of key chapters or sections of the

Guidelines:

• eriss

Overall coordination, Chapters 1 & 2, biological assessment for ecosystem

protection and monitoring and assessment;

• CRC for Freshwater Ecology

Physical and chemical stressors for ecosystem protection;

• NSW EPA

Toxicants for ecosystem protection;

• NIWA (NZ)

Metal data for toxicants;

• CSIRO Energy Technology

Sediment quality for ecosystem protection;

• QLD Dept Natural Resources

Primary Industries, Agriculture: irrigation, livestock & general water use;

• CSIRO Land and Water

Primary Industries, Agriculture: irrigation;

• PSM Group P/L and Dosaqua

Primary Industries, Aquaculture and harvesting of aquatic foods;

• NH&MRC

Recreation and aesthetics and drinking water.

A large number of individuals contributed to the review, some on an unpaid basis,

and by way of each of the disciplines and sections we would like to give particular

thanks to: Introduction, philosophical approach and general coordination: Kevin

McAlpine and Chris Humphrey; Aquatic ecosystems: Biological assessment — Chris

Humphrey, Leon Barmuta, Peter Davies (UTas), Jenny Davis, Bruce Mapstone,

Kerry Trayler, Trevor Ward and Barry Biggs, and for additional specific protocols

or regional information, James Boyden, Peter Davies (UWA), Terry Hume, Ross

Jeffree, Chris Madden, Rob Murdoch, Mika Peck, Bob Pidgeon, Abbie Spiers and

Rick van Dam; Physical and chemical stressors — Barry Hart, Ian Lawrence, Bill

Maher, Mark O’Donohue, Wendy van Dok and David Buckley; Toxicants — John

Chapman, Michael Warne, Scott Markich, Rick van Dam, Meenatchi Sunderam,

Fleur Pablo, Rebecca Rose, Graeme Batley, Simon Apte, Jenny Stauber, Paul

Brown and Chris Hickey; Sediment quality — Graeme Batley, Bill Maher and Chris

Hickey; Primary industries, aquaculture and harvesting of aquatic foods: Dos

O’Sullivan, Eva-Maria Bernoth, Susan Duda, Peter Montague, Barry Munday, Paul

Martin, Barbara Nowak, John Purser and Rick van Dam; Primary industries,
Acknowledgments

page xvi Version — October 2000

agriculture: Irrigation (salinity, sodicity, major ions, biological parameters and pesticides),

livestock and general water use — Heather Hunter, Rob DeHayr, Nicole Diatloff, Ian

Gordon and Roger Shaw; Irrigation (metals, metalloids, nitrogen and phosphorus) —

Steve Rogers, Mike McLaughlin and Daryl Stevens; Recreation and aesthetics:

Richard Lugg and Phil Callan; Drinking water: Richard Lugg and Phil Callan;

Monitoring and assessment: Leon Barmuta, Chris le Gras, David Fox, Bruce

Mapstone, Chris Humphrey, Peter Davies (UTas), Bill Maher, Graham Skyring and

Trevor Ward.

The basis of the risk-based approach used for applying the guidelines to aquatic

ecosystem protection was developed at a workshop attended by Barry Hart, Ian

Lawrence, Bill Maher, Wendy van Dok, Graeme Batley, Peter Cullen, Chris Bell,

Bill Dennison, Graham Harris, Graeme McBride, John Parslow and Eric Pyle, with

later input from Kevin McAlpine, Chris Humphrey and John Chapman. Additional

contributions to the Guidelines arose from the following agencies — MfE (NZ),

NIWA (NZ), Australian Radiation Laboratory, and QDPI. We also wish to thank

Ann Webb, Professor Arthur McComb and Ann Milligan for editing of the

documents and Paul Bainton, Peter Breen, Mike Burch, Stephanie Butcher, Bruce

Chessman, Geoff Coade, Malcolm Cooper, Carolyn Davies, Jim Fitzgerald, Terry

Hume, Gary Jones, Peter Liston, Ray Masini, Keith McGuinness, Robert

McLaughlan Rob Murdoch, Jon Nevill, Eric Pyle, Tim Riding, Jane Roberts, Liz

Rogers, Alan Thomas, Rod Thomas, Michael Tyler, Bob Zuur and Robin Wilson

for minor contributions and useful comments made upon early guidelines drafts.

During the public consultation phase held over July–October 1999, 96 submissions

were received from individuals and agencies from across Australia and New

Zealand. The comments received in these submissions were used as the basis for

the final drafting conducted in early 2000 and these resulted in substantial

improvements to the Guidelines. There are also many other organisations and

individuals who contributed to the review and, although they are too numerous to

list here, we would like to thank them all.

Finally, a special thanks to eriss who co-funded the review and whose staff coordinated technical aspects of the revision and provided logistical support.

Photography on ring-binder cover of the Guidelines (from top left)

Child drinking, Ministry for the Environment, New Zealand;

Logan River (Qld), Qld EPA;

Maori war boat, Photosource New Zealand Limited Image Library;

Hereford, Qld EPA;

Irrigation channels, Bruce Cooper, NSW DLWC;

Water quality monitoring, Bruce Cooper;

Bondi Beach (NSW), Brian Robson;

Rakaia River, South Island, New Zealand, Clint McCullough, eriss;

Aquaculture, Qld EPA;

Beenleigh Rum distillery (Qld), Qld EPA;

Aboriginal cultural ceremony on upper Katherine River (NT), Diane Lucas:

‘Because our great grandmothers and grandfathers been here before, their spirits are still here.

Now the spirits smell your sweat and it goes down to the deep water and makes it alright for you

to be here, without any harm’ — Margaret Oenpelli & Penny Long, Barunga, NT. The

washing of people by spraying water on their head is an Arnhem Land ceremony for

newcomers to country to keep away bad health for people and water.
Version — October 2000 page 1–1

1 Introduction

This document updates the Australian water quality guidelines for fresh and marine

waters released in 1992 (ANZECC 1992).

Specifically, this document:

• outlines the important principles, objectives and philosophical basis

underpinning the development and application of the guidelines;

• outlines the management framework recommended for applying the water

quality guidelines to the natural and semi-natural marine and fresh water

resources in Australia and New Zealand;

• provides a summary of the water quality guidelines proposed to protect and

manage the environmental values supported by the water resources;

• provides advice on designing and implementing water quality monitoring and

assessment programs;

• has been revised using data, relevant literature, and other information available

to at least 1996.

A note on the structure and other features of the Guidelines

Readers should note the following features of the Guidelines:

• Given the broad scope of the Guidelines, it has been necessary to load much of the

detailed rationale and reference information, including software, onto a CD-ROM, which

is in the pocket of the ring-bound folder.

• While many users will be satisfied with use of the default guideline values provided in

this volume, others will want to tailor guidelines for local conditions, or may simply seek

further reading. To assist users to refine the guidelines in this way or to acquire further

information, cross-reference to the support information referred to in point 1 above is

provided. These cross-references are indicated in the text by way of superscript letters

that link the relevant passage to the corresponding italicised notes in the left hand

margin of the page.

• The loose-leaf format of the Guidelines is a feature that will enable discrete subject

areas to be revised in future independently of other sections. To assist this, the page

numbering is independent for each of the short chapters (e.g. 2–1 to 2–xx) and to the

first subsection level of the longer, more complex chapters, i.e. chapters 3, 4 and 7 (this

volume), 8 (Volume 2) and 9 (Volume 3) (e.g. pages 3.1–1 to 3.1–xx, 3.2–1 to 3.2–xx

etc).

• A glossary of the main terms is provided at the end of this volume to assist readers

further in understanding the main issues. Users are encouraged to check the glossary

for all key terms because the terminology used by the various jurisdictions throughout

Australia and New Zealand is not always consistent with the terminology used in these

Guidelines.

These Guidelines should not be used as mandatory standards because there is

significant uncertainty associated with the derivation and application of water

quality guidelines. For example, data on biological effects are not available for all
Chapter 1 — Introduction

page 1–2 Version — October 2000

local species; there is uncertainty over the behaviour of contaminants in the field;

there is uncertainty in water quality measurements. The user should be aware of

this uncertainty when determining if an environmental value has been supported or

not. However, the Guidelines should provide a framework for recognising and

protecting water quality for the full range of existing environmental values.a The

Guidelines also provide risk-based decision frameworks wherever possible, simply

to help the user refine guideline trigger values for application at local and/or

regional scales.

Box 1.1 Water quality guidelines may be used to trigger action

The guidelines provided in this document are designed to help users assess whether the

water quality of a water resource is good enough to allow it to be used for humans, food

production or aquatic ecosystems (these uses are termed environmental values). If the

water quality does not meet the water quality guidelines, the waters may not be safe for

those environmental values and management action could be triggered to either more

accurately determine whether the water is safe for that use or to remedy the problem.

For some environmental values the guideline number provided may be an adequate guide to

quality (e.g. for recreation or drinking). For other specific environmental values the guideline

can be just a starting point to trigger an investigation to develop more appropriate guidelines

based on the type of water resource and inherent differences in water quality across

regions. For water whose environmental value is aquatic ecosystem protection, for example,

the investigation should aim to develop and adapt these guidelines to suit the local area or

region. This document incorporates protocols and quite detailed advice to assist users in

tailoring the water quality guidelines to local conditions. Invariably, the process of refining

these guidelines — ‘trigger values’ — to local conditions will result in numbers for toxicants

at least, that are less conservative and hence less constraining on surrounding activities.

Box 1.2 Application of the guidelines to groundwater

Groundwater is an essential water resource for many aquatic ecosystems, and for substantial

periods it can be the sole source of water to some rivers, streams and wetlands. Groundwater

is also very important for primary and secondary industry as well as for domestic drinking

water, particularly in low rainfall areas with significant underground aquifers.

Generally these Guidelines should apply to the quality both of surface water and of

groundwater since the environmental values which they protect relate to above-ground uses

(e.g. irrigation, drinking water, farm animal or fish production and maintenance of aquatic

ecosystems). Hence groundwater should be managed in such a way that when it comes to the

surface, whether from natural seepages or from bores, it will not cause the established water

quality objectives for these waters to be exceeded, nor compromise their designated

environmental values. An important exception is for the protection of underground aquatic

ecosystems and their novel fauna. Little is known of the lifecycles and environmental

requirements of these quite recently-discovered communities, and given their high

conservation value, the groundwater upon which they depend should be given the highest

level of protection.

As a cautionary note the reader should be aware that different conditions and processes

operate in groundwater compared with surface waters and these can affect the fate and

transport of many organic chemicals. This may have implications for the application of

guidelines and management of groundwater quality.

a

Environmental

values are

defined in

Section 2.1.3
Chapter 1 — Introduction

Version — October 2000 page 1–3

The present Water Quality Guidelines have been prepared under the auspices of

Australia’s National Water Quality Management Strategy (NWQMS) and relate to

New Zealand’s National Agenda for Sustainable Water Management (NASWM).

More information on the NWQMS is provided below and in Appendix 2.

Guidelines for the management of effluent discharges (including stormwater) and

other activities affecting water resources are covered in other NWQMS documents

(Appendix 2) and in the documents released by the NZ Ministry for the

Environment listed in Appendix 3. All of these guidelines are complementary, and

users are encouraged to take a holistic approach to water resource management by

integrating these documents with other considerations such as catchment

management and habitat related issues.

A 24-page introductory brochure that summarises the main features of the

guidelines is also available for users who are seeking a general overview.

1.1 Background

The current Guidelines, including this working volume, arise from a revision of the

NWQMS Guidelines published in 1992 (ANZECC 1992). The revision was

necessary to:

• incorporate current scientific, national and international information in a clear

and understandable format;

• ensure that the Guidelines complement major policy initiatives and directions

undertaken at the state and federal levels in the areas of ecologically

sustainable development and water resource management;

• promote a more holistic approach to aquatic ecosystem management;

• incorporate more detailed guidance on how to refine national or regional

guidelines for site-specific application.

Important input to the review process from Australia and New Zealand has

included: public submissions on the 1992 Guidelines and on an earlier draft of the

revised document; the most recent local and overseas scientific and resource

management documents and information; relevant overseas water quality guideline

documents and government submissions.

In keeping with the underlying philosophy of the 1992 Guidelines, the chapters in

this document describe how to apply state-of-the-art practices of water resource

management and assessment, for the protection of the environmental values. The

key changes in direction taken in revising the water quality guidelines are

summarised below.

Management strategy

• The management strategy adopted in the 1992 guidelines has been refined so

that it provides a greater focus on local environmental conditions, which

should allow the water quality guidelines to be tailored to specific sites or

regions.

Aquatic ecosystems

• Methods for deriving the physical and chemical water quality guidelines for

ecosystem management (now termed ‘guideline trigger values’) have also been
Chapter 1 — Introduction

page 1–4 Version — October 2000

updated in the light of an increased understanding of ecosystems, and improving

technologies.

• There is greater focus on issue-based management of water quality rather than

on the management of individual parameters. In practice, this means

integrating monitoring programs so that managers measure biological

parameters and related physical and chemical parameters, in both water and

sediment. Therefore guidelines have been developed for these other indicator

types (e.g. biological assessment, sediment quality and environmental flows).

Primary industries

• The Guidelines have amalgamated agriculture, aquaculture and human

consumption of aquatic foods into one environmental value called ‘Primary

Industries’.

Recreation and aesthetics

• At the time of publication of these Guidelines, the material for Australian users

on Guidelines for Recreational Water Quality and Aesthetics was still under

review. Until these Guidelines are revised and endorsed, users should apply the

guidelines from the Australian Water Quality Guidelines for Fresh and Marine

Waters (ANZECC 1992). In New Zealand, water managers should refer to the

Ministry for the Environment publication Recreational Water Quality

Guidelines (New Zealand Ministry for the Environment 1999).

Drinking water

• The Guidelines refer to the Australian NHMRC and ARMCANZ (1996)

Australian Drinking Water Guidelines and the New Zealand Ministry of Health

(1995) Drinking-Water Standards for New Zealand, to avoid duplication and

confusion.

Industrial water

• After extensive consultation with representative industrial groups, the current

Guidelines provide no specific guidance for industrial water use, because

industrial water requirements are so varied (both within and between

industries) and sources of water for industry have other coincidental

environmental values that tend to drive management of the resource. Industrial

water use continues to be a recognised environmental value that has high

economic benefit to the community. It must be given adequate consideration

during the planning and management of water resources.

Cultural issues

• The current Guidelines recognise that water resources have important cultural

and spiritual values, particularly for indigenous peoples. No specific guidance

for protection of these values is provided, but consideration must be given to

cultural issues in the planning and management of water resources, and as

required by existing legislation, regulations and guidelines.

Monitoring and assessment

• The Guidelines discuss the essential elements of water quality monitoring and

assessment programs, but with extensive reference to the recent NWQMS

Monitoring and Reporting Guidelines (ANZECC & ARMCANZ 2000).
Chapter 1 — Introduction

Version — October 2000 page 1–5

1.2 Guiding principles

The Australian and New Zealand Guidelines for Fresh and Marine Water Quality

are primarily based on the philosophy of ecologically sustainable development

(ESD). The Australian National Strategy for Ecologically Sustainable

Development (ESD Steering Committee 1992) defined ESD as:

[development] using, conserving and enhancing the community’s resources so that

ecological processes, on which life depends, are maintained, and the total quality of

life, now and in the future can be increased. Put more simply, ESD is development

which aims to meet the needs of Australians today, while conserving our ecosystems to

the benefit of future generations.

The need to comply with ESD principles is being included in statutes throughout

Australia, with the commitment to continuous environmental improvement through

comprehensive and integrated public policy.

In New Zealand, the Purpose and Principles in the Resource Management Act

(1991) (RMA) set out the philosophy and approach for water management. The

purpose of the RMA is to promote sustainable management, which is broadly

equivalent to the ESD philosophy.

The Guidelines are also based on the policies and principles of the Australian

National Water Quality Management Strategy which are explained in ANZECC

and ARMCANZ (1994). The principles include:

• ecologically sustainable development;

• an integrated approach to water quality management;

• community involvement in water resource management, including

establishment of the environmental values and development of management

plans;

• government endorsement of the water quality policy objectives.

Four further guiding principles have also been adopted:

• A coordinated and cooperative approach to water quality management is vital

and involves all spheres of government, the community, local and indigenous

groups and the private sector.

• The high variability and complexity inherent in natural water resources needs to

be recognised and taken into account when evaluating water quality or

developing management strategies.

• Water resources are special features of the environment and their quality and

integrity should be conserved and managed according to the intent of the

Australian National Strategy for Ecologically Sustainable Development, the

Wetlands Policy of the Commonwealth Government of Australia and the

National Strategy for the Conservation of Australia’s Biological Diversity.

• Ongoing research into the inter-relationships between ecological processes,

water quality and the biota, and the dissemination of these findings in a readily

usable form, are essential for effective management of water resources.
Chapter 1 — Introduction

page 1–6 Version — October 2000

1.3 Objectives

The primary objective of the Australian National Water Quality Management

Strategy (NWQMS) (ANZECC & ARMCANZ 1994) is based on ecologically

sustainable development of water resources. The main objective of the Guidelines

for fresh and marine water quality is intended to support this overall objective:

to provide an authoritative guide for setting water quality objectives required to sustain

current or likely future environmental values for natural and semi-natural water

resources in Australia and New Zealand.

It is recognised that a nationally consistent approach to water quality

management is underpinned by the development of high-status guidelines which

can provide guidance when issues arise. The adoption of national guidelines

provides a shared national objective while allowing flexibility of response to

different circumstances at regional and local levels. Where appropriate, state

and/or local jurisdictions can use their own legislative and regulatory tools to

refine these national water quality guidelines either into their own regional

guidelines or into specific water quality objectives.

The Guidelines are intended to provide government, industry, consultants and

community groups with a sound set of tools that will enable the assessment and

management of ambient water quality in a wide range of water resource types, and

according to designated environmental values. They are the recommended limits to

acceptable change in water quality that will continue to protect the associated

environmental values. They are not mandatory and have no formal legal status (e.g.

they are not National Environmental Standards as provided for in Section 43 of the

New Zealand Resource Management Act 1991). They also do not signify threshold

levels of pollution since there is no certainty that significant impacts will occur

above these recommended limits, as might be required for prosecution in a court of

law. Instead, the guidelines provide certainty that there will be no significant

impact on water resource values if the guidelines are achieved.

The management framework, guidelines, protocols and strategies set out here

complement other documents produced under the NWQMS umbrella (Appendix 2).
Version — October 2000 page 2–1

2 A framework for applying the guidelines

2.1 Water quality management framework

For the long-term management of any water resource, there must be:

• a designated and clearly stated set of environmental values;

• understanding of the links between human activity (including indigenous uses

and values) and environmental quality, at an acceptable level of confidence;

• unambiguous goals for management;

• appropriate water quality objectives; and

• effective management frameworks, including cooperative, regulatory, feed-back

and auditing mechanisms.

Management strategies that combine prediction, acknowledgment of uncertainty,

monitoring and review are sufficiently flexible to adapt as the knowledge base

improves. However, before management can decide on strategies that will ensure

ecologically sustainable development in the long-term, society must have a

collective vision of what it wants for each water resource, and there must be a good

scientific understanding of the impact of human activities on the resource.

Until recently, management of Australian and New Zealand water resources was

primarily focused on protecting environmental values based on human health, such

as quality of drinking water, agricultural water and water from which aquatic foods

are harvested. Maintenance of water quality to protect aquatic ecosystems was

often included, but based on a very deterministic view of ecosystems that assumed

that factors controlling ecosystem function could be identified and managed to

prevent problems. However, it is now well recognised that the relationships

between key ecological processes and their components are complex and variable

(probabilistic) and cannot be determined precisely. The guidelines provided in this

document attempt to take these factors into consideration.

2.1.1 The broad strategy

Australia and New Zealand both have a regional or local government framework in

place. The political boundaries imposed within Australia place most of the

responsibility for the management of natural resources with the states and

territories. In New Zealand primary responsibility for water management rests with

regional councils.

Water resource management is best implemented by integrating national, state and

regional powers and responsibilities, and by using complementary water quality

planning and policy tools. After all available and technical information has been

collated for a defined water body, the steps listed below (and shown in figure 2.1.1)

could be followed to implement a broad national management strategy at a local

level.

1. Identify the environmental values that are to be protected in a particular water

body and the spatial designation of the environmental values (i.e. decide what

values will apply where).
Chapter 2 — A framework for applying the guidelines

page 2–2 Version — October 2000

2. Identify management goals and then select the relevant water quality

guidelines for measuring performance. Based on these guidelines, set water

quality objectives that must be met to maintain the environmental values.

3. Develop statistical performance criteria to evaluate the results of the monitoring

programs (e.g. statistical decision criteria for determining whether the water

quality objectives have been exceeded or not).

4. Develop tactical monitoring programs focusing on the water quality objectives.

5. Initiate appropriate management responses to attain (or maintain if already

achieved) the water quality objectives.

(Note: Several of the key terms from the broad management strategy outlined

above, some of them in italics, are explained in the sections below.)

Define

PRIMARY MANAGEMENT AIMS

(including environmental values, management goals

and level of protection)

Define

WATER QUALITY OBJECTIVES

(specific water quality to be achieved)

• taking account of social, cultural, political and economic concerns where

necessary

Establish

MONITORING AND ASSESSMENT PROGRAM

(focused on water quality objectives)

• after defining acceptable performance or decision criteria

Initiate appropriate

MANAGEMENT RESPONSE

(based on attaining or maintaining water quality objectives)

Determine appropriate

WATER QUALITY GUIDELINES

(tailored to local environmental conditions)

Figure 2.1.1 Management framework for applying the guidelines
2.1.1 The broad strategy

Version — October 2000 page 2–3

The elements of this management strategy can be incorporated into comprehensive

planning practices such as integrated (or total) catchment management plans (ICM

or TCM) or can remain relatively small-scale plans for local areas. However, there

must be consultation with stakeholders and the effective use and integration of a

multi-disciplinary array of skills and knowledge to achieve success.

With respect to point 5 above, the management responses will depend on the issue of

concern, the cause(s) of the poor water quality and the available tools, and should be

negotiated and agreed upon by the local or regional stakeholders. In Australia,

strategic management can be in the form of catchment management plans or state or

national policies (e.g. statutory Environmental Protection Policies) and in New

Zealand, in the form of Regional Policy Statements, regional plans or National

Policy Statements, based on the agreed environmental values and their associated

water quality objectives. Regulation could be achieved through discharge consents

and codes of practice designed to ensure water quality objectives are not exceeded

and taking into account cumulative impacts from all sources.

The monitoring programs identified in point 4 above should be maintained during

and after implementation of the agreed management response(s), to evaluate their

performance in achieving the water quality objectives and hence the management

goals. This process should be iterative and on-going to ensure the environmental

values continue to be sustained.

2.1.1.1 Responsibilities

The NWQMS outlines a three-tiered approach to water quality management at:

• the national level — a vision of achieving sustainable use of water resources by

protecting and enhancing their quality while maintaining economic and social

development together with overarching national guidelines for minimum water

quality;

• state or territory level — implementation through state water quality planning

and environmental policy processes, to provide a planning and management

framework with goals and objectives consistent with the agreed national

guidelines;

• regional or catchment level — complementary planning, with local or

catchment management strategies developed and implemented by the relevant

stakeholders. Regional communities are encouraged to participate in identifying

the local environmental values and to monitor and report on progress and

performance of the plans.

To underpin water resource management at the national, state and territory levels in

Australia, a range of legislative and regulatory tools are being used. Examples

include state and territory water and land resources management Acts, environment

protection Acts, the development of water quality guidelines focused on state and

territory water resources, and the development of national environmental protection

measures. Each state or territory uses its own water quality planning and

environmental policy tools to establish a framework compatible and consistent with

the agreed national guidelines.

In New Zealand, these guidelines are designed to assist water managers with the

implementation of the Resource Management Act 1991 (RMA) which gives

regional councils primary responsibility for water management. The RMA
Chapter 2 — A framework for applying the guidelines

page 2–4 Version — October 2000

empowers councils to develop statutory plans and local laws for water

management. The RMA also enables central government to develop national policy

and standards on a statutory basis.

Overall responsibility for water resource management rests with the community.

The tools, strategies and policies developed to manage and protect environmental

values should be applied in this wider context. In effect, there must ultimately be

education and change in community behaviour toward a more environmentally

sustainable approach.

The responsibilities for monitoring water resource quality should not always rest

with government alone and ideally would be shared with the dischargers/users of the

environment in question (these shared responsibilities could extend to the waters

beyond the mixing zone of outfalls). Many community and catchment groups have

already become involved in, or taken responsibility for, water quality monitoring

programs and are developing management strategies to maintain or improve their

water resources.

2.1.2 Stakeholder involvement

Stakeholders need to be actively involved in many of steps 1–5 outlined above, to

help ensure that:

• community needs are accurately reflected;

• impacts on the community are well understood and incorporated into the

decision-making (e.g. cultural, social, economic and political);

• the costs (financial, amenity, etc.) associated with decision making will be

acceptable to the community;

• management strategies are appropriately targeted; and

• a shared ownership of catchment knowledge and commitment to action are

being developed.

Relevant stakeholders include individuals and groups that directly and/or indirectly

use, derive benefit from, and/or have an impact on the waterway being considered.

These may include indigenous groups, community groups, government agencies and

utilities, catchment and water managers, regulators, industry (urban and rural),

agricultural groups, pest control groups, environmental groups, recreational users

(e.g. fishers, swimmers) and individual residents.

The stakeholders can be involved at a number of different levels, depending on

their interest and expertise, and the mechanisms available for their involvement.

The latter in particular will vary depending on the approach taken by state, territory

and local governments.

Box 2.1 shows examples of stakeholder involvement in Australia and New

Zealand.
2.1.2 Stakeholder involvement

Version — October 2000 page 2–5

Box 2.1 Examples of stakeholder involvement in Australia and New

Zealand

• New South Wales — A six month public consultation program in 1998 identified interim

environmental values and objectives for various catchments in the State. The process

involved written submissions, and information and discussion forums located in central

and regional locations.

• Victoria — State Environment Protection Policies (SEPP) for water set out the

‘beneficial uses’ (or environmental values) to be protected in various parts of rivers,

lakes, estuaries and bays and related environmental values. The SEPP process

includes a legislative requirement for a period of at least three months for submissions

to be received.

• Queensland — The Queensland Environmental Protection Policy for Water requires

appropriate consultation with the community before environmental values and water

quality objectives for a water are decided. Community groups can set an example on

working to improve water quality, e.g. the Condamine Balonne Water Committee Inc.

took responsibility for establishing and implementing a comprehensive water quality

monitoring program in their catchment.

• Western Australia — In 1998, development of the proposed Environment Protection

(Marine Waters) Policy involved community consultation to set the environmental values

and environmental objectives of Perth’s coastal waters. The process included key

stakeholders, stakeholder reference groups and a two month consultation period.

• South Australia — Catchment goals, objectives and actions for the Torrens Catchment

were formulated using a consultation program involving the community, local

government, state and federal agencies and other stakeholders. A series of technical

papers was presented as background to the catchment plan.

• Australian Capital Territory — The ACT Environment Protection Policy on Water

Pollution requires public consultation with individuals, community groups, industry,

government agencies and other stakeholders.

• Tasmania — Local communities and other stakeholders have a key role in identifying

the water quality values for regional wetlands and waterways as part of the State Policy

on Water Quality Management 1997. Information provided on these values assists the

Board of Environmental Management and Pollution Control and local councils to finalise

Protected Environmental Values (PEVs) for surface waters. The process of setting

PEVs takes a minimum of 3 months. These values are reflected in management plans

for the regions and in local council planning schemes. The Water Management Act

1999 provides for enhanced stakeholder and community input into water allocation and

management.

• Northern Territory — Environmental value declarations, informed by extensive public

participation, have been used to establish the framework for water resource

management in the territory since 1994. For example Darwin Harbour waters were

declared under the Water Act in 1996 to have aquatic ecosystem and recreation and

aesthetics values protected. This followed an extensive public consultation phase with

public meetings, newspaper and other media promotions. Environmental objectives and

water quality targets for harbour waters will be developed over 2000/01 through a

further public consultation process steered by a committee with broad government

agency and community representation.

• New Zealand — Consultation is an integral part of natural resource management under

the Resource Management Act. All statutory plans require a period of consultation and

a submission process. Consultation with stakeholders also frequently occurs in nonstatutory management processes by government (central and local) and industry.
Chapter 2 — A framework for applying the guidelines

page 2–6 Version — October 2000

Possible forms of stakeholder involvement are listed below. Stakeholders could be

part of:

• statutory reviews of development proposals;

• community forums or discussions to identify broad community goals and

potential areas of conflict;

• specific groups that relate these broad goals to the environmental values that

need to be protected in a particular water body, decide where these may apply,

and evaluate the potential implications of different options;

• specific groups such as stakeholder advisory committees (as outlined in the

NWQMS Implementation Guidelines) which would bring together all major

interests in the one forum to discuss ideas, issues and proposals and provide a

broad-based sounding board;

• community and industry participation in processes for developing management

strategies (e.g. through catchment management planning) and assessing progress

against water quality objectives and management goals (through monitoring of

discharges and ambient conditions);

• public hearings, although this form of community forum is not commonly used

in Australia.

(The NWQMS Implementation Guidelines (ARMCANZ & ANZECC 1998)

provide more detail.)

2.1.3 Environmental values

Environmental values are particular values or uses of the environment that are

important for a healthy ecosystem or for public benefit, welfare, safety or health

and which require protection from the effects of pollution, waste discharges and

deposits. They were often called ‘beneficial uses’ in the water quality literature but

this term has lost favour because of its exploitative connotations. For this reason,

the term ‘environmental value’ has been adopted by the NWQMS.

The following environmental values are recognised in these guidelines:

• aquatic ecosystems,

• primary industries (irrigation and general water uses, stock drinking water,

aquaculture and human consumption of aquatic foods),

• recreation and aesthetics,

• drinking water,

• industrial water (no water quality guidelines are provided for this

environmental value), and

• cultural and spiritual values (no water quality guidelines are provided for this

environmental value — see box 2.2).
2.1.3 Environmental values

Version — October 2000 page 2–7

Box 2.2 Cultural importance of water

Water resources have important cultural and spiritual values, particularly for indigenous

peoples of New Zealand and Australia.

In New Zealand, water has enormous cultural importance for Maori. Water acts as a link

between the spiritual and physical worlds, and many waterbodies are associated with waahi

tapu (sacred sites). All elements of the natural environment (including people) are believed

to possess a mauri (life force) which Maori endeavour to protect. The well-being of an iwi

(tribe) is linked to the condition of the water in its rohe (territory). In addition, water provides

important mahinga kai (food collected from marine and freshwater areas). Supply and

exchange of mahinga kai forms part of the social fabric of Maori tribal life. The New Zealand

Resource Management Act (1991) recognises Maori values, such as through Sections 6(e),

the relationship of Maori and their culture with their ancestral lands, water, sites, waahi tapu,

and other taonga (treasures), and Section 7(a), Kaitiakitanga (guardianship).

Giving effect to these values may present a considerable challenge to water managers. For

example, in New Zealand, water managers require guidance on how to manage water for

values associated with (i) mahinga kai, (ii) waahi tapu, and (iii) mauri. These Guidelines do

not provide such guidance. The New Zealand Ministry for the Environment proposes:1

• preparing guidelines and case studies which develop practical methods for reflecting

the values of mahinga kai, waahi tapu and mauri in the management of water;

• incorporating mahinga kai values into the relevant ecosystem outcomes and actions.

Likewise, in Australia, indigenous cultural and spiritual values may relate to a range of uses

and issues including spiritual relationships, sacred sites, customary use, the plants and

animals associated with water, drinking water or recreational activities. Native title

legislation, and Commonwealth and state cultural heritage legislation, provide for recognition

and management of indigenous interests in water.

At this stage no water quality guidelines have been developed for the protection of cultural

and spiritual values in either New Zealand or Australia. Because of the lack of such

guidelines, in the water management framework, cultural values can be taken into account

through the process of establishing the specific water quality objectives for a particular water

resource (see figure 2.1.1).

Until further work is undertaken to better define cultural and spiritual value for users in both

Australia and New Zealand, managers in both countries, in full consultation and co-operation

with indigenous peoples, will need to decide how best to account for cultural values within

their own management frameworks. They will need to take account of existing legislation,

regulations and guidelines.

All water resources should be subject to at least one of the above environmental

values, and in most cases more than one could be expected to apply. Where two

or more agreed environmental values are defined for a water body the more

conservative of the associated guidelines should prevail and become the water

quality objectives. It is essential that the needs and wants of the community be

identified when environmental values are being defined for a particular water

resource.

 

1 NZ Ministry for the Environment 1999. Making every drop count — A draft National Agenda for

Sustainable Water Management. New Zealand Ministry for the Environment, Wellington.
Chapter 2 — A framework for applying the guidelines

page 2–8 Version — October 2000

It should be recognised that environmental values are often interdependent. For

example, all relevant environmental values need to be considered when evaluating

the quality of return water for any one user to ensure that all agreed values are

maintained; functioning ecosystems and ecosystem processes are essential for

supporting wild fish populations and can provide some protection to water quality

through chemical degradation or buffering capacity; there may also be situations

where the water quality required to support downstream environmental values (e.g.

lake, estuary or marine) will influence the establishment of water quality objectives

upstream. This will be particularly relevant where downstream ecosystems are

more sensitive to a particular contaminant (e.g. nitrogen in marine environments),

where there are cumulative effects from persistent discharges (e.g. nutrients), or

where persistent contaminants are accumulating in depositional areas downstream

(e.g. heavy metals).

Once the environmental values for a water body have been defined by the relevant

management authority, the level of environmental quality or water quality

necessary to maintain each value must be determined. It may be broadly defined

through the establishment of management goals that describe more precisely and

in greater detail what is to be protected. As with environmental values, the

management goals should be defined according to community needs and desires

and therefore will involve consultation with relevant stakeholder groups. They

should be structured so that they can become the key objectives to be achieved

through management plans and therefore should relate to particular parts of the

environment that can be measured. In particular, management goals should reflect

the specific problems and/or threats to the established values, the desired levels of

protection for aquatic ecosystems, and the key attributes of the resource that must

be protected (e.g. endemic or key species, key agricultural or aquacultural species,

primary or secondary recreation). From the management goals it should be obvious

which the key water quality indicators are, and therefore which guidelines should

be selected for establishing water quality objectives. The specific water quality

objectives more tightly define the desired level of water quality, and are compared

with the existing water quality to assess performance.

In some cases, the water quality needed to support the desired environmental value

may not be attainable immediately. Where restoration is possible, there may be

costs associated with restoring the level of quality that the community desires.

Once full costs of restoration are known, the community may choose to accept a

lower quality based on a full cost–benefit analysis. The environmental values and

management goals for a particular area need to be well thought out, with full

knowledge of the implications to the broader community. This is a process

involving broad consultation with representatives of the whole community, with

the aim of reaching a desirable, practical and agreed set of management goals, and

hence water quality objectives.

Guidance on how to undertake community consultation processes is provided in

the NWQMS Implementation Guidelines (ARMCANZ & ANZECC 1998).

In the absence of a clear and agreed set of environmental values for a particular

water resource, managers should take a conservative approach and assume that all

appropriate environmental values apply to the resource, by default. For example in

the case of a coastal marine embayment, ‘drinking water’ would not apply by
2.1.4 Water quality guidelines

Version — October 2000 page 2–9

default, but ‘ecosystem protection’, ‘recreation and aesthetics’, and ‘primary

industries — aquaculture’ would apply.

2.1.4 Water quality guidelines

A water quality guideline is a numerical concentration limit or narrative statement

recommended to support and maintain a designated water use. This document

includes guidelines for chemical and physical parameters in water and sediment, as

well as biological indicators. The guidelines are used as a general tool for assessing

water quality and are the key to determining water quality objectives that protect

and support the designated environmental values of our water resources, and

against which performance can be measured.

Water quality parameters can be divided into those that have direct toxic effects on

organisms and animals (e.g. insecticides, herbicides, heavy metals and temperature)

and those that indirectly affect ecosystems causing a problem for a specified

environmental value (e.g. nutrients, turbidity and enrichment with organic matter).

Whether the effects are direct or indirect has important implications for

management, and perhaps for how a guideline might be derived. Some physical

and chemical stressors can also indirectly modify the toxicity of other

contaminants. While specific guidelines are not provided for this mode of action,

guidance is provided in each relevant section on how it can be taken into account.

The guidelines have been derived with the intention of providing some confidence

that there will be no significant impact on the environmental values if they are

achieved. Exceedance of the guidelines indicates that there is potential for an

impact to occur (or to have occurred), but does not provide any certainty that an

impact will occur (or has occurred). In areas where protection of aquatic

ecosystems is a designated environmental value, the Guidelines recommend direct

assessment of the biological community to assess whether ecosystem integrity is

being maintained, threatened or compromised to a level that causes pollution.

Biological indicators should therefore be used to complement the use of physical

and chemical indicators for this value. These Guidelines describe indicators for

biological assessment and give guidance for determining an acceptable level of

change so that the relative condition of the ecosystem can be estimated.

For some environmental values it may not be feasible to protect all water resources

to the same level, and the community may wish to aim for different levels of

protection for different resources. Whatever the level of protection, it should be

reflected in the management goals and the water quality objectives determined for

a particular resource. In this document three levels of protection, based on

ecosystem condition, are recognised for aquatic ecosystems.a For aquatic

ecosystems the guidelines in this document have mainly been developed for use at

the second and third levels of protection: slightly to moderately disturbed

ecosystems and highly disturbed ecosystems. The highest level of protection is for

high conservation/ecological value systems where management would be expected

to ensure there is no change2 in biological diversity, relative to a suitable reference

 

2 ‘No change’: In practice and in the absence of information that would define the thresholds of

ecological change, refers to statistically conservative changes from a baseline mean or median

value, e.g. change of 10% or one standard deviation from a baseline mean — see sections 3.2.4.2

and 7.2.3.3 (Stage 1).

a See Section

3.1.3
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page 2–10 Version — October 2000

condition.a For highly disturbed ecosystems that cannot feasibly be returned to a

slightly to moderately disturbed condition, these Guidelines provide advice to

assist managers to derive alternative guidelines that give lower levels of protection.

The earlier guidelines (ANZECC 1992) acknowledge there is such inherent

variability within the environment that ‘site-specific’ environmental information

needs to be used to develop appropriate guidelines and indices of environmental

quality. For example, light availability is a key factor controlling the growth and

survival of benthic plants. In naturally turbid waters the biomass of a particular

species may decrease with depth to a limit beyond which there is insufficient light.

This limit would be deeper in less turbid waters. Thus the selection of a water clarity

guideline value (e.g. light attenuation coefficient) would need to take into account

these site-specific considerations.

Guideline numbers and decision frameworks

These Guidelines have adopted an innovative risk-based approach that is intended

to improve the application of guidelines to all Australian and New Zealand aquatic

environments. It uses decision frameworks (particularly for the protection of aquatic

ecosystems) that help users to tailor water quality guidelines to local environmental

conditions.b In this approach the old ‘single number’ guidelines (see ANZECC

1992) are regarded as guideline trigger values that can be modified into regional,

local or site-specific guidelines by taking into account factors such as the variability

of the particular ecosystem or environment, soil type, rainfall and level of exposure

to contaminants. Trigger values are concentrations that, if exceeded, would indicate

a potential environmental problem, and so ‘trigger’ a management response, e.g.

further investigation and subsequent refinement of the guidelines according to local

conditions. Thus these Guidelines have moved away from promoting single-number

guidelines that are applied universally, towards guidelines that can be determined

individually according to local environmental conditions.

It is not mandatory to use decision frameworks, but they can reduce the amount of

conservatism necessarily incorporated in the guideline trigger values, and so produce

values more appropriate to a particular water resource. Decision frameworks or tools

also allow more flexibility and scope for water managers. Hence guidelines that are

more relevant to a specific water resource and environmental value can be developed

where considered appropriate. However, it may take more time, expertise or

resources to implement the risk-based decision frameworks, particularly where

additional data collection is required to augment the data already collated.

Which stakeholder(s) are responsible for data collection and implementation of the

decision frameworks will depend on the issue (e.g. environmental impact assessment

process or management strategy development) and the jurisdictions’ legislative and

regulatory tools, and should therefore be decided on a case-by-case basis.

Management agencies with responsibility for a number of water resources may need

to prioritise their water resources based on factors such as condition of the system,

increasing land use pressures, data availability, public concern, conservation issues

and the outcomes of risk and cost-benefit analyses, so that limited resources can be

appropriately allocated.

Alternatively, where resources, data and/or time are significant constraints, users

can take a more conservative approach and initiate an appropriate management

response when either the initial trigger value or a partly modified trigger value

a See Section

3.1.4

b Sections

2.2.1.4 & 3.1.5
2.1.5 Water quality objectives

Version — October 2000 page 2–11

(only part of the decision framework applied) is exceeded. The availability of data,

expertise, resources and time will determine which steps in the frameworks are used.

Note: it is emphasised here, and elsewhere throughout the document, that the use of

the term ‘risk-based’ does not imply the need for a full (quantitative) risk

assessment. For example, the aquatic ecosystem guideline trigger values for

toxicants are risk-based in the sense that they are calculated to protect a predetermined percentage of species with a specified level of confidence,a while the

decision frameworks simply provide a site-specific estimate of whether low,

possible or high risk exists.b

2.1.5 Water quality objectives

A water quality guideline was defined above as a numerical concentration limit or

descriptive statement recommended for the support and maintenance of a

designated water use. Water quality objectives take this a step further. They are the

specific water quality targets agreed between stakeholders, or set by local

jurisdictions, that become the indicators of management performance. Normally,

only those indicators considered relevant to the environmental issues or problems

facing the resource are selected for deriving water quality objectives. They serve to

protect the designated environmental values of a resource and would normally be

based on the information from these Guidelines.

A water quality objective is a numerical concentration limit or descriptive

statement to be measured and reported back on. It is based on scientific water

quality criteria or water quality guidelines but may be modified by other inputs

such as social, cultural, economic or political constraints. Some of these inputs may

be intangible and therefore hard to quantify, but nevertheless they are valid inputs

to the management process. The relative weighting or importance placed on the

water quality guidelines and these other, potentially very important but less

tangible, considerations would be area specific, and therefore would be determined

on a case by case basis. The process of modifying guidelines to establish water

quality objectives would normally be carried out through cost–benefit analysis

programs involving input from stakeholders or local jurisdictions.

An additional consideration when setting water quality objectives in rivers and

streams is the water quality required to meet management goals and hence protect

the environmental values established further downstream, including estuaries and

coastal marine environments. The water quality required to support local

environmental values may not be sufficient to support downstream environmental

values, particularly for chemicals that persist in the environment or where

downstream ecosystems are more sensitive to the contaminant (e.g. heavy metals

or nutrients).

2.2 Application of the guidelines for water quality management

A primary aim of this document is to help users to develop management

frameworks for protecting the environmental values of Australian and New

Zealand natural and semi-natural water resources, and to derive appropriate water

and sediment quality guidelines for the ambient waters that will protect their

designated values.

a See Section

3.4.2

b Section 3.1.5
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page 2–12 Version — October 2000

The guidelines can:

• provide water resource managers with information that helps them identify and

prioritise key environmental issues (such as the loss of seagrass beds if light

intensity decreases below a critical level) and hence determine the management

goals;

• assist managers to establish management goals and water quality objectives

(preferably with appropriate baseline data);

• provide information that helps resource managers decide on the types of

management actions they need for achieving the desired goals and targets;

• provide a basis upon which to assess whether the management actions are

achieving the targets set for the management unit.

The preferred approach is to use the guidelines in a proactive way (where

management focuses on preventing change beyond some pre-determined level),

although in already degraded systems this may not be an option.

The purpose of a water quality management program should be to ensure that

environmental values will be supported through the management goals and by

meeting the agreed water quality objectives. It is recommended that this should be

done through a process of cooperative best management (involving all

stakeholders), and based on sound environmental arguments. Where the

environmental values are not being supported because the associated management

goals are not being met, remedial management programs, with appropriate

performance indicators and associated time frames, should be developed and

implemented to ensure the management goals will be met.

2.2.1 Philosophical approach to applying the guidelines

New ways of managing water quality have developed to match growing scientific

understanding of ecosystem complexity. Traditional scientific and management

approaches are now often inappropriate; instead there must be increasing reliance

on holistic best-practice approaches to ensuring sustainable use of water resources.

Key issues underpinning the new philosophy espoused in these Guidelines are

outlined below. Some of them were also fundamental to the previous (ANZECC

1992) Guidelines.

2.2.1.1 Sustainable use

The fundamental aims of the NWQMS in Australia and the Resource Management

Act in New Zealand are the sustainable use and management of each nation’s water

resources in environmental, economic and social contexts. To achieve these aims,

the concept of integrated catchment management3 (ICM) is promoted today. The

concept is consistent with the management framework outlined in these Guidelines,

and encompasses all aspects of environmental management within a catchment,

including water quality. Within the ICM framework, environmental values are

identified by all stakeholders of individual resources, namely landowners and the

community, in partnership with relevant government agencies.

 

3 Under section 30/1/a of the New Zealand Resource Management Act (1991) all regional councils are

required ‘to achieve integrated management of the natural and physical resources of the region’.
2.2.1 Philosophical approach to applying the guidelines

Version — October 2000 page 2–13

2.2.1.2 Cooperative best management

Formerly, deterioration in water quality was largely controlled by regulation and

management. While these command and control approaches successfully deal with

the obvious point source problems, they have produced an end-of-pipe and

minimum compliance culture. It is now also clear that a regulatory approach is

generally not an appropriate tool for resolving the problems of diffuse sources of

contamination which have just as much or more of an impact on water quality than

point sources.

Environmental regulation and management in Australia and New Zealand are

currently undergoing major change, adopting a more holistic and integrated

pollution-prevention approach to environment protection. This involves a shift

from control to prevention, from end-of-pipe regulation to cleaner production, from

a focus on prescriptive regulation to a focus on outcomes and on cooperation rather

than direction. This new approach is being increasingly adopted in formulating

water resource management policies and strategies. It requires the commitment of

industry and government and the involvement of the community to establish

cooperative best management and overall responsibilities for maintaining and

improving water resources. The NWQMS Implementation Guidelines (ARMCANZ

& ANZECC 1998) outlines a framework for involving all stakeholders in the

management of water resources.

The success of cooperative best management relies on negotiated agreements

developed through processes involving the entire community; they set boundaries

within which business can maintain the defined environmental values. Cooperative

best management provides a framework through which many sources of pollution

may be addressed in an equitable and effective manner. Sharing of responsibility,

cooperative action, and effective monitoring and reporting arrangements are key

aspects. Also important is the emphasis on flexibility and integrated management

to achieve the best feasible environmental outcomes. In practice there may be

issues over which stakeholders are unable to reach agreement. Local jurisdictions

may therefore need to consider establishing conflict resolution mechanisms to

facilitate the decision making process.

It is also important that communication networks be developed across whole

catchments to address broad-scale issues cooperatively. For example, when setting

the water quality objectives for upstream riverine ecosystems, effects on

downstream environmental values, including cumulative effects, must also be

considered.

Cooperative best management focuses on attaining goals of environmental quality

rather than on compliance per se. For example, licence conditions or agreed levels of

unacceptable environmental change in monitoring programs would be negotiated

between all the stakeholders, with the overriding objective of attaining the

established management goals for a water resource (and hence protecting its

environmental values), rather than simply regulating to meet individual water quality

parameters. The process would consider best management practices and the ability of

the industry to achieve adequate effluent quality within a reasonable time frame.

Where the agreed licence conditions were not met, or a trend toward a significant

change in ambient water quality was detected, there would be an attempt to resolve

the problem cooperatively before using a regulatory approach as a last resort. To

complement their cooperative approach, jurisdictions might need to introduce a
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number of tools for controlling diffuse and point source pollution (e.g. economic

incentives, emissions trading).

Cooperative best management involves monitoring and impact assessment.

Although risk assessment concepts are familiar to many water resource managers,

analogous concepts of the potential for errors in statistical inferences based on

monitoring data are often poorly understood, or neglected. An alternative approach

to statistical decision making (Mapstone 1995, 1996) is suggested (see box 2.3)

that should jointly involve all stakeholders. As mentioned above, even where water

quality meets the agreed water quality objectives and is ‘acceptable’ in a statistical

sense, stakeholders should work cooperatively to develop a clear understanding of

the issues associated with, and consequences of, water quality that is trending

towards the established objectives. In this way, intervention, including changes to

industry practices, can be set in place at an early stage if deemed necessary.

Box 2.3 An alternative approach to statistical decision making

(Mapstone 1995, 1996)

Traditionally, statistical analysis of monitoring data has only considered minimising the

probability of concluding that an environmental impact has occurred when, in fact, no impact

has occurred — a Type I error. However, to maximise protection of the environment it is

perhaps more important to consider Type II errors — the probability of concluding that an

impact has not occurred when, in fact, it has. The first step in the suggested alternative

approach is to decide on the size of effect that would cause concern (or constitute an early

warning). Then, with this critical effect size in mind, the stakeholders consider the possibility

that such an effect might either be missed (a Type II error) or inferred incorrectly (a Type I

error). Monitoring and data collection are then designed to keep the risks of both Type I and

Type II errors to the values agreed up-front by the stakeholders, given the stipulated critical

effect size. The significance criterion used in statistical tests should be that which ensures

that the agreed ratio of Type I and Type II errors is maintained.a

Consistent with the principle of cooperative best management, this approach should result in

benefits to all stakeholders. All parties would be aware of the targets for monitoring, the

statistical criteria by which statistical decisions will be made, what will trigger management

action, the level of safeguard built into the decision making process, and the risks of

expense or environmental impact arising from errors in the assessment and monitoring

process. Clear knowledge of these factors should reduce the risk of wrangling or litigation

when impacts have been inferred and should allow explicit planning for mitigation or

restoration actions that might arise in the future, but with some known minimum probability

of a wrong conclusion.

Other tools that might be considered in cooperative best management are

memoranda of understanding and catchment management plans. Non-point source

pollution problems, in particular, could be addressed through the development and

implementation of catchment management plans by landowners and the

community, in partnership with relevant government agencies. One of the main

objectives of these catchment management plans would be to achieve the

management goals set for the aquatic environment. Well designed and

appropriately focused monitoring programs could assess the effectiveness of

catchment management plans in meeting specified water quality targets.

a Type I and II

errors are

explained more

fully in Section

3.1.7
2.2.1 Philosophical approach to applying the guidelines

Version — October 2000 page 2–15

2.2.1.3 Management focus on issues not guidelines

The philosophical approach for using these revised Guidelines is this: protect

environmental values by meeting management goals that focus on concerns or

potential problems, e.g. toxicity. This is in contrast to previous approaches which

more often focused on simple management of individual water quality parameters,

e.g. toxicant concentration, to meet respective water quality guidelines or

objectives. First, identify the water quality concern (e.g. toxicity, algal blooms, soil

structure degradation, loss of animal vigour, deoxygenation, loss of biodiversity),

and establish and understand the environmental processes that most influence or

affect the particular concern. Then select the most appropriate water quality

indicators to be measured, and identify the relevant guidelines.

Usually a range of environmental problems is responsible for degradation of water

resources in Australia and New Zealand and so issues typically involve a range of

water quality parameters. An issue-based approach to management would focus on

the overall problem, and ensure an integrated approach to addressing relevant

biological, chemical and physical aspects of water quality. For example, in situations

where sediment contamination is likely, water managers should not focus solely on

whether the measured sediment concentrations are above or below a guideline. They

should also consider the bioavailability of the contaminant, and analyse trends and

consider risk factors to determine whether, under current or proposed management

regimes, guideline values are likely to be exceeded in the future.

2.2.1.4 Tailoring guidelines for local conditions

Optimum water quality characteristics differ between regions. There is a wide

range of ecosystem types and environments in Australia and New Zealand, and it is

not possible to develop a universal set of specific guidelines that apply equally to

all. (Some of the default guidelines, however, do now distinguish amongst several

different ecosystem types and regions making these values much more focused

than they were in the previous Guidelines.) Further, environmental factors can

significantly alter the toxicity of physical and chemical stressors at a site and these

factors can vary considerably among sites. The present Guidelines move away from

single number values that are mostly conservative, and emphasise guidelines that can

be determined individually, according to local environmental conditions. This is

done through the use of local reference data and ‘risk-based decision frameworks’.

Decision frameworks provide guideline trigger values (equivalent to the old

guideline default values) that refer to the concentration of the chemical available

for uptake by organisms. Guideline trigger values are concentrations that, if

exceeded, will indicate a potential environmental problem, and so ‘trigger’ further

investigation. The investigation aims to both assess whether exceedance of a

trigger value will result in environmental harm and refine a guideline value, by

accounting for environmental factors that can modify the effect of the chemical.a

Although in some cases this will require more work, it will result in much more

realistic goals for management and therefore has the potential to reduce both costs

for industry and confrontation.

2.2.1.5 Water or environmental quality

Water (and sediment) quality is only one aspect of maintaining some

environmental values. In many cases (e.g. for primary industries and aquatic

ecosystems) other factors are also important, e.g. flow, habitat, soil type, animal

a See also

Section 3.1.5
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diet, groundwater hydrology and barriers to recruitment. In many parts of Australia

and New Zealand, water quality is reasonably good but management goals for

maintaining aquatic ecosystems are not being met because of loss or degradation of

habitat, particularly riparian vegetation. In these situations, enhancement of water

quality is unlikely to result in any significant environmental benefit because

improvement in habitat is needed to achieve management goals and protect the

environmental value.

Before investing in water quality management strategies, managers need to be sure

that water quality is the key issue to be addressed in the water body under

consideration, and that resources would not be better spent on other aspects of the

water resource, such as riparian vegetation, habitat or hydrological regime.

2.2.1.6 Integrated water quality assessment

Water quality, environmental values and the surrounding environment are all

intimately connected and need integrated assessment. This should also

acknowledge that ecosystems and environmental values upstream and downstream

are linked and can affect each other.

These Guidelines include a substantial section on assessment of biological aspects

of aquatic ecosystems,a to accompany physical and chemical indicators in

assessing impacts on ecosystem integrity. Sediment quality guidelines are also

given.b This is important because pollutants become partitioned between water,

sediment and biota and move between them depending on prevailing

environmental conditions. These Guidelines also advise on suitable environmental

flows in rivers and streams.

Similarly, in assessing water quality for irrigation, the Guidelines include

consideration of soil and plant aspects of the production system, as well as the offfarm implications of water use.

2.2.1.7 Continual improvement

An overriding principle that should guide management should be continual

improvement. This is more obvious where water or sediment quality does not

match the water quality objectives. In badly polluted waters it might even be

necessary to set intermediate levels of water quality to be achieved in well defined

stages, each subsequent target closer to the required water quality objective, until it

is finally met. However, in waters that are of better quality than that set by the

water quality objectives, some emphasis could still be given to reducing the level

of contamination from all sources, particularly for highly modified water resources.

Wherever possible, ambient water quality should not be allowed to degrade to the

levels prescribed by the water quality objectives.

2.2.1.8 Guidelines not standards

The Guidelines recommend numerical and descriptive water quality guidelines to

help managers establish water quality objectives that will maintain the

environmental values of water resources. They are not standards, and should not be

regarded as such. The vast range of environments, ecosystem types and food

production systems in Australia and New Zealand require a critically discerning

approach to setting water quality objectives.

a Section 3.2

b Section 3.5
2.2.2 Mixing zones

Version — October 2000 page 2–17

State, territory or local jurisdictions will need to determine whether water quality

objectives should be enshrined in legislation based on the particular local

circumstances.

2.2.1.9 Ambient waters

The Guidelines have not been designed for direct application in activities such as

discharge consents, recycled water quality or stormwater quality, nor should they

be used in this way. (The exception to this may be water quality in stormwater

systems that are regarded as having some conservation value.) They have been

derived to apply to the ambient waters that receive effluent or stormwater

discharges, and protect the environmental values they support. In this respect, the

Guidelines have not been designed to deal with mixing zones, explicitly defined

areas around an effluent discharge where the water quality may still be below that

required to protect the designated environmental values. As such, the application

and management of mixing zones are independent but very important processes.

2.2.2 Mixing zones

Even when stringent effluent limits are set and strict waste minimisation is

practised, effluents may be of poorer quality than the receiving water. It has been

accepted practice to apply the concept of the mixing zone, an explicitly defined area

around an effluent discharge where certain environmental values are not protected

(see description in box 2.4).

Box 2.4 Mixing zones adjacent to effluent outfalls

Mixing zones are often defined as explicit areas around effluent discharges where the

management goals of the ambient waters do not need to be achieved and hence the

designated environmental values may not be protected. In this context mixing zones are

sometimes termed exclusion zones. Appendix 1 of Volume 2 provides some key references

and further information and advice on mixing zones. The following issues are covered there:

• the nature of mixing zones;

• difficulties with mixing zones;

• the management of mixing zones;

• best-practice effluent release and mixing zone management, as a case study; and

• mixing zone models.

Effective discharge controls that consider both the concentration and the total mass

of contaminants, combined with in situ dilution and waste treatment, should ensure

that the area of a mixing zone is limited and the values of the waterbody as a whole

are not jeopardised. The environmental conditions within a mixing zone, and its size,

are important concerns, particularly because degraded areas around effluent

discharges reduce environmental benefits. If mixing zones are to be applied, then

management should ensure that impacts are effectively contained within the mixing

zone, that the combined size of these zones is small and, most importantly, that the

agreed and designated values and uses of the broader ecosystem are not

compromised.
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2.2.3 Application of water quality prediction models

Development of a water quality management strategy depends on the quality of

available information and a capacity to predict the effects of various actions on

water quality. This can be done via conceptual models, which are often used to

predict effects of discharges on the environment. Conceptual models can be very

simple flow diagrams that illustrate the linkages between the components of the

system or they can be more complex models built up from information arising from

previous studies. They should indicate which key processes are influencing the

system and highlight those processes likely to be affecting the water quality

indicators that are of concern. These models may be conservative and rely on

worst-case conditions of dilution and degradation of effluents, or they may attempt

more complex analysis of cause and effect. Because of the unique features of each

system, it has generally been found that models developed for a particular

waterbody cannot be used for other waterbodies without significant modification.

A more detailed discussion of conceptual models is provided in the Australian

Guidelines for Water Quality Monitoring and Reporting, the Monitoring

Guidelines (ANZECC & ARMCANZ 2000).

The preferred approach for managing persistent contaminants that have

concentration-related toxicity potential is appropriately based on controlling

ambient concentration in the environment. Generic guidelines for the protection of

environmental values can be established for these types of contaminants, based on

modelled relationships between concentration/exposure of the contaminant and the

toxicity to test organisms, and applying safety margins designed to take account of

the uncertainty associated with transferring laboratory-derived data to the open

environment and the likelihood and pathways of bioaccumulation/

persistence/degradation. This model is suitable for managing toxicants in general,

but alternative approaches are needed for managing substances such as nutrients

that may stimulate rather than retard growth of particular species.

An example of a model that is raised frequently in the context of water pollution

control is that of assimilative capacity. The underlying philosophy of this concept

is that a natural system has the capacity to receive some level of human-induced

nutrient input without unacceptable changes occurring. This concept has been

defined using a variety of terms including: assimilative capacity or environmental

capacity (GESAMP 1986, WAEPA 1990, Masini et al. 1992); receiving capacity

or absorptive capacity (UNESCO 1988, WAWA 1994) and carrying capacity

(French 1991, Jenkins 1991). Regardless of what name is used this ecosystembased approach is now recognised as central to the principle of ecological

sustainability (IUCN, UNEP & WWF 1991, Jenkins 1991, Folke et al. 1993).

This ecosystem-based approach is based on establishing linkages between total

nutrient loadings to an ecosystem and the response of the most sensitive or

important component of that ecosystem. Once these relationships have been

quantified, and the desired management goals defined, regulatory agencies can set

ecologically-based maximum nutrient loadings consistent with maintaining the

desired environmental quality. An ecosystem-based approach linking nutrient

loadings to environmental response has been successfully applied to Perth’s coastal

waters (WADEP 1996).
2.2.4 Deriving guidelines for compounds where no guidelines currently exist

Version — October 2000 page 2–19

In many freshwater and estuarine systemsa the biological response to nutrient

additions can become confounded and less predictable because plant growth and

biomass may be significantly limited by factors other than nutrient availability.

Light availability is the dominant limiting factor in many naturally–turbid or tannin

stained waters for at least part of the year. Under these conditions nutrients can

behave more conservatively and concentrations can differ between seasons. For

effective management in this case, key pathways of nutrient transformation need to

be known, and seasonal and interannual variation in flow regimes need to be

understood. Under these circumstances, models of system behaviour employing

flow-weighted concentration-based approaches can be very useful and are

sometimes essential.

No matter which ‘predictive’ model is used, it is essential that environmental quality

and the attainment of management goals are regularly assessed through monitoring to

determine whether regulation or other management is necessary. Undesirable trends

and the necessity for proactive management can be identified if data are collected at

appropriate temporal scales. This relies on regular monitoring of indicators within

some or all of the environmental media (water, sediment and biota) and assessment

against appropriate guidelines or water quality objectives and reference sites. This

information improves our conceptual understanding of the ecosystem being managed

and in particular the pathways that underpin predictive models.

2.2.4 Deriving guidelines for compounds where no guidelines currently exist

The Guidelines focus on water quality management in Australia and New Zealand,

but situations will arise where there is not enough information to address an issue.

There are more than 70 000 chemicals in use around the world. It is not feasible to

develop guidelines for all of them, either because there are insufficient

toxicological studies available, or because the chemical is currently not available in

Australia or New Zealand or not considered a risk there. There could also be

situations where effluent contained a range of chemicals and complexes, and the

chemical make-up might not be well understood. In this instance the complex

chemistry might increase or reduce the toxicity of the overall mixture to an

unknown degree and so the guidelines would be irrelevant. A third possible

situation relating to the protection of aquatic ecosystems is where there is a well

founded suspicion that a particular natural community may have atypical

sensitivity to one or more contaminants.

Direct toxicity assessment is a useful tool that can be used in these circumstances,

although it is mainly used to assess the toxicity of complex effluents and to derive

guidelines for the amount of dilution required to safely discharge an effluent to

aquatic environments. It can also be used as a monitoring tool, testing the ambient

waters after they have received effluent discharges. The main advantage with using

direct toxicity assessment is that it is not necessary to know the exact chemical

make-up of the test effluent, and the interactions between the components, to

determine potential impacts.

a See Section

3.1.2
Chapter 2 — A framework for applying the guidelines

page 2–20 Version — October 2000

If water quality guidelines do not exist for a specific chemical, or if effluents

contain a complex range of chemicals, expert advice should be sought from the

relevant authorities on whether a current guideline exists or how a guideline might

be derived. These sorts of situations are most likely to arise for the protection of

aquatic ecosystems, and later chapters of the Guidelines give extensive guidance

for addressing these problems.
Version — October 2000 page 3.1–1

3 Aquatic ecosystems

3.1 Issues for all indicator types

This chapter specifies biological, water and sediment quality guidelines for

protecting the range of aquatic ecosystems, from freshwater to marine. As already

noted the guidelines are not sufficient in themselves to protect ecosystem integrity;

they must be used in the context of local environmental conditions and other

important environmental factors, for example, habitat, flow and recruitment. For

the protection of rare aquatic communities and/or species, guidelines for the

highest level of protection should be applied.a

The chapter is divided into five sections: Section 3.1 is introductory and covers

information common to all indicator types; Section 3.2 contains guidelines for the

biological assessment of ecosystem condition; Section 3.3, guidelines for physicochemical stressors; Section 3.4, guidelines for toxicants in water; and Section 3.5,

guidelines for toxicants in sediments.

The scientific rationale behind the guidelines, and other useful background

information for applying the guidelines, are provided in Volume 2 of the

Guidelines. Guidelines for the design and implementation of monitoring and

assessment programs involving the types of water quality indicators discussed in

this chapter, are contained in Chapter 7.

3.1.1 Philosophy and steps to applying the guidelines

Many benefits of aquatic ecosystems can only be maintained if the ecosystems are

protected from degradation. Aquatic ecosystems comprise the animals, plants and

micro-organisms that live in water, and the physical and chemical environment and

climatic regime with which they interact. It is predominantly the physical

components (e.g. light, temperature, mixing, flow, habitat) and chemical

components (e.g. organic and inorganic carbon, oxygen, nutrients) of an ecosystem

that determine what lives and breeds in it, and therefore the structure of the food

web. Biological interactions (e.g. grazing and predation) can also play a part in

structuring many aquatic ecosystems.

Humans have caused profound changes in Australian and New Zealand aquatic

ecosystems, particularly in the 200 years since European settlement of these

countries (ANZECC 1992) and the need to protect and even reverse degradation of

important aquatic ecosystems is now recognised. Commercial and recreational

harvests of fish and shellfish can only be obtained from waters where ecosystems

provide the food and habitat to support the growth and reproduction of the

harvestable species. Aquatic ecosystems are worthy of protection for their intrinsic

value. Effective conservation of endangered species can only be achieved by

conserving the ecosystems that support them (ANZECC 1992).

a See Section

3.1.3
Chapter 3 — Aquatic ecosystems

page 3.1–2 Version — October 2000

Box 3.1.1 Human activities affecting aquatic ecosystems

A wide range of human activities can cause variations in abiotic factors, which can lead to

biological changes more dramatic than those which occur naturally. The effects of human

activities include pollution from industrial, urban, agricultural and mining sources; regulation

of rivers through the construction of dams and weirs; salinisation; siltation and sedimentation

from land clearance, forestry and road building; clearance of stream bank vegetation; overexploitation of fisheries resources; introduction of alien plant and animal species; removal

and destruction of habitat; polluted discharges from industrial, urban, agricultural and mining

activities; over-exploitation of the biological resources of freshwater and marine systems;

recreation (e.g. lead shot in wetlands, hydrocarbons from boats and jet skis); cold water

from reservoirs and hot water from power plants; ship ballast water containing exotic

species; intentional introduction of non-native species for recreation or commercial

production; and eutrophication (nutrient enrichment that may stimulate the growth and

dominance of toxic cyanobacteria in freshwaters and estuaries, and toxic dinoflagellates in

marine waters).

The greatest threat to the maintenance of ecological integrity is habitat destruction

(Biodiversity Working Party 1991). The previous ANZECC (1992) guidelines

foreshadowed the need for a broader, more holistic approach to aquatic ecosystem

management, to consider all changes, not just those affecting water quality. Such

changes could include serious pollution of sediments, reduction in stream flow by

river regulation, removal of habitat (de-snagging, draining wetlands) or significant

changes in catchment land use, any of which could cause significant ecosystem

deterioration (ANZECC 1992). The guidelines for water quality management

documented here are therefore a necessary but only partially sufficient tool for

aquatic ecosystem management or rehabilitation.

The objective adopted in this document for the protection of aquatic ecosystems is:

to maintain and enhance the ‘ecological integrity’ of freshwater and marine

ecosystems, including biological diversity, relative abundance and ecological

processes.

Ecological integrity, as a measure of the ‘health’ or ‘condition’ of an ecosystem,

has been defined by Schofield and Davies (1996) as:

the ability of the aquatic ecosystem to support and maintain key ecological processes

and a community of organisms with a species composition, diversity and functional

organisation as comparable as possible to that of natural habitats within a region.

Depending on whether the ecosystem is non-degraded or has a history of

degradation the management focus can vary from simple maintenance of present

water quality to improvement in water quality so that the condition of the

ecosystem is more natural and ecological integrity is enhanced.

For the assessment of ecosystem integrity, these Guidelines focus on the structural

components of aquatic communities (biodiversity) and key ecological processes

(e.g. community metabolism) as defined in Section 3.2.1.1.

With or without biological assessment,a chemical and physical water quality

indicators continue to be important surrogates for assessing and/or protecting

ecosystem integrity. This document therefore provides guidelines for chemical and

physical water quality indicators as well as biological indicators.

a See Section

3.1.6
3.1.1 Philosophy and steps to applying the guidelines

Version — October 2000 page 3.1–3

Box 3.1.2 Protecting biodiversity

Biological diversity is defined as the variety of life forms, including the various plants,

animals and micro-organisms, the genes they contain and the ecosystems of which they are

a part (Biodiversity Unit 1994, DEST State of the Environment Advisory Council 1996).

Broadly, biodiversity is considered at three levels: genetic diversity, species diversity and

ecosystem diversity.

Great difficulty arises in establishing a level of protection for biodiversity so that its

maintenance is guaranteed. The Biodiversity Working Party (1991) suggested:

Ideally, it should be that level that guarantees the future evolutionary potential of

species and ecosystems. All development is likely to cause some loss of the genetic

component of biodiversity, to reduce overall populations of some species, and to

interfere to a greater or lesser extent with the ecosystem processes. Protecting

biodiversity means ensuring that these factors do not threaten the integrity of

ecosystems or the conservation of species.

Figure 3.1.1 shows a framework for applying the guidelines to the protection of

aquatic ecosystems.a The three parts are described below. Each of the first two

steps is common to the application of all the indicator types (biological, physicochemical, chemical and sediment).

Box 3.1.3 How to apply the guidelines

The following steps should be followed when applying the guidelines for the protection of

aquatic ecosystems; steps 1–3 are the first parts of the broad framework presented in

figure 3.1.1.

1. Define the primary management aims (Section 3.1.1.1)

2. Determine appropriate guideline trigger values for selected indicators (Section 3.1.1.2).

After determining a balance of indicator types, each of the remaining steps is common

to the application of physical and chemical stressors and toxicants in water and

sediment. For the biological indicators, the principles of the steps ‘Select relevant

indicators’ and ‘Select specific indicators ...’ should be applied to the general framework

for biological indicators (figure 3.2.1). At this stage, initial sampling can commence,

ideally in support of a pilot program.

3. Assess test site data and, where possible, refine trigger values to guidelines using

(i) the general framework for biological indicators (figure 3.2.1), and (ii) the decision

frameworks for other indicators. Frameworks for (ii) are described in Section 3.1.1.3

(‘Risk-based application of the guidelines’). Decision frameworks to apply to specific

indicators, and detailed guidance on applying these, may be found in the Guidelines

figures and sections as follows:

(a) physical and chemical stressors — figure 3.3.1, Section 3.3

(b) toxicants — figure 3.4.2, Section 3.4

(c) sediments — figure 3.5.1, Section 3.5.

4. Define water quality objectives (figure 2.1.1, Section 2.1.5)

5. Establish a monitoring and assessment program (figures 2.1.1 & 7.1, Chapter 7).

a See also box

3.1.3
Chapter 3 — Aquatic ecosystems

page 3.1–4 Version — October 2000

Define Primary Management Aims

• Define the water body (scientific information, monitoring

data, classify ecosystem type (section 3.1.2))

• Determine environmental values to be protected

• Determine level of protection (section 3.1.3)

• Identify environmental concerns

 e.g. — toxic effects

 — nuisance aquatic plant growth

 — maintenance of dissolved oxygen

 — effects due to changes in salinity

• Determine major natural and anthropogenic factors affecting

the ecosystem

• Determine ‘management goals’

 — often defined in biological terms (section 2.1.3)

Determine appropriate Guideline Trigger Values for

selected indicators

• Determine a balance of indicator types (based upon level of

protection and local constraints, section 7.2.1)

• Select indicators relevant to concerns and goals

• Determine appropriate guideline trigger values (low risk

concentrations of contaminants/stressors; may depend on level of

protection)

• Determine specific indicators to be applied

Determining appropriate guideline trigger values

Apply the Trigger Values using (risk-based) Decision Trees 

or Guideline ‘packages’

• Water quality monitoring data

• Site specific environmental information

• Effects of ecosystem-specific modifying factors.

(see fig 3.2.1 — biological assessment

 fig 3.3.1 — physical and chemical stressors

 fig 3.4.2 — toxicants

 fig 3.5.1 — sediments)

Figure 3.1.1 Flow chart of the steps involved in applying the guidelines

for protection of aquatic ecosystems

3.1.1.1 Primary management aims

Define the water body, from scientific information and monitoring data. Good

management can only be based on detailed information about the ecosystem being

protected. Information can be collected by site-specific studies. The previous

Guidelines (ANZECC 1992) also recommended that site-specific studies be

undertaken in many cases.

Define the water body by ecosystem classification. Using appropriate scientific

information the ecosystem can be classified into its corresponding type (up to six

types are recognised for the guidelines for physical and chemical stressors;a see

figure 3.1.3). The new Guidelines recognise the diverse range of ecosystem types

a See Sections

3.1.2

and 3.3
3.1.1 Philosophy and steps to applying the guidelines

Version — October 2000 page 3.1–5

in Australia and New Zealand, and the need to consider the particular attributes of

each ecosystem to achieve effective management.

Determine the environmental values. These have been described in Chapter 2.a

Determine the level of protection required. What condition should the ecosystem

be in, and what level of change would be regarded as acceptable? Three levels of

ecosystem condition are proposed as a basis for applying the guidelines.b

Identify environmental concerns. What are the main concerns or problems? For

most chemical contaminants the issue is generally toxicity,c but eight other

problems or issues can result from physical and chemical stressors.d

Determine the natural and human-induced factors affecting the ecosystem. It is

important to identify and collate information about the most important natural

processes and human activities that could influence the system being evaluated.

These processes and activities need to be taken into account when conceptual

models are being formulated to improve understanding of the system. They will

also guide subsequent management strategies developed to improve water quality

and designs for water quality monitoring programs.

Determine management goals. Next, define the management goals or targets, in

terms of measurable indicators of the condition (or state) of the ecosystem.

Indicators are usually biological parameters, but may also be physical and chemical

parameterse such as toxicant concentrations (in water column and in sediments)

and concentrations or loads of physical and chemical stressors. f

3.1.1.2 Determine appropriate guideline trigger values for selected indicators

The next exercise is predominately a desk-top study, using existing reference data

and other biological, physical and chemical information about the system. Some

preliminary analyses may be required to characterise the nature and dispersion

behaviour of contaminants. Four steps are involved:

1. Determine a balance of indicator types. The extent of the water quality

assessment program and the level of detail it must achieve will depend partly

upon the level of protection assigned to the water resource and the local

information constraints. More detailed investigation (and therefore additional

monitoring and assessment effort) would be expected for sites assigned high

levels of protection and for sites where serious constraints are identified, such

as lack of pre-disturbance data.g

2. Select relevant indicators. Determine indicators which will be relevant to the

environmental concerns and management goals. An indicator is a parameter4

that can be used as a measure of the quality of water.

3. Determine appropriate guideline trigger values. Determine guideline trigger

values for all indicators, taking into account level of protection. For physical

and chemical stressors and toxicants in water and sediment, the preferred

approach to deriving trigger values follows the order: use of biological effects

data, then local reference data (mainly physical and chemical stressors), and

finally (least preferred) the tables of default values provided in the Guidelines

(see figure 3.1.2). (While the default values are the least preferred method of

 

4 Readers who also read the Monitoring Guidelines (ANZECC & ARMCANZ 2000) should note

that there the term ‘indicator’ is only used to refer to parameters that, either severally or singly,

can indicate ecosystem condition.

a See Section

2.1.3

b Section 3.1.3

c Section 3.4

and 3.5

d Section 3.3

e Section 2.1.4

and 3.2

f Section 3.3.2

g Section 7.2.1
Chapter 3 — Aquatic ecosystems

page 3.1–6 Version — October 2000

deriving trigger values, it is conceded that these will be most commonly sought

and applied until users have acquired local information.)

4. Select specific indicators for inclusion in the monitoring and assessment

program. The choice of indicators will be based upon the level of protection

assigned to the water body, local information constraints, resource constraints,

availability of expertise and an initial hazard assessment. The hazard

assessment is based upon a comparison of estimated (first-pass) ambient

concentrations of indicators against the guideline trigger values determined

from the previous step.

Local biological effects data

(e.g. ecotoxicity tests, including multiple

species toxicity tests, mesocosms)

Preferred hierarchy for

deriving trigger values

Local or site-specific information

Local reference data

(mainly physical and chemical stressors; for

toxicants and sediments, applies only for the

case where background data exceed default

values from the box immediately below)

Default approach

Generic effects-based

guidelines

(Toxicants — Table 3.4.1

Sediments — Table 3.5.1)

Regional reference data

(Physical and chemical

stressors only — see Tables

3.3.2 to 3.3.11)

Decision trees

• Guideline packages for physical and chemical stressors (section 3.3.3)

• Applying guideline trigger values to sites for toxicants (section 3.4.3)

• Applying the sediment quality guidelines for sediments (section 3.5.5)

Most preferred

Least preferred

 YES, if potential guidelines exceedance is to be

assessed or if the trigger value can be refined a

a Local biological effects data and data from local reference site(s) that closely match test site generally not

required in the decision trees — see Section 3.1.5

Figure 3.1.2 Procedures for deriving and refining trigger values, and assessing test sites,

for physical and chemical stressors and toxicants in water and sediment. Dark grey shading

indicates most likely point of entry for users requiring trigger values.

3.1.1.3 Risk-based application of the guidelines

This is the final part of the framework for applying the guidelines. In summary, for

each issue (such as toxicity, algal blooms, deoxygenation) or type of water quality

indicator (physical/chemical stressor, toxicant and sediment) the Guidelines

provide detailed decision frameworks in the form of decision trees or guideline

‘packages’ for applying the guideline trigger (low risk) values, rather than
3.1.2 Features and classification of aquatic ecosystems in Australia and New Zealand

Version — October 2000 page 3.1–7

simplistic threshold numbers for single indicators. If data from a test site exceed

the trigger value, the decision trees are used to determine if the test values are

inappropriately (unnecessarily) ‘triggering’ potential risk and hence management

response. For this, ecosystem-specific modifying factors are introduced to assess

test data. The decision trees also enable the guideline trigger values to be adjusted

and refined. Further introduction to the use of decision trees in this assessment of

test site data and refinement of trigger values is provided in section 3.1.5.

While it is not mandatory to use decision frameworks, they are recommended so

that the resulting guidelines are relevant to the site. The guideline trigger values are

based on bioavailable concentrations, and hence are relatively conservative when

compared with total concentrations in the field, so the use of the decision

frameworks will increase guideline concentrations in most cases.

For biological indicators a general framework is applied, instead of a decision-tree

framework.

3.1.2 Features and classification of aquatic ecosystems in Australia and New

Zealand

3.1.2.1 Ecosystem features that may affect water quality assessment and ecosystem

protection

There is a diverse range of ecosystem types in Australia and New Zealand,

including tropical, temperate, arid, alpine and lowland. Within ecosystem types,

waterbodies may be static, flowing or ephemeral, deep or shallow, and fresh,

brackish or saline.

Variations in physical and chemical water quality variables can occur naturally

through droughts and floods, climatic conditions and erosion events, and can have

important consequences for the biota. Variations in climate, and, consequent

variations in rainfall, runoff and river flow, are particularly marked in Australia

(Finlayson & McMahon 1988, Harris & Baxter 1996, Harris 1996), and are

strongly linked to climate variability through mechanisms such as the El Niño–

Southern Oscillation or ENSO (Simpson et al. 1993).

Elsewhere in the Guidelines, a comprehensive account of the features of Australian

and New Zealand ecosystems is provided, together with some of the consequences

of these features that should be taken into account when considering water quality

assessment and ecosystem protection.a Table 3.1.1 summarises these issues.

3.1.2.2 Classifying the ecosystem

The wide range of geographic, climatic, physical and biological factors that can

influence a particular aquatic ecosystem makes it essential that ecosystem

management incorporates site-specific information together with more general

scientific information relating to ecosystem changes. This is the basis of the new

approach to the management of aquatic ecosystems,b involving the use of decision

frameworks to tailor water quality guidelines to local conditions. A first step in

tailoring guidelines to local conditions is to choose an appropriate category of

ecosystem; hence the need to classify the ecosystem being monitored.

a See Appendix 2

(Vol. 2)

b See outline

in Section 3.1.5
Chapter 3 — Aquatic ecosystems

page 3.1–8 Version — October 2000

Table 3.1.1 Some features of Australian and New Zealand ecosystems that have possible

consequences for water quality assessment and ecosystem protection.

Ecosystem feature Possible consequence

High degree of endemism

amongst the biota of many

Australian and New Zealand

ecosystems (fresh and marine)

Possible risks to natural heritage and conservation values

Naturally low nutrient status of

many of Australia’s fresh and

marine systems

• Ecosystems are adapted to low nutrient status; (natural)

lack of algal grazers for example may mean algal

growth/blooms proceed unchecked

• Greater accuracy and precision may be required for water

sampling programs where early detection of trends in

nutrient concentrations is important

Fresh water systems of Australia

often dominated by sodium and

chloride

Greater ‘softness’ of these systems places biota at risk from

classes of contaminants for which water hardness and acidbuffering capacity may ameliorate toxicity

Water temperatures in Australian

aquatic ecosystems are often

higher and more varied than

those in northern hemisphere

ecosystems

More often, toxicity of chemicals increases with increasing

temperature — an important consideration given that most

toxicity data used in the Guidelines are derived from northern

hemisphere studies.

Many of Australia’s fresh water

systems have only

periodic/episodic flow or water

availability

• Dilution of contaminants is reduced at low/recessional flow

or water levels

• After dry periods, oxidative processes can produce

degradation products such as acidity that may mobilise

deposited contaminants with ‘first flush’ flows (e.g.

oxidation of sulfide deposits)

• Classifications based on trophic status, and developed for

deep lakes of Northern Hemisphere, unlikely to be

applicable to shallow Australian standing waters

Over recent years, there has been considerable activity in classifying ecosystems or

parts of them, and this experience has been used to develop the general scheme for

these Guidelines. This is a hierarchical classification, with different levels of detail

applying to different categories of indicator. For future versions of the Guidelines it

is envisaged that this classification will be developed further as knowledge

increases, with specific guidelines and protocols being developed for each

combination of indicator and ecosystem type. The annex of Appendix 2, Volume 2,

describes some of the research in ecosystem classification, with some commentary

on recent applications of more detailed schemes in Victoria and New Zealand that

may be useful in future revisions of these Guidelines.

The ecosystem classification is given in figure 3.1.3. Note that each of the broad

categories of indicators has a different level of detail in terms of the ecosystem

classification. Thus for sediments, the guidelines make no distinction between

freshwater and marine systems, whereas for chemical and physical stressors there

are six categories of ecosystem. This approach has been adopted because different

levels of detail are available or applicable to each category of indicator:

information about sediment indicators is at a relatively early stage of development

whereas chemical and physical stressors have a much longer history of use in water

quality monitoring.
3.1.3 Assigning a level of protection

Version — October 2000 page 3.1–9

Sediments

(section

3.5)

Toxicants

(section

3.4)

Biological

indicators

(section

3.2)

 Physical

& chemical

stressors

(section

3.3)

All aquatic ecosystems

Marine

Marine

Estuarine Coastal &

marine

Freshwater

Standing waters Flowing waters

Lakes &

reservoirs Wetlands

Upland

rivers &

streams

Lowland

rivers &

streams

Figure 3.1.3 Classification of ecosystem type for each of the broad categories of indicators

(in grey boxes at left of the diagram)

The classification is necessarily coarse. There is no subdivision of estuaries, for

example, into those dominated by rivers or by marine influences, or those

permanently open to the sea, or temporarily or permanently closed (cf. Hodgkin

1994). Nor is there sufficient information to characterise the water quality

requirements of ephemeral rivers or saltwater lakes. Similarly, it should be possible

to subdivide these categories on the basis of climate (e.g. tropical vs. temperate),

but there is insufficient information available at present about the aquatic ecology

of tropical and temperate ecosystems in Australia and New Zealand to make such

subdivision meaningful.

Subsequent revisions of the Guidelines should further refine the broad ecosystem

classification scheme recommended here. Ideally, within an overall framework of

guiding principles and approaches, there should be a separate set of guidelines for

each ecosystem type — this should be the long-term aim of the Guidelines.

3.1.3 Assigning a level of protection

To define a level of protection this section describes a hierarchy of ecosystem

conditions, and recommends threshold levels of change that are acceptable for each.

The Guidelines also provide data or advice to assist relevant jurisdictions to make

their own informed decisions on alternative levels of protection where desired.

3.1.3.1 Ecosystem condition and levels of protection

The previous Guidelines (ANZECC 1992), in describing the concept of levels of

protection, recognised two categories of aquatic ecosystem condition: (i) pristine or

outstanding ecosystems for which maintenance of the existing water quality was
Chapter 3 — Aquatic ecosystems

page 3.1–10 Version — October 2000

deemed appropriate; and (ii) all remaining ecosystems to which the guidelines

would be applied to manage water quality. In this document the concept is

extended to acknowledge three categories of ecosystem condition, with a level of

protection ascribed to each.

Three ecosystem conditions are recognised.

1. High conservation/ecological value systems — effectively unmodified or other

highly-valued ecosystems, typically (but not always) occurring in national

parks, conservation reserves or in remote and/or inaccessible locations. While

there are no aquatic ecosystems in Australia and New Zealand that are entirely

without some human influence, the ecological integrity of high

conservation/ecological value systems is regarded as intact.

2. Slightly to moderately disturbed systems — ecosystems in which aquatic

biological diversity may have been adversely affected to a relatively small but

measurable degree by human activity. The biological communities remain in a

healthy condition and ecosystem integrity is largely retained. Typically,

freshwater systems would have slightly to moderately cleared catchments

and/or reasonably intact riparian vegetation; marine systems would have

largely intact habitats and associated biological communities. Slightly–

moderately disturbed systems could include rural streams receiving runoff

from land disturbed to varying degrees by grazing or pastoralism, or marine

ecosystems lying immediately adjacent to metropolitan areas.

3. Highly disturbed systems. These are measurably degraded ecosystems of lower

ecological value. Examples of highly disturbed systems would be some

shipping ports and sections of harbours serving coastal cities, urban streams

receiving road and stormwater runoff, or rural streams receiving runoff from

intensive horticulture.

The third ecosystem condition recognises that degraded aquatic ecosystems still

retain, or after rehabilitation may have, ecological or conservation values, but for

practical reasons it may not be feasible to return them to a slightly–moderately

disturbed condition.

A level of protection is a level of quality desired by stakeholders and implied by the

selected management goals and water quality objectives for the water resource. The

water quality objectives may have been derived from default guideline values

recommended for the particular ecosystem condition, or they may represent an

acceptable level of change from a defined reference condition; it can be formalised as

a critical effect size.a Where appropriate, the reference condition is defined from as

many reference sites as practicable using pre-impact data where appropriate.b The

reference condition could correspond to one of the three recognised condition levels

described above, depending upon the desired level of protection.

Key stakeholders in a region would normally be expected to decide upon an

appropriate level of protection through determination of the management goals and

based on the community’s long-term desires for the ecosystem. The philosophy

behind selecting a level of protection should be (1) maintain the existing ecosystem

condition, or (2) enhance a modified ecosystem by targeting the most appropriate

condition level. (Thus the recommended level of protection for ‘condition 1

ecosystems’ (above) would be no changec beyond any natural variability.) This is

a See box 2.3

& Section 3.1.7

b Section 3.1.4

c Footnote 2

on page 2-9
3.1.3 Assigning a level of protection

Version — October 2000 page 3.1–11

the starting point from which local jurisdictions might negotiate or select a level of

protection for a given ecosystem: in doing so, they might need to draw upon more

than the general scientific advicea provided in these Guidelines. A number of other

factors, such as those of a socio-economic nature, might need to be included in the

decision making process.

3.1.3.2 A framework for assigning a level of protection

When stakeholders are deciding upon an appropriate level of protection for

ecosystems, it is suggested that they consider the following framework based on

the three ecosystem conditions recognised above.

Some waters (e.g. many of those in national parks or reserves) are highly valued for

their unmodified state and outstanding natural values (condition 1 ecosystems).5 In

many countries and in some Australian states these waters are afforded a high degree

of protection by ensuring that there is no reduction in the existing water quality,

irrespective of the water quality guidelines (ANZECC 1992).

The present Guidelines recommend that for condition 1 ecosystems the values of

the indicators of biological diversity should not change markedly. To meet this

goal, the decision criteria for detecting a change should be ecologically

conservative and based on sound ecological principles.b Moreover, a precautionary

approach is recommended — management action should be considered for any

apparent trend away from a baseline, or once an agreed threshold has been reached.

Any decision to relax the physical and chemical guidelines for condition 1

ecosystems should only be made if it is known that such a degradation in water

quality will not compromise the objective of maintaining biological diversity in the

system. Therefore, considerable biological assessment data would be required for

the system in question, including biological effects and an ongoing monitoring

program based on sufficient baseline data. The nature of contaminants expected in

the receiving waters might also affect decisions on this issue.c Where there are few

biological assessment data available for the system, the management objective

should be to ensure no change in the concentrations of the physical and chemical

water quality variables beyond natural variation.

Where data for a reference/control site have only been collected for a limited

period and the reference condition cannot be clearly characterised, the power of

detection should be increased by using more indicators, and/or more

reference/control sites and/or more monitoring sites placed along any probable

disturbance gradients.

For slightly to moderately disturbed ecosystems (‘condition 2 ecosystems’), some

relaxation of the stringent management approach used for condition 1 ecosystems

may be appropriate. An increased level of change might be acceptable, or there

might be reduced inferential strength for detecting any change in biological

diversity. Nevertheless, as for condition 1 ecosystems, maintenance of biological

diversity relative to a suitable reference condition should be a key management

goal. The Guidelines provide specific guidelines for biological indicators for each

 

5 While waters in many remote and inaccessible locations may retain an unmodified condition, the

level of protection assigned to these systems is a jurisdictional decision made in consultation with

stakeholders. It does not automatically follow that these waters default to ‘condition 1

ecosystems’.

a See Section

2.1.3

b Sections

3.2.1.1, 3.1.7

and 7.2.3.3

c Section

3.1.3.3
Chapter 3 — Aquatic ecosystems

page 3.1–12 Version — October 2000

of the three ecosystem conditions.a For the other types of water quality indicator,

the default guidelines in Sections 3.3–3.5 provide a suitable level of protection for

condition 2 ecosystems.

The situation for highly disturbed ecosystems (‘condition 3 ecosystems’) can be

more flexible. The general objective might be to retain a functional, albeit

modified, ecosystem that would support the management goals assigned to it. In

most cases the ecological values of highly disturbed ecosystems can be maintained

by the direct application of the guidelines contained in this chapter. However, there

could be situations where these guidelines would be too stringent and a lower level

of protection would be sought. Some guidance to assist managers in these

situations is provided in the discussion of each indicator type.b

Table 3.1.2 summarises a general framework for considering levels of protection

across each of the indicator types for each of the ecosystem conditions.

The three levels of protection described above form just one practical but arbitrary

approach to viewing the continuum of disturbance across ecosystems. Inevitably,

stakeholders in different jurisdictions, catchments or regions will make different

judgements about ecosystem conditions. For example, an ecosystem that is

regarded as highly disturbed in one area could be regarded as only slightly to

moderately disturbed in a more populated region. This makes it imperative, as

emphasised in these Guidelines, that the setting of levels of protection is carried out

in an open and transparent way, involving all key stakeholders, so that a fair and

reasonable outcome is achieved.

Note that even though a system is assigned a certain level of protection, it does not

have to remain ‘locked’ at that level in perpetuity. The environmental values and

management goals (including level of protection) for a particular system should

normally be reviewed after a defined period of time, and stakeholders may agree to

assign it a different level of protection at that time. However, the concept of

continual improvement should be promoted always, to ensure that future options

for a water resource are maximised and that highly disturbed systems are not

regarded as ‘pollution havens’.

3.1.3.3 Alternative levels of protection

Local jurisdictions may negotiate alternative site-specific levels of protection after

considering factors such as:

• whether a policy of ‘no release’ (total containment) of contaminants applies;

• the nature of contaminants that might reach aquatic ecosystems. (Greater

consideration might be given to those ecosystems receiving contaminants or

effluents of potentially high toxicity and which are persistent in the environment,

e.g. metals. Alternatively, differing levels of protection could apply according to

the anticipated capacity of an ecosystem to readily recover from impact if

contamination is to be of short duration.)

• perceived conservation/ecological values of the system additional to those

recognised in the simple classification of ecosystem condition described in

Sections 3.1.2 and 3.1.3.1.

a See Section

3.2.4

b Sections

3.1.8

& 3.2 to 3.5
Table 3.1.2 Recommended levels of protection defined for each indicator type

Ecosystem Level of protectioncondition Biological indicators Physical & chemical stressors Toxicants 1 Highconservation/ecological value

• No change in biodiversity beyond

natural variability. Recommendecologically conservativedecision criteria for level ofdetection.

• Where reference condition is

poorly characterised, actions toincrease the power of detecting achange recommended.

• Precautionary approach

recommended for assessment ofpost-baseline data through trendanalysis or feedback triggers.

• No change beyond natural variability recommended, using

ecologically conservative decision criteria for detecting change.Any relaxation of this objective should only occur wherecomprehensive biological effects and monitoring data clearlyshow that biodiversity would not be altered.

• Where reference condition is poorly characterised, actions to

increase the power of detecting a change recommended.

• Precautionary approach taken for assessment of post-baseline

data through trend analysis or feedback triggers.

• For toxicants generated by 

detection at any concentratfor investigating their source management intervention1; 

occurring toxicants, backgroshould not be exceeded.Where local biological or chyet been gathered, apply thprovided in sec 3.4.2.4.Any relaxation of these objeoccur where comprehensive and monitoring data clearly biodiversity would not be alt

• In the case of effluent disch

assessment (DTA) should a

• Precautionary approach tak

post-baseline data through feedback triggers.

2 Slightly tomoderatelydisturbedsystems

• Negotiated statistical decision

criteria for detecting departurefrom reference condition.Maintenance of biodiversity still akey management goal.

• Where reference condition is

poorly characterised, actions toincrease the inferential strengthof the monitoring programsuggested.

• Precautionary approach may be

required for assessment of postbaseline data through trendanalysis or feedback triggers.

• Always preferable to use data on local biological effects to

derive guidelines.If local biological effects data unavailable, local or regionalreference site data used to derive guideline values usingsuggested approach in sec 3.3.2.3. Alternatives to the defaultdecision criteria for detecting departure from reference conditionmay be negotiated by stakeholders but should be ecologicallyconservative and not compromise biodiversity.Where local reference site data not yet gathered, apply default,regional low-risk trigger values from sec 3.3.2.5.

• Precautionary approach may be required for assessment of

post-baseline data through trend analysis or feedback triggers.

• Always preferable to use da

effects (including DTA) to dIf local biological effects dat

default, low-risk trigger valu

• Precautionary approach ma

assessment of post-baselinanalysis or feedback trigger

• In the case of effluent disch

required.

3 Highlydisturbedsystems

• Selection of reference condition

within this category based oncommunity desires. Negotiatedstatistical decision criteria fordetecting departure fromreference condition may be morelenient than the previous twocondition categories.

• Local or regional reference site data used to derive guideline

values using suggested approach in sec 3.3.2.3. Selection ofreference condition within this category based on communitydesires. Negotiated statistical decision criteria may be morelenient than the previous two condition categories.Where local reference site data not yet gathered, apply default,

regional low-risk trigger values from sec 3.3.2.5; or usebiological effects data from the literature to derive guidelines.

• Apply the same guidelines a

moderately’ disturbed systelower protection levels provGuidelines may be accepte

• DTA could be used as an a

for deriving site-specific gui

1 For globally-distributed chemicals such as DDT residues, it may be necessary to apply background concentrations, as for naturally-occurring toxicants.

Version — October 2000 page 3.1–13
Chapter 3 — Aquatic ecosystems

page 3.1–14 Version — October 2000

3.1.4 Defining a reference condition

For some water quality indicators, users will need to define a reference condition

that provides both a target for management actions to aim for and a meaningful

comparison for use in a monitoring or assessment program. The reference

condition is particularly appropriate to condition 2 or condition 3 ecosystems, and

is a key component of the framework provided in figure 3.1.1a for applying the

guidelines. For biological indicators, and for physical and chemical stressors where

no biological or ecological effects data are available, the preferred approach to

deriving guideline trigger values is from local reference data; for toxicants in water

or sediment this reference condition, sometimes called background data, may in

some situations supplant the default guideline values.b The next sections

summarise the sources of information that can be used for defining a reference

condition, and clarify the terminology of ‘controls’ and what constitutes a ‘site’,

respectively. Chapter 7 describes the design of monitoring programs, but also see

the Monitoring Guidelines (ANZECC & ARMCANZ 2000).

3.1.4.1 Sources of information

The reference condition for sites that may or may not be disturbed at present can be

defined in terms of these sources of information: historical data collected from the

site being assessed; spatial data collected from sites or areas nearby that are

uninfluenced (or not as influenced) by the disturbance being assessed; or data

derived from other sources.

1. Historical data collected from the site being assessed will usually represent

measurements made before a disturbance or before management actions. For

example, measurements of salinity collected from a river before the initiation

of an irrigation scheme may be used to set the reference condition for salinity

that stakeholders would hope to achieve in a rehabilitation program. For cases

where rehabilitation of degraded systems can only be achieved over long timescales, such benchmarks may be progressively stepped by way of a series of

targets intermediate between the existing and pre-disturbance condition.

2. Spatial data can be collected from reference sites or areas nearby that are

relatively uninfluenced by the disturbance being assessed. The sites include,

but are not restricted to, control sites which are identical in all respects to the

site being assessed (sometimes called the test site) except for the disturbance

(the distinction between control and reference sites is explained more fully

below). For example, the impact of an ocean outfall on marine benthos may be

judged relative to the values of the selected indicators in one or more reference

sites that are in the same vicinity but lack any influence of an outfall. For

modified ecosystems, ‘best-available’ reference sites may provide the only

choice for the reference condition.c

3. Data can be derived from other sources if there are neither suitable historical data

nor comparable reference sites. The reference condition may be identifiable from

the published literature, from models, from expert opinion, from detailed

consultations with stakeholders, or from some combination of all of these. For

example, when setting the reference condition for nutrient concentrations in a

series of wetlands, information on desirable and attainable concentrations may

come from published studies from similar regions overseas, from nutrient models

a See Section

3.1.1.2

b See also

Sections

3.4.3.2,

7.4.4.2, 7.4.4.4

c Section 3.1.8
3.1.4 Defining a reference condition

Version — October 2000 page 3.1–15

with appropriate local adaptations, from scientific advice about what levels of

nutrients result in undesirable end-points (e.g. blooms of toxic cyanobacteria) and

from input from community groups and landholders about their expectations of

what the wetlands should become. The necessary negotiations need considerable

technical and social skill. The reference condition should not be defined in terms

of ecological targets that are impossible to attain. Conversely, the reference

condition should represent a substantial achievement in environmental protection

that is agreeable to the majority of stakeholders.

Obviously, the best reference conditions are set by locally appropriate data. If the

disturbance to be assessed has not yet occurred, then pre-disturbance data provide a

valuable basis from which to define the reference condition. If the disturbance has

already occurred then data from reference sites and other appropriate sources can

be used to define the reference condition.a These issues are treated in more depth

in the Monitoring Guidelines (ANZECC & ARMCANZ 2000).

In summary, the reference condition must be chosen using information about the

physical and biological characteristics of both catchment and aquatic environment

to ensure the sites are relevant and represent suitable target conditions. Some of the

important factors that should be considered are these:

• data collected prior to the disturbance need to be of sufficient quality and

timespan to provide valid comparisons with post-disturbance data;b

• where possible, pre-disturbance data should be collected from appropriate

control or reference sites as well as from the site(s) subjected to the disturbance;

• the definition of a reference condition must be consistent with the level of

protection proposed for the ecosystem in question — unimpacted, or slightly

modified or relatively degraded (where the community does not wish to

rehabilitate a degraded ecosystem to such a high level);

• sites should be from the same biogeographic and climatic region;

• reference site catchments should have similar geology, soil types and

topography;

• reference sites should contain a range of habitats similar to those at the test

sites;

• reference and test sites should not be so close to each other that changes in the

test site due to the disturbance also result in changes in the reference sites, nor,

conversely, should changes in the reference sites mask changes that might be

occurring in the test site.

3.1.4.2 Clarification of the terms ‘control’ and ‘reference’

In the context of monitoring and assessing water quality, a disturbance (or

‘treatment’) is an event or occurrence which may or may not result in an effect on a

water body, and the ‘control’ refers to a set of observations taken from conditions

identical to the disturbed conditions except for the disturbance.

Controls may be defined in terms of space (‘spatial controls’) or time (‘temporal

controls’) or both. For example, if stakeholders had to assess the effect of

urbanisation on a wetland, they might be able to find similar wetlands nearby with

no urban development in their catchments, to act as spatial controls. If development

a See Section

3.1.8

b Section

7.2.3.1 & the

Monitoring

Guidelines
Chapter 3 — Aquatic ecosystems

page 3.1–16 Version — October 2000

had not commenced, the stakeholders could collect data from the wetland at this

stage to use as a temporal control, and the inferences that they could make about

the effects of urbanisation on the wetland would be strongest if they collected data

from the spatial controls before and after urbanisation as well.a

In environmental science, as in classical field experiments, ‘controls’ are unlikely

to be completely identical to ‘treatments’. If there is important systematic variation

between ‘controls’ and ‘treatments’, this can be incorporated into the sampling

program and statistical analysis via regression-related techniques. Analysis of

covariance is one classical technique for handling such differences. Some statistical

textbooks refer to these procedures as methods of statistical control (which should

not be confused with statistical process control or control charting).

Sometimes controls are impossible to find, but there are still sites or sets of temporal

observations that represent a desirable set of conditions that the disturbed site(s) could

ultimately match, if rehabilitated. Thus the term reference condition or reference site

denotes something more general than the ‘control’. In the wetland example above,

there may be no wetlands on similar soil types that are completely free of

urbanisation, and even those with little urbanisation may differ in the dominant landuse in their catchments. In this instance, stakeholders would need to negotiate over

which wetlands would provide the most appropriate reference conditions.

The use of reference sites to establish targets on a broader regional scale is

becoming increasingly popular. For example, this method is the basis of the

national rapid biological assessment procedure adopted for the AUSRIVAS

program (Schofield & Davies 1996). In this case, reference sites are usually

selected in ecosystems that are similar to and in the vicinity of a test ecosystem but

unimpacted or little changed.

3.1.4.3 What constitutes ‘a site’

For the purposes of these Guidelines, a site refers to a location which is being

monitored or assessed, and constitutes the smallest spatial unit that will be used in

judging whether an impact has occurred. Thus a site may vary in size from a few

square metres, as in the case of a stretch of an upland stream, to a few square

kilometres, as in the case of a large seagrass bed. In the case of the upland stream,

stakeholders may be interested in monitoring the water quality of the site and

comparing it with, for example, several other reference sites on other streams nearby.

For the large seagrass bed, selected indicators might be measured in that bed and

compared with measures from similar seagrass beds elsewhere on the coast.

Only rarely will sites be homogeneous internally. Concentrations of chemicals may

vary across a stream, and there may be differences in the sediments and species

composition across a seagrass bed. There are a number of strategies for dealing

with such within-site variation.b For large sites, this may involve sampling at more

than one spatial scale within the site. For example, in the seagrass bed, several

sampling locations of, say, 100 m2 may be selected, within which smaller ‘sublocations’ (e.g. 1 m2

 quadrats) may be selected. Care needs to be taken not to

confuse these within-site spatial units with the site itself. Note that in the literature

there is little consistency in the use of terms such as ‘site’, ‘location’, ‘area’, etc.,

so readers should not assume that the term ‘site’ in other publications automatically

equates with the term ‘site’ as it is used in these Guidelines and in the Monitoring

Guidelines (ANZECC & ARMCANZ 2000).

a See Section

7.2

b See Ch 7

and the

Monitoring

Guidelines
3.1.5 Decision frameworks for assessing test site data and deriving site-specific water quality guidelines

Version — October 2000 page 3.1–17

3.1.5 Decision frameworks for assessing test site data and deriving sitespecific water quality guidelines

The effect of a particular stressor or toxicant on biological diversity or ecological

integrity depends upon three major factors:

• the nature of the ecosystem, its biological communities and processes;

• the type of stressor;

• the influence of environmental factors which may modify the effect of the

stressor.

Aquatic ecosystems are variable and complex and difficult to manage. The previous

Guidelines recognised the need to address this variability and the influence of

environmental factors on stressors. This section introduces the concept of managers

using risk-based decision frameworks to assess test site data and to tailor guidelines

to suit regional, local or site-specific conditions. It provides a consistent framework

that can be used in New Zealand and the states and territories of Australia for

applying the guidelines in a meaningful way to the various types of aquatic

ecosystems in these regions. The approach addresses the issues of variability and

complexity, more realistically and effectively protecting biodiversity or ecological

integrity. As emphasised above, the approach does not constitute or require a full risk

assessment,a but simply assists in providing a site-specific estimate of whether a

stressor represents a low, possible or high risk to the aquatic ecosystem of interest.b

As already discussed, for non-biological indicators, these Guidelines recommend

guideline trigger values, which represent bioavailable concentrations or

unacceptable levels of contamination6 and are equivalent to the old single number

guidelines. If exceeded, these values trigger the incorporation of additional

information or further investigation to determine whether or not a real risk to the

ecosystem exists and, where possible, to adjust the trigger values into regional,

local or site-specific guidelines. The decision frameworks in Sections 3.3–3.5

demonstrate how this can be done.

Through the decision frameworks the ambient (existing) concentration of a

contaminant is compared with the guideline trigger value. The initial measurement

may be a relatively simple and therefore low-cost measurement (e.g. total

concentration). If the trigger value is not exceeded, the risk of an impact is low and no

further action is required. However, if the trigger value is exceeded there is some risk

of an impact occurring and successive, more complex steps should be taken to

account for environmental factors that modify the bioavailability, biological uptake or

toxicity of the stressor; this would also entail considering more complex monitoring

designs and negotiating effect sizes explicitly with stakeholders.c The final guideline

for that parameter should therefore reflect the real hazard to the particular ecosystem.

At each step in the process, a decision must be made on whether the adjusted trigger

value should be modified further or accepted. In general, the further one travels down

the series of steps the more resource-intensive the steps become; the user should

consider costs vs. benefits for each step. At any stage the decision tree process can be

 

6 Formally, the guideline trigger values are held to be a default, conservative statement of the

critical effect size as explained in section 3.1.7.

a See Section

2.1.4

b As indicated

in figures 3.3.1,

3.4.1, 3.5.1

c Sections 7.2

and 3.1.7
Chapter 3 — Aquatic ecosystems

page 3.1–18 Version — October 2000

terminated and the most recently modified trigger value applied as the guideline for

the particular situation. Because the default trigger values for toxicants at least are

conservative, a precautionary approach should be applied, using these values where

there is no background information on a particular system to which the guidelines are

to be applied, and no program for its acquisition. Alternatively the preferred option

might be to conduct toxicological studies or direct toxicity assessment relevant to the

site and use these data to derive a site-specific guideline.

Where a trigger value is refined using data gathered from a test site on a single or

limited sampling occasion(s), this does not automatically mean that this new value

applies henceforth in further test site/trigger value comparisons. More extensive

information is required before a guideline trigger value can be revised. For this, it

is important to distinguish two levels of refinement of guideline trigger values:

1. The first level applies to some indicators where guideline trigger values can be

adjusted and refined upfront, relatively simply, with fore-knowledge of the

range of values of some key physical and chemical parameters that occur in a

waterbody. This is particularly relevant to some toxicants. For example, the

toxicity and bioavailability of some metals (e.g. copper, zinc and cadmium) are

strongly influenced by water quality conditions such as hardness, dissolved

organic matter and pH, and recent literature has increased the understanding of

the toxicity of different metal species. The current state of knowledge limits

upfront revision of the trigger values for these metals to a hardness correction,

using the simple algorithms in table 3.4.3. There is also some scope for

modifying the trigger values for a few non-metallic inorganic and organic

toxicants, based on associated water quality parameters (e.g. pH, for the

ammonia trigger value). a

2. For most indicators and issues, however, trigger values are refined only after

continuous and extensive monitoring shows that test site data exceedances are

consistently assessed as posing no risk to the ecosystem, using the decision

trees. Trigger values can also be refined if longer term monitoring shows that

test site data are consistently below the trigger values or, for situations such as

naturally mineral-rich waters where the natural background total concentrations

of some metals exceed the new trigger values. For each of these cases, the

methods described in section 7.4.4.2/1 can be used to refine the guideline

trigger values for all (non-biological) indicator types.

It is not mandatory to use the decision frameworks, but they are important if

meaningful and appropriate guidelines are to be applied. Moreover, simple

adjustments and corrections such as those described in 1 above make this a costeffective exercise where data on key water quality parameters are available.

Generally, local biological effects data and data from local reference site(s) that

closely match the test site7 are not required in the decision trees. If test site data

exceed trigger values that have been derived from these local data, this would

 

7 This latter situation might be relevant to point-source disturbances in streams, where reference

sites are located upstream of test sites; the reference and test sites would be similar in all

appearances and there would be no confounding factors, apart from the disturbance and stressor in

question, occurring between the sites. Local reference sites even in an adjacent stream/tributary

might not necessarily closely match test sites.

a See Sections

3.4.3, 3.5.5
3.1.6 Using management goals to integrate water quality assessment

Version — October 2000 page 3.1–19

normally trigger management action because these locally-derived trigger values

already have ecosystem-specific modifying factors built into them. For the same

reason, these locally-derived trigger values do not require refinement themselves

through the decision trees, though if there was opportunity to derive guideline

values based upon sound local biological effects data, these should replace those

based upon local reference data.

These decision frameworks have not been developed for all specific indicators and

issues but are presented mainly to assist water managers explore some of the ways

in which the guidelines can be used in site-specific situations. Water managers and

regulators are encouraged to develop their own decision trees to address any

additional issues that may be encountered. General guidance on designing

monitoring and assessment programs is given in Chapter 7, with additional

background in the Monitoring Guidelines (ANZECC & ARMCANZ 2000).

3.1.6 Using management goals to integrate water quality assessment

In general, there is not enough scientific knowledge at present to allow anyone to make

confident predictions about the way in which a particular concentration of toxicant or

nutrient will affect species, habitats or ecosystems. It is therefore important to measure

the characteristics of the biological components of the ecosystem as well as the

physical and chemical water quality characteristics, to be able to confidently assess

whether an important change has occurred or is likely to occur.

Although there is a considerable body of toxicological knowledge that is very

important for use in specific circumstances, the overall effects of mixtures of

toxicants on a wide variety of species or habitats are not fully understood.

Environments are typically dynamic, as well as being subjected to natural stresses

like storms and floods, and little is known about the highly complex internal forces

that operate within them. Relatively accurate predictive models can be developed

for specific ecosystems,a but this generally entails sophisticated, resource-intensive

programs which may not be feasible. Use of unproven or overly simplistic causal

models to justify avoiding using biological indicators is dangerous.

The process of setting management goals,b as outlined earlier, is useful for

conceptualising the issues surrounding integration in aquatic ecosystem

management. The goals should be defined in a quantitative manner, need to be

comprehensively related to all valued attributes of the ecosystem, and, typically,

should be biologically based. In this sense, the biological variables themselves are

the management end-points, and chemical variables such as concentrations of

toxicants are the proximal causes in the cause–effect relationship. Management is

then directed to these management goals (such as maintaining a certain level of

species diversity). All management and assessment activities are integrated by an

explicit relationship to the management goals, in this case the maintenance and

improvement of species diversity. Hence biological diversity, or some other valued

aspect of the ecosystems, becomes the target for management and assessment, and all

activities are defined and implemented in terms of management of those ecosystem

attributes (Ward & Jacoby 1992).

Overall, the aim of a monitoring program should be to answer a discrete set of

questions (hypotheses)c which focus on whether the management goals are being

achieved. Conceptual models of the important biological and physical interactions

a See Section

2.2.3

b Section 2.1.3

c Sections

3.1.7, 7.1.2 and

7.2.3.3
Chapter 3 — Aquatic ecosystems

page 3.1–20 Version — October 2000

within the ecosystem will assist in choosing those indicators that could be potentially

useful for the monitoring or assessment program. This is important because monitoring

programs must be cost effective and in most circumstances it is not feasible to design

and implement a program that intensively monitors all aspects of water quality.

Another important aspect of integrated water quality assessment is the development

of communication networks across whole catchments to address broad-scale issues.

This is essential at two levels: first, because of the interdependent nature of the

environmental values themselves — the water quality of one value can potentially

affect others;a second, for protection of the whole aquatic ecosystem — while

water quality objectives might be met in riverine ecosystems upstream, the

cumulative effects of discharges and contaminant build up in depositional areas

downstream (e.g. wetlands, estuaries) must also be considered when setting water

quality criteria. This applies to a number of environmental values.b

3.1.7 Decision criteria and trigger values

Indicators used in these Guidelines are likely to respond continuously to the intensity

of a disturbance; an example is given in figure 3.1.4. At some point along this

continuum, the ecosystem will be deemed to have been adversely affected and the

value of the indicator at this point will be used as the criterion to make the decision

that ‘the ecosystem has been impacted’.

Strength of disturbance

Value of indicator

Value of indicator

at which

ecological

damage has

occurred

‘Threshold value’

of indicator which,

if exceeded,

should signal

management

action

Figure 3.1.4 Graphical depiction of the relationship between indicator response and

strength of disturbance, and threshold for management intervention

In most situations, we will need to make a decision before the ecosystem becomes

adversely affected so that management actions can be implemented in time to

prevent the ecosystem becoming damaged. In other words, we will need to select a

‘threshold value’ of the indicator that is smaller than that which indicates that the

a See Section

2.1.3

b Section 7.4.4.3

for related

discussion
3.1.7 Decision criteria and trigger values

Version — October 2000 page 3.1–21

ecosystem has been impaired. How much smaller this value needs to be depends on

the nature of the impact, the level of our understanding of the relationship between

changes in the indicator and ecological impact, and the lead-time necessary to

implement management actions.

For example, if the impact is likely to be irreversible or persistent then the

threshold value will need to be set at a very small value of the indicator so that

irreversible harm is avoided. Also, if there is only a very rudimentary

understanding of how a particular contaminant might affect an ecosystem then the

threshold value will need to be relatively small in case the ecosystem is more

sensitive to the contaminant than expected. Similarly, if there is a long lag between

detection that the threshold has been exceeded and implementation of some action

or decision, the threshold value will need to be set at a very small value.

Thus, the first task is to choose the threshold value for a given indicator. This is not

a trivial exercise, and requires all stakeholders to agree on these values before the

program of monitoring or assessment commences.

For the non-biological indicators in Sections 3.3–3.5, the guideline trigger values

represent the best currently-available estimates of what are thought to be

ecologically low-risk levels of these indicators for chronic (sustained) exposures.

a

For these indicators, the guideline trigger values provide the starting point for

negotiations about the threshold value and criterion for a management decision (i.e.

water quality objectives). Users should also be aware that short-term intermittent

(or pulse) exposures to very high contaminant or stressor values may also need to

be managed in certain situations. Negotiating the equivalent of a guideline trigger

value for the biological indicators in Section 3.2 is more complex, because the use

of these indicators has a shorter history in Australia and New Zealand and because

these indicators nearly always need to be used in a comparative fashion (e.g.

comparing values from the site(s) of interest with those in an appropriate reference

condition). This may also be true for the non-biological indicators in situations

where a reference condition is being used to establish the water quality objectives.

Thus, for all types of indicators, there will be situations in which simple guideline

trigger values of the chosen indicator will be inadequate as a threshold value or

criterion on which to activate management decisions and actions. In these

situations, stakeholders need to negotiate an effect size, which describes how much

deviation from the reference condition is tolerable before management has to

intervene. To understand what an effect size is, stakeholders need to appreciate the

following points:

1. the values of all indicators vary naturally, and

2. not all of this variation is ecologically important.

This means that some of the changes that can potentially be detected in an indicator

may be ecologically trivial; such small changes should not initiate management

action. The situation where we conclude that an important change has happened

when, in fact it has not, is technically referred to as a Type I error.

Conversely, many indicators are very variable naturally and intensive sampling

may be essential to detect ecologically important changes in the indicator. If the

sampling intensity is too small and the important change is missed, then a Type II

error is committed.

a See Section

7.4.4
Chapter 3 — Aquatic ecosystems

page 3.1–22 Version — October 2000

In the context of cooperative best management, stakeholders need to balance these

two types of ‘error’ and negotiate these issues before the monitoring or assessment

program commences.a

3.1.8 Guidelines for highly disturbed ecosystems

Apparently common problems in assessing water quality for highly disturbed

ecosystems of Australia and New Zealand include:

1. the difficulty in deciding upon suitable water quality guidelines and objectives

(and in particular, a level of acceptable ecological change);

2. the lack of suitable reference sites or data;

3. the lack of advice and guidelines for highly disturbed ecosystems in urban

regions.

These Guidelines offer the following advice and information on these issues.

3.1.8.1 Determining water quality guidelines and objectives

As discussed in Sections 1.2 and 2.2, the philosophy espoused in the Guidelines is

one of ‘continual improvement’ for places where water or sediment quality is poorer

than the agreed water quality objectives. For highly-disturbed ecosystems, the water

quality objectives can be seen as progressive and intermediate targets for long-term

ecosystem recovery. The Guidelines offer specific advice on assessing the success of

remediation programs.b

The Guidelines recommend that guideline trigger values for slightly–moderately

disturbed systems also be applied to highly disturbed ecosystems wherever possible.

If that is not possible, local jurisdictions and relevant stakeholders must negotiate

alternative values. For this situation, the Guidelines provide less conservative trigger

values for toxicants: the less conservative values suit two lower levels of ecosystem

protection (table 3.4.1). The Guidelines also offer the following advice, relevant to

all indicators (biological, physical and chemical, toxicants, sediments) when test

data are being compared with data from reference sites:c

1. Where reference sites of high quality are available, lower levels of protection

may be negotiated for the site under consideration, through selection of more

relaxed statistical decision criteria. This would not necessarily, and should not,

result in a water of lesser quality than that already prevailing.

2. Where no high quality reference sites are available, modified water bodies of

the best environmental quality in the region serve as reference targets (or

intermediate targets for ecosystem recovery). Where these data indicate that

certain toxicants occur naturally at levels exceeding the guideline trigger value,

the Guidelines make provision for the background level, if clearly established,

to become the site-specific guideline level.

Where a reference condition is used to define water or sediment (pore water)

quality targets, the bioavailable fraction must be determined and compared for

those toxicants that exceed the guideline trigger values.d For sediment

particulates, the dilute-acid-extractable (1M HCl) fraction is used as a

surrogate for bioavailability.e

c See also

Sections 3.1.4

and 3.1.8.2

a See also box

2.3; these issues

are expanded in

Section 7.2.3

b Sections

3.2.5 & 7.2.3.3

d Sections 3.4

and 3.5

e Section

3.5.5.2
3.1.8 Guidelines for highly disturbed ecosystems

Version — October 2000 page 3.1–23

Negotiating the ‘acceptable’ level of change for disturbed ecosystems, and hence

the level of protection of species, is a constant challenge faced by local

jurisdictions and relevant stakeholders (including the community).

As is recognised in the Guidelines, more research is needed to develop methods to

describe degrees of acceptable ecological change relative to reference conditions.a

The Guidelines give general advice for determining the size of ecological change

that would be considered important. It can be useful to examine data from existing

impacts elsewhere, especially if it is possible to compare impacts across a gradient

from mild to extreme. These can be used as yard-sticks to decide upon the degree

of ecological change or impact.

As a first step towards improvement in water quality, the Guidelines recommend

that local jurisdictions assess a range of options for determining site-specific

guideline values for highly disturbed ecosystems. One approach is to select

different levels of acceptable change (e.g. protection of 90% of species with 50%

confidence). Another is to assess the disturbed ecosystem against the best-available

reference water body in the region, as a benchmark for water quality.

Different site-specific guideline values developed using various methods can be

examined and weighted according to pre-determined criteria of quality and

relevance to the ecosystem. This should be done in a manner consistent with risk

assessment principles,b to arrive at an appropriate figure.

3.1.8.2 Lack of suitable reference sites or data

Often, water bodies over large continuous tracts of Australia and New Zealand are

highly disturbed and none of the adjacent water bodies is necessarily of better

quality than the water body(ies) of interest, insofar as serving as useful reference

sites. Nevertheless, even if water bodies of only slightly better quality can be

found, these provide useful reference data, particularly if these data serve as an

intermediate target for ecosystem recovery.

Where the issue is biological assessment of water quality in highly-disturbed inland

streams and rivers, rapid assessment using macroinvertebrate communities offers,

potentially and in practice, a most useful approach.c Recent findings from the

Australian Commonwealth-funded National River Health Program from which this

rapid assessment approach has been developed, indicate that macroinvertebrate

communities are very similar at the family level across vast tracts of inland

Australia. This means that relatively intact ecosystems in remote and less

developed parts of inland Australia (e.g. channel country of south-western

Queensland) may potentially provide useful reference data for highly disturbed

ecosystems in, say, north-western NSW, if family-level information about

macroinvertebrates serves as a suitable indicator of river health at this spatial scale.

3.1.8.3 Guidelines for highly disturbed ecosystems in urban regions

Most of the populace of Australia and New Zealand lives in large cities where

most, but not all, natural aquatic ecosystems are highly disturbed. Approaches from

Section 3.1.8.1 above, ‘Determining water quality guidelines and objectives’, are

applicable to the development of guidelines for highly disturbed ecosystems in

urban regions. Indeed, a great deal of work has been conducted in urban waterways

across Australia and New Zealand and on a variety of chemical and biological

monitoring and assessment programs — see box 3.1.4. Utilities in many of the

a See section

8.5.1 in Vol. 2

and Section

7.2.3.3

b Section 3.4.3

c Sections 3.2,

7.2.1 and 7.3.3
Chapter 3 — Aquatic ecosystems

page 3.1–24 Version — October 2000

smaller, and therefore less well-resourced, urban centres will be able to benefit

from these larger urban programs by applying the same principles of investigation

to their own situations.

Box 3.1.4 Examples of water quality assessment programs conducted

in major urban regions of Australia

These are some of the existing monitoring and research programs in streams, estuaries and

coastal systems in major urban centres.

For urban streams and wetlands:

• Sydney streams are monitored and studied through the Environmental Indicators

program of Sydney Water Corporation, and by NSW DLWC;

• Melbourne streams are monitored and studied by Melbourne Water, VIC EPA and the

CRC for Freshwater Ecology;

• a predictive model of the AUSRIVAS type for monitoring and assessing health of

streams in the Hobart region has been completed by the University of Tasmania

(Zoology Dept);

• wetlands of the Swan Coastal Plain.

For coastal marine areas and estuaries:

• water quality monitoring and assessment are included amongst the research programs

of the Centre for Research on Ecological Impacts of Coastal Cities (Sydney University);

• Port Phillip Bay Environmental Study;

• Moreton Bay;

• programs in and around Perth, such as the Perth Coastal Water Study, South

Metropolitan Coastal Water Studies, Perth Coastal Waters Management and

Consultative Process.

General:

• Thirteen studies on streams and estuaries were commissioned under the Urban subprogram of the National River Health Program, covering physical, chemical and

ecological aspects. Reports arising from the sub-program may be found at the

LWRRDC website (http://www.lwrrdc.gov.au).
Version — October 2000 page 3.2–1

3.2 Biological assessment

3.2.1 Introduction and outline

In broad terms, this section provides advice about the selection of biological

indicators to apply to various water quality problems,a and the analytical procedures

that should be used to monitor and assess change in these indicators.b The material in

this section is accompanied by little in the way of rationale or justification; those are

provided in other chapters of the guidelines. Generic issues of designing a program

for monitoring or assessment are given in Sections 7.1 and 7.2, with much

background material provided in the Australian Guidelines for Water Quality

Monitoring and Reporting (the Monitoring Guidelines, ANZECC & ARMCANZ

2000) (especially Chapters 3, 4 & 6). For substantiation of the recommended

approaches and additional guidance, an expanded discussion about the selection of

biological indicators is provided in Section 8.1 (Vol. 2), while a detailed account of

specific issues for biological monitoring and assessment is provided in Section 7.3. It

is important that the material presented in the current Section (3.2) is not read in

isolation of these other detailed accounts.

3.2.1.1 Philosophy and approach behind bioindicators of water quality

The following sections discuss the concepts and monitoring frameworks necessary

to assess aquatic biological communities. A key concept is that of ecological

integrity (health), defined in Section 3.1.1.

Biological assessment (bioassessment) can measure the desired management goals

for an ecosystem (e.g. maintenance of a certain diversity of fish species or certain

level of nuisance algae) as might be described in the management goals.

Bioassessment provides information on biological or ecological outcomes; these may

result from changes in water quality but may also result from changes in the physical

habitat (e.g. increased fine sediment deposition, or changes in hydrology) or from

changes in biological interactions (e.g. the introduction of exotic species or diseases).

Thus, bioassessment should be seen as a vital part of assessing changes in aquatic

ecosystems, and as a tool in assessing achievement of environmental values and

attainment of the associated water quality objectives. At the same time, the

resulting biological message provides an insight into a complex system which:

• integrates multiple natural and human changes in physico-chemical conditions;

• integrates disturbances over time;

• absorbs human effects into complex interacting biological communities and

processes;

• can give a signal from more than one component (e.g. multiple species or

community similarities or ecological processes).

The guidelines for biological assessment are intended to detect important departures

from a relatively natural, unpolluted or undisturbed state — the reference condition.c

An important departure is deemed to be one in which the ecosystem shows

substantial effects, including:

• changes to species richness, community composition and/or structure;

• changes in abundance and distribution of species of high conservation value or

species important to the integrity of ecosystems;

c Section 3.1.4

a See Sections

3.2.1.3 to 3.2.2.2

b Sections 3.2.3

to 3.2.4
Chapter 3 — Aquatic ecosystems

page 3.2–2 Version — October 2000

• physical, chemical or biological changes to ecosystem processes.

Important in this context does not mean mere statistical significance, which is only

a tool in the context of a specific monitoring design. Rather it means a change or

departure deemed practically significant, in relation to previously agreed

performance criteria, for failing to achieve a water quality objective.

The results of bioassessment may require interpretation using additional supporting

information on water quality and physical conditions at site, catchment or regional

scales. Bioassessment provides a window onto the condition of the ecosystem

being managed.

Bioassessment and biological indicators have come into use because the traditional

physical and chemical guidelines are too simple to be meaningful for biological

communities or processes. Strong variation in ecosystem processes and biological

community composition in time and space is characteristic of many surface water

environments, particularly in Australia.

Biological systems are very variable. It is important to understand that because of

this variability, sampling designs have a limited capacity to detect and quantify

change relative to an undisturbed or reference state. Any given sample size or

number of sample units taken during a monitoring or assessment program has

quantifiable constraints on its capacity to detect a change of a given magnitude.

There is a strong relationship between the power (in statistical terms) of a

monitoring program design, the magnitude of the effect that is detectable and the

sample sizes involved.

There is also a trade-off between a capacity to detect change, and the sample size,

and the chance of not detecting that change (or of detecting a change that has not

occurred). This trade-off is often negotiated on the basis of financial resources for

monitoring programs, since to increase sample sizes or numbers of sample units is

the most common way of increasing the power to detect a change.a

It is vital to recognise the need for high quality, comprehensive designs in

bioassessment and biological monitoring. Protocols are being developed for

bioassessment, with improved designs and rigour in site selection, sampling

approaches and analysis. Several examples of this are given in the following sections

on biological assessment.

3.2.1.2 A framework for biological assessment of water quality

Successful employment of a biological monitoring and assessment program for the

protection of aquatic ecosystems involves a series of steps:

1. define the primary management aims, including the level of protection desired

by the community and other stakeholders; define the management goals for

achieving protection of the ecosystem, and the environmental concerns;b

2. together with a balance of indicators, identify the biological assessment

objectives for protection of the water resource;c

3. select appropriate indicators and protocols to apply to the assessment

objectives;d

4. select the appropriate experimental design to apply to the indicator;e

b Section

3.1.1.1

c Sections

7.2.1 & 3.2.1.3

d Sections

3.2.2 and 3.2.3

e Section 3.2.3

and Ch.7

a See Sections

3.1.7 & 7.2.3.3
3.2.1 Introduction and outline

Version — October 2000 page 3.2–3

5. determine key decision criteria, i.e. acceptable level of change and statistical

sensitivity with which to detect such change;a

6. assess results from monitoring programs,b with feedback to management.

This framework of steps is also shown in figure 3.2.1.

Define primary management aims

(see figure 3.1.1)

Assessment objective

(section 3.2.1.3)

determined concurrently with a ‘balance of indicator types’ (from section 7.2.1)

e.g. How can we quickly

 determine the extent of

 the problem or potential

 problem?

Broad-scale assessment

e.g. How can we pre-empt or

 prevent irreversible

 damage, irretrievable

 habitat loss etc?

Early detection

e.g. How can we assess

 ecological importance of the

 impact or potential impact?

Biodiversity or 

ecosystem-level response

Select indicator and protocol

(sections 3.2.2 & 3.2.3)

Interpret results, assess whether 

WQOs1 are being achieved

(section 3.2.4.2)

Select appropriate design &

analysis

(section 3.2.3)

Determine management 

decision criteria

(section 3.2.4.1)

Applying the guidelines for biological assessment

Decision criteria met Decision criteria exceeded

(initiate remedial actions)

1 = Water Quality Objectives (section 2.1.5)

Figure 3.2.1 Decision tree for biological assessment of water quality

a See Sections

3.1.7, 7.2.3.3

and 3.2.4

b Section 3.2.4.2
Chapter 3 — Aquatic ecosystems

page 3.2–4 Version — October 2000

3.2.1.3 Biological assessment objectives for ecosystem protection

Having determined the level of protection required for an ecosystem, the

management goals for achieving that protection, and the environmental concerns

(fig 3.1.1), managers should identify assessment objectives for protection of the

water resource. The objectives will help managers select the most appropriate

biological indicators and protocols. Three broad assessment objectives are

described as follows:

1. Broad-scale assessment of ecosystem health (at catchment, regional or larger scales)

Resources will never be adequate to provide detailed, quantitative8 biological

monitoring and assessment of water quality over wide geographical areas of

Australia and New Zealand. Therefore, tools for rapid biological assessment

(RBA) are being developed that, while not providing detailed quantitative

information, are cost-effective and quick enough to generate adequate first-pass

data over large areas. The data may be adequate for management purposes or

they may help managers to decide what type of further information may be

required and from where.

Broad-scale assessment can be useful for the following applications:

• rapid, cost-effective and adequate first-pass determination of the extent of a

problem or potential problem, e.g. as applied to broad-scale land-use issues,

diffuse-source effluent discharges or information for State of Environment

Reporting;

• screening of sites to identify locations needing more detailed investigation;

• remediation programs being conducted over broad geographical areas

(catchment, regional or larger scales).

The most developed RBA method is AUSRIVAS, a method using macroinvertebrate

communities in rivers and stream. Rapid bioassesment protocols are also being

developed for riverine benthic algae (diatoms) and fish, as well as for

macroinvertebrate communities in wetlands and estuarine sediments.

2. Early detection of short- or longer-term changes

Prediction and early detection of possible effects are useful to any water quality

management program so that substantial and ecologically important disturbances

can be avoided. Early information enhances the options for management. For

example, where an effect is observed from a controlled discharge, it may be

possible to adjust the rate of release or of subsequent releases.

Predictive information and early detection in the field can result if specific and

sensitive programs are set up, incorporating study of sublethal responses of

organisms. If sampling sites for any indicator can be located in mixing zones

effectively creating spatial disturbance gradients, they will enhance early detection

and predictive capabilities.9

 

8 The adjective ‘quantitative’ from here on, in Section 3.2, refers to an indicator measurement program

that permits rigorous and fair tests of the potential disturbances under consideration; typically,

conventional statistical tools would be employed to attach formal probability statements to the

observations — see Section 3.2.3.

9 The purpose of sampling in mixing zones in this case is solely for enhancing inference about

disturbances in receiving waters, not for determining compliance in this zone.
3.2.1 Introduction and outline

Version — October 2000 page 3.2–5

Also, RBA programs operating over broad geographical regions may, through their

extensive coverage, pin-point potential ‘hot-spots’ that would otherwise be missed.

However, these programs do not incorporate very sensitive protocols.

Early detection can be important for:

• sites of special interest (e.g. sites of high conservation value, major

developments and/or point-sources of particular potential concern) where the

cost of failing to detect a disturbance in a timely manner may be too high;

• timely identification of water quality issues and problems that may exist over a

broad geographical region in response to a specific pressure;

• any situation where a management objective has been strongly linked to the

Precautionary Principle tenet of the National Strategy for Ecologically

Sustainable Development (ESD Steering Committee 1992).

3. Assessment of biodiversity

Often it is not sufficient simply to have detected change in an early detection

indicator because the information cannot easily be linked (if at all) to adverse

effects at population, community and ecosystem levels. To determine effects upon

the ecosystem as a whole and as important end-points in themselves, measures of

biodiversity, including ecosystem processes and the conservation status of sites,

should be key responses sought-after in monitoring programs.

Biodiversity and conservation status are best measured using species-level data

gathered from quantitative studies. Information gathered at higher levels of

taxonomic resolution will serve these needs if the data are correlated with

biodiversity or conservation status at species level (e.g. Wright et al. 1998). Even in

the best-resourced studies, it is inevitable that biodiversity assessment will usually be

limited to the measurement of ecosystem surrogates — communities/assemblages of

organisms, or habitat or keystone-species indicators where these have been closely

linked to ecosystem-level effects. Information on the ecological importance of effects

will best be met in programs that have regional coverage and encompass a full

disturbance gradient.

Whether the assessment objective is biodiversity, conservation status or ecosystemlevel responses for assessing ecological importance of disturbance (as measured by

community structure or ecosystem process attributes), this indicator is hereafter

termed biodiversity indicator.

The biodiversity assessment objective may be important for the following

applications:

• for sites of special interest where indicators are needed to measure biodiversity,

conservation status, and/or ecosystem-level effects for assessing ecological

importance of disturbance. Information gathered for such indicators is highly

complementary to that gathered for early detection indicators.

• through RBA programs, as a first-pass measure of biodiversity, conservation

status and/or ecosystem-level effects for assessing ecological importance of

disturbance, at sites and over a broader geographical region.

• in any situation where a management objective has been strongly linked to the

Ecologically Sustainable Development tenet of the ‘Maintenance of

biodiversity and ecological systems’ (National Strategy for Ecologically

Sustainable Development, ESD Steering Committee 1992).
Chapter 3 — Aquatic ecosystems

page 3.2–6 Version — October 2000

3.2.2 Matching indicators to problems

3.2.2.1 Broad classes of indicators and desired attributes

Desired or essential attributes of the broad indicator types (or methods) required to

meet the assessment objectives are listed in table 3.2.1. Each of the three

assessment objectives is discussed fully in Section 8.1.1 (Volume 2), but the main

points are summarised below.

1. Broad-scale assessment of ecosystem ‘health’

The indicator types relevant to a broad-scale assessment objective have these

attributes:

i. the measured response adequately reflects the ecological condition or integrity

of a site, catchment or region (i.e. ecosystem surrogate);

ii. where community or assemblage data are gathered, these and associated

environmental data can be analysed using multivariate procedures;

iii. approaches to sampling and data analysis are highly standardised;

iv. responses are measured rapidly, cheaply and with rapid turnaround of results;

v. results are readily understood by non-specialists;

vi. responses have some diagnostic value.

A range of studies of populations and communities could provide information about

the ecological condition or integrity of a site, catchment or region, but only rapid

biological assessment (RBA) methods would enable such information to be gathered

over wide geographical areas in a standardised fashion and at relatively low cost.

Resh and Jackson (1993), Lenat and Barbour (1994) and Resh et al. (1995) elaborate

upon features of RBA approaches as applied to stream macroinvertebrate

communities. Comment upon some RBA methods currently being applied to

freshwater fish communities is provided in Section 8.1.2.1 of Volume 2.

2. Early detection of short- or longer-term changes

To have a predictive or early detection capability, an indicator should ideally have

a response that is:

i. sensitive to the type of stressor;

ii. correlated with environmental effects (i.e. linked to higher-levels of biological

organisation);

iii. time- and cost-effective to measure;

iv. highly constant over time and space, which confers high power to detect small

changes;

v. regionally and socially relevant;

vi. broadly applicable.

These attributes are important because assessments of actual or potential

disturbances will only be as effective as the indicators chosen to assess them

(Cairns et al. 1993). However, the attributes are idealised characteristics only, and

in many cases some will conflict or will not be achievable. Therefore the more

important and achievable attributes must be decided upon, and appropriate

indicators must be chosen accordingly.
Table 3.2.1 Biological assessment objectives for different management situations and the recommended methods and indicator

Assessment objective Applications Recommended indicators Es

be 

1. Broad-scale assessment ofecosystem ‘health’ (catchment, regionalor larger scale)

Water quality on a catchment or regionalbasis (e.g. SoE reporting, catchmentmanagement indicators)

Rapid bioassessment (e.g. AUSRIVAS) • 

• 

• 

2. Early detection of short- or longerterm changes Sites of special interest (high conservation value, major developments

or point-sources of particular potentialconcern)

Laboratory: Direct toxicity assessmentField: Instream/riverside assays, biomarkers,bioaccumulation; spatial disturbance gradients inrelevant quantitative biological indicators

• 

• • 

Water quality on a regional basis inresponse to specific pressure

Rapid bioassessment • 

3. Biodiversity or ecosystem-levelresponse

Sites of special interest • Detailed quantitative, preferably regionallycomparative, investigations of communities

possibly with species-level taxonomicresolution

• Direct and preferably comparative

measurement of the ecosystem process ofconcern

• 

• 

• 

• 

Water quality at sites and on a regionalbasis

• Direct and preferably comparative

measurement of the ecosystem process ofconcern

• Rapid bioassessment (for biodiversity/

conservation status where this has beenshown to correlate well with biodiversity)

• 

Version — October 2000 page 3.2–7
Chapter 3 — Aquatic ecosystems

page 3.2–8 Version — October 2000

As mentioned earlier, methods of prediction and early detection fall into two

categories: 1) sub-lethal organism responses (e.g. growth, reproduction), and 2)

rapid biological assessment (RBA, e.g. AUSRIVAS). The potential of these

methods to meet the objective of early detection is discussed below.

Sub-lethal organism responses

Sub-lethal organism responses can generally be found to meet, in the same

measured response, important attributes (i), (iii), (iv) and (v) above. However, there

will inevitably be conflict and difficulty in meeting all six attributes. For example,

an indicator with good diagnostic value for a particular stressor may not be

particularly applicable to a broad range of stressors. Socially-relevant sub-lethal

organism responses are also often difficult to find. A more significant limitation,

however, is that in very few situations have indicators of exposure to a pollutant

been correlated to environmental effects.

Rapid biological assessment (RBA)

Rapid biological assessment (or RBA) methods are applied and measured in a way

that makes them poorly suited to a role of early detection. In particular, they are not

designed to detect subtle disturbances so may not have desirable attributes (i) and

(iv) above. Nevertheless, unlike other early detection methods, RBA procedures

can be carried out at relatively low cost at a large number of sites or over large

geographical areas, and will generally have greater ecological, regional and social

relevance, i.e. features (ii), (iii), (v) and (vi) above. Indeed, RBA methods such as

AUSRIVAS, in which site data are compared with regionally-relevant reference

conditions, via a predictive model, and reported using a standard index, are

particularly relevant. In their broad coverage they may also be able to locate

problems and stressors that would otherwise pass unnoticed.

Sub-lethal organism responses and RBA methods combine different predictive and

early detection needs, and in comprehensive monitoring programs may play highly

complementary roles. Nevertheless, in a balanced program that measures both early

detection and biodiversity indicators, attributes (i), (iii) and (iv) above are regarded

as the most important guides to the selection of types of indicator.

3. Biodiversity assessment

The biodiversity assessment objective is similar to the broad-scale assessment

objective (1) above because both provide information about the ecological

condition or integrity of a site. Two important features distinguish the two

objectives in practical monitoring programs: the provision of relatively detailed

quantitative and accurate assessments of biodiversity indicators — but at limited

spatial scales, for reasons of high cost; and the provision of less accurate first-pass

assessments of broad-scale indicators — but at greater spatial scales.

Biological indicators used for broad-scale assessment can also be used for

biodiversity assessment. Tradeoffs in costs, the level of accuracy and detail of

information required will ultimately determine which approach is used.

Desired or essential attributes of biodiversity indicator types include features (i) and

(vi) from broad-scale assessment above, as well as either (i) direct measures of

diversity (using species-level identification) and/or (ii) surrogate measures for

biodiversity where a relationship between surrogate and biodiversity has been

shown; and (iii) direct measures of ecosystem function (e.g. community metabolism).
3.2.2 Matching indicators to problems

Version — October 2000 page 3.2–9

Box 3.2.1 A cautionary note on the use of the AUSRIVAS RBA

approach for site-specific assessments

AUSRIVAS, the RBA method using stream macroinvertebrate communities, is at an

intermediate stage of development. It may be limited in its ability to detect minor water

quality disturbances on biota. This restriction is caused by:

• the low level of taxonomic resolution (family level) used in existing state/territory-level

(large-scale) models;

• the use of presence–absence data only;

• the need to factor temporal variability into AUSRIVAS assessments using reference

sites as controls.

In general, stronger inference and greater sensitivity to disturbance become more important

requirements as the spatial scale of a study narrows. Therefore, for specific assessments

conducted at small scales (within a catchment), AUSRIVAS should be conducted using a

sampling design that offers sufficient scope (viz site selection, spatial and temporal

replication) to meet the study requirements. For more reliable assessments at small scales it

may be necessary to combine the data gathered for two seasons (e.g. autumn and spring)

and to enter the data into the ‘combined-seasons’ models developed by many state

agencies. However, some of the RBA’s ‘rapid assessment’ aspect would be lost.

These issues are expanded upon in Chapters 7 and 8.

This bioassessment approach is in a phase of ongoing development and refinement. One

characteristic of that phase is the need to increase the spatial spread and density of

reference sites in various regions in Australia. At present, site numbers and densities may

not be sufficient to allow reliable bioassessment in some regions. (It should be noted that

existing support software for AUSRIVAS models screens out any data collected from sites

outside the geographic region for which the model was derived.)

While the sensitivity of AUSRIVAS for site-specific assessments is being improved,

Guidelines’ users should seek updates on developments in this area to determine whether

the method meets the bioassessment requirements for their particular situation and region.

Such updates, including details of the geographic spread of reference sites, may be

obtained from the AUSRIVAS homepage, http://ausrivas.canberra.edu.au/ausrivas.

One would expect quantitative biodiversity indicators to be restricted in application

to a relatively small region, e.g. a river of interest and sites from rivers in

catchments immediately adjacent. This would be less a limitation for broad-scale

RBA indicators. In monitoring programs, RBA indicators would not normally be

expected to provide direct measures of diversity. Further guidance on whether

RBA or quantitative ‘biodiversity’ indicators (or both) are appropriate for a

particular situation is provided in Section 8.1.1.3 of Volume 2.

3.2.2.2 Matching specific indicators to the problem

These Guidelines discuss several stressors, such as metals, suspended solids and/or

sedimentation, salinity, herbicides and nutrients, any environmental effects of

which can be identified, quantified and assessed by particular biological indicators.

Viable protocols (i.e. proven or near-proven) using diatoms and algae,

macrophytes, macroinvertebrates and fish populations and/or communities,

together with community metabolism, have been developed for use in streams and

rivers, wetlands and lakes, and estuarine and marine ecosystems to monitor and

assess changes associated with these stressors. The stressors (or water quality issues)

and biological indicators recommended to apply to the monitoring and assessment of
Chapter 3 — Aquatic ecosystems

page 3.2–10 Version — October 2000

water quality are listed in table 3.2.2. Background to the development of the

biological indicators, including rationale and justification, is provided in Section

8.1 of the Guidelines.

Development of protocols for the early detection of sediment toxicity using field

assessment procedures is at an early stage in Australia and elsewhere. Until suitable

indicators are identified and protocols for these are developed, a laboratory

assessment approach is recommended (method 2A, table 3.2.2).a For this, a

potentially contaminated sediment from the field is brought back to the laboratory

and standard sediment toxicity tests are conducted to determine its toxicity. A

suitable uncontaminated sediment, collected from an adjacent control site or from the

same site prior to disturbance, is tested as a reference.

3.2.3 Recommended experimental design and analysis procedures for

generic protocols

It is essential that protocols permit rigorous and fair tests of the potential disturbances

under consideration. The best protocols are those that have sufficient baseline data

collected before as well as after a potential disturbance.b There are two advantages of

such protocols. Firstly, the logical basis for inferring whether or not a disturbance has

occurred is stronger because the natural variation inherent in the indicator(s) is

incorporated into the inference; secondly, a properly-designed testing program

permits use of conventional statistical tools to attach formal probability statements to

the observations.c Where such data do not exist or cannot be collected, alternative

analytical procedures can be adopted. These two broad groups of procedures are

outlined here and described in more detail in Section 7.2 (Table 7.2.1D).

Protocols which rely on conventional statistical procedures (Appendix 3, Volume 2)

have two essential features. First, they require that baseline data be collected prior to

the supposed disturbance because seasonal and inter-annual variability in the

indicators need to be accounted for. Second, pre- and post-disturbance data need to

be collected from both the disturbed area and from comparable undisturbed areas.

These control areas provide a benchmark against which changes in the indicator in

the disturbed areas can be judged. With few exceptions, the more control areas that

can be incorporated into the design of the experiment or assessment, the stronger and

fairer will be the test of the effect of the disturbance. The conventional statistical

procedures that are used to analyse these designs belong to the family of general

linear models, which includes univariate and multivariate analysis of variance,

analysis of covariance and regression.

Not all situations permit the implementation of inferentially strong designs.

Appropriate control areas may be limited in number or not available at all. In this

case, statistical methods can be applied to data collected within appropriate designs,

but the strength of the inferences that can be drawn is much weaker and there is a

correspondingly higher risk of either missing a disturbance or erroneously

concluding that a disturbance has occurred. Accordingly these designs should not be

implemented merely as a cost-saving measure; they should only be chosen if

appropriate control areas cannot be found.

a e.g. Method

2A, Appendix 3,

Vol 2

b See Sections

7.2.2 and 7.2.3

c Sections

7.2.2 and 7.2.3
3.2.3 Recommended experimental design and analysis procedures for generic protocols

Version — October 2000 page 3.2–11

Table 3.2.2 Water quality issues and recommended biological indicators for different ecosystem types: S = streams

and rivers, W = wetlands, L = lakes and M = estuarine/marine. Letters or indicator in italics denote that while the

indicator is not presently available, it could be developed relatively quickly with additional resourcing.

Code Issue Suitable biological indicator or assessment

approach

Protocol1 Ecosystem

type

1A, B General inorganic (including

metals) and organic contaminants:

Early detection of short- or longerterm changes from substances in

solution/water column

1A Instream/riverside assays measuring sublethal

‘whole-body’ responses of invertebrate and/or fish

species;

1B Biomarkers (chemical/biochemical changes in

an organism)

Direct toxicity assessment

1A(i), (ii)

1B(i), (ii)

sec 8.3.6

(Vol 2)

S

S, W, L, M

S, W, L, M

2A, B General inorganic (including

metals) and organic contaminants:

Early detection of short- or longerterm changes from substances

deposited (sediments)

2A ‘Whole-sediment’ laboratory toxicity

assessment (where sediment tests are available)

2B Bioaccumulation/biomarkers (for organisms

that feed through ingestion of sediment); other

sublethal incl. behavioural responses where

protocols developed

2A, sec

8.3.6

2B(i), (ii)

S, W, L, M

S, W, L, M

3 General inorganic (including

metals) and organic contaminants:

Changes to biodiversity and/or

ecosystem processes

Structure of macroinvertebrate and/or fish

populations2, 3/communities3 using rapid, broadscale (RBA4) or quantitative (Q) methods

Stream community metabolism

3A(i)–(v)

3B

S, W

S

4 Suspended solids in the water

column

Structure of macroinvertebrate and/or fish

populations2/communities using RBA4 or Q

methods

Seagrass depth distribution

3A(i)–(v)

6

S

M

5 Sedimentation of river bed As for 4 as well as stream community metabolism 3A(i)–(v),

3B

S

6 Effects of organotins Imposex in marine gastropods 9 M

7 Salinity:

Changes to biodiversity

Structure of macroinvertebrate and/or fish

populations2, 3/ communities3 (RBA4 or Q

methods); remote sensing (changes to vegetation

structure);

3A(i)–(v),

5

W, S?

8 Herbicide inputs:

Changes to biodiversity

Structure of phytoplankton or benthic algal

communities; remote sensing (changes to

vegetation structure).

4(i), (ii), 5 W, S

9 Nutrient inputs:

Early detection of short- or longerterm changes from substances

deposited or in solution/water

column

Structure and/or biomass of benthic algal or

phytoplankton communities

Stream community metabolism

4(i)–(iii)

3B

S, W

S

10 Nutrient inputs:

Changes to biodiversity and/or

ecosystem processes

Structure and or biomass of phytoplankton,

benthic algal and/or macroinvertebrate

populations2/communities (Q or RBA4)

Stream community metabolism

3A(i)–(v),

4(i), (ii)

3B

S, W

S

11 Nutrient inputs 11a Seagrass depth distribution

11b Frequency of algal blooms

11c Density of capitellids

11d In-water light climate

11e Filter feeder densities

11f Sediment nutrient status

11g Coral reef trophic status

6

7

8

M

M

M

12 General effluents (non-specific)

and effects of hypoxia

Structure of macroinvertebrate communities

(Q or RBA4)

3A(i), (ii) S, W

13 Broad-scale assessment of

ecosystem ‘health’ (non-specific

degradation)

13A Composition of macroinvertebrate

communities using RBA methods

13B Habitat distributions

13C Assemblage distributions

3A(i), (ii) S, W

M

M

1. The codes listed in this column refer to protocols that are listed by title in Section 8.1.3 of Volume 2. Summary descriptions of these protocols,

with references to important source documents, are provided in Appendix 3, Volume 2. 2. Populations could serve as biodiversity surrogates if a

‘keystone’ role could be established for a species. 3. For pesticides, study of non-target organisms. 4. Cautionary notes on use of RBA methods

for site-specific assessments are provided in various sections of these Guidelines.
Chapter 3 — Aquatic ecosystems

page 3.2–12 Version — October 2000

With some indicators, such as certain highly specific chemical and biochemical

markers, it is possible to use designs that need only limited controls in time or space

or no controls at all. However, there must be conclusive evidence that such indicators

are unequivocally related to the disturbance before such designs are adopted.

For some situations, a disturbance may have occurred and there are no predisturbance data. Alternatively, a development may proceed with insufficient, if

any, baseline data. In these circumstances, the rigour of any inferences about the

disturbance is severely curtailed; the sometimes novel analytical procedures that

have been applied to such data do not compensate for the lack of pre-disturbance

data.a Where multiple control areas are available, they can be used to describe how

atypical the potentially disturbed areas appear.b These procedures require the user

to assume that the indicator responded similarly in control and disturbance areas

before the disturbance. Where multiple control areas are not available, questions

are often framed around the extent of the disturbance. As discussed below,c under

these circumstances it is best that data be collected from a comparatively larger

number of disturbance sites than would otherwise be gathered (e.g. along a mixing

zone gradient), so that stronger inferences may be drawn about disturbance by way

of disturbance gradients. Such additional data may also enhance predictive

capabilities of monitoring programs.

For all these procedures it is necessary to collect and collate exploratory data. The

aim is to define the spatial and temporal extent of sampling and to identify and

choose sampling locations within the control and disturbance areas.d Such

exercises can include use of simulation or other predictive tools to model currents or

sediment movements, and/or be new or pre-existing data on the flora or fauna. It is

difficult to prescribe protocols for exploratory collections because the amount of preexisting data or auxiliary models will vary from case to case. In novel or unfamiliar

situations such exploratory collections are even more desirable and could lead to

substantial savings in time and costs.

Table 3.2.3 summarises the designs that apply to the protocols listed in table 3.2.2.

The BACI class of design uses conventional statistical procedures while designs

using alternative analytical procedures must be applied if inference is based on

temporal change only or spatial pattern alone.

Preferred designs using conventional statistical procedures involve both predisturbance baseline data and multiple control areas (MBACI and ‘Beyond-BACI’

designs of table 3.2.3). Where pre-disturbance baseline data are available or can be

collected, but only a single control site can be found, BACIP designs are appropriate.

Designs where the length of pre-disturbance baseline and/or the number of control

areas are reduced (e.g. BACI) have less inferential rigour because more assumptions

need to be made about the similarity of the behaviour of the indicator in control and

disturbance areas prior to the onset of the potential disturbance.

It is important to consider using any descriptive and exploratory analytical tools that

would enhance interpretation of the analytical procedures employed. These might

include graphs and plots accompanying univariate and multivariate approaches, clear

tabulations of relevant descriptive statistics in univariate analyses (e.g. means and

confidence intervals), and ordination and classification of data in multivariate

studies.e Some of the specific requirements of biological indicators that need to be

considered while designing the monitoring program are detailed in Section 7.3.

a & b See

Sections 7.2.2

& 7.2.3

c Section

3.2.4.2/4

& 7.2.2

d Section

7.2.3.2

e Sections 7.2,

7.3 and the

Monitoring

Guidelines Ch.6
3.2.4 Guidelines for determining an unacceptable level of change

Version — October 2000 page 3.2–13

Table 3.2.3 Experimental design and analysis procedures to apply to generic protocols. The letters used to

identify the broad categories of design are those used in figure 7.2.1. Explanations of the possible designs and

references are supplied in Section 7.2.3. Letters and numerals in the protocol column correspond to those used

in Table 3.2.2 and Section 8.1.3 (Volume 2).

Broad category of

design (from

Section 7.2.2)

Possible designs

(Described in table 7.2.1)

Protocol (from Section 8.1.3, Vol 2)

MBACI

Modifications (e.g. MBACIP, inclusion of covariates)

All protocols wherever possible

Any protocol if applicable

‘Beyond BACI’ designs Any protocol if applicable.

BACIP (single control site)

Modifications to BACIP

1A, 1B

1A, 1B

A. Inference based

on the BACI

(Before, After,

Control, Impact)

family of designs

Simple BACI 1B

Intervention analysis 1B, 2B, 3B, 4, 6, 7, 8. Possibly 3A(ii) but

may prove very expensive; behaviour of

3A(i) in face of temporal variations

unknown and not recommended for this

protocol

Trend analysis 1B, 2B, 3B, 4, 6, 7, 8. Possibly 3A(ii) but

may prove very expensive; behaviour of

3A(i) in face of temporal variations

unknown and not recommended for this

protocol

B. Inference based

on temporal

change alone

A posteriori sampling Possibly 1B, 2B, but only if chemical or

toxicant is unequivocally related to the

effluent

Conventional statistical designs (e.g. ANOVA,

ANCOVA)

Any protocol based on univariate indicator

e.g. 1B, 2B, 3B, 4(i)A, 4(ii), 4(iii)A, 6, 8, 9.

Analysis of ‘disturbance gradients’ Any protocol if applicable; may be too

cumbersome for 1A

D. Inference based

on spatial pattern

alone

Predictive models based on spatial controls only 3A(i), 3A(ii)

3.2.4 Guidelines for determining an unacceptable level of change

3.2.4.1 Inferences, assessment of change, setting decision criteria

A priori decisions made between stakeholders (e.g. developer and regulator) about

effect size and the probability of making a Type I error (α) and Type II error (β)

(generally only ‘effect size’ needs to be decided upon for RBA) are an essential

aspect of the guidelines philosophy.a These decision criteria should be preestablished in the following four scenarios: for flexible decision-making; for

compliance assessment; when there are multiple lines of evidence; and when data

are to be assessed against predictive models.

1. Flexible decisions in the spirit of cooperative best practice

Flexible decisions are important where adherence to a precautionary approach has

been agreed or stipulated by a regulatory authority or dictated by legislation.

Adequate baseline data should be collected according to the design criteria discussed

above, given any unavoidable constraints. Integral to design considerations is the

principle that monitoring should provide a strong basis for management response

(through decisions and/or action) to any early indications of adverse disturbances.

The decisions about the criteria and about responsive action by management should

a See sections

2.2.1.2, 3.1.7,

7.2.3.3
Chapter 3 — Aquatic ecosystems

page 3.2–14 Version — October 2000

be made a priori, especially where a superficially positive response might result from

the early stages of an abnormal, and therefore undesired, change in environmental

conditions; e.g. increased taxonomic richness accompanying a slight increase in

eutrophication. Management intervention will depend on the management

objective(s) for the receiving waters, but two approaches are possible.

i. Management could make ‘super-precautionary’ responses, dictated by any

statistically significant trend from baseline of a magnitude agreed a priori to

be important. The probability criteria for statistical significance would be

determined under the flexible decision regime proposed by Mapstone (1995,

1996), with the result that α and β would be variable and determined from

time to time on the basis of the available data and the critical effect size

agreed a priori. The emphasis is on setting values for critical effect sizes that

would be expected to trigger an early management response to a potential

disturbance. It is assumed that it is more important to react quickly to

potential problems, even though the response would be to something which

had not yet become a major ecological threat. Such a position would be

appropriate for activities in particularly sensitive or valuable areas. The

precision with which one could specify the location of the baseline reference

point would depend on the amount of sampling during the baseline period.

Increasing the precision with which the reference point is specified, which

would presumably also mean increasing the precision of sampling after the

start of a development, would reduce the risk of responding to an erroneous

trigger caused by early indications of a shift from baseline conditions. Thus,

it becomes to everyone’s advantage to seek thorough monitoring.

ii. Management response could be triggered by ongoing feedback or a continuously

monitored variable exceeding some threshold value. Control charting techniques

such as those used in quality assurance/quality control programs might be

employed here. The trigger value for a particular variable might represent a level

at which that variable is known to have important biological consequences, or

might simply be a statistical parameter used to indicate that an observed event

would be considered an outlier under normal circumstances and therefore is

worthy of further investigation. As in (i) above, it is important that all parties have

agreed a priori to intervene when that trigger occurs.

2. Compliance, legal framework: data gathered under strict and rigorous hypothesis-testing

framework

In this case, the criteria to which sampling programs are designed are set

independently of the particular activity being monitored. Such criteria would not

normally be subject to negotiations between regulators and proponents or other

interested parties. These external criteria are the reference points that, if exceeded,

will trigger action. In these cases, negotiations between regulators, interest groups,

and proponents focus on the degree of risk involved in either failing to confidently

recognise that the standard has been violated (β) or that apparent violations will be

flagged in error (α). As in (i) from Section 3.2.4.1/1 above, the thoroughness of

sampling design will directly influence the likelihood of erroneous decisions.
3.2.4 Guidelines for determining an unacceptable level of change

Version — October 2000 page 3.2–15

3. Data gathered from multiple lines of evidence, where statistical power for each indicator may be

poor (lack of adequate temporal baseline)

For situations where there is a paucity of baseline information and/or adequate

spatial controls, it is recommended that users adopt a ‘weight-of-evidence’

approach (Suter 1996) to inference. The process is based on risk assessment

principles and draws on epidemiological precepts in interpreting test results; the

concept in various forms has been described by Hodson (1990), Stewart-Oaten

(1993) and Suter (1996), amongst others, with examples. There is an onus on those

conducting monitoring programs under these situations to enhance the set of

monitoring techniques used: it should include chemical monitoring, spatial

gradients for a number of biological monitoring protocols,a and toxicological and

other experimental data in which concordance is sought between field results and

controlled experimental findings. In this way, lack of baseline information may be

at least partially compensated for, so that conclusions can be confidently drawn

and, importantly, agreed upon by all parties.

4. Data assessed against bands of AUSRIVAS predictive models

Two complementary indices summarise the outputs from the analysis of

AUSRIVAS data:

i. O/E Family — the ratio of the number of families of macroinvertebrates at a site

to the number of families expected (predicted) at that site. (The expected number

of families is actually the sum of the probabilities of each taxon occurring at the

site as calculated from the model.)

ii. O/E SIGNAL which is the ratio of the observed SIGNAL10 value for a site to the

expected SIGNAL value. SIGNAL assigns a grade to each family based on its

sensitivity to pollution. The sum of the grades is divided by the number of

families involved to give an average grade for the site. A grade of 10 represents

high sensitivity to pollution, while a grade of 1 represents high tolerance of

pollution.

The values of both indices can range from a minimum of 0 (indicating that none of

the families expected at a site were actually found at that site) to a theoretical

maximum of 1.0, indicating a perfect match between the families expected and those

that were found. In practice, the maximum can exceed 1.0 indicating that more

families were found at that site than were predicted by the model. This can indicate

an unusually diverse site, but could also indicate mild enrichment by organic

pollution where the added nutrients have allowed families not normally found in that

site to establish. Conversely, an undisturbed, high-quality site may score an index

value less than 1.0 because of chance exclusions of families during sampling.

For reporting, the value of each index is divided into categories or bands. The width

of the bands is based on the distribution of index values for the reference sites in a

particular model. The width of the reference band, labelled ‘A’ in table 3.2.4, is

centred on the value 1.0 and includes the central 80% of the reference sites. Any site

with index within the 10% and 90% bounds around 1.0 is allocated to band A and is

described as being of ‘reference condition’. A site with an index value exceeding the

upper bound of these values (i.e. the index value is greater than the 90th percentile of

 

10 SIGNAL is a biotic index, Stream Invertebrate Grade Number — Average Level; see Section

8.1.2.1 and Chessman (1995).

a See Section

7.2.1.2
Chapter 3 — Aquatic ecosystems

page 3.2–16 Version — October 2000

the reference sites) is judged to be richer than the reference condition, and is

allocated to ‘band X’. A site whose index value falls below the lower bound (i.e. the

index value is smaller than the 10th percentile of the reference sites) is judged to

have fewer families and/or a lower SIGNAL score than expected and is allocated to

one of the lower bands according to its value. The widths of bands B and C are the

same as the width of band A, the reference band. The band D may be narrower than

these, depending on variability in the index values of the reference sites in the model.

In most cases, sites falling in band D on either index are severely depleted in terms of

the number of families expected.

In many cases the values of the indices will allocate a site to the same band. In

situations where the two indices differ in band allocation, the site will be allocated

to lower of the two bands if the index value is below reference condition, or to the

above reference band if one of the indices places the site in band X.

These factors should be taken into consideration by stakeholders and management

who are setting situation-specific guidelines.

Table 3.2.4 Division of AUSRIVAS O/E indices into bands or categories for reporting. The

names of the bands refer to the relationship of the index value to the reference condition

(band A). For each index, the verbal interpretation of the band is stated first, followed by

likely causes (dot-points).

Band

label

Band name Comments

O/E Families O/E SIGNAL

X Richer than

reference

More families found than

expected.

• Potential biodiversity ‘hotspot’

• Mild organic enrichment

Greater SIGNAL value than

expected.

• Potential biodiversity ‘hotspot’

• Differential loss of pollutiontolerant taxa (potential

disturbance unrelated to

water quality)

A Reference Index value within range of

central 80% of reference sites

Index value within range of central

80% of reference sites

B Below reference Fewer families than expected

• Potential disturbance either

to water quality or habitat

quality or both resulting in a

loss of families

Lower SIGNAL value than

expected

• Differential loss of pollutionsensitive families

• Potential disturbance to water

quality

C Well below

reference

Many fewer families than

expected

• Loss of families due to

substantial disturbance to

water and/or habitat quality

Much lower SIGNAL value than

expected

• Most expected families that

are sensitive to pollution have

been lost

• Substantial disturbance to

water quality

D Impoverished Few of the expected families

remain

• Severe disturbance

Very low SIGNAL value

• Only hardy, pollution-tolerant

families remain
3.2.4 Guidelines for determining an unacceptable level of change

Version — October 2000 page 3.2–17

It should be noted that the calculation of indices and allocation to a band for a

stream site are automatically performed as part of the AUSRIVAS procedure by

the AUSRIVAS software package. This software, downloaded over the internet

(website address: http://ausrivas.canberra.edu.au/ausrivas) performs all calculations

required for performing an RBA AUSRIVAS bioassessment of a site’s

macroinvertebrate community. Further documentation is provided via the

AUSRIVAS homepage, as well as additional aids in diagnosing the disturbance at

a site, depending upon the band in which it falls.

3.2.4.2 Situation-dependent guidelines

The following subsections provide guidelines for protection of each of the three

ecosystem conditions listed in Section 3.1, i.e. condition 1 ecosystems, of high

conservation/ecological value; condition 2, slightly to moderately disturbed systems;

and condition 3, highly disturbed systems. For condition 1 and condition 2

ecosystems, management involves tracking the intrinsic attributes of the ecosystems

(the key structural and functional components) to ensure they do not deviate outside

natural variability as determined from baseline knowledge or accruing knowledge.

For any of the ecosystem conditions, local jurisdictions could negotiate site-specific

guidelines alternative to those recommended below after considering site-specific

factors.a (Elsewhere, the Guidelines recommend the type and number of indicators

that should be incorporated in an environmental monitoring and assessment

program, depending upon the situation.b)

1. Sites of high conservation value (condition 1 ecosystems)

For most applications using bioindicators in Australia, there is insufficient

information about ecosystems upon which to make informed judgments about an

acceptable level of change. All stakeholders (e.g. developer and regulator) are

strongly encouraged to adopt the following strategy towards determining

appropriate guidelines for indicator responses: first, for collecting baseline data;

then, detecting and assessing environmental impacts.

Baseline data collection

Using an appropriate statistical design for the indicator response as prescribed in the

protocols,c parties should ensure an ‘adequate’ baseline is gathered for the indicators

measured. This may be achieved by setting ‘conservative’ α, β and effect size, where

the effect size is determined on the basis of statistical or other criteria. In the absence

of clear information from which to set decision criteria, it is recommended default

targets for ecologically conservative decisions be set at α = 0.1, β = 0.2 (power of

0.8) and effect size = 10% of, or 1 SD about, the baseline mean, whichever is

smaller. Whether these defaults are applied or not, the importance of sound and

numerous baseline data cannot be over-emphasised. It is strongly recommended that

baseline data be gathered from at least 3−5 control or reference locations (for

biodiversity indicators at least) over a period of at least three years (all indicators)

wherever possible. (See case study presented in Appendix 4, Vol 2, and Section 7.2

for rationale, justification and further discussion.) Guidelines are provided below for

those situations in which it is not possible to meet these baseline requirements.d

The default guidelines for α, β, and effect size, from above, should not be simply

accepted as a new convention (or dogma), but should be seen as the starting point for

considering (and negotiating) what is appropriate or reasonable for each case. The

setting of effect size should be an active and explicit decision, usually made on a

a See section

3.1.3.3

b Section 7.2.1

c App. 3, Vol. 2

for protocols

d Section

3.2.4.2/4
Chapter 3 — Aquatic ecosystems

page 3.2–18 Version — October 2000

case-by-case basis. Mapstone (1995, 1996), for example, provides additional case

studies describing the setting of statistical decision criteria. For some situations an

effect size as small as 10% is achievable and deemed necessary.a For many others of

the variables typically encountered in environmental work, it will be very difficult to

detect changes of 10% or less about some mean, and perhaps impossible. In some

cases, changes of 10% might be inconsequential, even in terms of an early warning

system. Seeking to enforce monitoring to arbitrary decision criteria under such

circumstances could result in a strong backlash against the principle of setting

decision criteria a priori. However, relaxation of precautionary values should always

be a clearly argued and thoroughly justified step. If insufficient information exists to

justify such changes but nominated monitoring variables cannot be sampled

rigorously enough to satisfy default criteria, then other candidate variables should be

investigated as the mainstays for inferential decisions.

It is not always sensible to set an effect size of 10% (or some other value) of the

time-averaged baseline mean. In some cases it may be necessary to stipulate an

effect size that reflects the dynamics of the control sites and how they are related to

the disturbance site during baseline monitoring. For example, say the measurement

variable has a seasonal periodicity but the future disturbance site and control sites

show different responses to seasonality. Then it would be necessary to model that

knowledge into the effect size. At its simplest, this might mean having different

effect sizes for tests in summer and winter.

The baseline data referred to above are for use in determining if change has

occurred. Much of the information used for environmental impact assessments

(EIAs) is required for ecosystem characterisation and impact prediction and whilst

not ‘baseline’ in the statistically rigorous sense described above, should be

adequate as pilot data to design monitoring programs used for impact detection.

Once an environmental impact statement (EIS) is accepted and a development

proposal is approved, either development should be delayed, or there should be a

guarantee that no disturbance to aquatic ecosystems would occur, until adequate

baseline are gathered. (Humphrey et al. (1999) are critical of aspects of the EIA

process in Australia at least, in that too often developments proceed without

adequate baseline data gathered to detect and assess potential disturbances.)

Detecting and assessing disturbances

The guidelines for detecting and assessing environmental impacts or disturbances

are determined from a priori decisions made between all parties.b In the case of

flexible decision-making in the spirit of cooperative best practice, intervention can

be either (i) ‘super-precautionary’, sought once any apparent trend away from a

baseline appears, or (ii) sought once a feedback ‘trigger’ or threshold has been

reached. In the first of these two situations, management action may or may not be

required when a ‘positive’ response is detected. The proponent/discharger may also

wish to corroborate the results for an indicator with water chemistry data and data

obtained for other biological indicators.

Alternatively, data may be being gathered for compliance assessment within a

legal framework, under strict and rigorous hypothesis-testing. Here, using the

default settings from (i) above, unless all parties have determined other values a

priori, an unacceptable disturbance has occurred if P < 0.1 in the statistical test

applied to the data.

a See App. 4,

Vol. 2

b Section

3.2.4.1
3.2.4 Guidelines for determining an unacceptable level of change

Version — October 2000 page 3.2–19

It is strongly recommended that parties adopt a precautionary approach and

respond wisely and in a timely manner to data gathered for ‘early detection’

indicators.

2. Slightly to moderately disturbed systems (condition 2 ecosystems)

Treat condition 2 ecosystems like condition 1 ecosystemsa acknowledging that

there may be negotiated deviations from default values prescribed for condition 1

ecosystems. Nevertheless, any decisions on effect size should be based on sound

ecological principles of sustainability rather than arbitrary relaxation of the default

values described above, or because of resource constraints.

3. Highly disturbed systems (condition 3 ecosystems)

The philosophy of the Guidelines for these systems is that at worst, water quality is

maintained. Ideally, the longer-term aim is towards improved water quality.

Normally, early detection indicators of sublethal toxicity would not be measured at

these sites.b For these sites, any decisions on effect size can be arbitrary relaxations

of the default values described above, although they should still be based on sound

ecological principles of sustainability. Guidelines from 3.2.4.2/5 below should be

applied for cases in which a rapid, broad-scale biodiversity indicator has been

selected. Where rapid assessment methods are applied to small-scale problems

(within a catchment), assessment of results must take into account the general

inability of the methods to detect all but large water quality problems. Approaches

recommended to enhance the general sensitivity of the methods are discussed in

box 3.2.1 and in Section 7.3.3.

4. Sites where an insufficient baseline sampling period is available to meet key default guideline

decision criteria

To compensate for an inability to gather sufficient baseline data, the Guidelines

recommend that additional monitoring be carried out, including a greater number

of indicators and/or sites for ‘early detection’ and biodiversity measurement (i.e.

the ‘multiple lines of evidence’ conceptc). Of course, resource constraints will limit

the number of additional indicators and sites that can be monitored, but these

resource constraints must be satisfactorily balanced with the need for unambiguous

and meaningful results.

For a development that is in the planning stage, if there are inadequate baseline

data against which to assess disturbance, it is recommended that data from all

monitoring programs be submitted to an independent expert (or panel of experts)

on a regular basis for assessment of acceptability. The same ethos of precaution

and ecological sustainability, as applied to guidelines in other situations listed here,

would influence the decisions made by the experts.

For existing developments for which adequate baseline data were never gathered, the

project approval phase probably pre-dated the more stringent discharge licensing

conditions that have subsequently been imposed by regulators. Apply the same

procedures as for (i) from above.

For a posteriori monitoring of accidental discharges, continue monitoring until

target indicator goals have been reached, as determined by an independent expert

(or panel of experts). d

a See Section

3.2.4.2/1

b Section

7.2.1.1/3

c Section

3.2.4.1

d Section 3.2.5
Chapter 3 — Aquatic ecosystems

page 3.2–20 Version — October 2000

5. Broad-scale assessment of ecosystem health

Broad-scale assessments of ecosystem health are used to assess water quality for

planning purposes, to set goals for remediation and rehabilitation programs, and to

monitor and assess broad-scale disturbances such as diffuse pollution.

If a site is found to be below reference condition on the AUSRIVAS banding scheme

(band B or lower), then it can be concluded that fewer invertebrate taxa have been

found than would be expected on the basis of the particular AUSRIVAS model. A

goal of subsequent management should be to improve the water and habitat quality

so as to move the site indices closer to reference conditions or into band A.

If a site is found to be above reference condition on the AUSRIVAS banding

scheme (band X), then further investigations are needed. The site may be naturally

more diverse than surrounding reference sites, and therefore warrants special

management to conserve that diversity. Alternatively, a naturally nutrient-poor site

has received organic or nutrient enrichment with successful establishment of

families of macroinvertebrates that would ordinarily not inhabit this site.a

3.2.5 Assessing the success of remedial actions

For aquatic ecosystems long degraded by human disturbances in Australia and

New Zealand, biological monitoring will be required to assess the success of

remedial works put in place to improve water quality and ecological condition. The

goals for remediation might be either restoration or rehabilitation. Restoration

refers to attempts to restore an ecosystem to its configuration prior to the

disturbance or disturbance. Rehabilitation refers to attempts to improve the

ecological status of some attributes of a disturbed ecosystem. The expected

management target would be improvement in the ecological condition or integrity

of a site (or sites) and specific biodiversity indicators could be selected for the

water quality problem identified.b

Invariably in these situations, there are no pre-disturbance data available to define a

target ecological condition, and because of this the scope for applying formal

statistical methods of inference is reduced.c The ecological target should then be

assumed to resemble that of appropriate control locations, where these are

available. The assumption being made in this process is that the indicator

responded similarly in the control and disturbance areas before the disturbance.

Simple hypotheses may be generated for these cases that test for likely indications

of improvement. In all likelihood, there are too few data and too many

uncertainties for formal statistical decision criteriad to be applied. Rather,

monitoring is continued until target indicator goals have been reached. Expert

panels can decide upon the goals and, if necessary, decide whether compliance has

been achieved. In determining goals for rehabilitation or restoration, stakeholders

and their consultants need to take into consideration the desired target ecosystem

condition e as well as experience elsewhere in achieving biological recovery for the

types of contaminants involved. f

d Section 3.2.4

a See Section

3.2.4.1/4

b Section

3.2.2.2

c Sections

7.2.1.2 and 7.2

e Section 3.1.3

f Sections

7.2.2 and 7.2.3
Version — October 2000 page 3.3–1

3.3 Physical and chemical stressors

3.3.1 Introduction

A number of naturally-occurring physical and chemical stressors can cause serious

degradation of aquatic ecosystems when ambient values are too high and/or too

low. In this section, the following physical and chemical stressors are considered:

nutrients, biodegradable organic matter, dissolved oxygen, turbidity, suspended

particulate matter (SPM), temperature, salinity, pH and changes in flow regime.

Other chemical stressors, such as ammonia, cyanide, heavy metals, biocides and

other toxic organic compounds, are covered in Sections 3.4 and 3.5.

Recommendations relating to the development of guidelines for the stressors not

covered in these Guidelines (e.g. introduced species and habitat modifications) are

contained in Section 8.5.2 of Volume 2.

The purpose of the guidelines provided in this section is to assist those involved in

managing water resources to ensure that condition 2 (slightly to moderately

disturbed) and condition 3 (highly disturbed) aquatic ecosystems are adequately

protected. For ecosystems requiring the highest level of protection (condition 1), the

objective of water quality management is to ensure that there is no detectable change

(beyond natural variability) in the levels of the physical and chemical stressors.a For

such highly-valued ecosystems, the statistical decision criteria for detecting any

change should be ecologically conservative and based on sound ecological

principles. This position should only be relaxed where there is considerable

biological assessment data showing that such changes will not affect biological

diversity in the system.b

Figure 3.3.1 is a flow chart of the steps involved in the detailed application of the

guidelines for the physical and chemical stressors using risk-based ‘guideline

packages’.

The steps consist of selecting key stressors, then guideline trigger values, and then,

where appropriate, a protocol for considering the effect of ecosystem-specific

modifiers in reducing the biological effects of individual stressors. The steps are

discussed in detail in this section.

The new approach for physical and chemical stressors recommended here differs

from that in the 1992 ANZECC Water Quality Guidelines (ANZECC 1992) in a

number of ways, the most significant being that:

• the guidelines are as specific as possible to each ecosystem. While not all of

the required information is available yet, a start has been made by increasing

the number of ecosystem types from two in the 1992 ANZECC Guidelines to

six in these Guidelines.c

• the focus here is on providing issue-based information, aimed at protecting

aquatic ecosystems from eight issues or problems caused by physical and

chemical stressors.d

• available biological effects data have been used to determine low-risk guideline

trigger values for toxic stressors for each ecosystem-type where sufficient data

exist — i.e. a risk-based approach. For non-toxic stressors, low-risk guideline

trigger values for key performance indicators have been determined by

comparison with suitable reference ecosystems.e

c Section 3.1.2

a Section 3.1.3

b Section

3.1.3.2

d Section

3.3.2.2

e Section

3.3.2.1
Chapter 3 — Aquatic ecosystems

page 3.3–2 Version — October 2000

• for each issue, the Guidelines give guideline packages (which are also riskbased) rather than simplistic threshold numbers for single indicators. These

packages consist of key performance indicators, guideline trigger values and,

where appropriate, a protocol for considering the effect of ecosystem-specific

modifiers in reducing the biological effects. The packages help managers

estimate whether low, possible or high risk exists at their sites as well as

providing them with a means of refining guideline trigger values. The steps

involved in applying the guideline packages are summarised in figure 3.3.1.

• guidelines for each issue are generally specified as concentrations, although it is

recommended that load-based guidelines be developed for nutrients,

biodegradable organic matter and suspended particulate matter.

The remainder of this section is divided into two parts: Section 3.3.2 outlines the

philosophy adopted in developing guidelines for physical and chemical stressors,

while Section 3.3.3 covers the detailed guideline packages for each of the eight

issues considered.

Low riskb High risk

(initiate remedial actions)

Low riskb

Test against guideline values

Compare key performance indicators with guideline ‘trigger’ 

values for specific ecosystem type

Further site-specific investigations:

• Consider effects of ecosystem-specific modifying factors

• Comparison with reference condition

• Biological effects data (e.g. direct toxicity assessment)

Define primary management aims (fig 3.1.1)

Decision framework for 

applying the trigger valuesa

Determine appropriate guideline trigger values

for selected indicators (fig 3.1.1)

Potential riskc

a Local biological effects data and some types of reference data (section 3.1.5) generally not required in the decision trees

b Possible refinement of trigger value after regular monitoring (section 3.1.5)

c Further investigations are not mandatory; users may opt to proceed to management/remedial action

Figure 3.3.1 Decision tree framework (‘guideline packages’) for assessing

the physico-chemical stressors in ambient waters
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–3

3.3.2 Philosophy used in developing guidelines for physical and chemical

stressors

3.3.2.1 Types of physical and chemical stressors

Physical and chemical stressors can be classified broadly into two types (fig 3.3.2)

depending on whether they have direct or indirect effects on the ecosystem.

Direct effects

Two types of physical and chemical stressors that directly affect aquatic ecosystems

can be distinguished: those that are directly toxic to biota, and those that, while not

directly toxic, can result in adverse changes to the ecosystem (e.g. to its biological

diversity or its usefulness to humans). Excessive amounts of direct-effect stressors

cause problems, but some of the elements and compounds covered here are essential

at low concentrations for the effective functioning of the biota — nutrients such as

phosphorus and nitrogen, and heavy metals such as copper and zinc, for example.

Types of physical and chemical

stressors

Stressors directly

toxic to biota

e.g.

• heavy metals

• ammonia

• salinity

• pH

• DO

• temperature

Stressors that are 

not toxic but can 

directly affect 

ecosystems & biota

e.g. 

• nutrients

• turbidity

• flow

• alien species

Stressors (or factors) that

can modify effects of other

stressors

e.g.

• pH — release metals

• DOC, SPM — complex metals and

 reduce toxicity

• temperature — increase physiological 

 rates

• DO — change redox conditions and 

 release P

Direct effect Indirect effect

Figure 3.3.2 Types of physical and chemical stressors

The trigger values of toxic stressors are generally determined from laboratory

ecotoxicity tests conducted on a range of sensitive aquatic plant and animal

species.a However, salinity, pH and temperature are three toxic direct-effect

stressors that are naturally very variable among and within ecosystem types and

seasonally, and natural biological communities are adapted to the site-specific

conditions. This suggests that trigger values for these three stressors may need to be

based on site-specific biological effects data.

Examples of non-toxic direct-effect stressors include:

• nutrients, that can result in excessive algal growth and cyanobacterial blooms;

• suspended particulate matter, that can reduce light penetration into a waterbody

and result in reduced primary production, possible deleterious effects on

a Section 3.4
Chapter 3 — Aquatic ecosystems

page 3.3–4 Version — October 2000

phytoplankton, macrophytes and seagrasses, or smother benthic organisms and

their habitats;

• organic matter decay processes, that can significantly reduce the dissolved

oxygen concentration and cause death of aquatic organisms, particularly fish;

• water flow, which can significantly affect the amount and type of habitats

present in a river or stream.

Indirect effects

Indirect stressors (or factors) are those that, while not directly affecting the biota,

can affect other stressors making them more or less toxic. For example, dissolved

oxygen can influence redox conditions and influence the uptake or release of

nutrients by sediments. Equally, pH, dissolved organic carbon (DOC) and

suspended particulate matter can have a major effect on the bioavailable

concentrations of most heavy metals.

Through the risk-based decision trees,a managers will consider these indirect

stressors, with ecosystem-specific modifying factors, during the assessment of each

issue. Although many effects of these modifying factors are reasonably well known

from a theoretical viewpoint, there are few quantitative relationships (or models)

that allow them to be used to develop more ecosystem-specific guidelines (Schnoor

1996). Recommendations made in Section 8.5.2 (Volume 2) cover the type of

research and development needed to develop these relationships.b

For both types of physical and chemical stressors (eliciting direct or indirect effects

on the ecosystem) background information is provided in Section 8.2.1 by way of

Fact Sheets.c Key indicators provided in the Fact Sheets are nutrients, dissolved

oxygen, turbidity and suspended particulate matter, salinity, temperature, optical

properties, environmental flows and hydrodynamics.

3.3.2.2 Issues affecting aquatic ecosystems that are controlled by the physical and chemical

stressors

Many aquatic ecosystems experience a range of problems that affect biodiversity or

ecological health. These problems mostly result from human activities.

This section focuses on the development of guideline ‘packages’ to address the

specific issuesd (summarised in table 3.3.1) likely to result from physical and

chemical stressors:

• nuisance growth of aquatic plants (eutrophication);

• lack of dissolved oxygen (DO; asphyxiation of respiring organisms);

• excess suspended particulate matter (SPM; smothering of benthic organisms,

inhibition of primary production);

• unnatural change in salinity (change in biological diversity);

• unnatural change in temperature (change in biological diversity);

• unnatural change in pH (change in biological diversity);

• poor optical properties of waterbodies (reduction in photosynthesis; change in

predator–prey relationships);

• unnatural flow (inhibition of migration; associated temperature modification of

spawning; changes in estuarine productivity).

d See Sections

3.3.3, 8.2.3

a See Section

3.1.5

b Section 8.5.2

(Volume 2)

c Section 8.2.1
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–5

Table 3.3.1 Summary of the condition indicators, performance indicators, and location of default trigger value

tables, for each issue

Issue Condition

indicator/target

Performance

indicators

Preferred method

for obtaining

trigger values a

Default trigger

value for each

ecosystem-type

Consider

ecosystemspecific

modifiers

1. Nuisance aquatic

plants

Species composition

Cell numbers

Chlorophyll a conc

TP conc

TN conc

Chl a conc

Reference data

Reference data

Reference data

Tables 3.3.2,

3.3.4, 3.3.6, 3.3.8,

3.3.10

Yes — Section

3.3.3.1

2. Lack of DO Reduced DO conc

Species composition/

abundance

DO conc Reference data Tables 3.3.2,

3.3.4, 3.3.6, 3.3.8,

3.3.10

Yes — Section

3.3.3.2

3. Excess of SPM Species composition/

abundance

SPM conc Reference data Tables 3.3.3,

3.3.5, 3.3.7, 3.3.9,

3.3.11

Yes — Section

8.2.3.2

4. Unnatural change

in salinity

Species composition/

abundance

EC (salinity) Reference data Tables 3.3.3,

3.3.5, 3.3.7, 3.3.9,

3.3.11

No

5. Unnatural change

in temperature

Species composition/

abundance

Temperature Reference data > 80%ile

< 20%ile

No

6. Unnatural change

in pH

Species composition/

abundance

pH Reference data Tables 3.3.2,

3.3.4, 3.3.6, 3.3.8,

3.3.10

No

7. Poor optical

properties

Species composition/

abundance

Turbidity

Light regime

Reference data

Reference data

Tables 3.3.3,

3.3.5, 3.3.7, 3.3.9,

3.3.11

No

8. Unnatural flow

regime

Species composition/

abundance

Habitat change

% wetted area

Flow regime

a

 Where local biological and ecological effects data are unavailable.

3.3.2.3 Defining low-risk guideline trigger values

The guideline trigger values are the concentrations (or loads) of the key

performance indicators, below which there is a low risk that adverse biological

effects will occur. The physical and chemical trigger values are not designed to be

used as ‘magic numbers’ or threshold values at which an environmental problem is

inferred if they are exceeded. Rather they are designed to be used in conjunction

with professional judgement, to provide an initial assessment of the state of a water

body regarding the issue in question. They are the values that trigger two possible

responses. The first response, to continue monitoring, occurs if the test site value is

less than the trigger value, showing that there is a ‘low risk’ that a problem exists.

The alternative response, management/remedial action or further site-specific

investigations, occurs if the trigger value is exceeded — i.e. a ‘potential risk’

exists.a The aim with further site-specific investigations is to determine whether or

not there is an actual problem. Where, after continuous monitoring, with or without

site-specific investigations, indicator values at sites are assessed as ‘low risk’ (no

potential impact), guideline trigger values may be refined.b The guidelines have

attempted as far as possible to make the trigger values specific for each of the

different ecosystem types.

Four sources of information are available for use when deriving low-risk trigger

values: biological and ecological effects data, reference system data, predictive

a See figure

3.3.1

b Section 3.1.5
Chapter 3 — Aquatic ecosystems

page 3.3–6 Version — October 2000

modelling, or professional judgment.a The guidelines for physical and chemical

stressors promote and focus principally on the derivation of low-risk trigger values,

from biological and ecological effects data and through the use of reference data.

Ecosystem condition

As already mentioned, the Guidelines recognise three levels of ecosystem condition

(1) high conservation/ecological value (condition 1 ecosystems), (2) slightly or

moderately disturbed (condition 2 ecosystems), and (3) highly disturbed (condition 3

ecosystems), each with an associated level of protection (table 3.1.2). For condition 1

ecosystems, the Guidelines advise that there should be no change from ambient

conditions, unless it can be demonstrated that such change will not compromise the

maintenance of biological diversity in the system. Where comprehensive biological

effects data are unavailable, a monitoring program is required to show that values of

physical and chemical stressors are not changing, using statistically conservative

decision criteria as the basis for evaluation.b Values of the criteria as recommended

for biological indicators might be used as a starting point in negotiations;c further

discussion of statistical error rates relevant to detecting change in physical and

chemical stressors is provided in Section 7.4.4.1.d

Box 3.3.1. Sources of information for use when deriving low-risk

trigger values

a) biological and ecological effects data — obtained either from biological effects testing

using local biota and local waters (e.g. information derived by eriss for water release

standards in Kakadu National Park), or from the scientific literature (preferably for

Australia and New Zealand). This method is most appropriate for stressors directly toxic

to biota (e.g. salinity, pH, DO, ammonia), but can also be applied to naturally-occurring

stressors such as nutrients (e.g. nutrient addition bioassays). Ecological effects data are

obtained through site- or ecosystem-specific laboratory and field experiments (see text

below for deriving low-risk trigger values).

b) reference system data — obtained either from the same (undisturbed) ecosystem (i.e.

from upstream of possible environmental impacts) or from a local but different system,

or from regional reference ecosystems (Section 3.1.4). This is particularly useful for

aquatic ecosystems where the management target is to maintain or restore the

ecosystem, and where there are sufficient resources to obtain the required information

on the reference ecosystem (see the text below for deriving low-risk trigger values).

c) predictive modelling — particularly useful for certain physical and chemical stressors

whose disturbance occurs through transformations in the environment (e.g. nutrients,

biodegradable organic matter). In these cases, because of the other factors involved,

there does not appear to be a direct relationship between the ambient concentration of

the stressor (e.g. total P concentration) and the biological response (e.g. algal

biomass). However, there is often a plausible relationship between loading (or flux) and

biological response.

d) professional judgement — may be used in cases where it will not be possible to obtain

appropriate data for a reference ecosystem because insufficient study has been

undertaken to provide an adequate data base. Such judgement should be supported by

appropriate scientific information (e.g. information from 1992 ANZECC guidelines or other

guideline documents, e.g. Hart 1974, Alabaster & Lloyd 1982, USEPA 1986, CCREM

1991), and the scientific literature.

b Sections

3.1.3.2, 3.1.7

& 7.2.3.3

c Section

3.2.4.2

d Section

7.4.4.1

a See box

3.3.1
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–7

Low-risk trigger values can be developed for condition 2 and condition 3

ecosystems:

• condition 2, slightly–moderately disturbed ecosystems, where the objective is to

maintain biological diversity, acknowledging that stakeholders may also decide

to allow some small change to biodiversity as well as improve or restore the

ecosystem to a substantially unmodified condition, depending upon the situation;

• condition 3, highly disturbed ecosystems, where the management target will be

to maintain, and preferably, improve the ecosystem, although in many cases the

possibility of restoring the system to a substantially natural ecosystem may not

be realistic. Urban aquatic systems (rivers, streams, wetlands, estuaries) are a

case in point. For most of these, the hydrology in particular has been so markedly

changed that at best a somewhat modified ecosystem can be achieved.

As suggested for high conservation/ecological value sites above, users also need to

negotiate statistical decision criteria that can apply to any monitoring program for

condition 2 or condition 3 ecosystems designed to detect change in values of physical

and chemical stressors. Where maintenance of biological diversity is an important

management goal, these criteria need to be set conservatively, but can be relaxed if

some change to the system is acceptable.

The following sections outline the preferred hierarchy for deriving low-risk trigger

values for aquatic systems (see figure 3.1.2). Where the preferred approach cannot

be immediately implemented, a default or interim approach has been outlined.

3.3.2.4 Preferred approaches to deriving low-risk guideline trigger values

Using ecological effects data

For low-risk trigger values, measure the statistical distribution of water quality

indicators either at a specific site (preferred), or an appropriate reference system(s),

and also study the ecological and biological effects of physical and chemical

stressors.a Then define the trigger value as the level of key physical or chemical

stressors below which ecologically or biologically meaningful changes do not occur,

i.e. the acceptable level of change.b Depending on the level of protection of the water

body, the trigger value can be defined more or less conservatively after consultation

with stakeholders, and using professional advice.c

Using reference data

Where there is insufficient information on ecological effects to determine an

acceptable change from the reference condition, use an appropriate percentile of the

reference data distribution to derive the trigger value. The percentile represents a

measure that can be applied to data whether they be normally or non-normally

distributed.

For naturally-occurring stressors, use data from appropriate reference systems to

determine the low-risk trigger value for each key indicator. For these Guidelines,

data collected after two years of monthly sampling are regarded as sufficient to

indicate ecosystem variability and can be used to derive trigger values.

Ideally, in ecosystems not characterised by large seasonal or event-scale effects,

develop trigger values for each month, i.e. a total of 12 low-risk trigger values.

However, in some ecosystems, the relationships between physical and chemical

indicators and key biological responses can be influenced by strong seasonal or

a See Sections

3.2.3, 8.1 &

Monitoring

Guidelines

b Sections

3.3.2.7 & 7.2.3.3

c Section 8.5.2
Chapter 3 — Aquatic ecosystems

page 3.3–8 Version — October 2000

event-scale effects. In these systems, it will be necessary to monitor so as to detect

these seasonal influences or events. For ecosystems where seasonal or event-driven

processes dominate (e.g. tropical wetlands), it is possible to group the data and derive

a number of trigger values corresponding to the key seasonal periods. For example,

in wet–dry tropical systems two trigger values can be derived, one for the wet season

and another for the dry season. In these instances, collect, partition and compare

reference and test data according to specific flow regimes and/or seasons, particularly

where biological responses to a particular stressor can be identified to be more

pronounced in a particular season or flow regime.a

Where few data are available (i.e. few reference sites or sampling times) and

seasonal and event influences are poorly defined, derive a single trigger value from

available data as an interim measure.

Define trigger values for physical and chemical stressors for condition 2

ecosystems, in terms of the 80th and/or 20th percentile values obtained from an

appropriate reference system. This choice is arbitrary (though reasonably

conservative),b and professional advice should be sought wherever possible in

selecting an appropriate point on the distribution curve for a system. For stressors

that cause problems at high concentrations (e.g. nutrients, SPM, biochemical

oxygen demand (BOD), salinity), take the 80th percentile of the reference

distribution as the low-risk trigger value. For stressors that cause problems at low

levels (e.g. low temperature water releases from reservoirs, low dissolved oxygen

in waterbodies), use the 20th percentile of the reference distribution as a low-risk

trigger value. For stressors that cause problems at both high and low values (e.g.

temperature, salinity, pH), the desired range for the median concentration is

defined by the 20th percentile and 80th percentile of the reference distribution.c

For condition 3 waterbodies, derive trigger values from site-specific biological or

ecological effects data or, when an appropriate reference system(s) has been

identified and there are sufficient resources to collect the necessary information,

from local reference data. In this latter case, depending on management objectives,

define trigger values using a conservative percentile value (e.g. 80th percentile

value) to improve water quality (preferred approach), or a less conservative

percentile (e.g. 90th percentile) to maintain water quality. Use professional

judgement to determine the most appropriate cutoff percentile.

For either condition 2 or condition 3 ecosystems, where there are insufficient

information or resources to undertake the necessary site-specific studies, use the

default values provided that are derived from regional reference data (see following

section).

3.3.2.5 Default approach to deriving low-risk guideline trigger values

The default approach to deriving trigger values has used the statistical distribution

of reference data collected within five geographical regions across Australia and

New Zealand. Here, depending on the stressor, a measurable perturbation in

slightly to moderately disturbed ecosystems has been defined using the 80th and/or

20th percentile of the reference data.d

First, New Zealand and Australian state and territory representatives used

percentile distributions of available data and professional judgement to derive

trigger values for each ecosystem type in their regions. Trigger values were then

collated, discussed and agreed for south-east Australia (VIC, NSW, ACT, southa See Sections

3.3.2.9 &

3.3.3.3

b Section 7.4.4

c Section

7.4.4.1

d Section

7.4.4.1
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–9

east QLD, and TAS), south-west Australia (southern WA), tropical Australia

(northern WA, NT, northern QLD), south central Australia — low rainfall area

(SA) and New Zealand (tables 3.3.2 to 3.3.11). Summaries of the data used to

derive guideline trigger values for each Australian state and territory and for New

Zealand are provided in Volume 2.a

The default trigger values in the present guidelines were derived from ecosystem data

for unmodified or slightly-modified ecosystems supplied by state agenicies.

However, the choice of these reference systems was not based on any objective

biological criteria. This lack of specificity may have resulted in inclusion of reference

systems of varying quality, and further emphasises that the default trigger values

should only be used until site- or ecosystem-specific values can be generated.

Default trigger values for temperature are not provided here. Managers need to

define their own upper and lower low-risk trigger values, using the 80th and 20th

percentiles, respectively, of ecosystem temperature distribution.

a See Section

8.2.2
Chapter 3 — Aquatic ecosystems

page 3.3–10 Version — October 2000

Tables 3.3.2–3.3.3 South-east Australia

The following tables outline default trigger values applicable to Victoria, New

South Wales, south-east Queensland, the Australian Capital Territory and

Tasmania. Where individual states or territories have developed their own regional

guideline trigger values, those values should be used in preference to the default

values provided below. (Upland streams are defined as those at >150 m altitude,

while alpine streams are those at altitudes >1500 m.)

Table 3.3.2 Default trigger values for physical and chemical stressors for south-east Australia for slightly

disturbed ecosystems. Trigger values are used to assess risk of adverse effects due to nutrients, biodegradable

organic matter and pH in various ecosystem types. Data derived from trigger values supplied by Australian

states and territories. Chl a = chlorophyll a, TP = total phosphorus, FRP = filterable reactive phosphate,

TN = total nitrogen, NOx = oxides of nitrogen, NH4

+

 = ammonium, DO = dissolved oxygen.

Ecosystem type Chl a TP FRP TN NOx NH4

+ DO (% saturation)l pH

(μg L-1) (μg P L-1) (μg P L-1) (μg N L-1) (μg N L-1) (μg N L-1) Lower limit Upper limit Lower limit Upper limit

Upland river naa 20b 15g 250 c 15h 13i 90 110 6.5 7.5m

Lowland riverd

5 50 20 500 40o 20 85 110 6.5 8.0

Freshwater lakes &

Reservoirs 5e 10 5 350 10 10 90 110 6.5 8.0 m

Wetlands no data no data no data no data no data no data no data no data no data no data

Estuariesp

4f 30 5j 300 15 15 80 110 7.0 8.5

Marinep

1n 25n 10 120 5k 15 k 90 110 8.0 8.4

na = not applicable;

a = monitoring of periphyton and not phytoplankton biomass is recommended in upland rivers — values for periphyton biomass

(mg Chl a m-2) to be developed;

b = values are 30 μgL-1 for Qld rivers, 10 μgL-1 for Vic. alpine streams and 13 μgL-1 for Tas. rivers;

c = values are 100 μgL-1 for Vic. alpine streams and 480 μgL-1 for Tas. rivers;

d = values are 3 μgL-1 for Chl a, 25 μgL-1 for TP and 350 μgL-1 for TN for NSW & Vic. east flowing coastal rivers;

e = values are 3 μgL-1 for Tas. lakes;

f = value is 5 μgL-1 for Qld estuaries;

g = value is 5 μgL-1 for Vic. alpine streams and Tas. rivers;

h = value is 190 μgL-1 for Tas. rivers;

i = value is 10 μgL-1 for Qld. rivers;

j = value is 15 μgL-1 for Qld. estuaries;

k = values of 25 μgL-1 for NOx and 20 μgL-1 for NH4

+

 for NSW are elevated due to frequent upwelling events;

l = dissolved oxygen values were derived from daytime measurements. Dissolved oxygen concentrations may vary diurnally and

with depth. Monitoring programs should assess this potential variability (see Section 3.3.3.2);

m = values for NSW upland rivers are 6.5–8.0, for NSW lowland rivers 6.5–8.5, for humic rich Tas. lakes and rivers 4.0-6.5;

n = values are 20 μgL-1 for TP for offshore waters and 1.5 μgL-1 for Chl a for Qld inshore waters;

o = value is 60 μgL-1 for Qld rivers;

p = no data available for Tasmanian estuarine and marine waters. A precautionary approach should be adopted when applying

default trigger values to these systems.
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–11

Table 3.3.3 Ranges of default trigger values for conductivity (EC, salinity), turbidity and suspended particulate

matter (SPM) indicative of slightly disturbed ecosystems in south-east Australia. Ranges for turbidity and SPM

are similar and only turbidity is reported here. Values reflect high site-specific and regional variability.

Explanatory notes provide detail on specific variability issues for ecosystem type.

Ecosystem

type

Salinity (μScm–1) Explanatory notes

Upland rivers 30–350

Conductivity in upland streams will vary depending upon catchment geology.

Low values are found in Vic. alpine regions (30 μScm-1) and eastern highlands

(55 μScm-1), and high values (350 μScm-1) in NSW rivers. Tasmanian rivers are

mid-range (90 μScm-1).

Lowland rivers 125–2200

Lowland rivers may have higher conductivity during low flow periods and if the

system receives saline groundwater inputs. Low values are found in eastern

highlands of Vic. (125 μScm-1) and higher values in western lowlands and

northern plains of Vic (2200 μScm-1). NSW coastal rivers are typically in the

range 200–300 μScm-1.

Lakes &

reservoirs 20–30

Conductivity in lakes and reservoirs is generally low, but will vary depending

upon catchment geology. Values provided are typical of Tasmanian lakes and

reservoirs.

Turbidity (NTU)

Upland rivers 2–25 Most good condition upland streams have low turbidity. High values may be

observed during high flow events.

Lowland rivers 6–50

Turbidity in lowland rivers can be extremely variable. Values at the low end of the

range would be found in rivers flowing through well vegetated catchments and at

low flows. Values at the high end of the range would be found in rivers draining

slightly disturbed catchments and in many rivers at high flows.

Lakes &

reservoirs 1–20

Most deep lakes and reservoirs have low turbidity. However, shallow lakes and

reservoirs may have higher natural turbidity due to wind-induced resuspension of

sediments. Lakes and reservoirs in catchments with highly dispersible soils will

have high turbidity.

Estuarine &

marine 0.5–10

Low turbidity values are normally found in offshore waters. Higher values may

be found in estuaries or inshore coastal waters due to wind-induced

resuspension or to the input of turbid water from the catchment. Turbidity is not a

very useful indicator in estuarine and marine waters. A move towards the

measurement of light attenuation in preference to turbidity is recommended.
Chapter 3 — Aquatic ecosystems

page 3.3–12 Version — October 2000

Tables 3.3.4–3.3.5 Tropical Australia

The following tables outline default trigger values applicable to northern

Queensland, the Northern Territory and north-west Western Australia. Where

states or territories have developed regional guideline trigger values those values

should be used in preference to the default values provided below. (Upland streams

are defined as those at >150 m altitude.)

Table 3.3.4 Default trigger values for physical and chemical stressors for tropical Australia for slightly disturbed

ecosystems. Trigger values are used to assess risk of adverse effects due to nutrients, biodegradable organic

matter and pH in various ecosystem types. Data derived from trigger values supplied by Australian states and

territories, for the Northern Territory and regions north of Carnarvon in the west and Rockhampton in the east.

Chl a = chlorophyll a, TP = total phosphorus, FRP = filterable reactive phosphate, TN = total nitrogen,

NOx = oxides of nitrogen, NH4

+

 = ammonium, DO = dissolved oxygen.

Ecosystem type Chl a TP FRP TN NOx NH4

+ DO (% saturation) f pH

(μg L-1) (μg P L-1) (μg P L-1) (μg N L-1) (μg N L-1) (μg N L-1) Lower limit Upper limit Lower limit Upper limit

Upland rivere

naa 10 5 150 30 6 90 120 6.0 7.5

Lowland rivere

5 10 4 200– 300h 10b 10 85 120 6.0 8.0

Freshwater lakes &

reservoirs 3 10 5 350c 10b 10 90 120 6.0 8.0

Wetlands 10 10–50g 5–25g 350–1200g 10 10 90b 120 b 6.0 8.0

Estuariese

2 20 5 250 30 15 80 120 7.0 8.5

Marine Inshore 0.7–1.4d 15 5 100 2–8 d 1–10 d 90 no data 8.0 8.4

Offshore 0.5–0.9 d 10 2–5 d 100 1–4 d 1–6 d 90 no data 8.2 8.2

na = not applicable

a = monitoring of periphyton and not phytoplankton biomass is recommended in upland rivers — values for periphyton

biomass (mg Chl a m-2) to be developed;

b = Northern Territory values are 5μgL-1 for NOx, and <80 (lower limit) and >110% saturation (upper limit) for DO;

c = this value represents turbid lakes only. Clear lakes have much lower values;

d = the lower values are typical of clear coral dominated waters (e.g. Great Barrier Reef), while higher values typical of turbid

macrotidal systems (eg. North-west Shelf of WA);

e = no data available for tropical WA estuaries or rivers. A precautionary approach should be adopted when applying default

trigger values to these systems;

f = dissolved oxygen values were derived from daytime measurements. Dissolved oxygen concentrations may vary diurnally

and with depth. Monitoring programs should assess this potential variability (see Section 3.3.3.2);

g = higher values are indicative of tropical WA river pools;

h = lower values from rivers draining rainforest catchments.
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–13

Table 3.3.5 Ranges of default trigger values for conductivity (EC, salinity), turbidity and suspended particulate

matter (SPM) indicative of slightly disturbed ecosystems in tropical Australia. Ranges for turbidity and SPM are

similar and only turbidity is reported here. Values reflect high site-specific and regional variability. Explanatory

notes provide detail on specific variability issues for groupings of ecosystem type.

Ecosystem

type

Salinity (μScm-1) Explanatory notes

Upland &

lowland rivers 20–250

Conductivity in upland streams will vary depending upon catchment geology.

Values at the lower end of the range are typical of ephemeral flowing NT rivers.

Catchment type may influence values for Qld lowland rivers (e.g. 150 μScm-1 for

rivers draining rainforest catchments, 250 μScm-1 for savanna catchments). The

first flush of water following early seasonal rains may result in temporarily high

values.

Lakes,

reservoirs &

wetlands

90–900

Values at the lower end of the range are found in permanent billabongs in the NT.

Higher conductivity values will occur during summer when water levels are reduced

due to evaporation. WA wetlands can have values higher than 900 μScm-1. Turbid

freshwater lakes in Qld have reported conductivities of approx. 170 μScm-1.

Turbidity (NTU)

Upland &

lowland rivers 2–15

Low values for base flow conditions in NT rivers. QLD turbidity and SPM values

highly variable and dependent on degree of catchment modification and

seasonal rainfall runoff.

Lakes,

reservoirs &

wetlands

2–200

Most deep lakes and reservoirs have low turbidity. However, shallow lakes and

reservoirs may have higher turbidity naturally due to wind-induced resuspension

of sediments. Lakes and reservoirs in catchments with highly dispersible soils

will have high turbidity. Wetlands vary greatly in turbidity depending upon the

general condition of the catchment or river system draining into the wetland,

recent flow events and the water level in the wetland.

Estuarine

& marine 1–20

Low values indicative of offshore coral dominated waters. Higher values

representative of estuarine waters. Turbidity is not a very useful indicator in

estuarine and marine waters. A move towards the measurement of light

attenuation in preference to turbidity is recommended. Typical light attenuation

coefficients (log10) in waters off north-west WA range from 0.17 for inshore

waters to 0.07 for offshore waters.
Chapter 3 — Aquatic ecosystems

page 3.3–14 Version — October 2000

Tables 3.3.6–3.3.7 South-west Australia

The following tables outline default trigger values applicable to southern Western

Australia. Where regional guideline trigger values have been developed, those

values should be used in preference to the default values provided below. The WA

EPA is currently developing site-specific environmental quality criteria for Perth’s

coastal waters. (Upland streams are defined as those at >150 m altitude.)

Table 3.3.6 Default trigger values for physical and chemical stressors for south-west Australia for slightly

disturbed ecosystems. Trigger values are used to assess risk of adverse effects due to nutrients, biodegradable

organic matter and pH in various ecosystem types. Data derived from trigger values supplied by Western

Australia. Chl a = chlorophyll a, TP = total phosphorus, FRP = filterable reactive phosphate, TN = total nitrogen,

NOx = oxides of nitrogen, NH4

+

 = ammonium, DO = dissolved oxygen.

Ecosystem type Chl a TP FRP TN NOx NH4

+ DO (% saturation) i pH

(μg L-1) (μg P L-1) (μg P L-1) (μg N L-1) (μg N L-1) (μg N L-1) Lower limit Upper limit Lower limit Upper limit

Upland riverf

naa 20 10 450 200 60 90 na 6.5 8.0

Lowland riverf

3–5 65 40 1200 150 80 80 120 6.5 8.0

Freshwater lakes &

reservoirs 3–5 10 5 350 10 10 90 no data 6.5 8.0

Wetlandsd

30 60 30 1500 100 40 90 120 7.0e 8.5e

Estuaries 3 30 5 750 45 40 90 110 7.5 8.5

Marineg,h Inshorec

0.7 20 b 5b 230 5 5 90 na 8.0 8.4

Offshore 0.3 b 20 b 5 230 5 5 90 na 8.2 8.2

na = not applicable

a = monitoring of periphyton and not phytoplankton biomass is recommended in upland rivers — values for periphyton

biomass (mg Chl a m-2) to be developed;

b = summer (low rainfall) values, values higher in winter for Chl a (1.0 μgL-1), TP (40 μg P L-1), FRP (10 μg P L-1);

c = inshore waters defined as coastal lagoons (excluding estuaries) and embayments and waters less than 20 metres depth;

d = elevated nutrient concentrations in highly coloured wetlands (gilven >52 g440m-1) do not appear to stimulate algal growth;

e = in highly coloured wetlands (gilven >52 g440m-1) pH typically ranges 4.5–6.5;

f = all values derived during base river flow conditions not storm events;

g = nutrient concentrations alone are poor indicators of marine trophic status;

h = these trigger values are generic and therefore do not necessarily apply in all circumstances e.g. for some unprotected

coastlines, such as Albany and Geographe Bay, it may be more appropriate to use offshore values for inshore waters;

i = dissolved oxygen values were derived from daytime measurements. Dissolved oxygen concentrations may vary diurnally

and with depth. Monitoring programs should assess this potential variability (see Section 3.3.3.2).
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–15

Table 3.3.7 Range of default trigger values for conductivity (EC, salinity), turbidity and suspended particulate

matter (SPM) indicative of slightly disturbed ecosystems in south-west Australia. Ranges for turbidity and SPM

are similar and only turbidity is reported here. Values reflect high site-specific and regional variability.

Explanatory notes provide detail on specific variability issues for ecosystem types.

Ecosystem

type

Salinity

(μScm-1)

Explanatory notes

Upland &

lowland rivers 120–300

Conductivity in upland streams will vary depending upon catchment geology.

Values at the lower end of the range are typically found in upland rivers, with higher

values found in lowland rivers. Lower conductivity values are often observed

following seasonal rainfall.

Lakes,

reservoirs &

wetlands

300–1500

Values at the lower end of the range are observed during seasonal rainfall events.

Values even higher than 1500 μScm-1 are often found in saltwater lakes and

marshes. Wetlands typically have conductivity values in the range 500−1500

μScm-1 over winter. Higher values (>3000 μScm-1) are often measured in wetlands

in summer due to evaporative water loss.

Turbidity

(NTU)

Upland &

lowland rivers 10–20

Turbidity and SPM are highly variable and dependent on seasonal rainfall runoff.

These values representative of base river flow in lowland rivers.

Lakes,

reservoirs &

wetlands

10–100

Most deep lakes and reservoirs have low turbidity. However, shallow lakes and

reservoirs may have higher turbidity naturally due to wind-induced resuspension of

sediments. Lakes and reservoirs in catchments with highly dispersible soils will

have high turbidity. Wetlands vary greatly in turbidity depending upon the general

condition of the catchment or river system draining into the wetland and to the

water level in the wetland.

Estuarine &

marine 1–2

Turbidity is not a very useful indicator in estuarine and marine waters. A more

appropriate measure for WA coastal waters is light attenuation coefficient. Light

attenuation coefficients (log10) of 0.05–0.08 m-1 are indicative of unmodified

offshore waters and 0.09–0.13 m-1 for unmodified inshore waters, depending on

exposure. Light attenuation coefficients (log10) for unmodified estuaries typically

range 0.3–1.0 m-1, although more elevated values can be associated with

increased particulate loading or humic rich waters following seasonal rainfall

events.
Chapter 3 — Aquatic ecosystems

page 3.3–16 Version — October 2000

Tables 3.3.8–3.3.9 South central Australia — low rainfall area

The following tables outline default trigger values applicable to South Australia.

Where regional guideline trigger values have been developed those values should

be used in preference to the default values provided below. (Upland streams are

defined as those at >150 m altitude.)

Table 3.3.8 Default trigger values for physical and chemical stressors for south central Australia — low rainfall

areas — for slightly disturbed ecosystems. Trigger values are used to assess risk of adverse effects due to

nutrients, biodegradable organic matter and pH in various ecosystem types. Data derived from trigger values

supplied by South Australia. Chl a = chlorophyll a, TP = total phosphorus, FRP = filterable reactive phosphate,

TN = total nitrogen, NOx = oxides of nitrogen, NH4

+

 = ammonium, DO = dissolved oxygen.

Ecosystem type Chl a TP FRP TN NOx NH4

+ DO (% saturation) pH

(μg L-1) (μg P L-1) (μg P L-1) (μg N L-1) (μg N L-1) (μg N L-1) Lower limit Upper limit Lower limit Upper limit

Upland river no data no data no data no data no data no data no data no data no data no data

Lowland river no data 100 40 1000 100 100 90 no data 6.5 9.0

Freshwater lakes

& reservoirs no data 25 10 1000 100 25 90 no data 6.5 9.0

Wetlands no data no data no data no data no data no data no data no data no data no data

Estuaries 5 100 10 1000 100 50 90 no data 6.5 9.0

Marine 1 100 10 1000 50 50 no data no data 8.0 8.5

Table 3.3.9 Ranges of default trigger values for conductivity (EC, salinity), turbidity and suspended particulate

matter (SPM) indicative of slightly disturbed ecosystems in south central Australia — low rainfall areas. Ranges

for turbidity and SPM are similar and only turbidity is reported here. Values reflect high site-specific and regional

variability. Explanatory notes provide detail on specific variability issues for groupings of ecosystem type.

Ecosystem

types

Salinity

(μScm-1)

Explanatory notes

Lowland rivers 100–5000 Salinity can be highly variable depending on flow.

Lakes,

reservoirs &

wetlands

300–1000

Wetlands can have substantially higher salinity due to saline groundwater intrusion

and evaporation.

Turbidity (NTU)

Upland &

lowland rivers 1–50 Turbidity and SPM are highly variable and dependent on seasonal rainfall runoff.

Lakes &

reservoirs/

wetlands

1–100

Shallow lakes and reservoirs may have higher turbidity naturally due to windinduced resuspension of sediments. Lakes and reservoirs in catchments with highly

dispersible soils will have high turbidity.

Estuarine &

marine 0.5–10 Higher values are representative of estuarine waters.
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–17

Tables 3.3.10–3.3.11 New Zealand

The following tables outline default trigger values applicable to New Zealand.

Where regional guideline trigger values have been developed, those values should

be used in preference to the default values provided below. (Upland streams are

defined as those at >150 m altitude.)

For streams and rivers, New Zealand is developing a five-category ecosystem

health categorisation system (A–E, with A being desirable and E undesirable). The

draft National Agenda for Sustainable Water Management (NZ Ministry for the

Environment 1999) proposes as a long-term goal that all streams are in C grade or

better. For lakes, New Zealand has developed a fine scale lakes trophic assessment

system, that enables water managers to objectively score the trophic condition of

the lake. This assessment system combines a number of physical and chemical

parameters. These parameters vary considerably across New Zealand, depending,

for example, on whether a lake drains a volcanic catchment, in which case nitrate is

a critical parameter, or whether the lake drains a hard rock catchment, in which

case phosphorus is a critical parameter. Because of this variability, and because

New Zealand has developed this trophic assessment system, it is not appropriate to

propose trigger values for individual parameters from lakes.

Further work is needed to develop a categorisation system for New Zealand estuarine

and marine ecosystems. Consideration should be given to the use of interim trigger

values for south-east Australian estuarine and marine ecosystems (tables 3.3.2–3.3.3)

until New Zealand estuarine and marine trigger values are developed.

Table 3.3.10 Default trigger values for physical and chemical stressors in New Zealand for

slightly disturbed ecosystems. Trigger values are used to assess risk of adverse effects due

to nutrients, biodegradable organic matter and pH in various ecosystem types. Chl a =

chlorophyll a, TP = total phosphorus, FRP = filterable reactive phosphate,d TN = total

nitrogen, NOx = oxides of nitrogen, NH4

+

 = ammoniacal nitrogen, DO = dissolved oxygen.

Ecosystem

type

Chl a TP FRP TN NOx NH4

+ DOe

(% saturation)

pHe

(μg L-1) (μg P L-1) (μg P L-1) (μg N L-1) (μg N L-1) (μg N L-1) Lower

limit

Upper

limit

Lower

limit

Upper

limit

Upland river naa 26b 9b 295b 167b 10b 99 103 7.3 8.0

Lowland river no data 33c 10c 614c 444c 21c 98 105 7.2 7.8

na = not applicable

a = monitoring of periphyton and not phytoplankton biomass is recommended in upland rivers — values

for periphyton biomass (mg Chl a m-2) to be developed. New Zealand is currently making routine

observations of periphyton cover.

b = values for glacial and lake-fed sites in upland rivers are lower;

c = values are lower for Haast River which receives waters from alpine regions;

d = commonly referred to as dissolved reactive phosphorus in New Zealand;

e = DO and pH percentiles may not be very useful as trigger values because of diurnal and seasonal

variation — values listed are for daytime sampling.
Chapter 3 — Aquatic ecosystems

page 3.3–18 Version — October 2000

Table 3.3.11 Default trigger values for water clarity (lower limit) and turbidity (upper limit)

indicative of unmodified or slightly disturbed ecosystems in New Zealand

Ecosystem types Upland riversa b Lowland rivers

Clarity (m-1)

c d Turbidity (NTU) c d Clarity (m-1) Turbidity (NTU)

0.6 4.1 0.8 5.6

a = Light availability is generally less of an issue in NZ rivers and streams than is visual clarity because, in

contrast to many of Australia's rivers, most NZ rivers are comparatively clear and/or shallow. Davies-Colley

et al. (1992) recommend that visual clarity, light penetration and water colour are important optical

properties of an ecosystem which need to be protected (see Volume 2). Neither turbidity nor visual clarity

provide a useful estimate of light penetration — light penetration should be considered separately to

turbidity or visual clarity. Clarity relates to the transmission of light through water and is measured by the

visual range of a black disk (see NZ Ministry for the Environment (1994)) or a Secchi disk.

b = Recent work has shown that at least some NZ indigenous fish are sensitive to low levels of turbidity;

however, it may also be desirable to protect the naturally high turbidities of alpine glacial lakes to prevent

possible ecological impacts, such as change in predator–prey relationships.

c = Note that turbidity and visual water clarity are closely and inversely related, and the 80th percentile for

turbidity is consistent with the 20th percentile for visibility and vice versa.

d = Clarity and turbidity values for glacial sites in upland rivers are lower and higher, respectively.
3.3.2 Philosophy used in developing guidelines for physical and chemical stressors

Version — October 2000 page 3.3–19

3.3.2.6 Comparison with the low-risk guideline trigger value

Where trigger values have been developed from reference data, it is advisable to

compare the median of replicate samples from a test site with the low-risk trigger

value. Statistically, the median represents the most robust descriptor of the test site

data, while the reference percentile value represents the degree of excursion that

the test median is permitted before triggering some action.

Two issues will influence the outcome of the comparison: the amount of data used

to calculate the trigger value (minimum two years of monthly sampling); and the

number of replicates used to calculate the median from the test site (minimum of a

single sample). A fuller discussion of these issues, with guidance on statistical

ramifications of changes in sample size, are provided in Section 7.4.4.1.

Control charting

It is best to continually compare the trigger values against the results gathered

during ongoing monitoring of the physical and chemical indicators, using control

charts. Control charting displays the data trends and gives early warning that the

test site may be trending towards a high-risk situation. Further discussion on the

applications of control charts may be found in Section 7.4.4.1 and in the

Monitoring Guidelines (ANZECC & ARMCANZ 2000). Excursion of the test site

value beyond the trigger value requires that further action be undertaken. This may

include, simply, an examination of data for errors, comparisons with previous

excursions, or the use of simple decision trees such as those outlined in the riskbased guideline packages.a Site specific investigations may also be required to

decide if there is an issue or problem to be addressed.

3.3.2.7 Measuring acceptable ecological change

Measurement of ‘acceptable’ ecological change is difficult (Keough & Mapstone

1995, Mapstone 1995). In very few situations is there enough scientific knowledge

to indicate if a certain minimum change from the prevailing or target condition will

cause an adverse ecological effect. To define this level of change (a) water quality

indicator distributions must be correlated with grades or levels of ecosystem health

or integrity indicators/indices, and (b) substantiating potential cause and effect

relationships must be identified through these correlations, using laboratory and

field-based biological and ecological effects research.

A number of recent studies are trying to link physical and chemical stressors with

ecological effects and thereby define meaningful criteria for monitoring ecosystem

health:

• As mentioned above, New Zealand is developing a five-category ecosystem health

classification for freshwater shingle streams draining hard rock catchments. These

categories are derived by comparison with a reference condition, and are based on

a number of desirable biological features such as trout spawning, presence of

sensitive native fish and no growth of benthic filamentous green algae. Fifty

streams have been graded, and the distribution of water quality stressors within

each grade will be used to define trigger values for physical and chemical

indicators (E Pyle, NZ Ministry for the Environment, pers. comm.).

• Four large-scale studies in Australia have aimed to determine the cause and

effect relationships between coastal ecosystem health and physical and

chemical stressors (Port Phillip Bay Study, Moreton Bay and Brisbane River

a See Sections

3.3.3 & 8.2.3
Chapter 3 — Aquatic ecosystems

page 3.3–20 Version — October 2000

Wastewater Management Study, and two Perth studies — the Perth Coastal

Water Study and the South Metropolitan Coastal Water Studies). These

multidisciplinary studies have led to an understanding of the influence of key

stressors on ecosystem structure (e.g. suspended sediment concentration effects

on seagrass distribution) and function (e.g. nitrogen loading effects on

denitrification). The design and implementation of further such studies will aid

in defining acceptable levels of ecological change.a

3.3.2.8 Load-based guidelines

Traditionally, water quality guidelines have been expressed in terms of the

concentration of the stressor that should not be exceeded if problems are to be

avoided (ANZECC 1992). Such concentration-based guidelines are based primarily

on the prevention of toxic effects. In other situations, guidelines are better expressed

in terms of the flux or loading (i.e. mass per unit time), rather than concentration.

While algal growth rate (or productivity) is related to the concentration of key

nutrients in the water column, the biomass is more controlled by the total mass of

these nutrients available to the growing algae (Wetzel 1975).11 In many cases, the

water column nutrient concentration is not a good indicator of algal biomass. For

example, the net water column nutrient concentration could be quite small in an

ecosystem with a high algal biomass but with rapid nutrient cycling. Load-based

guidelines for nutrients are covered in more detail below.b

The dissolved oxygen concentration in a waterbody depends on the balance between

the flux of bioavailable organic carbon and the rate at which heterotrophic bacteria

use up oxygen in decomposing this material, and the daily inputs of oxygen by

diffusion from the atmosphere (increased by mixing) and via photosynthesis by

macrophytes and phytoplankton (Stumm & Morgan 1996). Load-based guidelines

for bioavailable organic matter are covered below.c

Load-based guidelines are applicable also for assessing the effects of sedimentation

of suspended particulate matter in smothering benthic organisms. Both the rate of

sedimentation and the critical depth of the deposited material are load-based.d

A number of case studies are presented to show the types of approaches

(particularly those involving predictive modelling) that can be used to determine

the sustainable load of particular materials for a particular ecosystem. We

recommend that work in developing similar types of case studies be increased. A

number of key research areas are identified in Section 8.5.2 of Volume 2.e

3.3.2.9 Tropical ecosystems

Although the guideline packages address issues that can apply to all biogeographic

regions, the case studies in Sections 3.3.3 and 8.2.3 use examples from temperate

regions. There is a need for tropical, risk-based guideline packages to be developed

for Australian aquatic ecosystems which are characterised by elevated seasonal

temperatures and significant seasonal variability in rainfall and stream-flow patterns

(Finlayson & McMahon 1988). Algal blooms may be an issue in some tropical

marine and freshwater ecosystems. Extensive macrophyte assemblages can have

direct (e.g. smothering) and indirect (e.g. on dissolved oxygen, nutrients and light

 

11 Note: this assumes that growth is not limited by light and that losses of algae by zooplankton

grazing, sedimentation and ‘washout’ from the system are small.

b Section

3.3.3.1 & case

studies 1 & 2 in

section 3.3.3

c Section 3.3.3.2

& case study 4

in Section 8.2.3

(Vol. 2)

d Case study 5

(Vol. 2)

e Section 8.5.2

of Volume 2

a See Section

8.5.2
3.3.3 Guideline packages for applying the guideline trigger values to sites

Version — October 2000 page 3.3–21

availability) effects on tropical wetlands, and risk-based guideline packages are

needed to address the influences of key stressors on such systems.

Monitoring should be arranged so that it targets episodic events. For instance,

seasonally-variable stream flows can cease for large parts of the year. In some

streams and reservoirs, slow flowing or pooled water leads to thermal stratification,

which together with autochthonous organic loading, results in naturally low and

variable dissolved oxygen concentrations (MacKinnon & Herbert 1996, Townsend

1999). Seasonal rainfall events often produce ‘first-flush’ loads of stressors that can

cause rapid changes in stressor concentrations (Hart et al. 1987, Townsend et al.

1992) that may not be captured with routine monitoring programs.

There are few data for tropical water bodies; site- or ecosystem-specific reference

data need to be collected for tropical ecosystems. The approach recommended in

these Guidelinesa — studies of site-specific biological or ecological effects to

develop local trigger values — is also especially appropriate in ecosystems that

demonstrate such a high degree of variability in physical and chemical stressors (e.g.

wet and wet–dry tropics).

3.3.3 Guideline packages for applying the guideline trigger values to sites

3.3.3.1 Risk-based guideline packages

Ideally, a guideline package, consisting of low-risk trigger values and a protocol for

including effects of environmental modifiers, should be developed for each

ecosystem issue and each ecosystem type. At this stage, only a limited number of

packages can be recommended. Guideline packages are shown and discussed here

for two issues:

• nuisance growth of aquatic plants, and

• lack of dissolved oxygen.

Further guideline packages are provided in Section 8.2.3 for:

• excess suspended particulate matter (SPM),

• unnatural change in salinity,

• unnatural change in temperature,

• unnatural change in pH,

• poor optical properties,

• unnatural flow.

Each guideline package consists of two components (figure 3.3.1):

• a set of low-risk trigger values — A set of key stressors such as total phosphorus

concentration has been identified for each issue. These are used for an initial

decision about the risk of an adverse biological effect occurring. The low-risk

trigger values for these key stressors need to be established as outlined in box

3.3.1. These trigger values are concentration-based, but protocols for the

development of load-based guidelines are provided where these are more

relevant.

• a protocol for further investigating the risk where the trigger value is exceeded —

In these potential risk situations, ecosystem-specific modifying factors that may

a See Section

3.3.2.4
Chapter 3 — Aquatic ecosystems

page 3.3–22 Version — October 2000

alter the biological effect of the key stressor need to be considered before the final

risk can be assessed. The suggested protocol involves a decision tree or predictive

modelling approach where increasingly detailed investigations are undertaken

(figure 3.3.1). For example, where testing of the key stressor against the appropriate

trigger values suggests a potential risk of excessive cyanobacterial growth in a

particular lowland river, the steps involved in further investigating this situation

could be:

i. make a simple assessment of the possible effect of key ecosystem-specific

modifiers on the biological effect of the stressor. A simple decision tree

model for this type of assessment is provided in Case Study 1.

ii. if this simple assessment still suggests a potential risk of adverse biological

effects, then undertake more sophisticated site-specific investigations and

associated modelling. For example, a load-based model of the system to

predict the relationship between nutrient loads, key ecosystem variables and

aquatic plant growth,a or a more comprehensive ecosystem-based model of

the system (see Case Study 4, Harris et al. 1996) could be devised.

In many cases there is insufficient information to allow quantification of the

relationships between the key stressor and environmental factors controlling

bioavailability.b It is essential that these relationships be clarified in the

immediate future.

As discussed in Section 3.1.5, generally, local biological effects data and data from

local reference site(s) that closely match the test site are not required in the

decision trees.

3.3.3.2 Issue: Nuisance growth of aquatic plants

Background

High concentrations of nutrients, particularly phosphorus and nitrogen, and

sometimes silica, can result in excessive growth of aquatic plants such as

phytoplankton, cyanobacteria, macrophytes, seagrasses, and filamentous and

attached algae, in a range of ecosystems, fresh and marine (AEC 1987, CSIRO &

Melbourne Water 1996, WADEP 1996, DWR-NSW 1992, WAEPA 1988, Harris et

al. 1996, Johnstone 1994, Jones 1992, McComb & Davis 1993, McDougall & Ho

1991, MDBC 1994, NZ Ministry for the Environment 1992).

The excessive growth can lead to a number of problems including:

• toxic effects, particularly due to cyanobacteria in fresh and brackish waters, and

dinoflagellates in marine waters;

• reduction in dissolved oxygen concentrations when the plants die and are

decomposed;

• reduction in recreational amenity (phytoplankton blooms and macrophytes in

wetlands and lakes, seagrasses in estuaries and coastal lagoons);

• blocking of waterways and standing waterbodies by macrophytes;

• change in biodiversity.

Excessive growth of aquatic plants occurs when there are high concentrations and

loads of nutrients. Other factors play a part in limiting the growth of nuisance

species, particularly toxic cyanobacteria. The factors include hydraulic retention

time, mixing conditions, light, temperature, suspended solids, grazing pressure and

type of substrate.

a See Case

Study 3 in

Section 8.2.3,

Vol. 2

b Section 8.5.2

in Vol. 2
3.3.3 Guideline packages for applying the guideline trigger values to sites

Version — October 2000 page 3.3–23

Key indicators

Condition indicators chlorophyll a (Chl a), cell numbers, species

composition

Key stressors total phosphorus (TP) and total nitrogen (TN)

concentrations

Ecosystem modifiers depend upon the ecosystem type, but will include

hydraulic retention time (flows and volume of

waterbody), mixing regimes, light regime, turbidity,

temperature, suspended solids (nutrient sorption),

grazing rates, and type of substrate.

Performance indicators median (or mean) concentrations of Chl a, TP and TN

measured under low flow conditions for rivers and

streams and during the growth periods for other

ecosystems.12

Note that nutrients may also be remobilised and released from sediments. Sediment

nutrient releases are influenced by the composition of the sediments (particularly

their bioavailable organic matter, Fe, S, N, P, etc.), temperature, mixing regime of

the water body and oxygen transfer rates. At present we cannot recommend

quantitative relationships to estimate these releases. However, such relationships

should become available in the next few years, and it is essential that these be

incorporated into the guidelines as soon as possible.a

Low-risk trigger values

The method used to determine the low-risk trigger values will depend upon the

desired level of protection.b

Slightly to moderately disturbed ecosystems (condition 2 ecosystems)

Depending upon the importance and present condition of the ecosystem, two

approaches may be taken to derive the most appropriate trigger values for

condition 2 ecosystems.

a) For important ecosystems, where an appropriate local reference system(s) is

available, and there are sufficient resources to collect the necessary information

for the reference system, the low-risk trigger concentrations for the three key

performance indicators (TP, TN and Chl a) should be determined as the 80th

percentile of the reference system(s) distribution. Where possible, the trigger

value should be obtained for that part of the seasonal or flow period when the

probability of aquatic plant growth is most likely.

b) The default regional trigger values contained in tables 3.3.2, 3.3.4, 3.3.6, 3.3.8

and 3.3.10 should be used for those situations where either an appropriate

reference system is not available, or the scale of the operation makes it difficult

to justify the allocation of resources to collect the necessary information on a

reference system.

Highly disturbed ecosystems (condition 3 ecosystems)

a) For important waterbodies, and those in very poor condition, it is best to make

appropriate site-specific scientific studies, and to use the information, with

professional judgement and other relevant information, to derive trigger values.

 

12 In the future, it is recommended that sustainable nutrient loading rates be estimated for each major

ecosystem type (see Section 8.5.2, Volume 2, for research and development recommendations).

b Section

3.3.2.3

a See

recommendations

in Section 8.5,

Vol. 2
Chapter 3 — Aquatic ecosystems

page 3.3–24 Version — October 2000

Where local but higher-quality reference data are used, a less stringent cutoff

than the 80th percentile value may be used. The 80th percentile values, however,

should be used as a target for site improvement.

b) For highly disturbed waterbodies, where there is a lack of either information or

resources to undertake the necessary site-specific studies, it is best to use the

default, regional trigger values using professional judgement to derive a less

stringent value if this is agreed upon by stakeholders.

Use of the guideline package

Figure 3.3.1 shows the recommended approach for determining the risk of nuisance

aquatic plant growth occurring in a particular ecosystem. There are three steps.

• Test the three performance indicators (Chl a, TP, TN concentrations) for the

particular ecosystem against the appropriate low-risk trigger value for that

ecosystem type. Compare the trigger values with the median concentration for

each performance indicator measured under low flow or high growth conditions.

• If test values are less than trigger values, there is low risk of adverse biological

effects and no further action is required, except for regular monitoring of the key

performance and condition indicators. If after regular monitoring a ‘low risk’

outcome is consistently obtained, there is scope to refine the guideline trigger

value. If test values are higher than the trigger values, there is an increased risk

that adverse biological effects will occur, and either management/remedial action

or further ecosystem-specific investigation is required.a

• For some types of ecosystem, further investigation may be needed, to determine

the influence of ecosystem-specific factors on the key stressors. Case studies 1,

2 and 3b illustrate how these factors might be used to modify the effect of high

nutrient concentrations so that problems due to aquatic plants may not arise

even though nutrient concentrations suggest otherwise. Relatively few

quantitative relationships between these factors have been identified for

Australian systems. More work needs to be undertaken on these relationships.

Sustainable nutrient loads

Although nutrient concentrations are responsible (together with other factors) for

stimulating algal growth, it is the total load of the key nutrients in the ecosystem

that controls the final biomass of aquatic plants. The balance between the nutrients

(e.g. the N:P ratio) can also influence the composition of the algal community.

Transformation processes that occur in a waterbody release additional nutrients

(e.g. from sediments, and suspended particles). It is difficult to account for these

without a detailed knowledge of the system, and in many cases a predictive model

(Lawrence 1997 a,b).

In Australia and New Zealand a number of advances now have helped define the

‘sustainable nutrient loading’ for particular waterbodies. For example, sustainable

total phosphorus loads for the River Murray have been determined using a

simplified Vollenweider model;c Harris et al. (1996) estimated the sustainable

nutrient loads to Port Phillip Bay with particular emphasis on nitrogen; and

sustainable nutrient loading rates have been recommended for several Western

Australian estuaries and the coastal waters near Perth (Masini et al. 1992, 1994,

WAWA 1995, WADEP 1996).

a Section 3.1.5

b Case Studies

1 & 2 in Section

3.3.3; Case

Study 3 in

Section 8.2.3 in

Vol. 2

c See also

Case Study 4 in

Section 8.2.3,

Vol. 2
3.3.3 Guideline packages for applying the guideline trigger values to sites

Version — October 2000 page 3.3–25

Most of the models used to estimate sustainable loads rely on empirical

relationships between phosphorus or nitrogen loads and chlorophyll a

concentration. For example, Cary et al. (1995) found a significant linear

relationship between the known externally-derived summer inorganic nitrogen

loads to Cockburn Sound, WA, and the mean chlorophyll a concentration over a 13

year period. This relationship was used to define a total external nitrogen loading

of 2030 kgN/d needed to sustain a target chlorophyll a concentration of 0.8 μg/L

(WADEP 1996). Similarly, ‘sustainable’ total phosphorus loads in various sections

of the River Murray system have been defined by relating the annual TP load to the

water residence time in a particular reservoir or weir pool to estimate the TP

concentration during the summer growth period. Then using published (or

empirically derived) TP vs Chl a relationships, the chlorophyll a concentration that

would result from a particular TP load has been predicted. Using this information,

it has been possible to define a TP load for that waterbody that will sustain a

particular target chlorophyll a concentration.

3.3.3.3 Issue: Lack of dissolved oxygen

Background

Low dissolved oxygen (DO) concentration has an adverse effect on many aquatic

organisms (e.g. fish, invertebrates and microorganisms) which depend upon

oxygen dissolved in the water for efficient functioning. It can also cause reducing

conditions in sediments, so the sediments release previously-bound nutrients and

toxicants to the water column where they may add to existing problems.

The concentration of DO is highly dependent on temperature, salinity, biological

activity (microbial, primary production) and rate of transfer from the atmosphere.

Under natural conditions, DO will change, sometimes considerably, over a daily

(or diurnal) period, and highly productive systems (e.g. tropical wetlands, dune

lakes and estuaries) can become severely depleted in DO, particularly when these

systems are stratified.

Of greater concern is the significant decrease in DO that can occur when organic

matter is added (e.g. from sewage effluent or dead plant material). The depletion of

DO depends on the load of biodegradable organic material and microbial activity,

and re-aeration mechanisms operating. A number of predictive computer models

now exist for estimating the DO depletion in a particular ecosystem type, and so it

should be possible to estimate sustainable loads of biodegradable organic matter

for most situations.

The 1992 ANZECC Guidelines recommended that dissolved oxygen should not

normally be permitted to fall below 6 mgL–1 or 80−90% saturation, determined

over at least one diurnal cycle. These guidelines were based almost exclusively on

overseas data, since there were very few data on the oxygen tolerance of Australian

or New Zealand aquatic organisms. The Australian data are restricted to freshwater

fish, and suggest that DO concentrations below 5 mgL–1 are stressful to many

species (Koehn & O’Connor 1990).

Key indicators

Condition indicators: variation in DO concentration; species composition

Key stressor indicator: loading of biodegradable organic matter

(BOM, kg m–2 d–1)
Chapter 3 — Aquatic ecosystems

page 3.3–26 Version — October 2000

Modifiers: depend upon the ecosystem type, and include mixing

condition (atmospheric O2 transfer), photosynthetic O2

production, rate of microbial decomposition, flow,

temperature, pre-loading DO, mass of other O2

consuming materials (e.g. nitrate)

Performance indicators: median (or mean) DO concentration13 measured

under low flow conditions for rivers and streams and

during low flow and high temperature periods for

other ecosystems.

Low-risk trigger values

The method used to determine the low-risk trigger values will depend upon the

desired level of protection.a

Slightly to moderately disturbed ecosystems (condition 2 ecosystems)

Depending upon the significance and present condition of the ecosystem, two

approaches may be taken to derive the most appropriate trigger values for

condition 2 ecosystems.

a) For important ecosystems, where an appropriate reference system(s) is available,

and there are sufficient resources to collect the necessary information for the

reference system, the low-risk trigger concentrations for DO should be determined

as the 20th percentile of the reference system(s) distribution. Where possible the

trigger value should be obtained for low flow conditions for rivers and streams and

during low flow and high temperature periods for other ecosystems, when DO

concentrations are likely to be at their lowest.

b) The default trigger values contained in tables 3.3.2, 3.3.4, 3.3.6, 3.3.8 and 3.3.10

should be used where either an appropriate reference system is not available, or

the scale of the operation makes it difficult to justify the allocation of resources

to collect the necessary information on a reference system.

Highly disturbed ecosystems (condition 3 ecosystems)

a) For important waterbodies, and those in very poor condition, it is best to make

appropriate site-specific scientific studies, and to use the information, with

professional judgement and other relevant information, to derive trigger values.

Where local but higher-quality reference data are used, a less stringent cutoff

than the 20th percentile value may be used. The 20th percentile values, however,

should be used as a target for site improvement.

b) For highly disturbed waterbodies, where there is a lack of either information or

resources to undertake the necessary site-specific studies, it is best to use the

default, regional trigger values using professional judgement to derive a less

stringent value if this is agreed upon by stakeholders.

Sustainable loading rates for biodegradable organic matter should be estimated for

each major ecosystem type, and used to develop load-based trigger values.b

 

13 The median DO concentration for the period should be calculated using the lowest diurnal DO

concentrations.

a See Section

3.3.2.3

b See

recommendations

in Section 8.5.2,

Volume 2
3.3.3 Guideline packages for applying the guideline trigger values to sites

Version — October 2000 page 3.3–27

Use of the guideline package

Figure 3.3.1 shows the recommended approach for determining the risk of

dissolved oxygen depletion occurring in a particular ecosystem. The approach

involves three steps.

• Test the performance indicator (DO concentration) for the particular ecosystem

against the appropriate low-risk trigger value for that ecosystem type. Compare

the trigger values with the median (or mean) DO concentration measured under

low flow conditions for rivers and streams and during low flow and high

temperature periods for other ecosystems.

• If the test values are greater than the trigger values, there is low risk of adverse

biological effects occurring and no further action is required, except for regular

monitoring of the key performance indicators and condition indicators. If after

regular monitoring a ‘low risk’ outcome is consistently obtained, there is scope to

refine the guideline trigger value.a If test values are lower than trigger values,

there is an increased risk that adverse biological effects will occur, and further

ecosystem-specific investigation is required.

• Investigations to determine the influence of ecosystem-specific factors on the

key stressors will depend upon the ecosystem type. A possible approach to

calculate the sustainable load of biodegradable organic matter to waterbodies is

provided by Lawrence (1997 a,b).b

a See Section

3.1.5

b See also

Case Study 2

below
Chapter 3 — Aquatic ecosystems

page 3.3–28 Version — October 2000

Case Study 1. Assessing the risk of cyanobacterial blooms in a lowland river

We present here an example of the use of a rather simple but effective decision tree, for assessing the risk

of algal blooms arising from nutrients released to a lowland river in irrigation return drains. The protocol

was initially developed as part of an environmental audit protocol developed for Goulburn-Murray Water

(Hart et al. 1997; SKM 1997). More complex (and significantly more expensive) models have been

developed for Port Phillip Bay (Harris et al. 1996), Hawkesbury-Nepean river (Sydney Water 1995) and the

coastal waters off Perth (WAWA 1995, WADEP 1996).

The conceptual model for this case study (see figure below) assumes that algal growth in lowland rivers is

controlled by three major factors:

• the concentrations of the nutrients P and N;

• the light climate (turbidity is used as a surrogate for light intensity because of a lack of data);

• the flow conditions in the river that are required for algal growth to occur.

Low –

medium

risk

Test nutrient

conditions

Low

risk

Yes Test light

conditions

(turbidity)

No

No Test flow

conditions

Yes

Undertake detailed

study of system

Are there one Yes or more ‘growth

events’ of > 6 days

duration?

No Medium –

low risk

High risk

Is

TP > 15 μg/L

or TN > 150 μg/L?

Is turbidity

 > 30 NTU?

The ‘guideline package’ in this case includes values for the nutrient concentrations (TP, TN) as the key

stressors, and values for turbidity and flow as the modifiers. The numbers provided in the decision boxes

for TP, TN and turbidity should be taken as indicative only because they will depend upon the particular

ecosystem being considered.

The decision box for flow was based on the requirement that there be a sufficient period of low flow to

allow algal numbers to increase to an alert level of 5000 cells mL–1. A period of 6−10 days was estimated,

based on an algal doubling time of 2 days and an initial algal concentration of 10−100 cells mL–1. A ‘growth

event’ was then defined as a period consisting of at least 6 consecutive days when the flow was less than

the 25th percentile flow obtained from the long term flow record for the system.

For the system in the figure, a high risk situation is indicated if the TP concentration is >15 μgL–1, the

turbidity less than 30 NTU, and there is more than one ‘growth event’ of >6 days duration per year. In this

case, further investigation and appropriate management actions would be warranted.

Further refinement of this simple model could include:

• determining a more quantitative relationship between turbidity and the light climate for algal growth;

• validation of the assumption that the <25th percentile flows are the most appropriate low flow

conditions to use. The present simple protocol does not take into consideration stratification that is

now known to have a significant influence on cyanobacterial growth in lowland rivers (Webster et al.

1996);

• introduction of measures of the ‘bioavailable’ fractions of the nutrients rather than TP and TN

(Hart et al. 1998);

• including the possibility that sediment release of nutrients (particularly phosphorus) may occur under

low flow conditions;

• incorporation of the various decision ‘rules’ into a user-friendly computer program for ease of use by

managers.
3.3.3 Guideline packages for applying the guideline trigger values to sites

Version — October 2000 page 3.3–29

Case Study 2. Establishing sustainable organic matter loads for standing

waterbodies

Australian research has shown that most rivers transport most water, suspended particulate matter,

nutrients and organic matter during a small number of high flow events (Cosser 1989, Harris & Baxter

1996). In standing waterbodies, these event-driven loads can be augmented by point source discharges,

decay of ‘in-lake’ algae, and releases from the sediments. High flow events are often followed by long

periods of low flow conditions, when rapid decomposition of sedimented organic material by benthic

bacteria can occur (Harris & Baxter 1996).

In many ecosystems, this sequence of events is quite normal and actually defines the ecosystem type.

However, problems arise when an excess supply of organic material leads to de-oxygenation of the water

column and to remobilisation of sediment-bound nutrients (and possibly toxic heavy metals) in bioavailable

forms.

These processes may be accelerated if there is reduced transfer of oxygen from the atmosphere to the

water column resulting from thermal stratification during the low flow and calm wind conditions typical of

summer (Webster et al. 1996). This potential release of sediment-bound nutrients to the water column is of

concern because by far the largest amount of phosphorus is stored in the sediments.

Thus, controls on the loading of organic matter to waterbodies is crucial in the effective management of the

biological health and other uses of these waterbodies and, in particular, in controlling both dissolved

oxygen concentrations and the remobilisation of nutrients from anaerobic sediments.

In terms of the approach proposed in these Guidelines, a possible method for establishing sustainable

loads of organic matter to reservoirs, lakes and weir pools (and estuaries) is shown below (see also

Lawrence 1997 a,b).

Select key biological indicator and management targets

• Chlorophyll a conc <10μgL-1 for 9 in 10 years

• Dissolved oxygen concentration >60% saturation for 9 in 10 years

Identify key stressor and key performance indicator

• Key stressor Organic matter (BOD)

• Key performance indicator BOD loading (kg.m-2.yr-1)

Determine trigger value for key stressor

• Develop models relating BOD loading to water column DO

concentration and sediment nutrient release for range of

waterbodies and sediment types

• Validate model relationships using local reference and impacted

sites for which data are available

• Use models to determine trigger values (sustainable loads) for key

waterbodies throughout the catchment

Version — October 2000 page 3.4–1

3.4 Water quality guidelines for toxicants

3.4.1 Introduction

This section provides guidance on the application of water quality guideline trigger

values for toxicants. Toxicants is a term used for chemical contaminants that have

the potential to exert toxic effects at concentrations that might be encountered in

the environment. The risk-based decision scheme (Section 3.4.3) would be most

commonly applied in ecosystems that could be classified as slightly to moderately

disturbed (condition 2 ecosystemsa). The decision scheme, which is optional,

guides water managers on how to alter the trigger values for specific sites to

account for local environmental conditions.

The current NWQMS approach recommends moving away from relying solely on

chemical guideline values for managing water quality, to the use of integrated

approaches, comprising:

• chemical-specific guidelines coupled with water quality monitoring;

• direct toxicity assessment; and

• biological monitoring.

This approach will help to ensure that the water management focus keeps in view

the goal of protecting the environment, and does not shift to merely meeting the

numbers.

If more details are required, users may consult Volume 2 Section 8.3.2 on the type

of data used to derive guidelines, Section 8.3.3 on the general approaches and

methods used, Section 8.3.4 on the derivation procedure and requirements for data,

and Section 8.3.5 on application of the decision scheme. Section 8.3.6 provides

more information on direct toxicity assessment (i.e. whole effluent and ambient

water toxicity testing) and Section 8.3.7 outlines the data used to derive each

trigger value and summarises relevant scientific and technical information

currently available.

3.4.2 How guidelines are developed for toxicants

Numerical guidelines are an essential tool for the management of receiving waters

where discharge of toxicants to the environment cannot reasonably be avoided. The

guidelines aim to protect ambient waters from sustained exposures to toxicants,

i.e. from chronic toxicity. The derived trigger values are chemical-specific estimates

to help managers achieve this aim.

Most users of these guidelines will use the trigger values (table 3.4.1) either

directly or as part of the risk-based decision scheme outlined in Section 3.4.3, and

in most cases will not need to know how the figures were derived. However, a brief

summary is provided here.

3.4.2.1 Toxicity data for deriving guideline trigger values

The preferred data for deriving trigger values come from multiple-species toxicity

tests, i.e. field or model ecosystem (mesocosm) tests that represent the complex

interactions of species in the field. However, many of these tests are difficult to

interpret and there were few such data available that met screening requirements.

a See Section

3.1.3
Chapter 3 — Aquatic ecosystems

page 3.4–2 Version — October 2000

Most of the trigger values have been derived using data from single-species

toxicity tests on a range of test species, because these formed the bulk of the

concentration–response information. High reliability trigger valuesa were

calculated from chronic ‘no observable effect concentration’ (NOEC) data.

However the majority of trigger values were moderate reliability trigger values,

derived from short-term acute toxicity data (from tests ≤96 h duration) by applying

acute-to-chronic conversion factors.

3.4.2.2 Extrapolating from laboratory data to protect aquatic ecosystems

Most reliable trigger values (table 3.4.1) were derived using a statistical

distribution approach, modified from Aldenberg and Slob (1993). This approachb

has been adopted in The Netherlands and is recommended by the OECD (1992,

1995). The approach is based on calculations of a probability distribution of

aquatic toxicity end-points. It attempts to protect a pre-determined percentage of

species, usually 95%, but enables quantitative alteration of protection levels. The

95 percent protection level is most commonly applied in these Guidelines to

ecosystems that could be classified as slightly to moderately disturbed.

The traditional approach for extrapolating from single-species toxicity data to

protect ecosystems has been to apply arbitrary assessment factors to the lowest

toxicity value for a particular chemical (ANZECC 1992). There are deficiencies in

this approach (Warne 1998), and it has been used in the current Guidelines only

when there was an inadequate data set for the statistical distribution approach. The

smallest assessment factors (where they were used) were applied to a

comprehensive set of available chronic toxicity data, rather than acute data, when

there was a high degree of confidence that the values reflected the field situation.

The use of the statistically-based 95% protection provides a more defensible basis

for decisions than use of assessment factors.

For chemicals such as mercury, polychlorinated biphenyls (PCBs) and

organochlorine pesticides, the main issue of concern is not their direct short-term

toxic effect but the indirect risks associated with their longer-term concentration in

organisms and the potential for secondary poisoning. Dietary accumulation can be

an important route of uptake for some chemicals, and it will need to be addressed

in future revisions of the Guidelines. There is currently no formal and specific

international guidance for incorporating bioaccumulation into water quality

guidelines. For those chemicals that have the potential to bioaccumulate, the

decision scheme provides for site-specific re-assessment of this issue if suitable

data become available. Field investigations of residue levels in appropriate

organisms may provide additional evidence for whether or not bioaccumulation is

an issue at the site under study. In the absence of such local data, a higher level of

protection is recommended (e.g. 99% protection for slightly–moderately disturbed

systems instead of 95%). Chemicals that have the potential to bioaccumulate are

indicated in table 3.4.1 (footnote ‘B’).

3.4.2.3 Procedures for deriving trigger values for toxicants

Three grades of guideline trigger values are derived: high, moderate or low

reliability trigger values. The grade depends on the data available and hence the

confidence or reliability of the final figures (Warne 1998). Only high and moderate

reliability trigger values are reported in table 3.4.1.

a See Section

3.4.2.3

b Described in

Section 8.3.3.3

in Vol. 2
3.4.2 How guidelines are developed for toxicants

Version — October 2000 page 3.4–3

• High reliability guideline trigger values were derived from multiple-species

data or chronic NOEC data, using the risk-based statistical distribution method.

• Moderate reliability guideline trigger values, which reflect a lower confidence

in extrapolation methods, were derived from acute toxicity data. Again, where

possible, the statistical distribution method was used with the acute toxicity

data. It was then necessary to convert the figure from that calculation to a

chronic protection figure by application of either calculated or default acute-tochronic ratios.

• Low reliability guideline trigger values were derived, in the absence of a data

set of sufficient quantity, using larger assessment factors to account for greater

uncertainty. These are considered as interim or indicative working levels subject

to more test data becoming available. Low reliability figures should not be used

as default guidelines, although it is reasonable to use them in the risk-based

decision scheme to determine if conditions at the site increase or decrease the

potential risk. It is important to recognise the interim nature of the low reliability

figures and the inherent uncertainties in their derivation and to obtain more data

where appropriate. Hence they are only reported in Section 8.3.7.

It has not been possible to derive trigger values for every chemical. Section 8.3.4.5 of

Volume 2 provides some preliminary guidance for deriving preliminary working

levels for such chemicals, according to international guidance (OECD 1992, 1995).

3.4.2.4 Altering the level of protection for different ecosystem conditions

The trigger values derived using the statistical distribution method were calculated at

four different protection levels, 99%, 95%, 90% and 80% (table 3.4.1). Here,

protection level signifies the percentage of species expected to be protected. The

decision to apply a certain protection level to a specific ecosystem is the prerogative

of each particular state jurisdiction or catchment manager, in consultation with the

community and stakeholders. State jurisdictions or catchment managers can choose

to apply different levels of protection to different ecosystem conditions if there is

confidence that the disturbance is due to an overall physico-chemical disturbance and

not just structural alteration.

One way of viewing the continuum of disturbance is to apply the three ‘categories

of ecosystem condition’ for aquatic ecosystems, described in Section 3.1.3. The

recommended procedure for applying the different levels of protection to the

continuum of ecosystem conditions is summarised for toxicants in table 3.4.2. In

most cases, the 95% protection level trigger values (table 3.4.1) should apply to

ecosystems that could be classified as slightly–moderately disturbed, although a

higher protection level could be applied to slightly disturbed ecosystems where the

management goal is no change in biodiversity. For a few chemicals, higher levels

of protection are recommended as default levels for those ecosystems, and the

recommended trigger values for typical slightly–moderately disturbed ecosystems

are in the shaded boxes in table 3.4.1.

The highest protection level (99%) has been chosen as the default value for

ecosystems with high conservation value, pending collection of local chemical and

biological monitoring data. The 99% protection levels can also be used as default

values for slightly–moderately disturbed systems where local data are lacking on

bioaccumulation effects or where it is considered that the 95% protection level fails
Chapter 3 — Aquatic ecosystems

page 3.4–4 Version — October 2000

to protect key test species. This usually only occurs where trigger values have been

calculated from chronic data but fail to protect against acute toxicity or vice versa.

Those chemicals are shown in table 3.4.1. An example of this is endosulfan, with

which key Australian species show acute toxicity at or near the 95% protection

trigger value.

For ecosystems that can be classified as highly disturbed, the 95% protection

trigger values can still apply. However, depending on the state of the ecosystem,

the management goals and the approval of the appropriate state or regional

authority in consultation with the community, it can be appropriate to apply a less

stringent guideline trigger value, say protection of 90% of species, or perhaps even

80%. These values are provided as intermediate targets for water quality

improvement. If the trigger values have been calculated using assessment factors,

there is no reliable way to predict what changes in ecosystem protection are

provided by an arbitrary reduction in the factor.
3.4.2 How guidelines are developed for toxicants

Version — October 2000 page 3.4–5

Table 3.4.1 Trigger values for toxicants at alternative levels of protection. Values in grey shading are the trigger

values applying to typical slightly–moderately disturbed systems; see table 3.4.2 and Section 3.4.2.4 for guidance on

applying these levels to different ecosystem conditions.

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

METALS & METALLOIDS

Aluminium pH >6.5 27 55 80 150 ID ID ID ID

Aluminium pH <6.5 ID ID ID ID ID ID ID ID

Antimony ID ID ID ID ID ID ID ID

Arsenic (As III) 1 24 94 C 360 C ID ID ID ID

Arsenic (AsV) 0.8 13 42 140 C ID ID ID ID

Beryllium ID ID ID ID ID ID ID ID

Bismuth ID ID ID ID ID ID ID ID

Boron 90 370 C 680 C 1300 C ID ID ID ID

Cadmium H 0.06 0.2 0.4 0.8 C 0.7 B 5.5 B, C 14 B, C 36 B, A

Chromium (Cr III) H ID ID ID ID 7.7 27.4 48.6 90.6

Chromium (CrVI) 0.01 1.0 C 6 A 40 A 0.14 4.4 20 C 85 C

Cobalt ID ID ID ID 0.005 1 14 150 C

Copper H 1.0 1.4 1.8 C 2.5 C 0.3 1.3 3 C 8 A

Gallium ID ID ID ID ID ID ID ID

Iron ID ID ID ID ID ID ID ID

Lanthanum ID ID ID ID ID ID ID ID

Lead H 1.0 3.4 5.6 9.4 C 2.2 4.4 6.6 C 12 C

Manganese 1200 1900C 2500C 3600C ID ID ID ID

Mercury (inorganic) B 0.06 0.6 1.9 C 5.4 A 0.1 0.4 C 0.7 C 1.4 C

Mercury (methyl) ID ID ID ID ID ID ID ID

Molybdenum ID ID ID ID ID ID ID ID

Nickel H 8 11 13 17 C 7 70 C 200 A 560A

Selenium (Total) B 5 11 18 34 ID ID ID ID

Selenium (SeIV) B ID ID ID ID ID ID ID ID

Silver 0.02 0.05 0.1 0.2 C 0.8 1.4 1.8 2.6 C

Thallium ID ID ID ID ID ID ID ID

Tin (inorganic, SnIV) ID ID ID ID ID ID ID ID

Tributyltin (as μg/L Sn) ID ID ID ID 0.0004 0.006 C 0.02 C 0.05 C

Uranium ID ID ID ID ID ID ID ID

Vanadium ID ID ID ID 50 100 160 280

Zinc H 2.4 8.0 C 15 C 31 C 7 15 C 23 C 43 C

NON-METALLIC INORGANICS

Ammonia D 320 900 C 1430 C 2300 A 500 910 1200 1700

Chlorine E 0.4 3 6 A 13 A ID ID ID ID

Cyanide F 4 7 11 18 2 4 7 14

Nitrate J 17 700 3400 C 17000 A ID ID ID ID

Hydrogen sulfide G 0.5 1.0 1.5 2.6 ID ID ID ID

ORGANIC ALCOHOLS

Ethanol 400 1400 2400 C 4000 C ID ID ID ID

Ethylene glycol ID ID ID ID ID ID ID ID

Isopropyl alcohol ID ID ID ID ID ID ID ID

CHLORINATED ALKANES

Chloromethanes

Dichloromethane ID ID ID ID ID ID ID ID

Chloroform ID ID ID ID ID ID ID ID

Carbon tetrachloride ID ID ID ID ID ID ID ID

Chloroethanes

1,2-dichloroethane ID ID ID ID ID ID ID ID

1,1,1-trichloroethane ID ID ID ID ID ID ID ID
Chapter 3 — Aquatic ecosystems

page 3.4–6 Version — October 2000

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

1,1,2-trichloroethane 5400 6500 7300 8400 140 1900 5800 C 18000 C

1,1,2,2-tetrachloroethane ID ID ID ID ID ID ID ID

Pentachloroethane ID ID ID ID ID ID ID ID

Hexachloroethane B 290 360 420 500 ID ID ID ID

Chloropropanes

1,1-dichloropropane ID ID ID ID ID ID ID ID

1,2-dichloropropane ID ID ID ID ID ID ID ID

1,3-dichloropropane ID ID ID ID ID ID ID ID

CHLORINATED ALKENES

Chloroethylene ID ID ID ID ID ID ID ID

1,1-dichloroethylene ID ID ID ID ID ID ID ID

1,1,2-trichloroethylene ID ID ID ID ID ID ID ID

1,1,2,2-tetrachloroethylene ID ID ID ID ID ID ID ID

3-chloropropene ID ID ID ID ID ID ID ID

1,3-dichloropropene ID ID ID ID ID ID ID ID

ANILINES

Aniline 8 250 A 1100 A 4800 A ID ID ID ID

2,4-dichloroaniline 0.6 7 20 60 C ID ID ID ID

2,5-dichloroaniline ID ID ID ID ID ID ID ID

3,4-dichloroaniline 1.3 3 6 C 13 C 85 150 190 260

3,5-dichloroaniline ID ID ID ID ID ID ID ID

Benzidine ID ID ID ID ID ID ID ID

Dichlorobenzidine ID ID ID ID ID ID ID ID

AROMATIC HYDROCARBONS

Benzene 600 950 1300 2000 500 C 700 C 900 C 1300 C

Toluene ID ID ID ID ID ID ID ID

Ethylbenzene ID ID ID ID ID ID ID ID

o-xylene 200 350 470 640 ID ID ID ID

m-xylene ID ID ID ID ID ID ID ID

p-xylene 140 200 250 340 ID ID ID ID

m+p-xylene ID ID ID ID ID ID ID ID

Cumene ID ID ID ID ID ID ID ID

Polycyclic Aromatic Hydrocarbons

Naphthalene 2.5 16 37 85 50 C 70 C 90 C 120 C

Anthracene B ID ID ID ID ID ID ID ID

Phenanthrene B ID ID ID ID ID ID ID ID

Fluoranthene B ID ID ID ID ID ID ID ID

Benzo(a)pyrene B ID ID ID ID ID ID ID ID

Nitrobenzenes

Nitrobenzene 230 550 820 1300 ID ID ID ID

1,2-dinitrobenzene ID ID ID ID ID ID ID ID

1,3-dinitrobenzene ID ID ID ID ID ID ID ID

1,4-dinitrobenzene ID ID ID ID ID ID ID ID

1,3,5-trinitrobenzene ID ID ID ID ID ID ID ID

1-methoxy-2-nitrobenzene ID ID ID ID ID ID ID ID

1-methoxy-4-nitrobenzene ID ID ID ID ID ID ID ID

1-chloro-2-nitrobenzene ID ID ID ID ID ID ID ID

1-chloro-3-nitrobenzene ID ID ID ID ID ID ID ID

1-chloro-4-nitrobenzene ID ID ID ID ID ID ID ID

1-chloro-2,4-dinitrobenzene ID ID ID ID ID ID ID ID

1,2-dichloro-3-nitrobenzene ID ID ID ID ID ID ID ID

1,3-dichloro-5-nitrobenzene ID ID ID ID ID ID ID ID

1,4-dichloro-2-nitrobenzene ID ID ID ID ID ID ID ID

2,4-dichloro-2-nitrobenzene ID ID ID ID ID ID ID ID
3.4.2 How guidelines are developed for toxicants

Version — October 2000 page 3.4–7

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

1,2,4,5-tetrachloro-3-nitrobenzene ID ID ID ID ID ID ID ID

1,5-dichloro-2,4-dinitrobenzene ID ID ID ID ID ID ID ID

1,3,5-trichloro-2,4-dinitrobenzene ID ID ID ID ID ID ID ID

1-fluoro-4-nitrobenzene ID ID ID ID ID ID ID ID

Nitrotoluenes

2-nitrotoluene ID ID ID ID ID ID ID ID

3-nitrotoluene ID ID ID ID ID ID ID ID

4-nitrotoluene ID ID ID ID ID ID ID ID

2,3-dinitrotoluene ID ID ID ID ID ID ID ID

2,4-dinitrotoluene 16 65 C 130 C 250 C ID ID ID ID

2,4,6-trinitrotoluene 100 140 160 210 ID ID ID ID

1,2-dimethyl-3-nitrobenzene ID ID ID ID ID ID ID ID

1,2-dimethyl-4-nitrobenzene ID ID ID ID ID ID ID ID

4-chloro-3-nitrotoluene ID ID ID ID ID ID ID ID

Chlorobenzenes and Chloronaphthalenes

Monochlorobenzene ID ID ID ID ID ID ID ID

1,2-dichlorobenzene 120 160 200 270 ID ID ID ID

1,3-dichlorobenzene 160 260 350 520 C ID ID ID ID

1,4-dichlorobenzene 40 60 75 100 ID ID ID ID

1,2,3-trichlorobenzene B 3 10 16 30 C ID ID ID ID

1,2,4-trichlorobenzene B 85 170C 220C 300C 20 80 140 240

1,3,5-trichlorobenzene B ID ID ID ID ID ID ID ID

1,2,3,4-tetrachlorobenzene B ID ID ID ID ID ID ID ID

1,2,3,5-tetrachlorobenzene B ID ID ID ID ID ID ID ID

1,2,4,5-tetrachlorobenzene B ID ID ID ID ID ID ID ID

Pentachlorobenzene B ID ID ID ID ID ID ID ID

Hexachlorobenzene B ID ID ID ID ID ID ID ID

1-chloronaphthalene ID ID ID ID ID ID ID ID

Polychlorinated Biphenyls (PCBs) & Dioxins

Capacitor 21 B ID ID ID ID ID ID ID ID

Aroclor 1016 B ID ID ID ID ID ID ID ID

Aroclor 1221 B ID ID ID ID ID ID ID ID

Aroclor 1232 B ID ID ID ID ID ID ID ID

Aroclor 1242 B 0.3 0.6 1.0 1.7 ID ID ID ID

Aroclor 1248 B ID ID ID ID ID ID ID ID

Aroclor 1254 B 0.01 0.03 0.07 0.2 ID ID ID ID

Aroclor 1260 B ID ID ID ID ID ID ID ID

Aroclor 1262 B ID ID ID ID ID ID ID ID

Aroclor 1268 B ID ID ID ID ID ID ID ID

2,3,4’-trichlorobiphenyl B ID ID ID ID ID ID ID ID

4,4’-dichlorobiphenyl B ID ID ID ID ID ID ID ID

2,2’,4,5,5’-pentachloro-1,1’-biphenylB ID ID ID ID ID ID ID ID

2,4,6,2’,4’,6’-hexachlorobiphenyl B ID ID ID ID ID ID ID ID

Total PCBs B ID ID ID ID ID ID ID ID

2,3,7,8-TCDD B ID ID ID ID ID ID ID ID

PHENOLS and XYLENOLS

Phenol 85 320 600 1200 C 270 400 520 720

2,4-dimethylphenol ID ID ID ID ID ID ID ID

Nonylphenol ID ID ID ID ID ID ID ID

2-chlorophenol T 340 C 490 C 630 C 870 C ID ID ID ID

3-chlorophenol T ID ID ID ID ID ID ID ID

4-chlorophenol T 160 220 280 C 360 C ID ID ID ID

2,3-dichlorophenol T ID ID ID ID ID ID ID ID

2,4-dichlorophenol T 120 160 C 200 C 270 C ID ID ID ID
Chapter 3 — Aquatic ecosystems

page 3.4–8 Version — October 2000

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

2,5-dichlorophenol T ID ID ID ID ID ID ID ID

2,6-dichlorophenol T ID ID ID ID ID ID ID ID

3,4-dichlorophenol T ID ID ID ID ID ID ID ID

3,5-dichlorophenol T ID ID ID ID ID ID ID ID

2,3,4-trichlorophenol T ID ID ID ID ID ID ID ID

2,3,5-trichlorophenol T ID ID ID ID ID ID ID ID

2,3,6-trichlorophenol T ID ID ID ID ID ID ID ID

2,4,5-trichlorophenol T,B ID ID ID ID ID ID ID ID

2,4,6-trichlorophenol T,B 3 20 40 95 ID ID ID ID

2,3,4,5-tetrachlorophenol T,B ID ID ID ID ID ID ID ID

2,3,4,6- tetrachlorophenol T,B 10 20 25 30 ID ID ID ID

2,3,5,6- tetrachlorophenol T,B ID ID ID ID ID ID ID ID

Pentachlorophenol T,B 3.6 10 17 27 A 11 22 33 55 A

Nitrophenols

2-nitrophenol ID ID ID ID ID ID ID ID

3-nitrophenol ID ID ID ID ID ID ID ID

4-nitrophenol ID ID ID ID ID ID ID ID

2,4-dinitrophenol 13 45 80 140 ID ID ID ID

2,4,6-trinitrophenol ID ID ID ID ID ID ID ID

ORGANIC SULFUR COMPOUNDS

Carbon disulfide ID ID ID ID ID ID ID ID

Isopropyl disulfide ID ID ID ID ID ID ID ID

n-propyl sulfide ID ID ID ID ID ID ID ID

Propyl disulfide ID ID ID ID ID ID ID ID

Tert-butyl sulfide ID ID ID ID ID ID ID ID

Phenyl disulfide ID ID ID ID ID ID ID ID

Bis(dimethylthiocarbamyl)sulfide ID ID ID ID ID ID ID ID

Bis(diethylthiocarbamyl)disulfide ID ID ID ID ID ID ID ID

2-methoxy-4H-1,3,2-

benzodioxaphosphorium-2-sulfide

ID ID ID ID ID ID ID ID

Xanthates

Potassium amyl xanthate ID ID ID ID ID ID ID ID

Potassium ethyl xanthate ID ID ID ID ID ID ID ID

Potassium hexyl xanthate ID ID ID ID ID ID ID ID

Potassium isopropyl xanthate ID ID ID ID ID ID ID ID

Sodium ethyl xanthate ID ID ID ID ID ID ID ID

Sodium isobutyl xanthate ID ID ID ID ID ID ID ID

Sodium isopropyl xanthate ID ID ID ID ID ID ID ID

Sodium sec-butyl xanthate ID ID ID ID ID ID ID ID

PHTHALATES

Dimethylphthalate 3000 3700 4300 5100 ID ID ID ID

Diethylphthalate 900 1000 1100 1300 ID ID ID ID

Dibutylphthalate B 9.9 26 40.2 64.6 ID ID ID ID

Di(2-ethylhexyl)phthalate B ID ID ID ID ID ID ID ID

MISCELLANEOUS INDUSTRIAL CHEMICALS

Acetonitrile ID ID ID ID ID ID ID ID

Acrylonitrile ID ID ID ID ID ID ID ID

Poly(acrylonitrile-co-butadiene-costyrene)

200 530 800 C 1200 C 200 250 280 340

Dimethylformamide ID ID ID ID ID ID ID ID

1,2-diphenylhydrazine ID ID ID ID ID ID ID ID

Diphenylnitrosamine ID ID ID ID ID ID ID ID

Hexachlorobutadiene ID ID ID ID ID ID ID ID

Hexachlorocyclopentadiene ID ID ID ID ID ID ID ID
3.4.2 How guidelines are developed for toxicants

Version — October 2000 page 3.4–9

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

Isophorone ID ID ID ID ID ID ID ID

ORGANOCHLORINE PESTICIDES

Aldrin B ID ID ID ID ID ID ID ID

Chlordane B 0.03 0.08 0.14 0.27 C ID ID ID ID

DDE B ID ID ID ID ID ID ID ID

DDT B 0.006 0.01 0.02 0.04 ID ID ID ID

Dicofol B ID ID ID ID ID ID ID ID

Dieldrin B ID ID ID ID ID ID ID ID

Endosulfan B 0.03 0.2 A 0.6 A 1.8 A 0.005 0.01 0.02 0.05 A

Endosulfan alpha B ID ID ID ID ID ID ID ID

Endosulfan beta B ID ID ID ID ID ID ID ID

Endrin B 0.01 0.02 0.04 C 0.06 A 0.004 0.008 0.01 0.02

Heptachlor B 0.01 0.09 0.25 0.7 A ID ID ID ID

Lindane 0.07 0.2 0.4 1.0 A ID ID ID ID

Methoxychlor B ID ID ID ID ID ID ID ID

Mirex B ID ID ID ID ID ID ID ID

Toxaphene B 0.1 0.2 0.3 0.5 ID ID ID ID

ORGANOPHOSPHORUS PESTICIDES

Azinphos methyl 0.01 0.02 0.05 0.11 A ID ID ID ID

Chlorpyrifos B 0.00004 0.01 0.11 A 1.2 A 0.0005 0.009 0.04A 0.3 A

Demeton ID ID ID ID ID ID ID ID

Demeton-S-methyl ID ID ID ID ID ID ID ID

Diazinon 0.00003 0.01 0.2 A 2 A ID ID ID ID

Dimethoate 0.1 0.15 0.2 0.3 ID ID ID ID

Fenitrothion 0.1 0.2 0.3 0.4 ID ID ID ID

Malathion 0.002 0.05 0.2 1.1 A ID ID ID ID

Parathion 0.0007 0.004 C 0.01 C 0.04 A ID ID ID ID

Profenofos B ID ID ID ID ID ID ID ID

Temephos B ID ID ID ID 0.0004 0.05 0.4 3.6 A

CARBAMATE & OTHER PESTICIDES

Carbofuran 0.06 1.2 A 4 A 15 A ID ID ID ID

Methomyl 0.5 3.5 9.5 23 ID ID ID ID

S-methoprene ID ID ID ID ID ID ID ID

PYRETHROIDS

Deltamethrin ID ID ID ID ID ID ID ID

Esfenvalerate ID 0.001* ID ID ID ID ID ID

HERBICIDES & FUNGICIDES

Bypyridilium herbicides

Diquat 0.01 1.4 10 80 A ID ID ID ID

Paraquat ID ID ID ID ID ID ID ID

Phenoxyacetic acid herbicides

MCPA ID ID ID ID ID ID ID ID

2,4-D 140 280 450 830 ID ID ID ID

2,4,5-T 3 36 100 290 A ID ID ID ID

Sulfonylurea herbicides

Bensulfuron ID ID ID ID ID ID ID ID

Metsulfuron ID ID ID ID ID ID ID ID

Thiocarbamate herbicides

Molinate 0.1 3.4 14 57 ID ID ID ID

Thiobencarb 1 2.8 4.6 8 C ID ID ID ID

Thiram 0.01 0.2 0.8 C 3 A ID ID ID ID

Triazine herbicides

Amitrole ID ID ID ID ID ID ID ID

Atrazine 0.7 13 45 C 150 C ID ID ID ID
Chapter 3 — Aquatic ecosystems

page 3.4–10 Version — October 2000

Chemical Trigger values for freshwater

(μgL-1)

Trigger values for marine water

(μgL-1)

Level of protection (% species) Level of protection (% species)

99% 95% 90% 80% 99% 95% 90% 80%

Hexazinone ID ID ID ID ID ID ID ID

Simazine 0.2 3.2 11 35 ID ID ID ID

Urea herbicides

Diuron ID ID ID ID ID ID ID ID

Tebuthiuron 0.02 2.2 20 160 C ID ID ID ID

Miscellaneous herbicides

Acrolein ID ID ID ID ID ID ID ID

Bromacil ID ID ID ID ID ID ID ID

Glyphosate 370 1200 2000 3600 A ID ID ID ID

Imazethapyr ID ID ID ID ID ID ID ID

Ioxynil ID ID ID ID ID ID ID ID

Metolachlor ID ID ID ID ID ID ID ID

Sethoxydim ID ID ID ID ID ID ID ID

Trifluralin B 2.6 4.4 6 9 A ID ID ID ID

GENERIC GROUPS OF CHEMICALS

Surfactants

Linear alkylbenzene sulfonates (LAS) 65 280 520 C 1000 C ID ID ID ID

Alcohol ethoxyolated sulfate (AES) 340 650 850 C 1100 C ID ID ID ID

Alcohol ethoxylated surfactants (AE) 50 140 220 360 C ID ID ID ID

Oils & Petroleum Hydrocarbons ID ID ID ID ID ID ID ID

Oil Spill Dispersants

BP 1100X ID ID ID ID ID ID ID ID

Corexit 7664 ID ID ID ID ID ID ID ID

Corexit 8667 ID ID ID ID ID ID ID

Corexit 9527 ID ID ID ID 230 1100 2200 4400 A

Corexit 9550 ID ID ID ID ID ID ID ID

Notes: Where the final water quality guideline to be applied to a site is below current analytical practical quantitation limits, see Section 3.4.3.3 for

guidance.

Most trigger values listed here for metals and metalloids are High reliability figures, derived from field or chronic NOEC data (see 3.4.2.3 for reference to

Volume 2). The exceptions are Moderate reliability for freshwater aluminium (pH >6.5), manganese and marine chromium (III).

Most trigger values listed here for non-metallic inorganics and organic chemicals are Moderate reliability figures, derived from acute LC50 data (see

3.4.2.3 for reference to Volume 2). The exceptions are High reliability for freshwater ammonia, 3,4-DCA, endosulfan, chlorpyrifos, esfenvalerate,

tebuthiuron, three surfactants and marine for 1,1,2-TCE and chlorpyrifos.

* = High reliability figure for esfenvalerate derived from mesocosm NOEC data (no alternative protection levels available).

A = Figure may not protect key test species from acute toxicity (and chronic) — check Section 8.3.7 for spread of data and its significance. ‘A’ indicates

that trigger value > acute toxicity figure; note that trigger value should be <1/3 of acute figure (Section 8.3.4.4).

B = Chemicals for which possible bioaccumulation and secondary poisoning effects should be considered (see Sections 8.3.3.4 and 8.3.5.7).

C = Figure may not protect key test species from chronic toxicity (this refers to experimental chronic figures or geometric mean for species) — check

Section 8.3.7 for spread of data and its significance. Where grey shading and ‘C’ coincide, refer to text in Section 8.3.7.

D = Ammonia as TOTAL ammonia as [NH3-N] at pH 8. For changes in trigger value with pH refer to Section 8.3.7.2.

E = Chlorine as total chlorine, as [Cl]; see Section 8.3.7.2.

F = Cyanide as un-ionised HCN, measured as [CN]; see Section 8.3.7.2.

G = Sulfide as un-ionised H2S, measured as [S]; see Section 8.3.7.2.

H = Chemicals for which algorithms have been provided in table 3.4.3 to account for the effects of hardness. The values have been calculated using a

hardness of 30 mg/L CaCO3. These should be adjusted to the site-specific hardness (see Section 3.4.3).

J = Figures protect against toxicity and do not relate to eutrophication issues. Refer to Section 3.3 if eutrophication is the issue of concern.

ID = Insufficient data to derive a reliable trigger value. Users advised to check if a low reliability value or an ECL is given in Section 8.3.7.

T = Tainting or flavour impairment of fish flesh may possibly occur at concentrations below the trigger value. See Sections 4.4.5.3/3 and 8.3.7.
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–11

Table 3.4.2 General framework for applying levels of protection for toxicants to different

ecosystem conditions

Ecosystem

condition

Level of protection

1 High

conservation/

ecological

value

• For anthropogenic toxicants, detection at any concentration could be grounds

for source investigation and management intervention; for natural toxicants

background concentrations should not be exceeded.a

Where local biological or chemical data have not yet been gathered, apply the

99% protection levels (table 3.4.1) as default values.

Any relaxation of these objectives should only occur where comprehensive

biological effects and monitoring data clearly show that biodiversity would not

be altered.

• In the case of effluent discharges, Direct Toxicity Assessment (DTA) should

also be required on the effluent.

• Precautionary approach taken to assessment of post-baseline data through

trend analysis or feedback triggers.

2 Slightly to

moderately

disturbed

ecosystems

• Always preferable to use local biological effects data (including DTA) to derive

guidelines.

If local biological effects data unavailable, apply 95% protection levels (table

3.4.1) as default, low-risk trigger values.b 99% values are recommended for

certain chemicals as noted in table 3.4.1.c

• Precautionary approach may be required for assessment of post-baseline data

through trend analysis or feedback triggers.

• In the case of effluent discharges DTA may be required.

3 Highly

disturbed

ecosystems

• Apply the same guidelines as for slightly–moderately disturbed systems.

However, the lower protection levels provided in the Guidelines may be

accepted by stakeholders.

• DTA could be used as an alternative approach for deriving site-specific

guidelines.

a This means that indicator values at background and test sites should be statistically indistinguishable. It is

acknowledged that it may not be strictly possible to meet this criterion in every situation.

b For slightly disturbed ecosystems where the management goal is no change in biodiversity, users may prefer to

apply a higher protection level.

c 99% values recommended for chemicals that bioaccumulate or for which 95% provides inadequate protection for

key test species. Jurisdictions may choose 99% values for some ecosystems that are more towards the slightly

disturbed end of the continuum.

Modified values for this lowest level of protection should not approach levels that

may cause acute toxicity. Footnotes in table 3.4.1 indicate where the figures at any

protection level may cause acute or chronic toxicity but it is better to view the

chemical descriptionsa to gain the full context. The data distribution curvesb

illustrate the spread of the data (either acute or chronic) used to derive each trigger

value. As indicated above, the emphasis should be on improvement of the highly

disturbed ecosystem, not just maintenance of a degraded condition.

3.4.3 Applying guideline trigger values to sites

The guideline trigger values (table 3.4.1) were mostly derived primarily according to

risk assessment principles, using data from laboratory tests in clean water. They

represent the best current estimates of the concentrations of chemicals that should

have no significant adverse effects on the aquatic ecosystem. They focus on direct

toxic effects of individual chemicals, but it is intended that they be applied at specific

sites, where possible, using the decision tree. This does not imply that application of

the guidelines requires a full (quantitative) risk assessment.c

a See Section

8.3.7

b See toxicant

databases on

the CD-Rom

c See last

paragraph of

Section 2.1.4
Chapter 3 — Aquatic ecosystems

page 3.4–12 Version — October 2000

These trigger values should not be considered as blanket guidelines for national

water quality, because ecosystem types vary so widely throughout Australia and

New Zealand. Such variations, even on a smaller scale, can have marked effects on

the bioavailability, transport and degradation of chemicals, and on their toxicity.

The trigger values may not apply to every aquatic ecosystem in Australia or New

Zealand and in some instances adequate protection of the environment may require

less or in some cases more stringent values.

3.4.3.1 Underlying philosophy for applying the guidelines

The general approach to use of the decision scheme is outlined in Section 3.1.5. If a

trigger value listed in table 3.4.1 is exceeded at a site, further action results. The

action can be:

i. Incorporation of additional information or further site-specific investigation to

determine whether or not the chemical is posing a real risk to the environment.

The investigation may determine the fraction of the chemical in the water that

organisms can take up (the bioavailable fraction) to use for comparing with the

trigger value. The investigation and/or regular monitoring may also result in

refinement of the guideline figure to suit regional or local water quality

parameters and other conditions. Such refinement would occur where

exceedance of the trigger value was shown to have no adverse effects upon the

ecosystem; alternatively

ii. Accept the trigger value without change as a guideline applying to that site and

initiate management action or remediation.

The appropriate state or regional authority can often provide guidance and

direction for implementing the decision scheme. Even if no other steps of the

scheme are undertaken, it is important at least to adjust the trigger values for the six

hardness-related metals (tables 3.4.3 and 3.4.4) to account for the local water

hardness (step 9 of the scheme below). The trigger values for these metals have

been derived at low water hardness, corresponding to high toxicity. In some cases,

either the potential for environmental harm or the economic importance of the

chemical may be sufficiently significant to warrant more intensive work to define a

concentration that would adequately protect the environment.

Although the calculated site-specific guideline figure represents a concentration of

toxicant that will not degrade the environmental value at the site, it should not be

inferred that the environment could be contaminated up to this level (ANZECC 1992).

Where the site-specific guideline for a toxicant is exceeded, management action

will normally result. However, this should be addressed under the processes of the

individual states/territories or regions. It is important that toxicant guidelines are

not interpreted in isolation from other ecosystem factors such as habitat, flow etc,

as well as chemical fate. If the chemical is likely to be deposited in sediment, then

consult the sediment guidelines.a

In practice, not all site-specific data on a particular chemical are equivalent in

extent, detail or method. If a manager were to apply the strict requirements used in

deriving the original guideline trigger value, much valuable local information

would be lost. Differing site-specific trigger values developed using various

methods can be examined and weighted according to pre-determined criteria of

quality and relevance to the ecosystem. This should be done in a commonsense

a See Section

3.5
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–13

manner consistent with commonly applied principles of risk assessment to arrive at

an appropriate figure (e.g. Menzie et al. 1996). The result can provide water

managers with a way of integrating varying information on a particular site; if it is

provided during assessments by the proponent, it can assist in maintaining

consistent professional judgement. Inclusion of these multiple lines of evidence

strengthens the overall result.a

3.4.3.2 Decision tree for applying the guideline trigger values

The decision scheme outlined below gives step by step guidance on how to assess

test site data and tailor the guideline trigger values according to site-specific

environmental conditions. A simplified diagrammatic version of the decision tree is

shown in figure 3.4.1.b The decision scheme is not mandatory and at any time a

water manager can default to the original trigger value or use only those steps that

are relevant to the situation and chemical at hand. The scheme is designed to

determine if the conditions at a specific site reduce (or occasionally, increase) the

risk to the environment of the study chemical.

The process of deriving water quality guidelines for a specific site begins with

determination of the management aims, including a decision on the appropriate

level of protection.c The next step is to assess the factors at the site that modify

toxicity and bioavailability of the chemical. The measured or calculated

bioavailable fraction can then be compared with the trigger value, or in some cases

a site-specific guideline can be developed on the basis of known relationships

between some physical or chemical parameters and the original trigger value.

Examples of the latter include corrections for the effects of hardness for metals, the

effects of pH for ammonia, or the effects of temperature for other chemicals. In the

absence of quantitative data for such relationships, it may be possible to

qualitatively estimate the likely trends in toxicity of a chemical, and hence risk, at a

particular site. This is where professional judgement may be necessary,

strengthened by examining multiple lines of evidence.

Ultimately, it is biological measurement that will provide confirmation of the sitespecific guideline, so the decision scheme directs users to the option of direct

toxicity assessment (DTA) if the guideline is exceeded or if there is low confidence

in desktop assessments.d When no default trigger value is provided, where the

trigger value is not applicable to a specific site, or if the chemical is one of a

complex mixture, DTA is also useful. Further, DTA may provide the required link

between chemical levels and biological effects or establish concentrations that are

unlikely to cause adverse environmental effects. Field biological assessments can

be undertaken also.e

The stepwise procedure for applying the decision scheme is outlined below. The

cross-references to Volume 2 provide background information on each step. Sitespecific trigger values can be derived at each step and compared with the

concentration of chemical measured at the site. Note that at any stage the

stakeholders may wish to accept the lowest original or partially modified trigger

value and institute management actions to reduce contamination or pollution, if that

value is exceeded. However, if a trigger value is accepted without completing the

decision tree, the value may not be the most appropriate for the site.

a See Section

8.3.5.1

b Section

8.3.5.1

c Section 3.1.3

d Section 8.3.6

e Section 3.2
Chapter 3 — Aquatic ecosystems

page 3.4–14 Version — October 2000

 Determine appropriate guideline trigger

values for selected indicators (figure 3.1.1)

Test against guideline values

Compare contaminant concentration (total) with relevant 

guideline ‘trigger’ value

Low risk

Above Below

(Potential riskb)

Consider site-specific factors that may

modify the guideline trigger value,

calculate a site specific guideline

e.g.

• background

• analytical limits

• locally important species

• chemical/water quality modifiers

• mixture interactions

Test against guideline values

Compare contaminant concentration with new 

guideline value

Below

Low risk

Above

Perform biological effects 

assessment (e.g. DTA)

Non toxic Toxic

High risk

(initiate remedial actions)

Low risk

Decision tree framework for

applying guideline trigger values

for toxicantsa

Define primary management aims (figure 3.1.1)

Potential riskb

a Local biological effects data not required in the decision trees (see section 3.1.5)

b Further investigations are not mandatory; users may opt to proceed to management/remedial action.

Figure 3.4.1 Simplified decision tree for assessing toxicants in ambient waters

Application of the decision tree

1. On advice of the water management authority, select the appropriate target

ecosystem condition (Section 3.1.3) for the particular site or region under

study.a This may determine which trigger value is used.b Alternative levels of

protection are also given in table 3.4.1. The concept of three ecosystem

conditions in Section 3.1.3 is for management guidance only. Users need to

a See Section

8.3.5.2

b Section

3.4.2.4
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–15

view these as examples that represent a continuum of ecosystem conditions.

Table 3.4.2 summarises the approaches and default trigger values

recommended for each ecosystem condition. For highly disturbed (condition 3)

ecosystems, it may be appropriate to negotiate a lower level of protection for

toxicants in some instances and hence to use a less stringent trigger value for

ensuing calculations. Initial decisions are also made about whether the sample

is freshwater or saline because different trigger values may apply, and whether

the chemical is a metal, which may affect which of the following steps apply.

2. Collect and analyse water samples. Design, implement and organise the

logistics of sampling protocols, filter samples and mathematically process

data.a

Judgement on whether a chemical concentration exceeds a guideline value

should not rely on results of analysis of a single sample, except possibly if the

concentration is high enough to potentially cause acute toxicity. It is better to

collect a number of samples and to compare the median value with the

guideline value.

Should the samples be filtered in the field? Samples do not normally need to be

filtered unless the user is studying metals and considers that field filtration is

cost-effective. Often, users will find it easier and most economical to compare

total unfiltered concentrations initially. Comparison of total concentrations will,

at best, overestimate the fraction that is bioavailable. The major toxic effect of

metals comes from the dissolved fraction, so it is valid to filter samples (e.g. to

0.45 μm) and compare the filtered concentration against the trigger value. If

other measurements of metal bioavailability are being pursued (e.g. step 10),

filtration will be necessary but chemical preservation is not advised.

There are few bioavailability measurements for organic chemicals and expert

advice should be sought on the appropriateness of this step for organic

chemicals.

The present guidelines do not prescribe specific methods for chemical analyses.b

Users must satisfy themselves that analysis methods are appropriate and

sufficiently accurate, that the laboratories are suitably accredited and that quality

control procedures have been adhered to.

If users intend to follow this decision scheme, it will also be necessary to analyse

for the water quality parameters that may affect the chemical toxicity and hence

the site-specific trigger value. Measures of pH, organic carbon and hardness (e.g.

for metals) will also assist some steps.

3. Consider the analytical practical quantitation limit (PQL)14 using the best

available technology.c If the PQL is above the trigger value (i.e. PQL >TV)

there are three options, on advice of the appropriate state regulator:

i) accept that any validated detection implies that guidelines have been

exceeded; or

 

14 The practical quantitation limit (PQL) is the lowest level achievable among laboratories within

specified limits during routine laboratory operations. The PQL represents a practical and routinely

achievable detection level with a relatively good certainty that any reported value is reliable

(Clesceri et al. 1998). The PQL is often around 5 times the method detection limit.

a See Section

3.4.3.3; see

also Section

8.3.5.3 and the

Monitoring

Guidelines

b See the

Monitoring

Guidelines

c Section

8.3.5.4
Chapter 3 — Aquatic ecosystems

page 3.4–16 Version — October 2000

ii) examine the decision scheme to see if site-specific factors reduce the

environmental risk; or

iii) proceed directly to direct toxicity assessment (DTA) where one of the

following two approaches can be adopted:

• site-specific toxicity testing of the toxicant in question, using local

species under local conditions, to derive a site-specific trigger

value (step 7). Note that some judgement is required (ideally,

based on existing information) about whether adverse effects can

be expected at concentrations below the PQL, in which case this

option is not appropriate.

• DTA of the ambient water (step 12) to ascertain whether adverse

effects are being observed at the present concentration of toxicant.

If effects are observed, management action is initiated. This can

include the use of toxicity identification and evaluation (TIE)

techniques, which assist in identifying the unmeasured toxicant

source (Burkhard & Ankley 1989, Manning et al. 1993).a

Water regulators may also recommend DTA if the chemical cannot

be measured and the issue is of high significance.

4. Consider the natural background concentration (or range) of the toxicant at the

site.b This applies mostly for metals and some non-metallic inorganics. The

only organic chemicals to which this will commonly apply will be some

phenols or globally distributed contaminants such as DDT. Table 8.3.2

(Volume 2) provides some general literature guidance on commonly

encountered background levels. If background concentrations cannot be

measured at the site, measurement at an equivalent high-quality reference site

that is deemed to closely match the geology, natural water quality etc of the

site(s) of interest is suggested.

If the background concentration has been clearly established and it exceeds the

trigger value (it is preferable to compare filtered background concentrations for

metals), the 80th percentile of the background concentration can be accepted as

the site-specific trigger value for ensuing steps.c In addition, users may apply

DTA to background or reference waters (Step 12) using locally adapted

species, to confirm that there is no toxicity. In the unlikely event that adverse

effects are observed, management action must be initiated, and again this can

include the use of TIE to begin to identify the compound(s) causing toxicity.

5. Examine if transient exposure is relevant and if it can be incorporated into the

decision scheme.d At present, there is little international guidance on how to

incorporate brief exposures into guidelines, and it may not yet be possible to do

this. A number of chemicals can cause delayed toxic effects after brief exposures,

so it has been considered unwise to develop a second set of guideline numbers

based on acute toxicity to account for brief exposures. Concentrations at which

acute toxicity is likely to occure may not necessarily bear any resemblance to the

concentrations that should protect against transient exposure. New information

about transient exposure, published in the peer-reviewed literature, may assist

users to take transient exposure into account for some chemicals.

a See Section

8.3.6.3

b See Sections

7.4.4.2, 8.3.5.5;

table 8.3.2

c Section

7.4.4.2

e Section 8.3.7

d Section

8.3.5.6
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–17

6. Determine if the chemical bioaccumulates in organisms and if it is likely to

cause secondary poisoning (i.e. biomagnify).a For some chemicals (e.g.

mercury and PCBs), this is the main issue of concern, rather than direct effects

of toxicants.b Chemicals that have the potential to bioaccumulate and cause

harm are identified by ‘B’ in table 3.4.1. Some metals, such as copper, can

accumulate in shellfish without causing harm.

The decision scheme provides the opportunity to examine whether the identified

chemicals may actually be bioaccumulating at the study site. This can be

validated by relating tissue residues in local organisms to chemical levels in

water. If data are available, it may be possible to refine the trigger value to

account for these phenomena.c Alternatively the Canadian approach (CCME

1997) can give guidance on what levels of chemicals in food may accumulate in

water-associated wildlife.d Appendix 3, Method 1B(i) of Volume 2 may also

provide some guidance here. If there are no local data for such chemicals to

enable these approaches to be used, users are advised to apply the 99%

protection level trigger values for ecosystems that could be classified as slightly

to moderately disturbed. However, this derivation is precautionary, and is not

directly related to bioconcentration effects.

7. Consider whether there are locally important species or genera, either

ecologically or economically, which were not adequately evaluated in

calculating the original default trigger value. It will be necessary to examine

the original data set used to calculate the trigger value, available on the

enclosed CD-Rom (under the title, The ANZECC & ARMCANZ Water Quality

Guideline Database for Toxicants), insert any new and appropriate data and

recalculate the trigger value by the same method as used originally.e If

considering this step, seek expert advice. In most situations it is reasonable to

accept the original suite of test species as an adequate surrogate for untested

species in the environment but there may be specific instances where it is

worthwhile to consider particular species. In some cases it may be valid to

check whether the original trigger value has been calculated using species that

are locally inappropriate and if these data can be substituted by new data from

more appropriate species which have an equivalent role in the ecosystem. Data

should only be deleted in exceptional circumstances. It is important in all cases

to maintain the integrity of the trigger values by adhering to the requirements

for data quality and quantity. It is also important to ensure that a

comprehensive overseas data set is not substituted by a native data set that

does not cover the necessary breadth of taxa.

f

8. Consider whether chemical or water quality parameters at the site may

increase or decrease the toxicity of the chemical and hence potentially alter the

site-specific trigger value. This applies for organic or non-metallic inorganic

chemicals, as the hardness calculations for metalsg also cover all these

parameters except temperature and dissolved oxygen.

These parameters may include differences in the proprietary formulation of the

chemicalh and variations in water quality parametersi

 such as suspended

matter, dissolved organic matter, salinity, pH, temperature, hardness and

dissolved oxygen. Specific guidance on which parameters are known to affect

toxicity of each chemical is given in Section 8.3.7. In some cases, there are

simple numerical factors or algorithms linking the water quality parameter and

the toxicity of the chemical. If so, this can be applied to the original data or to

a See Section

8.3.5.7

b Section

8.3.3.4

c Section

8.3.3.4

e Section

8.3.5.8

d Section

8.3.3.4

f Section

8.3.4.2

g Section

8.3.5.15

h Section 8.3.5.9

i Sections

8.3.5.10 to

8.3.5.17 for detail

on respective

parameters
Chapter 3 — Aquatic ecosystems

page 3.4–18 Version — October 2000

the trigger value to derive a site-specific guideline that accounts for these

parameters, as below (using temperature as an example). Thus:

• Check back to the original data and apply factors to convert all the data to a

single (say) temperature that better represents the site. Re-calculate the sitespecific guideline according to the method used to derive the original trigger

value; or

• if all the original data have been calculated at a standard (say)

temperature, apply the factor directly to the trigger value.

Remember that when the parameter increases toxicity, the factor is <1 and

when it decreases toxicity, the factor is >1. Tables for temperature and/or pH

conversions are available in Volume 2 for ammonia, cyanide and sulfide. If

there is not, a simple quantitative relationship, seek expert advice. For

instance, the equilibrium between many organic chemicals and suspended

matter is poorly understood and requires well-designed studies, e.g. DTA

(Step 12) under appropriate conditions. It may be possible to make a

qualitative estimate of whether the parameters increase or decrease the risk.

9. For metals or metalloids in fresh waters (up to 2500 mgL-1 salinity), consider the

effect of hardness, pH and alkalinity on toxicity and derive a hardness-modified

trigger value (HMTV)a using the appropriate algorithm from table 3.4.3. Table

3.4.4 indicates how the trigger values vary with different ranges of hardness but

extra care is needed for waters with hardness below 25 mgL-1 CaCO3. If samples

have been filtered, for comparison with the HMTV, this will also take into

account suspended organic matter. The hardness algorithms (table 3.4.3) also

account for pH. The recommended decision scheme for metals is illustrated in

figure 3.4.2 but steps beyond the initial hardness adjustment are optional.

If the total metal concentration in the unfiltered sample exceeds the HMTV,

then users may choose one or more of four steps:

(i) compare metal concentration with the HMTV after filtering the original unacidified sample through a 0.45 μm membrane filter. An alternative is to

proceed directly to measuring filtered concentrations instead of totals initially.

(ii) proceed to more complex estimates of metal bioavailability (step 10)

relating to the study site;

(iii) accept that the guideline has been exceeded and institute management

action;

(iv) proceed to DTA (step 12).

10. Examine the concentration of the metal or metalloid to determine the

concentration of the bioavailable species, i.e. the concentration that is most

likely to exert a biological effect. This uses speciation modelling or chemical

techniques for metal speciation analysisb to account for the effects of factors

such as dissolved organic matter, pH and redox potential on the bioavailable

fraction of the metal. Seek professional advice for this step.

If the bioavailable metal concentration exceeds the HMTV or the trigger value

(if a hardness algorithm is not available), consider these two options, with

guidance from the regulatory authority:

• use direct toxicity assessment (DTA) to confirm the results or develop a

new site-specific guideline; or

• develop management options to reduce contamination.

a See Section

8.3.5.15

b Section

8.3.5.16
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–19

Above

(Potential risk)a

Test against guideline value

Compare metal total concentration (acid soluble) with relevant

guideline ‘trigger’ value

Low risk

Below

Correct guideline for hardness

Applies to freshwaters only

Test against new guideline value

Low risk

Above Below

(Potential risk)a

Test dissolved metal concentration

against new guideline value

(filtration through a 0.45 μm or smaller membrane)

Low risk Consider metal speciation

Speciation 

modelling

Chemical 

measurement

Biological effects

assessment (eg DTA)

Test ‘bioavailable’ metal concentration

against new guideline value

Above Below

Non toxic Toxic

High riska

High risk

(initiate remedial actions)

Low risk

Low risk

aFurther investigations are not mandatory; users may opt to proceed to management/remedial action

Above

(Potential risk)a

Below

Figure 3.4.2 Decision tree for metal speciation guidelines
Chapter 3 — Aquatic ecosystems

page 3.4–20 Version — October 2000

11. Consider the effect of mixtures and chemical interactions on overall toxicity.a If

the chemical occurs as a component of a simple mixture, and the mixture

interactions are simple and predictable (i.e. usually two–three components and

additive toxicity) the mixture toxicity can be modelled using the mixtures

equation in Section 8.3.5.18.

12. If there is any degree of complexity in the mixture interactions, proceed to direct

toxicity assessment (DTA) on the ambient waters at the site.b Use an

appropriate battery of test species and chronic end-points to ascertain whether

toxicity is being observed. If adverse effects are observed, initiate management

action and use TIE to assist in identifying the compound(s) that are causing

toxicity. Use DTA also to assess toxicity of ambient waters when background

levels are high (step 3), when guideline values are lower than analytical PQLs

(step 4), or to quantify the effects of water quality parameters or proprietary

formulations on the chemical toxicity (step 8).

Where a chemical is to be used in an environment of particular sociopolitical or ecological importance, it is better to undertake toxicity testing

with that chemical on species relevant to that environment (i.e. step 7). It is

best to do this before the chemical is introduced. Such data can be used to

develop new guideline values relevant to that region; for example, to collect

a suite of tropical data for a development affecting tropical freshwaters.

When using DTA to examine toxicity of a chemical to locally important

species (step 7) or for pre-release effluents (see table 3.4.2), determine

chronic effects at a range of concentrations of the chemical or effluent. For

dilution, use the local reference dilution waters. Determine NOEC values for

the chemical or effluent and use them for calculating site-specific guidelines.

The method used for these calculations will depend on the number of data

points, but use the statistical distribution method if the data requirements

have been met (at least five species from four different taxonomic groups).c

Otherwise it is best to divide the lowest chronic NOEC by 10. Follow the

general methods for calculation of trigger values.d

The DTA can comprise in situ field and/or laboratory ecotoxicity tests

(Chapman 1995), preferably chronic or sub-chronic tests on appropriate

species using local dilution waters, satisfying all sampling, test and analysis

conditions.e

To aid interpretation of results, analyse the chemicals concurrently with

biological assessment, unless there is a biological marker of toxicity.

For already existing discharges and for chemicals that have a high potential

to disturb the environment, it will be necessary to measure and assess the

biological health of potentially disturbed sites.f

b Section

8.3.5.19

c Section 8.3.4.2

d Section 8.3.4.4

a See Section

8.3.5.18

e Section 8.3.6

f Section 3.2
3.4.3 Applying guideline trigger values to sites

Version — October 2000 page 3.4–21

Table 3.4.3 General form of the hardness-dependent algorithms describing guideline values

for selected metals in freshwaters

Metal Hardness-dependent algorithm

Cadmium HMTV = TV (H/30)0.89

Chromium(III) HMTV = TV (H/30)0.82

Copper HMTV = TV(H/30)0.85

Lead HMTV = TV(H/30)1.27

Nickel HMTV = TV(H/30)0.85

Zinc HMTV = TV(H/30)0.85

HMTV, hardness-modified trigger value (μg/L); TV, trigger value (μg/L) at a hardness of 30 mg/L as

CaCO3; H, measured hardness (mg/L as CaCO3) of a fresh surface water (≤2.5‰). From Markich et

al (in press).

Table 3.4.4 Approximate factors to apply to soft water trigger values for selected metals in

freshwaters of varying water hardnessa

Hardness categoryb

(mg/L as CaCO3)

Water hardnessc

(mg/L as CaCO3)

Cd Cr(III) Cu Pb Ni Zn

Soft (0–59) 30 TV TV TV TV TV TV

Moderate (60–119) 90 X 2.7 X 2.5 X 2.5 X 4.0 X 2.5 X 2.5

Hard (120–179) 150 X 4.2 X 3.7 X 3.9 X 7.6 X 3.9 X 3.9

Very hard (180–240) 210 X 5.7 X 4.9 X 5.2 X 11.8 X 5.2 X 5.2

Extremely hard (400) 400 X 10.0 X 8.4 X 9.0 X 26.7 X 9.0 X 9.0

a Trigger values from table 3.4.1;

b Range of water hardness (mg/L as CaCO3) for each category as defined by CCREM (1987);

c Mid-range value of each water hardness category. For example, a copper trigger value of 1.4 μg/L (from table

3.4.1) with 95% protection level chosen (e.g. slightly–moderately disturbed system) is applied to a site with very

hard water (e.g. 210 mg/L as CaCO3) by multiplying the trigger value by 5.2 to give a site-specific trigger value of

7.3 μg/L. If the hardness is away from the mid-range, it may be preferable to use the algorithm.

3.4.3.3 Comparing monitoring data with trigger values

Wherever there is concern about toxicants in a waterbody, data must be gathered to see

if there are accompanying adverse ecological effects. This process has many steps, and

it is beyond the scope of these Guidelines to address all of them in detail. Those which

are not elaborated in Chapter 7 of this volume are discussed in detail in the Monitoring

Guidelines (ANZECC & ARMCANZ 2000). The purpose of this section is to direct

readers to the appropriate places to learn more about the necessary procedures for a

chemical monitoring program.

• The design of sampling protocols. The Monitoring Guidelines (Chapter 3)

advises on: study type, temporal and spatial considerations, site selection and

identification, sampling precision, timing and frequency, and considerations for

selecting indicators (measurement parameters).

• The implementation of sampling protocols. Chapter 4 of the Monitoring Guidelines

discusses procedural issues in sample acquisition. Specifically it addresses ways

for ensuring that samples are sufficiently numerous, well-documented and

representative, and with appropriate analytical integrity, to enable strong inferences

to be made about water quality. It also offers advice on logistical issues and OH&S

considerations. Specific topics include: the mechanics of sampling; maintenance of

sample integrity; field QA and QC; and OH&S requirements.
Chapter 3 — Aquatic ecosystems

page 3.4–22 Version — October 2000

• The elucidation of the ‘biologically-relevant’ (usually bioavailable) fraction.

Chapter 7 of these Guidelines provides some information on this topic.

Chapter 4 of the Monitoring Guidelines makes recommendations about sample

filtration, but mainly from the perspective of sample preservation. Section

7.4.2 of the present Guidelines discusses filtration with an emphasis on

speciation considerations. That section also describes other steps in calculating

the relevant indicator concentration, such as thermodynamic modelling, while

section 8.3.5 describes the application of algorithms designed to account for

the modifying effect of indicators such as water hardness.

• The mathematical (including statistical) processing of raw or speciationadjusted data. Chapter 6 of the Monitoring Guidelines offers a detailed and

very useful primer on data management and interpretation, including summary

statistics, methods of inference, multivariate analysis, power analysis,

regression techniques, trend analysis, and non-parametric statistics. It also

contains useful discussions on water quality modelling, outlier detection and

the treatment of data below the analytical detection limit.

• The comparison of test data with background data and default trigger values.

Whether or not a study area has adequate water quality is decided by comparing

monitoring data with a guideline recommendation.a This assessment of whether

the guideline has been exceeded is embodied in the concept of an ‘attainment

benchmark’. The default trigger value can be structured as a comparison between

reference (or background) and test-site data or as a comparison with a single

default trigger value. Statistical decision criteria can be used to compare test data

with background data or default trigger values.b In general, the greater the

amount of reference data (if applicable) and test data gathered, the smaller will

be the error rates associated with detecting change in toxicant concentrations in

the field. Wherever maintenance of biological diversity is a key management

goal — e.g. sites of high conservation value (condition 1) or slightly disturbed

systems (condition 2), statistical decision criteria should be set as conservatively

as possible. Values of the criteria as recommended for biological indicators

might be used as a starting point in negotiations.c

a See Section

7.4.4.2

b Section 3.1.7

(statistical

decision criteria);

section 7.4.4.2

(default trigger

values); Section

7.4.4.2 (detecting

change in toxicant

concentrations in

the field); See also

the Monitoring

Guidelines

Chapter 6.

c Section 3.2.4.2
Version — October 2000 page 3.5–1

3.5 Sediment quality guidelines

3.5.1 Introduction

The Australian Water Quality Guidelines for Fresh and Marine Waters (ANZECC

1992) provided a framework for managing receiving water quality. Those

Guidelines recognised that total load and fate of contaminants, particularly to

enclosed systems, should also be considered. Sediments are important, both as a

source and as a sink of dissolved contaminants, as has been recognised for some

time. As well as influencing surface water quality, sediments represent a source of

bioavailable contaminants to benthic biota and hence potentially to the aquatic food

chain. Therefore it is desirable to define situations in which contaminants

associated with sediments represent a likely threat to ecosystem health. While

costly remediation or restoration might not represent a management option,

sediment guidelines can usefully serve to identify uncontaminated sites that are

worthy of protection. Sediment quality guidelines are being actively considered by

regulatory agencies worldwide.

This section reviews the current state of knowledge on environmental effects of

contaminants in sediments, and the approaches being used to formulate sediment

quality guidelines. On the basis of these, it outlines a procedure for the

development of appropriate sediment quality guidelines for Australia and New

Zealand. The guidelines would apply to slightly to moderately disturbed

(condition 2) and highly disturbed (condition 3) aquatic ecosystems.a

Consideration of sediment quality follows the decision-tree approach being

adopted in these Guidelines, with a focus on identifying the issues and the

protection necessary to manage them.

For aquatic ecosystems considered to be of high conservation/ecological value

(condition 1) a precautionary approach is recommended. In these ecosystems,

chemicals originating from human activities should be undetectable, and naturally

occurring toxicants (e.g. metals) should not exceed background sediment

concentrations.b This approach should only be relaxed when there are considerable

biological assessment data showing that such a change in sediment quality would

not disturb the biological diversity of the ecosystem.

3.5.2 Underlying philosophy of sediment guidelines

It is important to understand why sediment guidelines are being developed and how

and where they might be applied. The establishment of guidelines will serve three

principal purposes:

• to identify sediments where contaminant concentrations are likely to result in

adverse effects on sediment ecological health;

• to facilitate decisions about the potential remobilisation of contaminants into

the water column and/or into aquatic food chains;

• to identify and enable protection of uncontaminated sediments.

Many urban and harbour sediments fall into the first category, usually being

contaminated by heavy metals and hydrophobic organic compounds resulting from

both diffuse and point-source inputs. They are not easily remediated. At present,

a See Section

3.1.3

b Section

3.1.3.2
Chapter 3 — Aquatic ecosystems

page 3.5–2 Version — October 2000

ex situ treatment or dredging and disposal are the most cost-effective options. If a

site is known to have highly contaminated sediments with potential for biological

uptake, it may be possible to control the collection of benthic organisms for human

consumption. For the most part, because of the enormous costs involved, there is

unlikely to be large-scale sediment remediation, unless it is driven by human health

risk assessments. Contaminated sediments can be remediated naturally when fresh

sediments, able to support viable biological populations, settle on top of them. This

can occur through water column inputs and can be managed through controls on

inputs via water quality guidelines. Management conflicts can arise when natural

sediment accumulation restricts navigation.

It is possible to adopt measures to protect unmodified areas from further

contamination by managing inputs. This is where the application of sediment

quality guidelines will be of greatest value. Just as for water quality guidelines, the

application of sediment guidelines will involve a decision-tree approach. It is

important to reiterate that the guidelines should not be used on a pass or fail basis.

The guideline numbers are trigger values that, if exceeded, prompt further action as

defined by the decision tree. The first-level screening compares the trigger value

with the measured value for the total contaminant concentration in the sediment. If

the trigger value is exceeded, then this triggers either management/remedial action

or further investigation to consider the fraction of the contaminant that is

bioavailable or can be transformed and mobilised in a bioavailable form.

In the case of metals, the dilute-acid-soluble metal concentration is likely to be a

more meaningful measure than the total value. The derivation of future trigger

values might ultimately be based on this measurement. Non-available forms will

include mineralised contaminants that require strong acid dissolution. For metals

that form insoluble sulfides, the role of amorphous iron sulfide (FeS), measured as

so-called acid volatile sulfides (AVS), can be an important factor in reducing metal

bioavailability. This exchangeable sulfide is able to bind released metals in nonbioavailable forms. Changes in redox potential and pH also affect the availability

of metals and other contaminants, and should be considered.

It is important to consider both sediment pore waters and the sediment particles as

sources of contaminants. The importance of these sources varies for various classes

of sediment dwelling organisms, as discussed elsewhere.a

3.5.3 Approach and methodology used in trigger value derivation

The many approaches adopted internationally to derive sediment quality guidelines

are more fully described in Section 8.4 (Volume 2). By far the most widely used

method is an effects database for contaminated and uncontaminated sites, based on or

derived from field data, laboratory toxicity testing and predictions based on

equilibrium partitioning of contaminants between sediment and pore water. There are

few reliable data on sediment toxicity for either Australian or New Zealand samples

from which independent sediment quality guidelines might be derived, and without a

financial impetus there is little likelihood that further data will be forthcoming in the

immediate future. Because of this, and as has been done in many other countries, the

option selected for the sediment quality guidelines is to use the best available

overseas data and refine these on the basis of our knowledge of existing baseline

concentrations, as well as by using local effects data as they become available.

a See Section

8.4.3.2
3.5.4 Recommended guideline values

Version — October 2000 page 3.5–3

The recommended guideline values are tabulated as interim sediment quality

guideline (ISQG) values (table 3.5.1), and the low and high values correspond to

the effects range-low and -median used in the NOAA listing (Long et al. 1995).

3.5.4 Recommended guideline values

3.5.4.1 Metals, metalloids, organometallic and organic compounds

The recommended guideline values for a range of metals, metalloids,

organometallic and organic sediment contaminants are listed in table 3.5.1.a Values

are expressed as concentrations on a dry weight basis. This does not imply that

samples should be dried before analysis, resulting in potential losses of some

analytes, but that results should be corrected for moisture content. For organic

compounds, values are normalised to 1% organic carbon, rather than being

expressed as mg/kg organic carbon as is sometimes done. If the sediment organic

carbon content is markedly higher than 1%, the guideline value should be relaxed

(i.e. made less stringent), because additional carbon binding sites reduce the

contaminant bioavailability.

The issue of uncertainties is often overlooked and is worth re-emphasising. The

database underpinning the guidelines (Long et al. 1995) was originally designed to

rank sediments. The values represent a statistical probability of effects (10% or 50%)

when tested against only one or two species, principally amphipods. This is not

analogous to the Aldenberg and Slob (1993) approach to water quality guidelines

that are protective of 95% of the species, based on tests on a large range of aquatic

species of varying sensitivities. Note that some tests use sea urchin fertilisation,

while for organic compounds the tests apply Microtox® luminescent bacteria to

solvent extracts of sediments. The ecological relevance of these is questionable.

There are added uncertainties about how well the effects of multiple toxicants have

been dealt with. The data do not consider antagonism or synergism between

chemicals, and, as originally derived, they are based only on disturbances to

biological receptors and do not relate to human health disturbances.

3.5.4.2 Ammonia, sulfide, nutrients and other sediment contaminants

No specific guideline values are provided in any of the overseas databases for

ammonia or nutrients such as phosphate and nitrate, yet it is important to identify

when these represent a threat to benthic communities.

The major disturbance of ammonia will be seen in pore waters, and it is best that

these be sampled and the measured ammonia concentrations compared against

water quality guidelines.b

The biological effects of sulfide in sediments are poorly understood. The decision

tree acknowledges the role of sulfide in reducing metal toxicity, but sulfide can affect

animal behaviour which in turn can alter the toxicity of both sulfide and also other

sediment contaminants (Wang & Chapman 1999). Both sulfide and ammonia can

potentially be released in any sediment studies. This may require the refining of

appropriate TIE protocols for use with sediments.

a See Section

8.4.3

b Section 8.4
Chapter 3 — Aquatic ecosystems

page 3.5–4 Version — October 2000

Table 3.5.1 Recommended sediment quality guidelinesa

Contaminant ISQG-Low

(Trigger value)

ISQG-High

METALS (mg/kg dry wt)

Antimony 2 25

Cadmium 1.5 10

Chromium 80 370

Copper 65 270

Lead 50 220

Mercury 0.15 1

Nickel 21 52

Silver 1 3.7

Zinc 200 410

METALLOIDS (mg/kg dry wt)

Arsenic 20 70

ORGANOMETALLICS

Tributyltin (μg Sn/kg dry wt.) 5 70

ORGANICS (μg/kg dry wt) b

Acenaphthene 16 500

Acenaphthalene 44 640

Anthracene 85 1100

Fluorene 19 540

Naphthalene 160 2100

Phenanthrene 240 1500

Low Molecular Weight PAHs c 552 3160

Benzo(a)anthracene 261 1600

Benzo(a)pyrene 430 1600

Dibenzo(a,h)anthracene 63 260

Chrysene 384 2800

Fluoranthene 600 5100

Pyrene 665 2600

High Molecular Weight PAHs c 1700 9600

Total PAHs 4000 45000

Total DDT 1.6 46

p.p’-DDE 2.2 27

o,p’- + p,p’-DDD 2 20

Chlordane 0.5 6

Dieldrin 0.02 8

Endrin 0.02 8

Lindane 0.32 1

Total PCBs 23 –

a Primarily adapted from Long et al. (1995);

b Normalised to 1% organic carbon;

c Low molecular weight PAHs are the sum of concentrations of acenaphthene,

acenaphthalene, anthracene, fluorene, 2-methylnaphthalene, naphthalene and phenanthrene;

high molecular weight PAHs are the sum of concentrations of benzo(a)anthracene,

benzo(a)pyrene, chrysene, dibenzo(a,h)anthracene, fluoranthene and pyrene.
3.5.5 Applying the sediment quality guidelines

Version — October 2000 page 3.5–5

For nutrients, the need to define sediment guidelines is debatable. In this case, the

disturbance that we are seeking to protect against is algal or macrophyte blooms,

whereas the proposed guidelines address biological disturbances, based in part on

equilibrium partitioning to sediment pore waters and ultimately the water column.

It should theoretically be possible to derive a guideline value based on the

undesirable release of nutrients to the water column and their subsequent

undesirable ecosystem disturbances. This would require some measure or

prediction of pore water nitrogen and phosphorus and a judgement as to what

concentration of bioavailable nutrient constitutes a threat, logically based on water

quality guidelines.

There are methods that purport to measure bioavailable phosphorus, for example

bioassays or the use of iron strips, but there are factors such as redox potential that

will be important in defining this. Indeed, control of bioavailable carbon inputs is

more important than the concentration of phosphorus itself. The application of

water quality guidelines to pore waters is possible, although prior use of the

nutrients by benthic organisms may have already reduced the pore water

concentrations. It is generally thought that development of nutrient guidelines is

too difficult at this stage, and must await further research developments.

3.5.4.3 Absence of guidelines

In some instances, no guidelines will be specified for a contaminant of interest.

This generally reflects an absence of an adequate data set for that contaminant. An

interim approach is required to provide some guidance as well as to ensure

environmental protection in situations where guidelines would apply. The approach

suggested is to derive a value on the basis of natural background (reference)

concentration multiplied by an appropriate factor. A factor of two is recommended,

although in some highly disturbed ecosystems a slightly larger factor may be more

appropriate, but no larger than three. An alternative approach is to apply the water

quality guideline values to sediment pore waters.

3.5.5 Applying the sediment quality guidelines

A protocol is provided to summarise key aspects of collection and laboratory

analysis of sediment samples a while the Monitoring Guidelines provide full

details.

3.5.5.1 Sediment sampling

The use of appropriate sampling techniques is a prerequisite for chemical or

toxicity testing of sediments or sediment pore waters. The depth of sampling will

be dictated by the issue being investigated, and this in turn will determine whether

corers or grab sampling is preferable. Full details on sampling methodology are

provided in the Monitoring Guidelines.

3.5.5.2 Applications of chemical testing

It is important to recognise the limitations applicable to the guideline values in

table 3.5.1 as discussed above. They nevertheless form a good basis for sediment

quality assessment, if applied using a decision tree approach as illustrated in

figure 3.5.1.

a See App. 8,

Volume 2
Chapter 3 — Aquatic ecosystems

page 3.5–6 Version — October 2000

Determine appropriate guideline trigger 

values for selected indicators (figure 3.1.1)

Sediment contaminant characterisation

Measure total then dilute acid-soluble metals, organics plus

TOC, grain size.

Decision tree framework

for applying the sediment

quality guidelinesa

 Test against guideline values

Compare contaminant/stressor concentration with lower and upper guideline values

Below lower value Above upper valueb

Between upper and

lower valuesb

Low risk

(no action)

Check background 

concentrations

Low risk

(no action)

Below Aboveb

Examine factors controlling 

bioavailability (optional)

eg. AVS

 pore water concentrations

 sediment speciation

 organic carbon

Below Aboveb

Acute toxicity testing

Not toxicb Toxic

Low risk

(no action)

Chronic toxicity testing Highly contaminated

(initiate remedial actions)

Not toxic Toxic

Moderately contaminated

(initiate remedial actions)

Test against guideline value

Compare bioavailable concentration with lower guideline value

Define primary management aims (figure 3.1.1)

a Local biological effects data not required in the decision trees (see section 3.1.5) b Further investigations are not mandatory; users may opt to proceed to management/remedial action

Low risk

(no action)

Figure 3.5.1 Decision tree for the assessment of contaminated sediments

The general approach to use of the decision scheme is outlined in Section 3.1.5.a If

the lower sediment quality guideline, the trigger value, for a particular contaminant

is not exceeded, it is unlikely that it will result in any biological disturbance for

organisms inhabiting that sediment. If the trigger value is exceeded, either

management (including remedial) action is taken, or additional site-specific studies

are conducted to determine whether this exceedance poses a risk to the ecosystem.

a Section 3.1.5
3.5.5 Applying the sediment quality guidelines

Version — October 2000 page 3.5–7

Should a ‘low risk’ outcome result after continuous monitoring, there is scope to

refine the guideline trigger value. Note that in the consideration of guideline values

for metals, total metals concentrations are used, however, acid-soluble metals, are

more representative of a bioavailable fraction and it is envisaged that ultimately

trigger value compliance will be based on this measurement, as discussed later.

Comparison with background concentrations

The next step in the decision tree involves a comparison with background

concentrations. Exceedance of a trigger value is acceptable if it is at or below the

normal background concentration for a site. The selection of background or

reference no-effects sites should, where possible, use sediments of comparable

grain sizes. Similarly, the analysis of sediment cores must ensure that fluctuations

in contaminant concentrations with depth are not the result of grain size changes, or

in the case of organics, to changes in the organic carbon content.

For metals, a reliable determination of ‘natural’ levels of contaminants is best done

on the basis of trace element ratios determined for a range of uncontaminated sites.

Usually the contaminant element is referred to naturally occurring elements such as

lithium, iron or aluminium (e.g. Loring & Rantala 1992).

The theoretical background concentration of most synthetic organic compounds is

zero, but from a practical viewpoint, ubiquitous contamination has occurred far

from point sources. Reference sites removed from such sources are appropriate for

determining background concentrations.

Consideration of factors controlling bioavailability

If both the lower guideline trigger value and the background or reference site

concentrations are exceeded, the next level evaluation will be to consider whether

there are any factors which might lower the potential bioavailability of

contaminants. The methods of sampling of sediments and sediment pore waters

will be critical if meaningful data (especially for metals) are to be obtained, to

ensure that the natural chemical conditions, especially redox conditions, salinity

and pH, are not altered. If such changes are allowed to occur, erroneous analytical

data on contaminant bioavailability may be obtained.a

For metals, the speciation considerations might be:b

a) Sediment speciation — dilute-acid-extractable metals concentrations below

lower guideline value. It is recommended that this should involve treatment of

the sample with 1 M hydrochloric acid for 1 hour (Allen 1993).

Since a considerable fraction of the total metal concentration in sediments may

be present in detrital mineralised phases that are not bioavailable, a better

estimate of the bioavailable fraction is desirable. Although the capacity of

chemical extractions to selectively remove only this fraction is limited, a diluteacid-extraction will not remove the mineralised fractions and will therefore

provide more appropriate metal concentration data for use in new effects

databases. During extraction of carbonate- or sulfide-containing sediments,

allowance must be made for acid consumed by reaction with these phases.

Note that, except for spiked sediment toxicity tests where ionic metal additions

are made, the field data used to derive the guidelines are likely to be based on

total concentrations. Therefore a judgement against these measurements using

a See Section

4.3.5 of the

Monitoring

Guidelines

b See

discussion in

Section 8.4,

Vol. 2
Chapter 3 — Aquatic ecosystems

page 3.5–8 Version — October 2000

speciation cannot be fully justified. Rather, such considerations should be

applied in new guideline values developed from an NWQMS database.

b) Acid volatile sulfides, AVS: Σi [SEM] < [AVS]

If the concentration of acid volatile sulfide (AVS), released by dilute acid

treatment of the moist sediment, exceeds the sum of the heavy metal

concentrations released by the same treatment (referred to as simultaneously

extracted metals (SEM)), then this excess sulfide is able to bind heavy metals in

insoluble and non-bioavailable forms, and therefore the metals will not cause

toxicity.a This applies particularly to lead, zinc and cadmium. Its application to

copper, nickel and possibly cobalt is suspect.

Recent reports urge caution in the application of the AVS binding model,

particularly because of concern for its relevance in longer-term and community

level effects (IMO 1997). Other limitations are discussed in Section 8.4. A

description of the methods for measuring AVS and SEM may be found in

Allen et al. (1992).

c) Pore water: Σi [Mi,d]/[WQGi,d ] <1, where [Mi,d] is the total dissolved pore

water concentration for each metal and [WQGi,d ] is the water quality guideline

value for each metal.

Assuming that pore water represents the major exposure route to sediment

toxicants, then if pore water concentrations for any metal are below the water

quality guideline concentration, there is unlikely to be an adverse biological

disturbance. The correct methods should be used for sampling pore waters, to

avoid losses or changes in redox status. Note that there is the possibility of

seasonal variations in pore water contaminant concentrations as well as in AVS.

For organic compounds, the use of guidelines normalised to total organic carbon

(TOC) is a first stage. The effects of natural sediment and water chemistry on the

equilibrium partitioning of the particular organic compounds are moderating

factors requiring consideration. This may mean separate measurements of the

partitioning into natural waters of appropriate salinity or the measurement of pore

water concentrations. Analytical detection with the small volumes generally

encountered creates problems, so this is often a difficult area. Such

considerations as rates of degradation, either chemical, physical or biological,

can be important for hydrophilic and for some hydrophobic organics.

If on the basis of any of the above considerations the trigger value is still

exceeded, and further investigation is sought rather than management/remedial

action, toxicity tests will be required. The tests will further characterise the

nature of sediment as either moderately or highly contaminated. Alternatively,

toxicity testing might be employed in lieu of more detailed chemical

investigations when the trigger value is exceeded.

The guidelines discussed above have been derived on the basis of the toxicity of

contaminants in sediments and associated pore waters, to benthic biota. An

additional factor that needs to be taken into consideration, especially for riverine

sediments, is mobility. Dynamic zones can be created in rivers during periods of

high flow that lead to erosion and sediment mobilisation. Finer, contaminant-rich

particles will be the most mobile, although larger particles will also be moved in

storm flows. Two considerations arise under these conditions.

a See Section

8.4.3.2, Vol. 2
3.5.5 Applying the sediment quality guidelines

Version — October 2000 page 3.5–9

First there is the concern for enhanced contaminant release, either resulting from

the disturbance of surface sediments and pore waters, or as a consequence of

chemical transformations, such as oxidation of previously anoxic sediments. The

former is not important, since pore water concentrations will be diluted. The

possibility of oxidative release especially of metals is more a concern. In this case

the kinetics of oxidation of metal sulfides is important. Elutriate tests with

overlying saline or freshwaters can be used to demonstrate a worst case release

scenario.

Secondly there is the possibility that the deposition process will lead to particle

sorting, and if this were to result in a greater concentration of clay/silt particles at a

particular site, there is a real possibility that in some cases the guideline

concentrations for the whole sediment could now be exceeded because of removal

of the diluent effect of coarser particles. If sorting is believed to be a possibility, it

would be appropriate to assess the sediment on the basis of analyses on the <63 μm

size fraction only.

In the absence of sediment guideline values for a particular contaminant, the first

recourse is to the water quality guideline values. Sampling and analysis of

sediment pore water can be undertaken, and water quality values can be employed

to judge its acceptability. Care must be taken that the chemistry of the pore waters

is not altered during the sampling process. This means squeezing, or centrifuging

the sediment under nitrogen to minimise oxidation. Often it is very difficult to

obtain sufficient sample to undertake a pore water analysis, especially for organic

contaminants. In these cases, toxicity testing of the sediment or pore water is the

only option.

In relation to water quality, different levels of protection have been considered

for particular ecosystem conditions (namely high conservation value, slightly to

moderately disturbed and highly disturbed). It is not appropriate at this stage to

provide guidelines for different levels of protection for sediments, until more data

are available. The provision of low and high guideline values, in combination

with the decision-tree approach, should nevertheless provide useful guidance

about the potential ecological effects of sediment contaminants that can guide

management actions, as indicated in table 3.1.2.

Application of toxicity testing

The decision-tree allows for toxicity testing as the ultimate means of assessing

sediment quality. Although this is shown at the bottom of the tree, mainly on the

basis of its greater cost compared to chemical analyses, it may be applied at any

stage. Appropriate methods may include examining the water extractable

contaminants (elutriate testing), pore water testing, or whole sediment bioassays.

Whole sediment testing with infaunal species has the greatest ecological relevance.

Marine and freshwater testing with amphipods have been most widely used,

although tests using midge larvae, insects and worms have been reported.a

As with chemical testing, is important that the sample used for toxicity testing has

the same chemistry as it did in the field situation. Oxidation of sediments during

manipulations may significantly alter metal bioavailability.

Normally toxicity testing will be used to demonstrate the absence of toxicity when

the guideline for a particular contaminant is exceeded. If toxicity is observed, its

origins cannot necessarily be attributed to the contaminant of interest, because of

a See Section

8.4; also

Method 2A

(App. 3, Vol. 2),

table 3.2.2
Chapter 3 — Aquatic ecosystems

page 3.5–10 Version — October 2000

the possibility of other contaminants either contributing to the observed toxicity or

being the primary cause. Under these conditions, it will be necessary to apply TIE

procedures (USEPA 1991) which successively separate classes of contaminants

and identify any toxicity that they may have caused. Despite a large number of

applications of the TIE approach, it is most often ammonia or common pesticides

that have been found to be the source of toxicity.
Version — October 2000 page 4.1–1

4 Primary industries

4.1 Introduction

Both the quality and the quantity of water resources are critical issues for

agriculture and aquaculture in Australia and New Zealand. Water quality is also of

major importance for the protection of human consumers of food products. Growth

of these major primary industries, together with expanding urbanisation and other

industrial development, has increased the demand for good quality water but at the

same time exerted escalating pressures on the quality of the water resources that

are available. Therefore, to assess water quality for primary industries, not only

must productivity issues be considered but also the possible adverse effects of these

enterprises on downstream water quality and activities.

In recent years it has been recognised that pollution-related issues should be

addressed by approaching the conservation, management and use of water

resources in a holistic manner, according to the principles of integrated catchment

management. Key strategies for achieving ecologically sustainable development

include the involvement of stakeholders in decision-making processes and the

development and adoption by industry of best management practice guidelines.

This is the first occasion on which water quality guidelines have been provided for

aquaculture industries in Australia and New Zealand. Most of the guidelines

presented for aquaculture should be used with some caution because few are based

on a critical assessment of a wide data set.a This chapter also discusses issues

concerning water quality guidelines for the protection of human consumers of

aquatic foods. Recreational and commercial fisheries are based on wild populations

of fish, crustacea and shellfish species, which are supported by natural habitats and

food webs. Accordingly, for the protection of wild animal stocks, the reader is

referred to the water quality guidelines for the protection and maintenance of

aquatic ecosystems (Chapter 3).

Irrigation and livestock watering are the major agricultural uses of water. Minor

amounts are used for other production purposes, such as the mixing of pesticide,

fertiliser and veterinary formulations, and livestock dietary supplements. In

Australia particularly, both the irrigation and livestock industries rely heavily on

the use of groundwater, as well as surface water resources. Groundwater is also an

important source of stock water in parts of New Zealand. Thus the guidelines

provided for these industries are applicable (where appropriate) to both surface and

groundwater quality.

Guidelines for general on-farm water use are included with the irrigation guidelines

and cover topics such as corrosion and fouling of pipes and fittings. Certain issues

concerning water quality for use by agriculture are also discussed in other documents

published in conjunction with the National Water Quality Management Strategy; for

example, the Guidelines for Sewerage Systems — Use of Reclaimed Water

(ARMCANZ, ANZECC & NHMRC 2000). Note, however, that occasional

discrepancies may occur in the information provided by different NWQMS

documents; for example, when revision of the documents is out of step. All the

a See Sections

4.4 and 9.4.4
Chapter 4 — Primary industries

page 4.1–2 Version — October 2000

guideline documents are based on the best scientific information available at the time

of publication.

For information on the quality of farmstead water supplies for domestic use in

Australia, the reader is referred to Chapter 6 of these Guidelines and Section 7.7 of

the Australian Drinking Water Guidelines (NHMRC & ARMCANZ 1996).

Readers in New Zealand are referred to the Drinking-water Standards for New

Zealand (New Zealand Ministry of Health 1995a) and the Guidelines for Drinkingwater Quality Management (New Zealand Ministry of Health 1995b). Issues such

as water quality for washing of farm produce or for dairy water supplies are outside

the scope of the present guidelines and the reader is referred to local health and

hygiene regulations and the proposed food safety standards of the Australian and

New Zealand Food Authority.

An important first step in using these guidelines is to consider the management

framework for their application. This includes defining the primary management

aims, determining appropriate trigger values, defining water quality objectives, and

establishing a monitoring and assessment program to address these objectives.a

The type of monitoring and assessment program required will be specific to each

situation, but there are several broad principles or procedures that are common to

all programs. For details see Chapter 7, particularly noting figure 7.1 which gives a

generic flow chart of the procedural framework for monitoring and assessment, and

Section 7.4 which discusses specific issues for physical and chemical indicators.

a See Section

2.1
Version — October 2000 page 4.2–1

4.2 Water quality for irrigation and general water use

Agricultural practice in Australia and New Zealand is often dependent on

irrigation, because of climatic constraints on crop demand. In Australia

particularly, agriculture is predominantly based in areas of limited rainfall, and

there is heavy reliance on the use of surface and groundwaters for irrigation of

crops and pastures. Approximately 70% (nearly 12 000 giga-litres) of Australia’s

developed water is used for irrigation, 21% for urban or industrial purposes and 9%

for rural water supply (DEST State of the Environment Advisory Council 1996).

Irrigated agriculture contributes very significantly to the Australian economy, with

an annual production value of commodities such as cotton, rice, cereals, sugar,

horticulture and irrigated fodder, of over $7 billion (Cape 1997).

In New Zealand irrigation is playing an increasingly important role in agricultural

production. The area of irrigated land is doubling approximately every 10 years.

Around 80% of allocated water in New Zealand is used for irrigation, with the

remaining 20% for urban and industrial uses. Irrigated agriculture makes a significant

contribution to the New Zealand economy, with irrigation being worth an extra $800

million ‘at the farm gate’ and possibly three times this in export earnings.

An important goal of these Water Quality Guidelines is to maintain the

productivity of irrigated agricultural land and associated water resources, in

accordance with the principles of ecologically sustainable development and

integrated catchment management.a This should be a key consideration in any

irrigation strategy, alongside maximum yield and economic viability.

4.2.1 Philosophy

In developing the guidelines, emphasis has been placed on sustainability in

agricultural practice (DEST State of the Environment Advisory Council 1996),

which aims to ensure that:

• the supply of necessary inputs is sustainable;

• the quality of natural resources is not degraded;

• the environment is not irreversibly harmed;

• the welfare and options of future generations are not jeopardised by the

production and consumption activities of the present generation; and

• yields and produce quality are maintained and improved.

In terms of water quality, the focus for sustainable farming systems is on adopting

management practices that maintain productivity and minimise the off-farm movement

or leaching of potential aquatic contaminants. Key issues include soil erosion,

landscape salinity, fertiliser and pesticide management, livestock access to streams, and

safe disposal of effluent from intensive animal industries (Hunter et al. 1996).

4.2.2 Scope

Soil, plant and water resource issues that have been taken into account in developing

the water quality guidelines for irrigation use are summarised in table 4.2.1. Factors

affecting irrigation water quality concern physical, chemical and biological

characteristics that may affect the soil environment and crop growth.

a See Section

4.2.1
Chapter 4 — Primary industries

page 4.2–2 Version — October 2000

Table 4.2.1 Key issues concerning irrigation water quality effects on soil, plants and water

resources

Key issues

Soil Root zone salinity

Soil structural stability

Build-up of contaminants in soil

Release of contaminants from soil to crops & pastures

Plants Yield

Salt tolerance

Specific ion tolerance

Foliar injury

Uptake of toxicants in produce for human consumption

Contamination by pathogens

Water resources Deep drainage & leaching below root zone

Movement of salts, nutrients & contaminants to groundwaters & surface

waters

Important

associated factors

Quantity and seasonality of rainfall

Soil properties

Crop and pasture species and management options

Land type

Groundwater depth and quality

Guidelines are also included for general on-farm water use dealing with the

corrosion and fouling potential of waters. These characteristics are important for

the maintenance of farm equipment (pumps, pipes, etc.). The guidelines may also

be applied more widely where corrosion and fouling are of concern.

Specific irrigation water quality guidelines for intensive horticultural activities (e.g.

hydroponics and glass-house growing) are not included in this document.

Guidelines for irrigation water quality are given here for biological parameters,

salinity and sodicity, inorganic contaminants (i.e. specific ions, including heavy

metals and nutrients), organic contaminants (i.e. pesticides) and radiological

characteristics. The guidelines are trigger values below which there should be

minimal risk of adverse effects. Further investigation is recommended if a trigger

value is exceeded, to determine the level of risk.

A more detailed discussion of all water quality parameters included in the

guidelines is given in Volume 3, Section 9.2.

4.2.3 Biological parameters

4.2.3.1 Algae

No trigger value for algae in irrigation waters is recommended; however,

excessive algal growth may indicate nutrient pollution of the water supply.

Algae are commonly found in most water sources and do not generally cause

problems in irrigation waters unless there is excessive growth due to factors such
4.2.3 Biological parameters

Version — October 2000 page 4.2–3

as suitable flow regime, temperature, abundant nutrients and adequate sunlight.

The main problem associated with excessive algal growth in irrigation waters is the

blockage of distribution and irrigation equipment. This can result in reduced or

uneven flow throughout the irrigation system which may reduce crop yield and

increase overall maintenance costs.

4.2.3.2 Cyanobacteria (blue-green algae)

No trigger values for cyanobacteria in irrigation waters are recommended at

this time.

Cyanobacteria (blue-green algae) form part of the natural microbial population in

most waterbodies. Under certain natural or human-induced circumstances, toxic

blooms can occur and these may adversely affect the suitability of waters for

irrigation, particularly because toxin residues can potentially accumulate on

produce for human or animal consumption. If an algal bloom occurs, it is

recommended that an alternative source of irrigation water be used, and that the

water be tested for microbial composition and (if necessary) toxicity. There is

presently insufficient information available for use in deriving trigger values for

cyanobacteria in irrigation water.

4.2.3.3 Human and animal pathogens

Trigger values for thermotolerant coliforms in irrigation waters are provided

in table 4.2.2.

Table 4.2.2 Trigger values for thermotolerant coliforms in irrigation waters used for food and

non-food cropsa

Intended use Level of thermotolerant

coliformsb

Raw human food crops in direct contact with irrigation water (e.g.

via sprays, irrigation of salad vegetables) <10 cfuc

 / 100 mL

Raw human food crops not in direct contact with irrigation water

(edible product separated from contact with water, e.g. by peel,

use of trickle irrigation); or crops sold to consumers cooked or

processed

<1000 cfu / 100 mL

Pasture and fodder for dairy animals (without withholding period) <100 cfu / 100 mL

Pasture and fodder for dairy animals (with withholding period of 5

days) <1000 cfu / 100 mL

Pasture and fodder (for grazing animals except pigs and dairy

animals, i.e. cattle, sheep and goats) <1000 cfu / 100 mL

Silviculture, turf, cotton, etc. (restricted public access) <10 000 cfu / 100 mL

a Adapted from ARMCANZ, ANZECC & NHMRC (1999)

b Median values (refer to text)

c cfu = colony forming units

It is generally not feasible nor warranted to test irrigation water for the presence of

the wide range of water-borne microbial pathogens that may affect human and

animal health. The guidelines recommended here are based on the practicable testing

of irrigation waters for the presence of thermotolerant coliforms (also known as

faecal coliforms), which gives an indication of faecal contamination and thus the

possible presence of microbial pathogens (NHMRC & ARMCANZ 1996). However,

the test does not specifically indicate whether pathogenic organisms are present.
Chapter 4 — Primary industries

page 4.2–4 Version — October 2000

It is recommended that a median value of thermotolerant coliforms be used, based

on a number of readings generated over time from a regular monitoring program.

Investigations of likely causes are warranted when 20% of results exceed four

times the median guideline value (ARMCANZ, ANZECC & NHMRC 2000).

For helminths, a trigger value of ≤1 helminth egg per litre is proposed for the

protection of crop consumers in areas where helminth infections are known to be

endemic. A lower value of 0.5 eggs per litre may be required to protect farm

workers and their families in situations of direct exposure to the water

(ARMCANZ, ANZECC & NHMRC 2000). Insufficient information is available

for use in setting guidelines for protozoa and viruses in irrigation water.a

4.2.3.4 Plant pathogens

No trigger values for plant pathogens in irrigation waters are recommended

at this time. As a general precaution, disinfestation treatment is advisable for

water that contains plant pathogens and is to be used for irrigating

potentially susceptible plants.

Agricultural crops and pastures can be affected by various plant pathogens

transmitted through a number of different pathways including irrigation water,

although it is believed that the risk from pathogens in irrigation water is low under

most circumstances. However, plant pathogens in irrigation water used for

intensive agricultural and horticultural industries (particularly where wastewaters

are reused) can potentially lead to crop damage and economic loss.

A great deal of work needs to be done before guidelines can be developed, particularly

regarding the efficacy of water-borne plant pathogens on a wide range of crops.

4.2.4 Irrigation salinity and sodicity

4.2.4.1 Salinity and sodicity

To assess the salinity and sodicity of water for irrigation use, a number of

interactive factors must be considered. As outlined in this section, these

include irrigation water quality, soil properties, plant salt tolerance, climate,

landscape (including geological and hydrological features), and water and

soil management.

Salinity is the presence of soluble salts in or on soils, or in waters. High salinity

levels in soils may result in reduced plant productivity or, in extreme cases, the

elimination of crops and native vegetation. Salinity related issues are of concern in

many parts of Australia but salinisation is currently considered to be only of minor

importance in New Zealand.

Sodicity is the presence of a high proportion of sodium (Na+

) ions relative to

calcium (Ca2+) and magnesium (Mg2+) ions in soil or water. Sodicity degrades soil

structure by breaking down clay aggregates, which makes the soil more erodible

and less permeable to water, and reduces plant growth.

The effects of salinity and sodicity in irrigation waters are very situation-specific,

making it inappropriate to set water quality trigger values for general application.

Factors which need to be considered include: the type of crop being cultivated and

its salt tolerance, the characteristics of the soil under irrigation, soil management

and water management practices, climate and rainfall (figure 4.2.1).

a See also

Section 9.2.2.3
4.2.4 Irrigation salinity and sodicity

Version — October 2000 page 4.2–5

Water quality

salinity (EC)

sodicity (SAR)

Soil properties

clay % (average root zone)

 cation exchange capacity (average root zone)

 exchangeable sodium % (at bottom of root zone)

Rainfall

mm/year

Irrigation

mm/year

Leaching fraction

calculated using all input information

Average root zone salinity

calculated

Crop salt tolerance

impact threshold & yield decline

Management practices

application methods

amelioration techniques

Plant response managing variable quality water supplies

relative yield

!

"

#

$

% Broader landscape issues

e.g. land use and watertable management

Figure 4.2.1 Flow diagram for evaluating salinity and sodicity impacts of irrigation water

There are five key steps to determining the suitability of irrigation water with

respect to salinity and sodicity (figure 4.2.1).a

Step 1. Identify the soil properties, water quality, climate (rainfall) and management

(irrigation application rates) practices for the site in question.

Step 2. Estimate the leaching fraction under the proposed irrigation regime using

approaches outlined in this section.

Step 3. Estimate the new average root zone salinity as outlined in this section.

Average root zone salinity is considered the key limitation to plant growth in

response to salinity and sodicity levels in irrigation water. However, poor soil

structure can also reduce plant yields by limiting aeration, water infiltration

and root growth. The likelihood of soil structural problems induced by

irrigation can be predicted from trigger values derived in this section.

Step 4. Estimate relative plant yield (although note that the impact of salinity and

sodicity can be modified by management practices as discussed later in this

section).

Step 5. Consider salinity and sodicity problems within the framework of broader

catchment issues such as regional watertables, groundwater pollution and

surface water quality. Watertable salinity develops in response to excess

water and salts accumulating in sensitive parts of the landscape. Excess

water can percolate to groundwaters as a result of changing climatic

patterns (e.g. frequency and duration of rainfall events), land use or land

management (including irrigation). Before an irrigation scheme is

developed, the planning process should include investigation of the

regional hydrogeology to avoid development of watertable salinity. The

guidelines given here concentrate on localised effects of irrigation, but

broader salinity issues should not be ignored.

a See details in

Section 9.2.3
Chapter 4 — Primary industries

page 4.2–6 Version — October 2000

Software SALF PREDICT is now available. It estimates the parameters necessary

for a detailed assessment of irrigation water quality in relation to soil properties,

rainfall, water quality and plant salt tolerance. The software is based on summer

rainfall areas and should be used with some caution in winter rainfall areas. It

incorporates many of the detailed algorithms presented in Volume 3, Section 9.2.3.

The software is provided on the CD ROM provided with these Guidelines and is

also available from the Queensland Department of Natural Resources.

A simple initial assessment can be made by measuring the electrical conductivity

(ECi) and concentrations of sodium (Na+

), calcium (Ca2+) and magnesium (Mg2+)

in irrigation water. Note that EC is expressed in units of dS/m throughout Section

4.2.4 (1 dS/m = 1000 μS/cm).

Determining the suitability of irrigation water salinity for a crop

Calculate the average root zone salinity (ECse) from ECi and the average root zone

leaching fraction (LF), to see if a crop is likely to be affected by the irrigation

water salinity. First, estimate the LF of the soil being irrigated (i.e. the proportion

of applied water that leaches below the root zone). Approximate average LF values

for four broad soil types are listed in table 4.2.3. Then calculate ECse using the

following equation:

LFx2.2

EC EC i

se= (4.1)

where:

ECse= average root zone salinity in dS/m

ECi = electrical conductivity of irrigation water in dS/m

LF = average leaching fraction.

Table 4.2.3 Soil type and average root zone leaching fractiona

Soil type Average root zone LF

Sand 0.6

Loam 0.33

Light clay 0.33

Heavy clay 0.2

a From DNR (1997a), adapted from DNR (1997b)

Then use the ECse value to assess the general level of crop tolerance to the irrigation

water salinity by comparing it with the values in table 4.2.4. Alternatively, compare

the ECse with the relative salt tolerances of specific crop and pasture species provided

here in table 4.2.5 and in Volume 3, Section 9.2.3, table 9.2.10.
4.2.4 Irrigation salinity and sodicity

Version — October 2000 page 4.2–7

Table 4.2.4 Soil and water salinity criteria based on plant salt tolerance groupingsa

Plant salt tolerance groupings Water or soil salinity rating Average root zone salinity, ECse

(dS/m)b

Sensitive crops

Moderately sensitive crops

Moderately tolerant crops

Tolerant crops

Very tolerant crops

Generally too saline

Very low

Low

Medium

High

Very high

Extreme

<0.95

0.95−1.9

1.9−4.5

4.5−7.7

7.7−12.2

>12.2

a Adapted from DNR (1997b)

b 1 dS/m = 1000 μS/cm

A list of the relative salt tolerances of a limited selection of common field crop,

pasture and horticulture species is provided in table 4.2.5. Information in this table

is derived from data currently available in the literature, but preference should be

given to locally derived data where available. This gives approximate values of

average root zone salinities at the threshold level (the level causing yield

reduction). It also shows electrical conductivity of irrigation water at the threshold

level for a range of soil types, but it is meant as a general guide only.a

If at all uncertain about salt tolerance or the effect of irrigation water quality on soil

structure, submit a soil sample for analysis and seek expert advice.

Determining the risk of soil structure degradation caused by irrigation water quality

Calculate the sodium adsorption ratio (SAR) and use it (with ECi) to predict soil

structure stability in relation to irrigation water. The SAR value measures the

relative concentration of sodium (Na+) to calcium (Ca2+) and magnesium (Mg2+)

and can be calculated from the following equation:

2

MgCa

Na SAR

22 ++

+

+

= (4.2)

Where Na+, Ca2+ and Mg2+ are expressed in mmolec/L (where subscript c

indicates change).

Evaluate the quality of the irrigation water by superimposing its ECi

 and SAR

values on figure 4.2.2, to see if it will affect soil structure (through clay aggregate

breakdown). Water quality that falls to the right of the dashed line is unlikely to

cause soil structural problems. Water quality that falls to the left of the solid line is

likely to induce degradation of soil structure; corrective management will be

required (e.g. application of lime or gypsum). Water that falls between the lines is

of marginal quality and should be treated with caution.

a See also

Section 9.2.3
Chapter 4 — Primary industries

page 4.2–8 Version — October 2000

Table 4.2.5 Tolerance of plants to salinity in irrigation watera

ECi threshold for crops

growing in

Common name Scientific name Average root

zone salinity

threshold

(ECse) (dS/m)b sand loam clay

Field Crops

Barley, grain Hordeum vulgare 8 12.6 7.2 4.2

Cotton Gossypium hirsutum 7.7 12.1 6.9 4.0

Beet, sugar Beta vulgaris 7 11.0 6.3 3.7

Sorghum Sorghum bicolor 6.8 9.4 5.3 3.1

Wheat Triticum aestivum 6 9.4 5.3 3.1

Sunflower Helianthus annuus 5.5 7.5 4.3 2.5

Oats Avena sativa 5 7.0 4.0 2.3

Soybean Glycine max 5 7.0 4.0 2.3

Peanut Arachis hypogala 3.2 4.4 2.5 1.5

Rice, paddy Oryza sativa 3 4.8 2.7 1.6

Corn, grain, sweet Zea mays 1.7 3.2 1.8 1.1

Sugarcane Saccharum officinarum 1.7 4.3 2.5 1.4

Fruits

Olive Olea europaea 4 5.1 2.9 1.7

Macadamia seedling 3.6 4.6 2.6 1.5

Peach Prunus persica 3.2 4.7 2.7 1.6

Rockmelon Cucumis melo 2.2 4.6 2.6 1.5

Grapefruit Citrus paradisi 1.8 3.0 1.7 1.0

Orange Citrus sinensis 1.7 2.9 1.7 1.0

Grape Vitis spp. 1.5 3.3 1.9 1.1

Avocado Persea americana 1.3 2.3 1.3 0.8

Apple Malus sylvestris 1 2.0 1.2 0.7

Pastures

Wheatgrass, tall Agropyron elongatum 7.5 12.5 7.2 4.2

Rhodes grass, Pioneer Chloris gayana 7 12.8 7.3 4.2

Couch grass Cynodon dactylon 6.9 10.8 6.1 3.6

Buffel grass, Gayndah Cenchrus ciliaris var Gayndah 5.5 8.2 4.7 2.7

Phalaris Phalaris tuberosa (aquatica) 4.2 5.3 3.0 1.8

Fescue Festuca clatior 3.9 7.3 4.2 2.4

Green panic, Petri Panicum maximum 3 5.6 3.2 1.8

Townsville stylo Stylosanthes humilis 2.4 3.7 2.1 1.2

Clover, Berseem Clover Trifolium alexandrinum 2 3.8 2.2 1.3

Lucerne, Hunter River Medicago sativa 2 4.7 2.7 1.6

Clover, strawberry (Palestine) Trifolium fragiferum 1.6 3.3 1.9 1.1

Snail medic Medicago scutellata 1.5 2.9 1.7 1.0

Clover, white (New Zealand) Trifolium repens 1 2.5 1.4 0.8

Vegetables

Zucchini Cucurbita pepo melopepo 4.7 7.3 4.2 2.4

Beet, garden Beta vulgaris 4 6.5 3.7 2.1

Broccoli Brassica oleracea 2.8 4.9 2.8 1.6

Cucumber Cucumis sativus 2.5 4.2 2.4 1.4

Pea Pisum sativum L. 2.5 3.2 1.8 1.1

Tomato Lycopersicon esculentum 2.3 3.5 2.0 1.2

Potato Solanum tuberosum 1.7 3.2 1.8 1.1

Pepper Capsicum annum 1.5 2.8 1.6 0.9

Lettuce Lactuca sativa 1.3 2.7 1.5 0.9

Onion Allium cepa 1.2 2.3 1.3 0.8

Eggplant Solanum melongena 1.1 3.2 1.8 1.1

Bean Phaseolus vulgaris 1 1.9 1.1 0.6

Carrot Daucus carota 1 2.2 1.2 0.7

a From DNR (1997a), adapted from DNR (1997b); b 1 dS/m = 1000 μS/cm
4.2.5 Major ions of concern for irrigation water quality

Version — October 2000 page 4.2–9

Figure 4.2.2 Relationship between SAR and EC of irrigation water for prediction of soil

structural stability (from DNR 1997a, adapted from DNR 1997b; note that 1 dS/m = 1000 μS/cm)

4.2.5 Major ions of concern for irrigation water quality

4.2.5.1 Bicarbonate

No trigger value is recommended for bicarbonate in irrigation waters.

Elevated levels of bicarbonate in irrigation waters can adversely affect irrigation

equipment, soil structure and crop foliage. These problems occur when the

bicarbonate (or carbonate) in solution with calcium is sufficient to exceed the

solubility of calcium carbonate. The precipitation of calcium carbonate can lead to

white scale formation on leaves and fruit and may clog irrigation equipment.

The same process can give rise to precipitates of calcium carbonate in soil. This will

effectively increase the sodium adsorption ratio (SAR) or exchangeable sodium

percentage (ESP) and may lead to soil structural problems. An overview of the effect

of irrigation with waters of high SAR is given in Volume 3, Section 9.2.3.

4.2.5.2 Chloride

Issues concerning chloride in irrigation waters relate to the risk of: (1) foliar injury

to crops; and (2) increased uptake by plants of cadmium from soil. These are

discussed more fully in Volume 3, Section 9.2.4.2.

1 Foliar injury

Trigger values for prevention of foliar injury due to chloride in irrigation

water from sprinkler application are provided in table 4.2.6.

Chloride in irrigation water can also reduce the quality of tobacco leaf. Chloride

concentrations >40 mg/L are considered unsuitable for irrigation of tobacco and

some reduction in quality may occur with concentrations in the range 25–40 mg/L

(Gill 1986).
Chapter 4 — Primary industries

page 4.2–10 Version — October 2000

Table 4.2.6 Chloride concentrations (mg/L) causing foliar injury in crops of varying

sensitivitya

Sensitive

<175

Moderately sensitive

175–350

Moderately tolerant

350–700

Tolerant

>700

Almond Pepper Barley Cauliflower

Apricot Potato Maize Cotton

Citrus Tomato Cucumber Sugar beet

Plum Lucerne Sunflower

Grape Safflower

Sorghum

a After Maas (1990)

2 Interaction between chloride in irrigation water and cadmium in soil

Trigger values for assessing chloride levels in irrigation water with respect to

increased cadmium uptake by crops are provided in table 4.2.7.

Table 4.2.7 Risks of increasing cadmium concentrations in crops due to chloride in

irrigation watersa

Irrigation water chloride concentration (mg/L) Risk of increasing crop cadmium concentrations

0–350 Low

350–750 Medium

>750 High

a McLaughlin et al. (1999)

If high chloride concentrations are present in irrigation water, it is recommended

that produce is tested for cadmium concentration in the edible portions (e.g. tubers

for potatoes, leaves for leafy vegetables, grain for cereals, etc.).

4.2.5.3 Sodium

Trigger values for prevention of foliar injury due to sodium in irrigation

water from sprinkler application are provided in table 4.2.8. Trigger values

for specific toxicity effects are provided in table 4.2.9.

Table 4.2.8 Sodium concentration (mg/L) causing foliar injury in crops of varying sensitivitya

Sensitive

<115

Moderately sensitive

115–230

Moderately tolerant

230–460

Tolerant

>460

Almond Pepper Barley Cauliflower

Apricot Potato Maize Cotton

Citrus Tomato Cucumber Sugar beet

Plum Lucerne Sunflower

Grape Safflower

Sesame

Sorghum

a After Maas (1990)
4.2.6 Heavy metals and metalloids

Version — October 2000 page 4.2–11

Table 4.2.9 Effect of sodium expressed as sodium adsorption ratio (SAR) on crop yield and

quality under non-saline conditionsa

Tolerance to SAR and range at

which affected

Crop Growth response under field conditions

Extremely sensitive

SAR = 2–8

Avocado

Deciduous fruits

Nuts

Citrus

Leaf tip burn, leaf scorch

Sensitive

SAR = 8–18

Beans Stunted growth

Medium

SAR = 18–46

Clover

Oats

Tall fescue

Rice

Dallis grass

Stunted growth, possible sodium toxicity,

possible calcium or magnesium deficiency

High

SAR = 46–102

Wheat

Cotton

Lucerne

Barley

Beets

Rhodes grass

Stunted growth

a After Pearson (1960); SAR = Sodium Adsorption Ratio (see Section 4.2.4.1)

4.2.6 Heavy metals and metalloids

Long-term trigger values (LTV) and short-term trigger values (STV) for

heavy metals and metalloids in irrigation water are presented in table 4.2.10.

Concentrations in irrigation water should be less than the recommended

trigger values.

Table 4.2.10 Agricultural irrigation water long-term trigger value (LTV), short-term trigger

value (STV) and soil cumulative contaminant loading limit (CCL) triggers for heavy metals

and metalloidsa

Element Suggested

soil CCLb

LTV in irrigation water (longterm use — up to 100 yrs)STV in irrigation water (shortterm use — up to 20 yrs)

(kg/ha) (mg/L) (mg/L)

Aluminium ND 5 20

Arsenic 20 0.1 2.0

Beryllium ND 0.1 0.5

Boron ND 0.5 Refer to table 9.2.18 (Volume 3)

Cadmium 2 0.01 0.05

Chromium ND 0.1 1

Cobalt ND 0.05 0.1

Copper 140 0.2 5

Fluoride ND 1 2

Iron ND 0.2 10

Lead 260 2 5

Lithium ND 2.5

(0.075 Citrus crops)

2.5

(0.075 Citrus crops)

Manganese ND 0.2 10

Mercury 2 0.002 0.002

Molybdenum ND 0.01 0.05

Nickel 85 0.2 2

Selenium 10 0.02 0.05

Uranium ND 0.01 0.1

Vanadium ND 0.1 0.5

Zinc 300 2 5

a Trigger values should only be used in conjunction with information on each individual element and the potential for

off-site transport of contaminants (Volume 3, Section 9.2.5)

b ND = Not determined; insufficient background data to calculate CCL
Chapter 4 — Primary industries

page 4.2–12 Version — October 2000

The long-term trigger value (LTV) is the maximum concentration (mg/L) of

contaminant in the irrigation water which can be tolerated assuming 100 years of

irrigation, based on the irrigation loading assumptions described in Volume 3,

Section 9.2.5.

The short-term trigger value (STV) is the maximum concentration (mg/L) of

contaminant in the irrigation water which can be tolerated for a shorter period of time

(20 years) assuming the same maximum annual irrigation loading to soil as for LTV.

The LTV and STV values have been developed: (1) to minimise the build-up of

contaminants in surface soils during the period of irrigation; and (2) to prevent the

direct toxicity of contaminants in irrigation waters to standing crops. Where LTV

and STV have been set at the same value, the primary concern is the direct toxicity

of irrigation water to the standing crop (e.g. for lithium and citrus crops), rather

than a risk of contaminant accumulation in soils and plant uptake.

The trigger value for contaminant concentration in soil is defined as the cumulative

contaminant loading limit (CCL). The CCL is the maximum contaminant loading

in soil defined in gravimetric units (kg/ha) and indicates the cumulative amount of

contaminant added, above which site-specific risk assessment is recommended if

irrigation and contaminant addition is continued.

Once the CCL has been reached, it is recommended that a soil sampling and

analysis program be initiated on the irrigated area, and an environmental impact

assessment of continued contaminant addition be prepared. As background

concentrations of contaminants in soil may vary with soil type, and contaminant

behaviour is dependent on soil texture, pH, salinity, etc., it should be noted that

CCLs may be overly protective in some situations and less protective in others. The

CCL is designed for use in soils with no known history of contamination from

other sources. When it is suspected that the soil is contaminated before

commencement of irrigation, background levels of contaminants in the soil should

be determined and the CCL adjusted accordingly.

The trigger values assume that irrigation water is applied to soils and that soils may

reduce contaminant bioavailability by binding contaminants and reducing

concentrations in solution. They may not be suitable for plants grown in soil-less

media (hydroponics or similar methods). The trigger values should only be used in

conjunction with the discussion in Volume 3 on each individual element and the

potential for off-site transport of contaminants.a The assumptions underlying these

trigger values are recognised internationally as a basis for developing irrigation

water quality guidelines.

4.2.7 Nitrogen and phosphorus

Long-term trigger values (LTV) and short-term trigger values (STV) for

nitrogen and phosphorus in irrigation water are presented in table 4.2.11.

They are based on maintaining crop yield, preventing bioclogging of

irrigation equipment and minimising off-site impacts. Concentrations in

irrigation water should be less than the recommended trigger values.

a See Section

9.2.5 for full

details of

methods used
4.2.8 Pesticides

Version — October 2000 page 4.2–13

Table 4.2.11 Agricultural irrigation water long-term trigger value (LTV) and short-term

trigger value (STV) guidelines for nitrogen and phosphorus

Element LTV in irrigation water

(long-term — up to 100 yrs)

(mg/L)

STV in irrigation water

(short-term — up to 20 yrs)

(mg/L)

Nitrogen 5 25–125 a

Phosphorus 0.05

(To minimise bioclogging of irrigation equipment

only)

0.8–12 a

a Requires site-specific assessment (see Section 9.2.6)

The concepts of long-term trigger value (LTV) and short-term trigger value (STV)

developed for metals and metalloids have also been used to develop guidelines for

phosphorus (P) and nitrogen (N).

Excess quantities of N can lead to leaching of N into groundwater and surface

water, over-stimulation of plant growth (decreasing yields) and stimulation of algal

growth in surface water. The LTV for nitrogen has been set at a concentration low

enough to ensure no decreases in crop yields or quality occur. The STV range for

nitrogen has been set to minimise the risk of contaminating groundwater and

surface water and requires site-specific informationa which considers the crop that

is being grown, environmentally significant concentrations, and gaseous losses.

Phosphorus is often the nutrient that stimulates rapid growth of many microorganisms

(i.e. algal blooms). The LTV for P has been set to prevent algal growth in irrigation

water. The STV range for P has been set as an interim range due to the limited data

currently available. Calculation of the interim range considers the fertiliser value of

phosphorus in water, the phosphorus removed from irrigation sites through harvest,

fertiliser inputs, and phosphorus sorption/retention capacities of soils.b

The trigger values provided in table 4.2.11 should only be used in conjunction with

the discussion contained in Volume 3, Section 9.2.6.

4.2.8 Pesticides

Trigger values for pesticides in irrigation water are listed in table 4.2.12.

They consider likely adverse effects of herbicides on crop growth but do not

consider potential impacts on aquatic ecosystems. They are based on

relatively limited information and include only a subset of herbicides (and no

other pesticides) that might be found in irrigation waters.

4.2.9 Radiological quality of irrigation water

Trigger values for the radiological quality of agricultural waters are given in

table 4.2.13.

Radioactive contaminants can originate from both natural and artificial sources and

can potentially be found in surface waters and groundwaters. The main risks to

human health due to radioactivity in irrigation water arise from the transfer of

radionuclides to crop and animal products for human consumption. Cancer is a

potential health hazard for humans associated with exposure to radionuclides in

irrigation water.

a See Section

9.2.6

b An interim

method of

calculating a sitespecific STV is

outlined in Section

9.2.6
Chapter 4 — Primary industries

page 4.2–14 Version — October 2000

Table 4.2.12 Interim trigger value concentrations for a range of herbicides registered in

Australia for use in or near watersa

Herbicide Residue

limits in

irrigation

water (mg/L)b

Hazard to

crops from

residue in

waterc

Crop injury threshold in irrigation water

(mg/L)

Acrolein 0.1 + Flood or furrow: beans 60, corn 60, cotton

80, soybeans 20, sugar-beets 60. Sprinkler:

corn 60, soybeans 15, sugar-beets 15

AF 100 + Beets (rutabaga) 3.5, corn 3.5

Amitrol 0.002 ++ Lucerne 1600, beans 1200, carrots 1600,

corn 3000, cotton 1600, grains sorghum 800

Aromatic solvents

(Xylene)

+ Oats 2400, potatoes 1300, wheat 1200

Asulam ++

Atrazine ++

Bromazil +++

Chlorthiamid ++

Copper sulfate + Apparently above concentrations used for

weed control

2,4-D ++ Field beans 3.5–10, grapes 0.7–1.5, sugarbeets 1.0–10

Dicamba ++ Cotton 0.18

Dichlobenil ++ Lucerne 10, corn 10, soybeans 1.0, sugarbeets 1.0–10, corn 125, beans 5

Diquat +

Diuron 0.002 +++

2,2-DPA (Dalapon) 0.004 ++ Beets 7.0, corn 0.35

Fosamine +++

Fluometuron ++ Sugar-beets, alfalfa, tomatoes, squash 2.2

Glyphosate +

Hexazinone +++

Karbutilate +++

Molinate ++

Paraquat + Corn 10, field beans 0.1, sugar-beets 1.0

Picloram +++

Propanil ++ Alfalfa 0.15, brome grass (eradicated) 0.15

Simazine ++

2,4,5-T ++ Potatoes, alfalfa, garden peas, corn, sugarbeets, wheat, peaches, grapes, apples,

tomatoes 0.5

TCA (Trichloroacetic

acid)

+++

Terbutryne ++

Triclopyr ++

a From ANZECC (1992). These should be regarded as interim trigger values only.

b Guidelines have not been set for herbicides where specific residue limits are not provided, except for a general limit

of 0.01 mg/L for all herbicides in NSW.

c Hazard from residue at maximum concentration likely to be found in irrigation water: + = low, ++ = moderate,

+++ = high
4.2.10 General water uses

Version — October 2000 page 4.2–15

Table 4.2.13 Trigger values for radioactive contaminants for irrigation water

Radionuclide Trigger concentration

Radium 226

Radium 228

Uranium 238

Gross alpha

Gross beta (excluding K-40)

5 Bq/L

2 Bq/L

0.2 Bq/L

0.5 Bq/L

0.5 Bq/L

4.2.10 General water uses

4.2.10.1 pH

To limit corrosion and fouling of pumping, irrigation and stock watering

systems, pH should be maintained between 6 and 8.5 for groundwater

systems and between 6 and 9 for surface water systems.

The pH of water is a measure of its acidity or alkalinity. Generally, pH itself is not a

water quality issue of concern, but it can indicate the presence of a number of related

problems. The greatest hazard with high or low pH is the potential for deterioration

as a result of corrosion or fouling. Values between 4 and 6 should be regarded with

caution and a pH >6 should be maintained to reduce the potential for corrosion. The

upper pH limit for groundwaters should be slightly lower than for surface waters

because of the increased potential for encrustation and fouling. Soil and animal

health will not generally be affected by water with pH in the range of 4–9.

4.2.10.2 Corrosion

Trigger values for assessing the corrosiveness of water are given in table

4.2.14.

Table 4.2.14 Corrosion potential of waters on metal surfaces as indicated by pH, hardness,

Langelier index, Ryznar index and the log of chloride:carbonate ratio

Parametera Value Comments

pH <5

5 to 6

>6

High corrosion potential

Likelihood of corrosion

Limited corrosion potential

Hardness <60 mg/L CaCO3 Increased corrosion potential

Langelier Index <-0.5

-0.5 to 0.5

Increased corrosion potential

Limited corrosion potential

Ryznar Index <6

>7

Limited corrosion potential

Increased corrosion potential

Log of chloride to carbonate ratio >2 Increased corrosion potential

a For further information on these parameters refer to Volume 3, Section 9.2.9.1

Corrosion of pumping, irrigation and stock watering equipment is a common

problem in many agricultural areas of Australia, particularly where groundwater

sources are used. It often results in the deterioration of well and pumping

equipment, pipelines, channels, sprinkler devices and storage tanks, leading to

decreased or uneven water distribution. Corrosion can be caused by chemical,

physical or microbiological processes acting on metal surfaces in contact with
Chapter 4 — Primary industries

page 4.2–16 Version — October 2000

water. Plastics and concrete may also deteriorate, through processes similar to

corrosion, if elevated levels of certain constituents are present.

4.2.10.3 Fouling

Trigger values for assessing the fouling potential of water are given in

table 4.2.15.

Table 4.2.15 Fouling potential of waters as indicated by pH, hardness, Langelier index,

Ryznar index and the log of chloride:carbonate ratio

Parametera Value Comments

pH <7

7 to 8.5

>8.5

Limited fouling potential

Moderate fouling potential (groundwater)b

Increased fouling potential (groundwater)c

Hardness >350 mg/L CaCO3 Increased fouling potential

Langelier Index >0.5

-0.5 to 0.5

Increased fouling potential

Limited fouling potential

Ryznar Index <6

>7

Increased fouling potential

Limited fouling potential

Log of chloride to carbonate ratio <2 Increased fouling potential

a For further information on these parameters refer to Volume 3, Section 9.2.9.1

b For surface waters, pH range 7 to 9

c For surface waters, pH >9

Fouling of agricultural water systems can lead to decreased water quality and yield

as a result of clogging, encrustation and scaling. All parts of the system can be

affected including wells, pumping equipment, pipes and sprinklers. The main

causes of fouling in agricultural water systems can be attributed to physical,

chemical and biological properties of the water.

4.2.10.4 Agricultural chemical preparation

Insufficient information is available to set trigger values for water used to

prepare agricultural chemicals.

Water is the most common additive and diluent used in the preparation of

agricultural chemicals (e.g. pesticides, stock dips and fertilisers) for on-farm use.

Although some agricultural chemicals can withstand a range of water qualities

before performance is substantially affected, it is recommended that good quality

water be used to ensure the desired result.

To check that a particular water is suitable for use with an agricultural chemical, it is

best to make up and test a trial solution first. Specific details on water quality

requirements should be noted from the product label or by contacting the

manufacturer.
Version — October 2000 page 4.3–1

4.3 Livestock drinking water quality

Good water quality is essential for successful livestock production. Poor quality

water may reduce animal production and impair fertility. In extreme cases, stock may

die. Contaminants in drinking water can produce residues in animal products (e.g.

meat, milk and eggs), adversely affecting their saleability and/or creating human

health risks. Animal industries themselves may impair water quality downstream

(e.g. through faecal contamination), highlighting the need for an integrated approach

to land and water management in rural catchments.

Daily water intake varies widely among different forms of livestock and is also

influenced by factors such as climate and the type of feed being consumed.

Average and peak daily water requirements for a range of livestock are given in

Volume 3, Section 9.3.1.

4.3.1 Derivation and use of guidelines

Many factors influence the suitability of waters for livestock watering.

Requirements may differ between animal species (generally tolerances decrease in

the order sheep, cattle, horses, pigs, poultry), and between different stages of

growth and animal condition, and between monogastric and ruminant animals.

Moreover, stock accustomed to good quality water can initially suffer ill effects or

refuse to drink water of poorer quality, but may adjust if introduced gradually.

A review of the scientific literature reveals that most trigger values tend to be based

on field observations rather than rigorous experimentation, although there are notable

exceptions. In the present guidelines, several new trigger values have been calculated

using data on chronic and toxic effect levels on animals. Since derivation of most

trigger values for livestock drinking water needs further validation, they should be

considered interim guidelines at this stage. Further details on the derivation of each

trigger value and a more detailed discussion of all water quality parameters included

in the guidelines are given in Volume 3, Section 9.3.

The scope of the guidelines for livestock drinking water includes biological,

chemical and radiological characteristics that may affect animal health. The

guidelines are trigger values below which there should be minimal risk to animal

health. If the water quality exceeds a trigger value, it is advisable to investigate

further to determine the level of risk.

4.3.2 Biological parameters

4.3.2.1 Cyanobacteria (blue-green algae)

An increasing risk to livestock health is likely when cell counts of Microcystis

exceed 11 500 cells/mL and/or concentrations of microcystins exceed 2.3 μg/L

expressed as microcystin-LR toxicity equivalents. There are insufficient data

available to derive trigger values for other species of cyanobacteria.

Diagnostic procedure

The presence of an algal bloom does not necessarily mean that animals will be

poisoned, so the following steps should be taken to assess the risk from such a

bloom (after Carmichael & Falconer 1993).
Chapter 4 — Primary industries

page 4.3–2 Version — October 2000

1. Establish that animals are drinking the water or eating algal mats from the area

where there is a substantial bloom.

2. Indentify the algae associated with the bloom to determine whether

cyanobacteria are present in numbers large enough to constitute a risk.

3. If necessary, chemically analyse a sample of the bloom to identify and quantify

toxins present.

Since all blooms of cyanobacteria have the potential to be toxic and all livestock are

susceptible, it is prudent to consider all scums toxic until proven safe, as described

above. In the interim, stock should be withdrawn from the water supply and an

alternative source used. Where an alternative source is not available and the bloom is

localised, it may be possible to allow stock to drink from an area on the upwind side

of the bloom. In the long term, prevention of blooms is by far the best strategy, and

water supplies should be managed so that nutrient inputs are minimal.a

4.3.2.2 Pathogens and parasites

Drinking water for livestock should contain less than 100 thermotolerant

coliforms per 100 mL (median value).

It is generally not feasible nor warranted to test livestock drinking water for the

presence of the wide range of water-borne microbial pathogens (bacteria, viruses

and protozoa) and parasites that may affect stock health. In practice, water supplies

are more commonly tested for the presence of thermotolerant coliforms (also

known as faecal coliforms), to give an indication of faecal contamination and thus

the possible presence of microbial pathogens (NHMRC & ARMCANZ 1996).

However, the test does not specifically indicate whether pathogenic organisms are

present or not. Testing for specific organisms may be necessary in these situations

if animal health is affected.

It is recommended that a median value of thermotolerant coliforms is used, based

on a number of readings generated over time from a regular monitoring program.

Investigations of likely causes are warranted when 20% of results exceed four

times the median trigger value (ARMCANZ, ANZECC & NHMRC 1999).b

4.3.3 Major ions of concern for livestock drinking water quality

Many inorganic salts are essential nutrients for animal health, but elevated

concentrations of certain compounds may cause chronic or toxic effects in

livestock. Unless otherwise stated, the trigger values relate to the total

concentration of the constituent, irrespective of whether it is dissolved, complexed

with an organic compound, or bound to suspended solids.c

4.3.3.1 Calcium

Stock should tolerate concentrations of calcium in water up to 1000 mg/L, if

calcium is the dominant cation and dietary phosphorus levels are adequate.

In the presence of high concentrations of magnesium and sodium, or if

calcium is added to feed as a dietary supplement, the level of calcium

tolerable in drinking water may be less.

Calcium is an essential element in the animal diet. However, high calcium

concentrations may cause phosphorus deficiency by interfering with phosphorus

absorption in the gastrointestinal tract.

a See also

Section 9.3.3.1

b Section

9.3.3.2

c Section 9.3.4
4.3.3 Major ions of concern for livestock drinking water quality

Version — October 2000 page 4.3–3

4.3.3.2 Magnesium

Insufficient information is available to set trigger values for magnesium in

livestock drinking water.

Magnesium is an essential element for animal nutrition. In high doses magnesium

can cause scouring and diarrhoea, lethargy, lameness, decreased feed intake and

decreased performance. Drinking water containing magnesium at concentrations up

to 2000 mg/L has been found to have no adverse effects on cattle.a

4.3.3.3 Nitrate and nitrite

Nitrate concentrations less than 400 mg/L in livestock drinking water should

not be harmful to animal health. Stock may tolerate higher nitrate

concentrations in drinking water, provided nitrate concentrations in feed are

not high. Water containing more than 1500 mg/L nitrate is likely to be toxic

to animals and should be avoided.

Concentrations of nitrite exceeding 30 mg/L may be hazardous to animal

health.

Both nitrate and nitrite can cause toxicity to animals, with nitrite being far more

toxic than nitrate. Symptoms of acute poisoning include increased urination,

restlessness and cyanosis, leading to vomiting, convulsions and death.

Confusion can arise concerning trigger values for nitrate and nitrite because

concentrations are sometimes reported on the basis of their respective nitrogen (N)

contents, i.e. as nitrate-N and nitrite-N. Note that trigger values in the present

guidelines are expressed as nitrate and nitrite. The conversions are as follows:

1 mg/L nitrate-N = 4.43 mg/L nitrate, (4.3)

1 mg/L nitrite-N = 3.29 mg/L nitrite. (4.4)

4.3.3.4 Sulfate

No adverse effects to stock are expected if the concentration of sulfate in

drinking water does not exceed 1000 mg/L. Adverse effects may occur at

sulfate concentrations between 1000 and 2000 mg/L, especially in young or

lactating animals or in dry, hot weather when water intake is high. These

effects may be temporary and may cease once stock become accustomed to

the water. Levels of sulfate greater than 2000 mg/L may cause chronic or

acute health problems in stock.

Sulfur is essential for animal nutrition. Excessive concentrations of sulfate in water

typically cause diarrhoea in stock, but animals generally avoid water containing

high sulfate concentrations.

4.3.3.5 Total dissolved solids (salinity)

Recommended concentrations of total dissolved solids in drinking water for

livestock are given in table 4.3.1.

a See Section

9.3.4.2
Chapter 4 — Primary industries

page 4.3–4 Version — October 2000

Table 4.3.1 Tolerances of livestock to total dissolved solids (salinity) in drinking watera

Livestock Total dissolved solids (mg/L)

No adverse

effects on

animals

expected

Animals may have initial

reluctance to drink or there may

be some scouring, but stock

should adapt without loss of

production

Loss of production and a decline

in animal condition and health

would be expected. Stock may

tolerate these levels for short

periods if introduced gradually

Beef cattle 0–4000 4000–5000 5000–10 000

Dairy cattle 0–2500 2500–4000 4000–7000

Sheep 0–5000 5000–10 000 10 000–13 000b

Horses 0–4000 4000–6000 6000–7000

Pigs 0–4000 4000–6000 6000–8000

Poultry 0–2000 2000–3000 3000–4000

a From ANZECC (1992), adapted to incorporate more recent information

b Sheep on lush green feed may tolerate up to 13 000 mg/L TDS without loss of condition or production

Total dissolved solids (TDS) is a measure of all inorganic salts dissolved in water

and is a guide to water quality. For convenience, TDS is often estimated from

electrical conductivity (EC). An approximate conversion of EC to TDS is:

EC (dS/m) x 670 = TDS (mg/L) or, (4.5)

EC (μS/cm) x 0.67 = TDS (mg/L) (4.6)

Salinity is used as a convenient guide to the suitability of water for livestock

watering. If a water has purgative or toxic effects, especially if the TDS

concentration is above 2400 mg/L, the water should be analysed to determine the

concentrations of specific ions.

4.3.4 Heavy metals and metalloids

Many metal elements are essential nutrients for animal health, but elevated

concentrations of certain compounds may cause chronic or toxic effects in

livestock. Stock can tolerate many metal elements in drinking water if they are not

ingesting them in quantity in the diet, because accumulation in the body depends

on the amount ingested from both food and water sources. The trigger values in

table 4.3.2 are the metal concentrations below which there is a minimal risk of

toxic effects. If these values are exceeded the situation should be investigated

further. In some cases higher concentrations may be tolerated, depending on factors

such as total dietary exposure to the metal or levels of other compensating

elements.a Unless otherwise stated, the trigger values relate to the total

concentration of the constituent, irrespective of whether it is dissolved, complexed

with an organic compound, or bound to suspended solids.

a See also

Section 9.3.5
4.3.5 Pesticides and other organic contaminants

Version — October 2000 page 4.3–5

Table 4.3.2 Recommended water quality trigger values (low risk) for heavy metals and

metalloids in livestock drinking water a

Metal or metalloid Trigger value (low risk)a,b

(mg/L)

Aluminium 5

Arsenic 0.5

up to 5c

Beryllium ND

Boron 5

Cadmium 0.01

Chromium 1

Cobalt 1

Copper 0.4 (sheep)

1 (cattle)

5 (pigs)

5 (poultry)

Fluoride 2

Iron not sufficiently toxic

Lead 0.1

Manganese not sufficiently toxic

Mercury 0.002

Molybdenum 0.15

Nickel 1

Selenium 0.02

Uranium 0.2

Vanadium ND

Zinc 20

a Higher concentrations may be tolerated in some situations (details provided in Volume 3, Section 9.3.5)

b ND = not determined, insufficient background data to calculate

c May be tolerated if not provided as a food additive and natural levels in the diet are low

4.3.5 Pesticides and other organic contaminants

In the absence of adequate information derived specifically for livestock

under Australian and New Zealand conditions, it is recommended that the

drinking water guidelines for human health be adopted.

A major concern in rural environments is the potential for pesticide residues to

contaminate water supplies by spray drift, deep percolation, surface runoff, accidental

spillage, or by direct application to water supplies for controlling aquatic weeds. In the

absence of guidelines derived specifically for livestock, the reader is referred to the

Australian Drinking Water Guidelines (NHMRC & ARMCANZ 1996). Readers in

New Zealand are referred to the Drinking-water Standards for New Zealand (New

Zealand Ministry of Health 1995a) and the Guidelines for Drinking-water Quality

Management for New Zealand (New Zealand Ministry of Health 1995b).
Chapter 4 — Primary industries

page 4.3–6 Version — October 2000

4.3.6 Radiological quality of livestock drinking water

Trigger values for the radiological quality of livestock drinking water are

given in table 4.3.3.

Table 4.3.3 Trigger values for radioactive contaminants in livestock drinking water

Radionuclide Trigger value

Radium 226

Radium 228

Uranium 238

Gross alpha

Gross beta (excluding K-40)

5 Bq/L

2 Bq/L

0.2 Bq/L

0.5 Bq/L

0.5 Bq/L

Radioactive contaminants can originate from both natural and artificial sources and

can potentially be found in surface waters and groundwaters. For livestock, the

main water-related risks due to radioactivity arise from the transfer of

radionuclides from irrigation or stock drinking water to animals and animal

products for human consumption. Cancer is a potential health hazard for humans

associated with exposure to radionuclides.
Version — October 2000 page 4.4–1

4.4 Aquaculture and human consumption of aquatic foods

4.4.1 Background

Aquaculture involves the production of food for human consumption, fry for

recreational fishing and natural fisheries, ornamental fish and plants for the

aquarium trade, raw materials for energy and biochemicals, and a number of items

for the fashion industry. With wild fisheries approaching maximum sustainable

levels and many already being over exploited, aquaculture is increasingly

important worldwide as a source of aquatic food and other products.

During 1997–98, almost 31 000 tonnes of product and around 9.3 million juveniles

(mostly finfish fry and ornamental fish) were produced in Australia at an estimated

farm gate value in excess of $517.4 million (O’Sullivan & Roberts 1999). This

represents approximately 25% of total aquatic food production in Australia. The

pearl oyster, southern bluefin tuna, salmonid, edible oyster and prawn industries

represent the major commercial aquaculture sectors economically, totalling more

than 90% of overall aquaculture production.

The main culture species in New Zealand are green shell mussels, Pacific salmon

and Pacific oysters. According to the New Zealand Fishing Industry (Treyton

Maldoc, pers. com. 1999), annual production of these species totalled almost

50 000 tonnes, with an estimated value of around $160 million. Aquaculture now

contributes over 13% of all New Zealand aquatic food exports.

Within the growing aquaculture industry, it is well accepted that satisfactory water

quality is needed for maintaining viable aquaculture operations. Poor water quality

can result in loss of production of culture species, and can also reduce the quality

of the end product. Production is reduced when influent water contains enough

contaminants to impair development, growth or reproduction, with the ultimate

result being death. Quality is reduced when low levels of a contaminant cause no

obvious adverse effects but gradually accumulate in the culture species to the point

where it poses a potential health risk to human consumers. Thus, both these issues

needed to be considered if useful and usable guidelines are to be provided for the

aquaculture industry.

This section provides water quality guidelines for influent (i.e. water that is

entering the aquaculture operation) or source water quality, and it also addresses

the safety of aquatic foods for human consumers, whether the foods be produced

by aquaculture, or commercial, or recreational or indigenous fishing. It is the first

set of joint guidelines to have been provided for the protection of aquaculture in

Australia and New Zealand. Note that these guidelines for protecting the health of

commercial fish speciesa do not apply to recreational and commercial fisheries

based upon wild populations of aquatic organisms. Wild fish stocks are dependent

on healthy ecosystems to support them thoughout their life cycle (e.g. for feeding,

breeding, habitat). Hence, for the protection of wild fish stocks it is best to apply

the water quality guidelines for managing aquatic ecosystems.b

b Chapter 3

a See Section

4.4.4
Chapter 4 — Primary industries

page 4.4–2 Version — October 2000

4.4.2 Philosophy

In developing these guidelines, the objective was to provide information and

guidance that would:

• promote the quality of water necessary for use by the aquaculture industry; and

• protect human consumers of harvested aquatic food species.

4.4.2.1 Protection of cultured fish, molluscs and crustaceans

The guidelines for protecting aquaculture species have been developed to assist

water managers to maintain an appropriate level of water quality for existing and

future aquaculture activities. The water quality guidelines will provide a basis for

aquaculture management decisions, such as:

• environmental planning and management,

• environmental assessment and monitoring requirements,

• appropriate environmental zoning and legislation,

• appropriate species and suitable site selection,

• site capacity,

• farm design criteria,

• stocking densities and feeding regimes,

• production schedules.

4.4.2.2 Protection of human consumers of aquatic foods

Standards for the protection of human consumers of aquatic foods are of paramount

importance to the viability of the aquaculture industry. To maintain demand, the

aquaculture and fishing industries must ensure the highest quality of their products,

both from a visual and, more importantly, from a human health perspective. Under a

treaty between Australia and New Zealand (ANZFA 1996), the Australia New Zealand

Food Authority (ANZFA) develops and administers uniform (statutory) standards for

chemical contamination in foods (including aquatic foods) that are likely to affect

human health. Unlike the water quality guidelines, the ANZFA food standards are

enforceable through legislation. Guidelines are also provided in this section against

biological contaminants and against the tainting of aquatic animal flesh.

4.4.3 Scope

As the aquaculture guidelines for Australia and New Zealand are a new

development, they have drawn extensively on recent overseas guidelines for

aquaculture as well as on the personal experiences of a number of local industry

specialists. The guidelines address the following issues:

• protection of the health of culture species from water-borne contaminants

(chemicals, elements, microorganisms, toxins, etc.) during the growing period

(pre-harvest), but not during post-harvest processes (e.g. slaughter, processing,

transport, marketing);

• the effects of water quality on adult forms of cultured species, recognising that

larval and juvenile stages may have lower tolerance levels than the adult stages;

• the protection of human consumers of harvested aquatic food species from the

toxic effects of chemical and biological contaminants and from tainted flesh.
4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and crustaceans

Version — October 2000 page 4.4–3

The guidelines do not address effluent water quality from aquaculture activities;

however, aquaculturists need to manage their operations with downstream water

quality in mind. Effluent water quality is regulated by state and federal government

legislation and regulations in Australia, and through the Resource Management Act

and Industry Agreed Implementation Standards in New Zealand. In addition, as

stated above, the guidelines in Section 4.4.4 are only concerned with the protection

of cultured, not wild species.

Given the limited information on contaminant accumulation in aquaculture species,

it has not been possible to provide water quality guidelines that will guarantee that

the Australian and New Zealand food standards will be achieved. Therefore, the

guidelines for the protection of human consumers of aquatic foods are intended to

be used in conjunction with the Food Standards Code (ANZFA 1996, and updates)

to protect the health of human consumers of aquatic foods from the aquaculture

industry. These standards are continually under review and can be examined on the

appropriate web sites (for Australia: www.anzfa.gov.au; for New Zealand:

www.anzfa.govt.nz).

Precautionary comments and discussion on the limitations of the guidelines are

provided below in Section 4.4.6.a

4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and

crustaceans

4.4.4.1 Overview of approach

There are many aquaculture species in Australia and New Zealand and information

is generally lacking on most of them, so all finfish, mollusc and crustacean species

were divided into eight indicative groups. Then toxicity and tolerance data were

reviewed for one or two representative species within those groups, with the

species being chosen according to the level of production and availability of

scientific data. Where discrepancies in the data were identified, the more

conservative data were generally used. The species groups and representative

species are summarised in table 4.4.1.

Justification for selecting the representative species is provided in Section 9.4.1.4

(Volume 3). As indicated in table 4.4.1, a range of aquatic plants, reptiles and

invertebrates that are cultured were not included in the list of representative

species. In 1997/98 the production of these species contributed less than 1.5% of

the total value of aquaculture production in Australia (O’Sullivan & Roberts 1999),

with the amount of relevant literature or information about them being

correspondingly small.

Guideline values were determined in several ways, depending on the quantity and

quality of information. Where they were available, appropriate guidelines for the

protection of aquaculture from other countries (e.g. DWAF 1996, Zweig et al.

1999) were applied. In some cases, guideline values were based on acceptable

risks, according to the value judgements or professional judgements of local

aquaculture specialists. When neither of the above approaches could be used, the

water quality requirements for the eight indicative species groups were reviewed to

determine a guideline value.b Discussion of the confidence levels for these

guidelines is provided in Section 9.4.1.5 (Volume 3).

b Sections

9.4.1.4, 9.4.1.5

a See Section

9.4.1 for more

detail
Chapter 4 — Primary industries

page 4.4–4 Version — October 2000

Table 4.4.1 Representative aquaculture species, occurrence and culture status

Species group Representative species1 Occurrence Aquaculture status2

Freshwater fish rainbow trout

silver perch

Australia/New Zealand

Australia

commercial/none

commercial

Marine fish snapper

flounder/whiting

Australia/New Zealand

Australia/Australia

commercial/commercial

experimental/experimental

Brackish water or

euryhaline fish

barramundi

black bream

Australia

Australia

commercial

experimental

Freshwater

crustaceans

marron

yabbies

red claw

freshwater shrimp

Australia

Australia

Australia

Australia/New Zealand

commercial

commercial

commercial

experimental/commercial

Marine crustaceans black tiger prawns

kuruma prawns

Australia

Australia

commercial

commercial

Edible bivalves Sydney rock oysters

Pacific oysters

blue mussels

green shell mussels

Australia

Australia/New Zealand

Australia/New Zealand

New Zealand

commercial

commercial/commercial

commercial/none

commercial

Pearl oysters golden lip Australia commercial

Gastropod/molluscs abalone/paua

trochus

Australia/New Zealand

Australia

commercial/commercial

experimental

1 The groups of aquaculture species not included in this list are: seaweeds and aquatic plants; crocodiles; a range of

live feed and microalgal species; sea cucumbers (beche-de-mer), sponges and other invertebrates.

2 commercial = products offered for sale; experimental = production but no sales; none = species occurs but no

culture is undertaken

The guidelines are provided in the following four categories:

• physico-chemical stressors,

• inorganic toxicants,

• organic toxicants,

• pathogens and biological contaminants.

General guideline values for the aquaculture of freshwater and saltwater (brackish

and marine water) are recommended. In addition, specific guideline values are

provided for species groups for which information is available on their water quality

requirements. Information sources used to derive the water quality guidelines for

protection of aquaculture species are listed in Section 9.4.1.4 (Volume 3).

4.4.4.2 Using the guidelines

The water quality guidelines can be used with reasonable confidence to assess

ambient water quality for aquacultural uses. Where specific water quality

guidelines cannot be given for the protection of aquaculture species, use the

guidelines for the protection of aquatic ecosystems.a

Many different aquaculture production systems and species are used in Australia

and New Zealand across a wide range of environmental conditions, so it should not

be assumed that one set of specific values will apply equally in all situations.

Local, site-specific information will be needed to supplement the broad

information provided in this chapter. This might include information on specific

culture species, or local water quality variables that could affect the bioavailability

and toxicity of metals (e.g. hardness, dissolved organic matter, pH, temperature).

a See Chapter 3
4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and crustaceans

Version — October 2000 page 4.4–5

Details of factors that could affect toxicant bioavailability are provided in Section

8.3.5 (Volume 2).

Figure 4.4.1 is a decision tree for determining water quality guidelines for the

protection of aquaculture species; it includes a number of factors that might modify

the guideline values. Specialist assistance may be required to complete the steps

which involve chemical speciation/complexation, and likewise to conduct toxicity

tests should they become necessary.a

Note that a user can make a decision on the risk-based framework and leave the

process at any level. However, the further through the process one moves, the

greater the confidence in the level of risk. A worked example of the use of the

decision tree for an aquaculturist planning to culture prawns is provided in Section

9.4.2 (Volume 3).

If ambient water quality exceeds the guideline value for any parameter then there

could be a significant risk of an impact on aquacultural activities, and further

investigations should be undertaken, in accordance with the decision framework in

figure 4.4.1. If ambient water quality remains below the guideline values, risk can

be deemed to be low. However, this cannot be taken as a guarantee that problems

will not occur in the future.

It is unrealistic to expect an aquaculture operation to measure all of the water

quality parameters. However, knowledge of activities upstream of the operation

that may be contributing to contaminants in the influent water should serve to

identify which of the parameters might be of particular concern.

4.4.4.3 The guideline values

Tables 4.4.2 and 4.4.3 provide the recommended water quality guideline values for

physico-chemical parameters and toxicants, respectively, to be applied for use in

general freshwater and saltwater (brackish and marine water) aquaculture. Where

guideline values are available for some or all of the species groups outlined in table

4.4.1, they have been incorporated in Section 9.4.2 (Volume 3), and can be used

where guidance is sought for a particular species group. A short summary for each

category (i.e. physico-chemical, inorganic, etc.) is also provided after the tables.

Section 9.4.2 (Volume 3) also contains further background information on each

water quality parameter, including a description of how the recommended

guideline value was determined.

a See Section

3.4.3, Vol. 1;

Section 8.3.6,

Volume 2
Chapter 4 — Primary industries

page 4.4–6 Version — October 2000

Water quality parameter characterisation

Test against general guideline values

(see tables 4.4.2 and 4.4.3)

within range outside range

Low risk

(water quality acceptable) Test against specific guideline

values for species group

(see tables in section 9.4.2)

Low risk

(water quality acceptable) Examine factors influencing 

toxicity (see section 8.3.5)

Low risk

(water quality acceptable)

within range outside range

Conduct acute toxicity testing

Not toxic Toxic

Culture not 

recommended

(water quality unacceptable)

Test against guideline values

(bioavailable concentration )

within range

outside

 range

Conduct chronic toxicity testing

Not toxic Toxic

Culture not 

recommended

(water quality unacceptable)

Low risk

(water quality acceptable)

Figure 4.4.1 Decision tree for determining if water quality is acceptable

for the protection of aquaculture species
4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and crustaceans

Version — October 2000 page 4.4–7

Table 4.4.2 Physico-chemical stressor guidelines for the protection of aquaculture species

Measured parameter Recommended guideline (mg/L)

Freshwater production Saltwater production

Alkalinity ≥205 >203

Biochemical oxygen demand

(BOD5)

<151 ND

Chemical oxygen demand (COD) <401 ND

Carbon dioxide <10 <15

Colour and appearance of water 30–402 (Pt-Co units) 30–402 (Pt-Co units)

Dissolved oxygen >53 >53

Gas supersaturation <100%6 <100%6

Hardness (CaCO3) 20–1005 NC6

pH 5.0–9.0 6.0-9.0

Salinity (total dissolved solids) <30006 33 000–37 0006

(3000–35 000 Brackish)6

Suspended solids <40 <10

(<75 Brackish)

Temperature <2.0°C change over 1 hour4 <2.0°C change over 1 hour4

1 Schlotfeldt & Alderman (1995)

2 O’Connor pers. comm.

3 Meade (1989)

4 ANZECC (1992)

5 DWAF (1996)

6 Lawson (1995)

Others are based on professional judgements of the project team.
Chapter 4 — Primary industries

page 4.4–8 Version — October 2000

Table 4.4.3 Toxicant guidelines for the protection of aquaculture species

Measured parameter Guideline (μg/L)

Freshwater production Saltwater production

INORGANIC TOXICANTS (HEAVY METALS AND OTHERS)

Aluminium <30 (pH >6.5)1

<10 (pH <6.5)

<101

Ammonia (un-ionised) <20 (pH >8.0) coldwater2

<30 warmwater2

<100

Arsenic <501,2 <301,2

Cadmium (varies with hardness) <0.2–1.82 <0.5–51

Chlorine <31 <31

Chromium <202 <20

Copper (varies with hardness) <52 <53

Cyanide <51 <51

Fluorides <204 ND

Hydrogen sulfide <12 <2

Iron <101 <101

Lead (varies with hardness) <1–74 <1–74

Magnesium <15 0001 ND

Manganese <101,5 <101,5

Mercury <1 <1

Nickel <1001 <1001

Nitrate (NO3

-

) <50 0006 <100 0003,7

Nitrite (NO2) <1001,7 <1001,7

Phosphates <1002 <50

Selenium <101 <101

Silver <31 <31

Tributyltin (TBT) <0.0261 <0.011

Total available nitrogen (TAN) <10001 <10001

Vanadium <1001 <1001

Zinc <51 <51

ORGANIC TOXICANTS (NON-PESTICIDES)

Detergents and surfactants <0.18 ND

Methane <65 0009,10 <65 0009,10

Oils and greases (including petrochemicals) <3006 ND

Phenols and chlorinated phenols <0.6–1.76 ND

Polychlorinated biphenyls (PCBs) <21 <21

PESTICIDES

2,4-dichlorophenol <4.02 ND

Aldrin <0.012,3,8 ND

Azinphos-methyl <0.012 ND

Chlordane <0.0111 0.00411

Chlorpyrifos <0.0012 ND

DDT (including DDD & DDE) <0.00152 ND

Demton <0.0111 ND

Dieldrin <0.0052 ND

Endosulfan <0.0032,11 0.00111

Endrin <0.0022 ND

Gunthion (see also Azinphos-methyl) <0.0111 ND

Hexachlorobenzole <0.000016 ND

Heptachlor <0.0052 ND

Lindane <0.0111 0.00411

Malathion <0.15,11 ND

Methoxychlor <0.0311 ND

Mirex <0.0012,11 ND

Paraquat ND <0.01

Parathion <0.0411 ND

Toxaphene <0.0022 ND

ND: Not determined — insufficient information; NC: Not of concern; 1. Meade (1989); 2. DWAF (1996); 3. Pillay (1990); 4. Tebbutt (1972);

5. Zweig et al. (1999); 6. Schlotfeldt & Alderman (1995); 7. Coche (1981); 8. Langdon (1988); 9. McKee & Wolf (1963); 10. Boyd (1990);

11. Lannan et al. (1986). Others are based on professional judgements of the project team.
4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and crustaceans

Version — October 2000 page 4.4–9

1 Physico-chemical stressors

A number of naturally-occurring physico-chemical stressors can cause adverse

effects on aquaculture operations when influent water values are too high and/or

too low. These guidelines address 11 physico-chemical stressors that are

considered of importance to aquaculture operations. Many of these should also be

regularly monitored in the culture system to ensure that the aquatic organisms are

being held in conditions conducive to survival and growth. Some of the major

stressors are summarised below.a

Dissolved oxygen (DO) is a basic requirement for aquaculture species (Zweig et al.

1999). The amount of oxygen required by aquatic animals is quite variable and

depends on species, size, activity, water temperature, condition, and the DO

concentration itself (Boyd 1990). Thus, some species are more sensitive to low

levels of oxygen than others. Daily fluctuations of DO in impounded waters are

much higher than those in the open sea or running waters, with low levels often

occurring at dawn, and high levels in the late afternoon (Boyd 1990). The most

common cause of low DO levels in an aquaculture operation is contamination by

biodegradable organic substances resulting in a high BOD; the problem is further

exacerbated at higher temperatures.

Water hardness, a total measure of the major cations (predominantly calcium and

magnesium), is an important parameter in freshwaters, mostly because it can have a

major effect on the toxicity of metals. In addition, some aquaculture species have

specific calcium requirements for bone or exoskeleton formation, and calcium is also

necessary for proper osmoregulation. Water hardness (measured as mg CaCO3/L)

can range from <1 (very soft) to >400 mg/L (very hard).

The pH of influent water refers to the log10 of the hydrogen ion concentration, or,

more simply, how acidic or basic the water is. The pH is interdependent with a

number of other water quality parameters including carbon dioxide, alkalinity and

hardness. It is known to influence the toxicity of hydrogen sulfide, cyanides, heavy

metals, and ammonia (Klontz 1993), and it can also be toxic in its own right. The

pH levels in natural waters vary enormously and the aquaculturist should ensure

that culture species are adapted to living in the conditions existing in the

aquaculture operation.

Salinity is an important limiting factor in the distribution of many aquatic animals,

and therefore it is an important parameter for aquaculture. In addition, salinity

requirements can vary for particular species depending on their life cycle stage.

Outside their natural salinity ranges, aquatic animals must expend considerable

energy on osmoregulation at the expense of other processes such as growth. Salinity

ranges are 0.05–1.0 gL-1 for freshwaters, 0.5–>30 gL-1 for estuarine waters, 30–40

gL-1 for marine waters, and can exceed 40 gL-1 for hypersaline/brackish waters.

Suspended solids and turbidity can have major effects on aquaculture operations.

Suspended solids include phytoplankton, zooplankton and bacterial blooms,

suspended organic and humic acids, and suspended silt and clay particles. All these

components contribute to some extent to increased turbidity. In some instances this

is advantageous, because it inhibits the growth of nuisance algae and macrophytes.

However, suspended solids can cause gill irritations and tissue damage to aquatic

animals, while they can also shield food organisms and clog filters (Zweig et al.

1999). Smothering effects caused by suspended solids settling on sessile

a See also

Section 9.4.2.1
Chapter 4 — Primary industries

page 4.4–10 Version — October 2000

aquaculture species (e.g. mussels, oysters) can also present problems (Duchrow &

Everhart 1971).

In summary, it should be highlighted that physico-chemical parameters vary widely

in natural waters, and aquatic organisms have a wide range of tolerances and

adaptive capacities. Thus, it is extremely difficult to recommend broadly applicable

guidelines.

2 Inorganic toxicants (heavy metals and others)

A wide range of inorganic toxicants, particularly heavy metals, can be a problem in

freshwater, brackish water and inshore marine aquaculture, especially in areas of

human habitation that may be polluted. Trace quantities of metals are present in

natural waters; however, their concentrations are generally greater in the vicinity of

industrial processes (ore mining and processing, smelting plants, rolling sheet

metal mills, textile and leather industries) and exhaust gases of motor vehicles and

burning of other fossil fuels. These guidelines provide information on 27 inorganic

toxicants. Those of greatest concern to fisheries (including aquaculture) include

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

zinc (Svobodova et al. 1993). Other inorganic toxicants include ammonia, chlorine,

cyanide, fluoride, hydrogen sulfide, nitrite, nitrate and phosphates. As mentioned

above, the levels of calcium and magnesium are also important because they

influence the hardness of the waters.a

Speciation of metals is important in determining toxicity to aquatic organisms

because it influences metal bioavailability. Water quality guidelines for metals in

aquatic ecosystems have typically been based on total concentrations; yet it is now

well established that the chemical form or speciation of metals critically influences

their bioavailability (i.e. their ability to penetrate a biological cell membrane) and

toxicity to aquatic organisms.b

Most studies of the toxicity of heavy metals to fish and other aquatic organisms have

shown that the free (hydrated) metal ion is the most toxic form, and that toxicity is

related to the activity of the free (dissolved) metal ion (e.g. Cu2+ or Zn2+) rather than

to total metal concentration (including adsorbed, chelated or complexed forms)

(Florence & Batley 1988, Boyd 1989). Heavy metal toxicity also can be affected by

pH, hardness, alkalinity, dissolved oxygen, temperature and turbidity (SECL 1983).

In pond water, heavy metals can be adsorbed onto clay particles and chelated by

organic matter so that they remain in solution but may not have an adverse effect on

fish or crustaceans (Boyd 1990). Duration of exposure, interaction with other toxic

agents and species can affect the biological response to these toxic metals

significantly (e.g. mercury and methane give rise to methyl mercury).

Guidelines based on total concentrations may be over-protective, since only a

fraction of the total concentration will generally be bioavailable, especially in

samples containing appreciable concentrations of particulate matter. Thus, it is

important to measure the bioavailable metal fraction.c Importantly, Svobodova et

al. (1993) noted that the toxic action of metals is particularly pronounced in the

early stages of development of the fish.

3 Organic toxicants

Organic toxicants can present a problem to all types of aquaculture operations. The

types of organic chemicals considered in these guidelines are detergents and

surfactants, hydrocarbons derived from human activities (namely petroleum

a See Section

9.4.2.1/7

b Sections

8.3.5.16 and

9.4.2.2

c Section 8.3.5
4.4.4 Water quality guidelines for the protection of cultured fish, molluscs and crustaceans

Version — October 2000 page 4.4–11

hydrocarbons), a large number of pesticides, phenolic compounds, and

polychlorinated biphenyls. Most of these originate from domestic, agricultural or

industrial activities, and some are also used by aquaculture operations.

No data were available to provide guidelines for antibiotics and antimicrobials, but

it is best to take due care when using such chemicals in aquaculture operations.

Detergents and surfactants are widely used in domestic and industrial operations,

and can often be detected in natural waters receiving domestic and industrial

effluent (Svobodova et al. 1993), while on-farm activities may also be major

sources of such chemicals. There is limited toxicity information for detergents and

surfactants, although a general guideline value was derived for freshwaters.

Petroleum hydrocarbons are among the most widely processed and distributed

chemical products in the world (Zweig et al. 1999). Although high levels of

petroleum hydrocarbons can result in mortalities and major losses of production,

the major concern to the aquaculture industry is the tainting of culture animals with

off-flavours (Zweig et al. 1999a). Given the large number of petroleum-derived

hydrocarbons and their wide ranges of toxicities, it is difficult to derive meaningful

guidelines (SECL 1983), although some general guidelines have been

recommended.

The pesticides represent a large and complex group of organic toxicants because

they incorporate insecticides, acaricides, herbicides, algicides and fungicides. In

addition, the behaviour (e.g. persistence, partitioning) and toxicity of pesticides

varies greatly, making it difficult to generalise about risks. Pesticides generally

enter water from sources in the primary industry sector, including aquaculture, but

primarily agriculture. Table 4.4.2 presents guideline values only for those

pesticides for which a general freshwater or saltwater value can be recommended.

A more comprehensive list of pesticide guideline values for specific species groups

is provided in Section 9.4.2.3/4 (Volume 3). Given the limited information on the

effects of pesticides on culture species, it is also worthwhile consulting the

guidelines for aquatic ecosystem protection.b

Other organic compounds of concern include phenols and polychlorinated biphenyls

(PCBs). Phenolic compounds originate from the distillation of fossil fuels, the

degradation of pesticides, natural (SECL 1983) and other sources. They can result in

effects ranging from toxicity to the tainting of flesh. Guideline values are

recommended for freshwater and saltwater, while some guideline values for specific

phenols are recommended for freshwater fish culture. The PCBs are extremely

persistent lipid soluble chemicals that are of great environmental concern

(Svobodova et al. 1993). It is extremely difficult to recommend guidelines for PCBs

because of their large number and the wide spectrum of toxicity they exhibit.

However, general guideline values are recommended for freshwater and saltwater.

4 Pathogens and biological contaminants

Pathogens and biological contaminants also need to be considered for aquaculture

operations, and include algal blooms and algal toxins, bacteria, viruses and

parasites. As noted by Zweig et al. (1999), high concentrations of pathogenic

organisms are commonly found in waters polluted by human sewage and animal

wastes. No guidelines are provided for pathogens and biological contaminants

because their effects can vary considerably between the type of contaminant or

species of pathogen, and the culture species. Nevertheless, Section 9.4.2.4

a Also see

Section

4.4.5.3/3 below

b Chapter 3,

Volume 1
Chapter 4 — Primary industries

page 4.4–12 Version — October 2000

(Volume 3) provides useful background information and some guidance on how to

manage for them. Brief summary information is provided below.

Algal blooms arise from a series of processes but commonly from eutrophication

(addition of excess nutrients). Direct and indirect results of algal blooms include

increased pH, depleted oxygen (anoxia), the production and release of algal toxins,

and gill obstruction and irritation in fish. Algal toxins can also accumulate in

culture species, resulting in potential risks to human consumers.a

It has been suggested that culture organism mortality due to disease poses a more

direct threat to the aquaculture industry than pollutants (Handlinger 1996).

Aquaculture source waters contain a certain number of bacteria, viruses, fungi,

parasites and other organisms, which, given certain environmental conditions, can

contribute to impaired health of the culture species. Thus, the maintenance of optimal

water quality appears to be the best defence against infections by these organisms

(DWAF 1996). Some equipment that reduces the amount of incoming potential

pathogens includes inflow filters that retain particles (to which most of the bacteria

will be attached) and ultra-violet (UV) sterilisers. Reducing the level of infectious

organisms contributes to better culture health, reduced need to treat animals with

chemicals and drugs, and lower production costs.

4.4.5 Water quality guidelines for the protection of human consumers of

aquatic foods

4.4.5.1 Overview of approach

Although guidelines are provided for biological contaminants and for the tainting of

animal flesh, a search of the available data has produced insufficient information for

deriving water quality guidelines that will ensure the Australian and New Zealand

food standards will be met. Consequently, relevant food standards from the Food

Standards Code (ANZFA 1996, and updates) established by the Australia New

Zealand Food Authority (ANZFA) are provided as guidance and discussed below.b

4.4.5.2 Using the guidelines

The guidelines for the protection of human consumers of aquatic foods are

intended to be used in conjunction with the Food Standards Code (ANZFA 1996,

and updates) to protect the health of human consumers of aquatic foods from the

effects of toxicants, whether the foods be derived from aquaculture, recreational

fishing, commercial fishing or indigenous fishing. Essentially, they provide useful

background information and some guidance to complement the ANZFA food

standards. In particular, they give detailed information on measures for predicting

the tissue concentrations of contaminants before, rather than after, harvest. Such

approaches may form the basis for the future development of guidelines for the

protection of human consumers of aquatic foods.

The ANZFA food standards for contamination of aquatic foods are enforceable

through legislation and must be adhered to. However, it is important to note that at

the time of publication of these Water Quality Guidelines, the ANZFA food

standards were under review and subject to change. Thus, aquaculturists and other

users of these guidelines should ensure they obtain the most recent ANZFA

information (for Australia: www.anzfa.gov.au; for New Zealand:

www.anzfa.govt.nz).

a See Section

4.4.5.3/2

b Section 9.4.3
4.4.5 Water quality guidelines for the protection of human consumers of aquatic foods

Version — October 2000 page 4.4–13

4.4.5.3 The guidelines

The food standards developed by ANZFA and published in the Food Standards

Code (ANZFA 1996, and updates) aim to protect consumers from chemically

contaminated foods, including aquatic species. Standards for aquatic species are

based on the notion of acceptable daily intake (ADI) or acceptable weekly intake

(AWI). See Zweig et al. (1999) for the World Health Organization (WHO)

provisional tolerable weekly intakes for selected elements, as well as import

regulations for residues. Guidelines are also provided for biological contaminants

and for the tainting of animal flesh.

1 Chemical contaminants (toxicants)

Chemical contaminants can be categorised into three broad groups:a

i) Inorganic toxicants (mostly heavy metals)

Inorganic toxicants (mainly heavy metals) are a potential problem for human

health, particularly through bivalved molluscs in which bioaccumulation increases

the concentrations of inorganic toxicants. The rate of accumulation is speciesspecific and depends on the mechanisms of absorption and tissue distribution.

ii) Organic toxicants (e.g. hydrocarbons, pesticides)

The broad group comprising organic toxicants such as hydrocarbons and

pesticides includes synthetic compounds which through either bioaccumulation or

residue concentrations are potentially toxic to human consumers of contaminated

aquatic foods.

iii) Radionuclides (radioactive elements)

At present, ANZFA does not specify maximum permitted concentrations (MPCs)

for radionuclides in edible tissues. Many countries have limits set on imported

foods, particularly for caesium-137 (Cs-137). Environmental levels of Cs-137 are

considerably lower in the southern hemisphere than in the northern hemisphere,

and exporters in Australia and New Zealand should not generally experience

difficulty in meeting such limits.

2 Biological contaminants

There are a number of biological contaminants that can affect human consumers of

aquatic foods. The guidelines for biological contaminants are based on either a

concentration of the contaminant in the water (e.g. cells/L) or the level which is

considered safe in edible soft tissue of fish, crustaceans and molluscs (e.g. mg/kg,

number/g). Summary information on the major biological contaminants is provided

below.b

i) Bacteria

Aquatic bacterial food-borne diseases in humans can originate either from bacteria

naturally present in water and/or sediments, or from bacteria introduced into

aquatic environments through human and/or animal faeces. Aquatic foods can

become contaminated with bacteria from exposure within the aquatic environment

and/or during post-harvest activities. The present guidelines only deal with

exposure within the aquatic environment.

The guidelines in table 4.4.4 are provided to assist managers to minimise the

exposure of human consumers of aquatic food species (e.g. recreational fishermen)

to bacterial borne disease.

b Section

9.4.3.3 for

ANZFA

standards

a See Section

9.4.3.2 (Vol. 3)

for ANZFA

standards
Chapter 4 — Primary industries

page 4.4–14 Version — October 2000

Table 4.4.4 Guidelines for the protection of human consumers of fish and other aquatic

organisms from bacterial infection

Toxicant Guideline in shellfishing water Standard in edible tissue

Faecal (thermotolerant)

coliforms

The median faecal coliform bacterial

concentration should not exceed

14 MPN/100 mL, with no more than

10% of the samples exceeding

43 MPN/100 mL

Fish destined for human

consumption should not exceed a

limit of 2.3 MPN E. coli /g of flesh

with a standard plate count of

100 000 organisms/g

MPN: Most probable number

The guideline for faecal (thermotolerant) coliforms should only be used in conjunction with the data from a sanitary

survey of the shellfish harvesting areas for the purpose of harvesting area classification. Source: USEPA (1986),

NAS/NAE (1973), IWBDE (1972).

A two-tiered approach is usually used to reduce bacterial loads in cultured species:

• risk-based classification of waters to allow only certain waters and times for

rearing or harvesting of shellfish;a

• treatment of shellfish to remove or destroy the bacteria (e.g. heat treatment or

irradiation).

Depuration is an integral part of removing bacteria from shellfish, and is a statutory

requirement in NSW only.b

ii) Viruses

Viruses that infect humans following consumption of aquatic food are of human

origin, having entered aquatic ecosystems in sewage effluent. These enteric viruses

are able to remain viable in the aquatic environment for long periods (Goyal et al.

1984).

Shellfish are able to accumulate viruses in their gastrointestinal tracts, digestive

glands and other tissues, but the rate of accumulation is dependent on the viral

species and the shellfish species. Viruses are very difficult to detect, and other

species (e.g. Escherichia coli, faecal coliforms) are usually used to indicate

exposure to sewage-related pollution. While such sanitary surveys may not be as

reliable as once thought, they are still relevant and are used in Australia and New

Zealand as well as a number of other countries.c

Heat treatment and depuration are generally not as efficient at reducing viral loads

as they are bacterial loads. Normal cooking/steaming times for shellfish may not be

sufficient to inactivate viruses (University of California, Davis 1997). Similarly,

depuration may not remove all viruses from shellfish (Jackson & Ogburn 1998).

iii) Parasites

There is no evidence of transmission of parasites to humans following aquatic food

consumption in Australia or New Zealand. Thus, no guidelines are provided.

However, the presence of parasites, cysts and necrotic tissue resulting from

parasitic infections will reduce the marketability of product.

iv) Marine biotoxins

A number of marine biotoxins, most of them associated with marine algae, represent

a threat to human consumers of aquatic foods. Aquatic animals accumulate the toxins

when they graze on the algae or on other consumers of the algae.

a See Section

4.4.5.3/4

b Section

9.4.3.3/1

c See part 4

below

& Section

9.4.3.3/2
4.4.5 Water quality guidelines for the protection of human consumers of aquatic foods

Version — October 2000 page 4.4–15

There are five recognised types of microalgal toxins:

• paralytic shellfish poisoning (PSP),

• diarrhetic shellfish poisoning (DSP),

• amnesic shellfish poisoning (ASP),

• neurotoxic shellfish poisoning (NSP),

• ciguatera fish poisoning (CFP).

Three naturally-occurring toxins that are not related to algae are (University of

California, Davis 1997):

• gempylotoxin,

• tetramine,

• tetrodotoxin.

Important background information on the above biotoxins is provided in Section

9.4.3.3 (Volume 3), including guidelines for water and standards for edible tissue

(MBMB 1996, K Jackson pers. comm. 2000). For a detailed discussion of

biotoxins in New Zealand, refer to MBMB (1996). University of California, Davis

(1997) also provides useful guidance and background information.

3 Off-flavour compounds

Off-flavour compounds, also known as tainting substances, can seriously affect the

palatability of aquatic food. They can result in major adverse impacts to the

aquaculture and wild-capture fishing industries. Table 4.4.5 lists threshold

concentrations at which tainting will occur for a variety of off-flavour

compounds.a

a See also

Section 9.4.3.4
Chapter 4 — Primary industries

page 4.4–16 Version — October 2000

Table 4.4.5 Guidelines for chemical compounds in water found to cause tainting of fish flesh and other aquatic organisms

Parameter Estimated threshold level in water (mg/L)

Acenaphthene 0.02

Acetophenone 0.5

Acrylonitrile 18.0

Copper 1.0

m-cresol 0.2

o-cresol 0.4

p-cresol 0.1

Cresylic acids (meta, para) 0.2

Chlorobenzene 0.02

n-butylmercaptan 0.06

o-sec. butylphenol 0.3

p-tert. butylphenol 0.03

o-chlorophenol 0.0001–0.015

p-chlorophenol 0.0001

2,3-dinitrophenol 0.08

2,4,6-trinitrophenol 0.002

2,4-dichlorophenol 0.0001–0.014

2,5-dichlorophenol 0.02

2,6-dichlorophenol 0.03

3,4-dichlorophenol 0.0003

2-methyl-4-chlorophenol 2.0

2-methyl-6-cholorophenol 0.003

3-methyl-4-chlorophenol 0.02–3.0

o-phenylphenol 1.0

Pentachlorophenol 0.03

Phenol 1.0–10.0

Phenols in polluted rivers 0.15–0.02

2,3,4,6-tetrachlorophenol 0.001

2,3,5-trichlorophenol 0.001

2,4,6-trichlorophenol 0.002

2,4-dimethylphenol 0.4

Dimethylamine 7.0

Diphenyloxide 0.05

B,B-dichlorodiethyl ether 0.09–1

o-dichlorobenzene <0.25

Ethylbenzene 0.25

Ethanethiol 0.2

Ethylacrylate 0.6

Formaldehyde 95.0

Gasoline 0.005

Guaicol 0.08

Kerosene 0.1

Kerosene plus kaolin 1.0

Hexachlorocyclopentadiene 0.001

Isopropylbenzene <0.25

Naphtha 0.1

Naphthalene 1.0

Naphthol 0.5

2-Naphthol 0.3

Nitrobenzene 0.03

a-methylstyrene 0.25

Oil, emulsifiable >15.0

Pyridine 5–28

Pyrocatechol 0.8–5

Pyrogallol 20–30

Quinoline 0.5–1

p-quinone 0.5

Styrene 0.25

Toluene 0.25

Outboard motor fuel as exhaust 7.2

Zinc 5.0

Source: Reproduced from ANZECC (1992), an adaptation of NAS/NAE (1973)
4.4.5 Water quality guidelines for the protection of human consumers of aquatic foods

Version — October 2000 page 4.4–17

According to Zweig et al. (1999), sophisticated analytical equipment is usually not

necessary for detecting tainting substances; water that tastes or smells unusual may

result in off-flavours, and sensory assessments (i.e. taste, smell) are often

preferable to chemical analyses.

In addition to the chemical contaminants, a number of freshwater blue-green

microalgae and bacteria can cause off-flavours in native fish. The most common is

the earthy or musty flavour often referred to as ‘muddy’ taste, which often occurs

in silver perch (Bidyanus bidyanus). Decaying organic matter can also cause offflavour. The incidence of off-flavours is highest in warmer months, during blooms

of blue-green algae and in ponds with high stocking and feeding rates. Most offflavours can be readily purged by placing fish in clean water such as underground

or spring water, domestic (dechlorinated) or rainwater.

4 Preventative and management approaches

It is generally accepted that food species should not be grown in, or harvested

from, waters likely to be exposed to contamination. If a contamination event should

occur, the aquatic organisms should be regularly analysed to ensure that the

ANZFA standards are not exceeded in harvested product. However, chemical

analysis for the detection of contaminants in aquatic food can be an expensive

process. For planning purposes a method of product quality prediction would be

preferable. This problem may be illustrated by the following examples:

• The viability of the setup of an aquaculture business is being investigated. How

can the investors predict whether, on harvesting, the product will be suitable

for sale for human consumption?

• It is proposed to start up an industrial/sewage plant upstream of a commercial

fishery. How can we predict whether effluent from the plant will have a

significant adverse effect on the fishery product quality?

Section 9.4.3.5 (Volume 3) provides detailed information and guidance on several

approaches for predicting water quality or safety of the aquatic food product. Due

to the complexities involved, uncertainties will be associated with any prediction.

Predictions cannot replace product testing, but they may enable problems to be

identified and resolved before they affect an industry. Summaries of four predictive

approaches are provided below.

i) Bioconcentration factor approach

Bioaccumulation can be predicted using the bioconcentration factor approach.

Since circumstances will vary enormously from case to case, this approach is only

intended as a general guide, not as a set of prescriptive rules; it has several

limitations. The underlying principle of the bioconcentration factor approach is that

where the uptake of a chemical is not controlled by the organism’s metabolism, a

concentration of the chemical in the organism will be proportional to the

concentration of the chemical in the water or food (or sediment).

ii) Area classification approach

The area classification approach is used by the Australian Shellfish Quality

Assurance Program (ASQAP) and the New Zealand Shellfish Quality Assurance

Program (NZSQAP) to identify safe shellfish-growing areas to permit commercial

harvesting for the domestic market and/or for export. The programs provide a risk-
Chapter 4 — Primary industries

page 4.4–18 Version — October 2000

based system of procedures and guidelines for regulating shellfish-growing areas,

harvesting, processing and distribution of shellfish. In general, they cover:

• classification and survey of growing areas,

• relaying (relocation) and harvesting controls,

• post-harvest handling, storage, processing and transportation.

The shellfish harvesting area classification systems rely on the Sanitary Survey

approach to ensure that molluscan shellfish harvested for human consumption are

safe. The Sanitary Survey consists of:

• the identification and evaluation of all potential and actual pollution sources

(i.e. Shoreline Survey),

• the monitoring of growing waters and shellfish to determine the most suitable

classification for the shellfish harvesting area (i.e. Bacteriological Survey).

The categories of classification are based on levels of contamination from sewage,

poisonous or deleterious substances, other pathogenic organisms of non-faecal

origin and biotoxin-producing organisms, radionuclides, and toxic wastes (ASSAC

1997). A number of classifications can result from the Sanitary Surveys, but they

differ slightly between countries.a

iii) Phytoplankton monitoring

The purpose of phytoplankton monitoring is to predict marine biotoxins in

shellfish. In New Zealand, phytoplankton monitoring is mandatory for all

commercial harvested areas under the marine biotoxin monitoring program, while a

similar program is operated by the Ministry of Health for all recreational shellfish

harvesting sites. A combination of phytoplankton and flesh tests are used to

monitor for biotoxin activity. Commercial areas are sampled weekly for biotoxin

activity and if mandated trigger values are reached for a number of species, flesh

testing is invoked immediately. Little such monitoring is undertaken in Australia.

Trigger values for a number of phytoplankton species under the New Zealand

program (MBMB 1996) are provided in Section 9.4.3.5/3 (Volume 3).

iv) Three-phased screening approach

The three-phased screening approach is a tiered process designed for aquaculture

operations to evaluate source water quality in a step-by-step process of increasing

detail and complexity, in order to minimise costs (Zweig et al. 1999). Phase I

screening involves the analysis of basic physico-chemical properties necessary to

sustain culture species. Phase II is designed to screen source water for

anthropogenic contaminants (chemical and biological). Phase III involves field

assessments of the capacity of the source water to culture the selected species,

using management/culture techniques similar to those of the proposed operation

(i.e. a pilot study).b

4.4.6 Some precautionary comments

Section 9.4.4 (Volume 3) provides a detailed discussion of the limitations of the

current guidelines for the protection of aquaculture species and human consumers

of aquatic foods, and it is strongly recommended that it be read. A brief summary

of the major issues is given below.

b Section

9.4.3.5/4

a See Section

9.4.3.5/2
4.4.7 Priorities for research and development

Version — October 2000 page 4.4–19

Two of the major limitations of the current guidelines are the lack of data and the

variability of the data. Data variability can be attributed to several factors, including

the use of different test methods (e.g. time and duration of exposure, size and age of

fish, test conditions) over time, and analytical advances over time. Where differences

in acceptable or tolerated concentrations are extreme between different guideline

documents, it is suggested that the general/recommended guideline value provided in

the current guidelines be applied, exercising some caution.

To relate laboratory toxicity studies to aquaculture operations is not a

straightforward process. Many of the limitations and uncertainties are similar to

those that apply when extrapolating laboratory toxicity data to natural aquatic

ecosystems.a Some that are more specific to aquaculture operations include:

• aquaculture environments possess very different characteristics to natural

environments (e.g. avoidance is not an option, feed is often derived from

external sources, culture species may be regularly handled, stocking densities

may be higher than in natural environments);

• very few ecotoxicological studies test aquaculture species;

• tolerance to individual contaminants is very variable between aquaculture

species, even within the species groups outlined in table 4.4.1;

• toxicity test durations (i.e. usually ≤96 h) are not applicable to aquaculture

operations, where organisms are constrained to an area and particular water

quality for periods longer than toxicity test durations.

4.4.7 Priorities for research and development

As these guidelines are the first synthesis of water quality information for the

aquaculture industry in Australia and New Zealand, a substantial number of

information gaps and research needs have been identified. These are described in

full in Section 9.4.5 (Volume 3).

a See Section

9.4.4 for more

detail
Version — October 2000 page 5–1

5 Guidelines for recreational water quality

and aesthetics

Water-based recreational activities are popular with Australians and New

Zealanders. Although each country has an extensive coastline, much of it is

inaccessible for recreational purposes, resulting in highly localised pressures on

accessible coastline. The same is true for estuarine and freshwater rivers and lakes,

especially those close to urban centres. Therefore, water quality guidelines are

necessary to protect these waters for recreational activities such as swimming and

boating, and to preserve the aesthetic appeal of water bodies. Water quality

guidelines are used in the monitoring and management of a range of

microbiological, physical and chemical characteristics that determine the suitability

of a water resource for recreational purposes.

5.1 Guidelines for users in New Zealand

In New Zealand, water managers should refer to Recreational Water Quality

Guidelines (NZ Ministry for the Environment 1999). This document and the draft

supporting manual can be downloaded from:

http://www.mfe.govt.nz/about/publications/water_quality/beaches-guidelines.htm

The revised New Zealand guidelines were trialed over the 1999/2000 bathing season.

This trial period will be followed by a consultation round similar to that carried out

for the 1998 Bacteriological Water Quality Guidelines for Marine and Fresh Water.

The extent of further revisions, if any, will depend upon the response to the revised

guidelines. Any recommendation to the Minister for the Environment regarding a

National Environmental Standard will be made after the round of consultation.

5.2 Guidelines for users in Australia

The material for Australian users of Guidelines for Recreational Water Quality and

Aesthetics is currently being prepared. When completed, it will replace this section,

in accordance with NWQMS requirements and National Health and Medical

Research Council (NHMRC) statutory procedures. The NHMRC, ANZECC and

ARMCANZ all recognise the need for a single guideline document to supplant

earlier sets of guidelines for recreational water quality, published separately by the

NHMRC and NWQMS (Australian Guidelines for Recreational Use of Water

(NHMRC 1990) and ANZECC (1992) respectively).

It is intended that the new guidelines should be largely based on recommendations

from the World Health Organization (WHO) including draft WHO Guidelines for

Safe Recreational-water Environments: Coastal and Fresh-waters (WHO 1998)

and WHO Health-based Monitoring of Recreational Waters: The Feasibility of a

New Approach (The ‘Annapolis’ Protocol) (WHO 1999). These documents will

provide the impetus to develop a single Australian guideline document. It will be

part of the revised NWQMS Guidelines and will also be available as a separate

NHMRC/ARMCANZ/ANZECC publication. The basis of the proposed guidelines

for recreational water quality and aesthetics in Australia is provided in Appendix 5.
Chapter 5 — Guidelines for recreational water quality and aesthetics

page 5–2 Version — October 2000

Until these Guidelines are revised and endorsed, users should apply the guidelines

from the Australian Water Quality Guidelines for Fresh and Marine Waters

(ANZECC 1992). These guidelines are reproduced below. While these (1992)

guidelines are interim, the eventual guidelines that result from the NHMRC’s

current revision will be the definitive guidelines.

5.2.1 Introduction

Recreational guidelines accommodate two categories of sporting activity:

• sports in which the user comes into frequent direct contact with water, either as

part of the activity or accidently; for example, swimming or surfing (primary

contact);

• sports that generally have less-frequent body contact with the water; for

example, boating or fishing (secondary contact).

A third recreational category concerns the passive recreational use of waterbodies,

mainly as pleasant places to be near or to look at (no body contact). The relevance

of the different water quality guidelines to the three recreational categories is

shown in table 5.2.1. The detailed water quality guidelines for recreational water

are summarised in table 5.2.2.

Table 5.2.1. Water quality characteristics relevant to recreational use

Characteristics Primary contact

(e.g. swimming)

Secondary contact

(e.g. boating)

Visual use

(no contact)

Microbiological guidelines x x

Nuisance organisms (e.g. algae) x x x

Physical and chemical guidelines:

Aesthetics x x x

Clarity x x x

Colour x x x

pH x

Temperature x

Toxic chemicals x x

Oil, debris x x x

The first part of this section on Australian guidelines provides a brief summary of

the most important aspects of the above categories, while the second section

contains details on the specific guidelines. Many of the guidelines necessary for the

maintenance of certain aspects of recreational water quality (e.g. preservation of

aquatic life and wildlife) are discussed in other chapters and will only be briefly

mentioned here. The recommended guidelines rely on the guidelines developed by

NHMRC (1990), with additional indicators included where appropriate.
5.2.2 Recreational categories

Version — October 2000 page 5–3

Table 5.2.2 Summary of water quality guidelines for recreational waters

Parameter Guideline

Microbiological

Primary contact* The median bacterial content in fresh and marine waters taken over

the bathing season should not exceed 150 faecal coliform

organisms/100 mL or 35 enterococci organisms/100 mL. Pathogenic

free-living protozoans should be absent from bodies of fresh water.**

Secondary contact* The median value in fresh and marine waters should not exceed 1000

faecal coliform organisms/100 mL or 230 enterococci

organisms/100 mL.**

Nuisance organisms Macrophytes, phytoplankton scums, filamentous algal mats, sewage

fungus, leeches, etc., should not be present in excessive amounts.*

Direct contact activities should be discouraged if algal levels of

15 000–20 000 cells/mL are present, depending on the algal species.

Large numbers of midges and aquatic worms should also be avoided.

Physical and chemical

Visual clarity & colour To protect the aesthetic quality of a waterbody:

• the natural visual clarity should not be reduced by more than

20%;

• the natural hue of the water should not be changed by more than

10 points on the Munsell Scale;

• the natural reflectance of the water should not be changed by

more than 50%.

To protect the visual clarity of waters used for swimming, the horizontal

sighting of a 200 mm diameter black disc should exceed 1.6 m.

pH The pH of the water should be within the range 5.0–9.0, assuming that

the buffering capacity of the water is low near the extremes of the pH

limits.

Temperature For prolonged exposure, temperatures should be in the range 15–35°C.

Toxic chemicals Waters containing chemicals that are either toxic or irritating to the skin

or mucous membranes are unsuitable for recreation. Toxic substances

should not exceed values in tables 5.2.3 and 5.2.4.

Surface films Oil and petrochemicals should not be noticeable as a visible film on the

water nor should they be detectable by odour.

* Refer to Section 3.3 of these revised Guidelines relating to nutrient concentrations necessary to limit excessive

aquatic plant growth.

** Sampling frequency and maximum values are given in Section 5.2.3.1.

5.2.2 Recreational categories

5.2.2.1 Primary contact

Water used for primary contact activities, such as swimming, bathing and other

direct water-contact sports, should be sufficiently free from faecal contamination,

pathogenic organisms and other hazards (e.g. poor visibility or toxic chemicals) to

protect the health and safety of the user. The general guidelines desirable for

aquatic scenery are also applicable for water used for primary contact.

5.2.2.2 Secondary contact

Water used for secondary contact activities, such as boating and fishing, should

also meet the guidelines suggested for aquatic scenery. Since there is less body

contact with the water, the microbiological guidelines can generally be lower,

although not in cases when shellfish might be taken from from the waterbody. To

protect water-skiers from injury and boating vessels from damage, the water should

be free from floating or submerged logs and stumps and excessive growth of algae

and other aquatic plants. The quality of the water should be maintained so that

there is minimal alteration of the fish habitat.a a See Ch 3
Chapter 5 — Guidelines for recreational water quality and aesthetics

page 5–4 Version — October 2000

5.2.2.3 Visual use

Surface waters used for visual recreational use (no-contact activity) should not be

altered in any way that reduces their ability to support aesthetically valuable flora

and fauna. Such alteration could be physical, such as dredging and dam

construction, or could be due to the addition of wastes to the water. Visual impact

of the surface waters is important; they should be free from:

• floating debris, oil, grease and other objectionable matter;

• substances that produce undesirable colour, odour, taste or foaming;

• undesirable aquatic life, such as algal blooms, or dense growths of attached

plants or insects.

All these factors have to be considered in areas used for aquatic scenery.

5.2.3 Detailed water quality guidelines

5.2.3.1 Microbiological characteristics

Primary contact

The median bacterial content in samples of fresh or marine waters taken

over the bathing season should not exceed:

• 150 faecal coliform organisms/100 mL (minimum of five samples taken

at regular intervals not exceeding one month, with four out of five

samples containing less than 600 organisms/100 mL);

• 35 enterococci organisms/100 mL (maximum number in any one sample:

60–100 organisms/100 mL).

Pathogenic free-living protozoans should be absent from bodies of fresh

water. (It is not necessary to analyse water for these pathogens unless the

temperature is greater than 24°C.)

Secondary contact

The median bacterial content in fresh and marine waters should not exceed:

• 1000 faecal coliform organisms/100 mL (minimum of five samples taken

at regular intervals not exceeding one month, with four out of five

samples containing less than 4000 organisms/100 mL);

• 230 enterococci organisms/100 mL (maximum number in any one

sample: 450–700 organisms/100 mL).

There is a long international experience of disease outbreaks associated with

contaminated water (McNeill 1985, Cabelli 1989). Disease-causing microorganisms (pathogens) associated with bathing areas include salmonellae,

shigellae, enteropathogenic Escherichia coli, cysts of Entamoeba histolytica,

parasite ova, enteroviruses and infectious hepatitis (Hart 1974, McNeill 1985).

Generally, the most common types of diseases that have been associated with

swimming areas are eye, ear, nose and throat infections, skin diseases and

gastrointestinal disorders. McNeill (1985) has reviewed epidemiological studies

associated with recreational waters.

Direct detection of pathogens is not a feasible option for routine assessment, since

they occur intermittently and are difficult to recover from water. For this reason,

‘indicator’ micro-organisms are generally used to assess the health risks associated
5.2.3 Detailed water quality guidelines

Version — October 2000 page 5–5

with pathogens in recreational waters (Elliot & Colwell 1985). A number of

organisms have been considered as indicators of health risks for swimming areas

(McNeill 1985, Daly 1991).

NHMRC (1990) favours the use of faecal coliform bacteria, a sub-group of the

total coliform population that are easy to measure and are present in virtually all

warm-blooded animals. Faecal coliform bacteria in human faeces comprise about

97% E. coli, around 2% Klebsiella, and a further 2% Enterobacter and

Citrobacter together. However, McBride et al. (1991) have documented a

number of deficiencies with the use of faecal coliforms as indicator organisms of

health risks in recreational waters and waters used for shellfish growing. Recent

epidemiological studies have shown poorer relationships between faecal coliform

densities and illness rates in bathers than are obtained using enterococci (marine

waters: Cabelli 1983a,b, Cabelli et al. 1982, 1983) and using either enterococci or

E. coli (fresh waters: Dufour 1984). Further, there is now considerable evidence

that faecal coliforms die off faster than pathogens under certain circumstances;

therefore, they may go undetected during beach monitoring programs, resulting

in the disease risks being underestimated.

New Zealand (McBride et al. 1991), Canada (CCREM 1991) and the United

States (USEPA 1986) now recommend guidelines for recreational waters in terms

of either enterococci or E. coli (or the non-faecal indicator Pseudomonas

aeruginosa). For example, the New Zealand guidelines recommend that the

median bacterial content of samples taken over the bathing season should not

exceed 33 enterococci/100 mL (or 126 E. coli/10 mL) for fresh waters, and

35 enterococci/100 mL for marine waters (McBride et al. 1991). The guidelines

recommended here are based on the levels recommended by NHMRC (1990) in

terms of faecal coliforms, and those recommended by McBride et al. (1991) in

terms of enterococci.

5.2.3.2 Nuisance organisms

Macrophytes, phytoplankton scums, filamentous algal mats, blue-green

algae, sewage fungus and leeches should not be present in excessive

amounts. Guidelines relating to nutrient concentrations necessary to limit

excessive aquatic plant growth are given in Section 3.3 of these revised

Guidelines.

Direct contact activities should be discouraged if algal levels of 15 000–

20 000 cells/mL are present, depending upon the algal species. Large

numbers of midges and aquatic worms should be avoided.

Biological factors that influence the recreational value of surface waters include

those that endanger the health or physical comfort of people and animals, and those

that render water aesthetically objectionable. In the first category are non-biting

midges, phantom midges, caddis flies and mayflies, which can emerge in large

numbers and cause serious nuisance to people picnicking, camping or living near

the shoreline. More serious are biting insects that can cause irritation from their

bites, respiratory allergic reactions or quite serious diseases. Common diseases

transmitted by aquatic invertebrates are encephalitis, malaria and schistosome

dermatitis (swimmer’s itch).

Excessive growths of aquatic plants can also cause problems in recreational areas.

Rooted and non-rooted macrophytes may obstruct the view of swimmers and
Chapter 5 — Guidelines for recreational water quality and aesthetics

page 5–6 Version — October 2000

obscure underwater hazards. They can also entangle swimmers and induce panic if

encountered unexpectedly. If the growth is very dense, boating and fishing may

also be restricted. Dislodged or free-floating plants may also drift on to beaches,

decay and cause objectionable odours as well as provide breeding areas for

nuisance organisms.

Algal blooms, particularly if dominated by blue-green algae (cyanobacteria), can

impair the recreational values of a waterbody by reducing the clarity and by

accumulating along shorelines with effects similar to those cited for macrophytes.

In addition, several species of blue-green algae can produce toxic substances that

may kill fish, birds and domestic animals (Shilo 1981, Codd 1990, Falconer 1990).

Species of blue-green algae have also been responsible for contact dermatitis in

humans and influenza-like symptoms in swimmers (Codd 1990). Primary contact

activities in waters containing high levels of cyanobacteria should be discouraged.

Ingestion of cyanobacterial-infested water has been associated with gastrointestinal

disorders in swimmers, and lipopolysaccharides found in certain cyanobacteria

have been identified as causing skin irritations, dermatitis and allergy reactions

observed in swimmers using cyanobacterial-infested waters (A McNeill, Victorian

Rural Water Corporation, pers. comm., June 1992). As an interim guide, direct

contact should be avoided when 15 000–20 000 cells/mL are present, depending on

the algal species.

Periphyton growing on the bed of rivers and streams can also reduce the usefulness

of these systems for contact recreation. Quinn (1991) recommended that to protect

contact recreational areas:

... the seasonal maximum cover of stream or river bed by periphyton as filamentous

growths or mats (greater than about 3 mm thick) should not exceed 40%, and/or

biomass should not exceed 100 mg chlorophyll a/m2

.

Quinn also called for additional research to define the level of periphyton that

constitutes a nuisance.

Excessive aquatic plant growth is most often caused by high nutrient

concentrations (mostly phosphorus and nitrogen) entering the waterbody.

Guidelines for limitations on nutrients can be found in Section 3.3.

5.2.3.3 Physical and chemical characteristics

Visual clarity and colour

To protect the aesthetic quality of a waterbody:

• the natural visual clarity should not be reduced by more than 20%;

• the natural hue of the water should not be changed by more than 10

points on the Munsell Scale;

• the natural reflectance of the water should not be changed by more than

50%.

To protect the visual clarity of waters used for swimming, the horizontal

sighting of a 200 mm diameter black disc (Secchi disc) should exceed 1.6 m.

Guidelines relating to visual clarity and colour are required for two reasons: first, to

ensure that the aesthetic quality of the waterbody is maintained and that there is no
5.2.3 Detailed water quality guidelines

Version — October 2000 page 5–7

obvious change in the colour or visual clarity; and second, that the visual clarity of

the water is not so low that it is unsuitable for swimming.

As discussed in Section 8.2.3 (Vol. 2), the optical quality of water, primarily its

colour and clarity, is determined by the attenuation of light, particularly by SPM

but also by dissolved matter (Kirk 1983, 1988). Visual clarity, defined in Section

8.2.3, is of considerable importance because it affects the recreational and aesthetic

quality of water.

Panel studies undertaken by Davies-Colley and Smith (1990) in New Zealand

showed that almost all people can detect a change of 30% in visual clarity. DaviesColley (1991) used these results to recommend that reduction in visual clarity

should be limited to less than 20%. This value is also used here.

In addition to aesthetic values, visual clarity of water is also important so that

swimmers can estimate depth and see subsurface hazards easily (Thornton &

McMillon 1989, Smith et al. 1991). Most guidelines require that the substrate

should be visible in areas that are of wadeable depth, the water clarity usually

being specified in terms of Secchi depth (NHMRC 1990, CCREM 1991).

However, as Davies-Colley (1991) points out, a just-visible Secchi disc on the

bottom means that potential hazards, such as snags and broken bottles, will not be

visible because the Secchi disc has a higher contrast than the hazards. DaviesColley (1991) recommended that a better guideline for the visual clarity relevant to

swimmer safety in wadeable areas would be to require that the black disc visibility

should be not less than 1.6 m, which is equivalent to the bottom of the waterbody

being visible at an adult chest height of around 1.2 m. For diving areas, the water

clarity would need to be considerably greater than this.

Water colour is the perception of light backscattered from within the waterbody as

observed when viewed downwards at a near-vertical angle. Typically, about 3% of

the incident light will re-emerge from the waterbody as backscattered light, although

this ratio can vary widely. Colour of water has three aspects: hue, brightness and

saturation or colour purity (Davies-Colley 1991). New Zealand research has shown

that people value blue and green hues in water, but not yellows and reds (Smith &

Davies-Colley 1992). Davies-Colley (1991) recommended that the natural hue of a

waterbody should not be changed by more than 10 points on the Munsell Scale.

Further, he recommended that the natural reflectance should not be changed by more

than 50% to protect the brightness of the waterbody. New Zealand studies have

shown that people are not particularly sensitive to water brightness.

pH

The pH of the water should be within the range 5.0–9.0, assuming that the

buffering capacity of the water is low near the extremes of the pH limits.

Ideally, the pH of the water for swimming purposes should be approximately the

same as the lacrimal fluid of the eyes, which is about pH 7.4. However, lacrimal

fluids have a high buffering capacity when in contact with solutions of different pH

levels. They are able to maintain their pH within limits until their buffering

capacity is exhausted. A deviation as small as 0.1 unit of the normal pH of the

lacrimal fluid causes irritation of the eyes (Mood 1968).
Chapter 5 — Guidelines for recreational water quality and aesthetics

page 5–8 Version — October 2000

Temperature

For human survival in cold water, the critical problem is to maintain body

temperature. There is considerable variation from one individual to another in the

rate of body cooling; it is primarily a function of body size, fat content, prior

acclimatisation and overall physical fitness. Body heat is lost primarily by

conduction from the inner organs through the trunk. Water cooler than 15°C is

extremely stressful to swimmers not wearing appropriate protective clothing.

Extended periods of continuous immersion at these temperatures may cause death.

Thermal stress can be induced by temperatures exceeding the normal skin

temperature of 33°C, and there is a risk of injury with prolonged exposure to

temperatures above 34–35°C (Health & Welfare Canada 1983).

Toxic chemicals

Waters containing chemicals that are either toxic or irritating to the skin or

mucous membranes are unsuitable for recreation. In general, toxic

substances should not exceed the concentrations provided in tables 5.2.3 and

5.2.4.

In general, there are two kinds of human exposure in swimming areas: contact with

the waterbody and ingestion of the water. Recreational water should contain no

chemicals that can irritate the skin of the human body. To protect swimmers from

harmful effects through ingestion, the guidelines from tables 5.2.3 and 5.2.4 should

be applied for other toxicants. Special care must be taken to check for substances

that can enter the body by absorption through the skin. Higher concentrations of

toxicants may be tolerated occasionally if it is assumed that no person will ingest

more than a maximum of 100 mL water during a normal swimming session

(NHMRC 1990) compared with 2 L/d for potable water.

Surface films

Oil and petrochemicals should not be noticeable as a visible film on the

water nor should they be detectable by odour.

The presence of oil and petrochemicals makes water aesthetically unattractive.

They can form deposits on shorelines, and bottom sediments that are detectable by

sight and odour. Some organic compounds can be absorbed directly from the water

through the skin (CCREM 1991), making these substances even more undesirable

in recreational areas.
5.2.3 Detailed water quality guidelines

Version — October 2000 page 5–9

Table 5.2.3 Summary of water quality guidelines for recreational purposes: general

chemicals

Parameter Guideline values (μg/L, unless otherwise stated)

Inorganic:

Arsenic 50

Asbestos NR

Barium 1000

Boron 1000

Cadmium 5

Chromium 50

Cyanide 100

Lead 50

Mercury 1

Nickel 100

Nitrate-N 10 000

Nitrite-N 1000

Selenium 10

Silver 50

Organic:

Benzene 10

Benzo(a)pyrene 0.01

Carbon tetrachloride 3

1,1-Dichloroethene 0.3

1,2-Dichloroethane 10

Pentachlorophenol 10

Polychlorinated biphenyls 0.1

Tetrachloroethene 10

2,3,4,6-Tetrachlorophenol 1

Trichloroethene 30

2,4,5-Trichlorophenol 1

2,4,6-Trichlorophenol 10

Radiological:

Gross alpha activity 0.1 Bq/L

Gross beta activity (excluding activity of 40K) 0.1 Bq/L

Other chemicals:

Aluminium 200

Ammonia (as N) 10

Chloride 400 000

Copper 1000

Oxygen >6.5 (>80% saturation)

Hardness (as CaCO3) 500 000

Iron 300

Manganese 100

Organics (CCE & CAE) 200

pH 6.5–8.5

Phenolics 2

Sodium 300 000

Sulfate 400 000

Sulfide 50

Surfactant (MBAS) 200

Total dissolved solids 1 000 000

Zinc 5000

NR = No guideline recommended at this time; MBAS Methylene blue active substances
Chapter 5 — Guidelines for recreational water quality and aesthetics

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Table 5.2.4 Summary of water quality guidelines for recreational purposes: pesticides

Compound Maximum

concentration

(μg/L)

Compound Maximum

concentration

(μg/L)

Acephate 20 Fenvalerate 40

Alachlor 3 Flamprop-methyl 6

Aldrin 1 Fluometuron 100

Amitrol 1 Formothion 100

Asulam 100 Fosamine (ammonium salt) 3000

Azinphos-methyl 10 Glyphosate 200

Barban 300 Heptachlor 3

Benomyl 200 Hexaflurate 60

Bentazone 400 Hexazinone 600

Bioresmethrin 60 Lindane 10

Bromazil 600 Maldison 100

Bromophos-ethyl 20 Methidathion 60

Bromoxynil 30 Methomyl 60

Carbaryl 60 Metolachlor 800

Carbendazim 200 Metribuzin 5

Carbofuran 30 Mevinphos 6

Carbophenothion 1 Molinate 1

Chlordane 6 Monocrotophos 2

Chlordimeform 20 Nabam 30

Chlorfenvinphos 10 Nitralin 1000

Chloroxuron 30 Omethoate 0.4

Chlorpyrifos 2 Oryzalin 60

Clopzralid 1000 Paraquat 40

Cyhexatin 200 Parathion 30

2,4-D 100 Parathion-methyl 6

DDT 3 Pendimethalin 600

Demeton 30 Perfluidone 20

Diazinon 10 Permethrin 300

Dicamba 300 Picloram 30

Dichlobenil 20 Piperonyl butoxide 200

3,6-Dichloropicolinic acid 1000 Pirimicarb 100

Dichlorvos 20 Pirimiphos-ethyl 1

Diclofop-methyl 3 Pirimiphos-methyl 60

Dicofol 100 Profenofos 0.6

Dieldrin 1 Promecarb 60

Difenzoquat 200 Propanil 1000

Dimethoate 100 Propargite 1000

Diquat 10 Propoxur 1000

Disulfoton 6 Pyrazophos 1000

Diuron 40 Quintozene 6

DPA 500 Sulprofos 20

Endosulfan 40 2,4,5-T 2

Endothal 600 Temephos 30

Endrin 1 Thiobencarb 40

EPTC 60 Thiometon 20

Ethion 6 Thiophanate 100

Ethoprophos 1 Thiram 30

Fenchlorphos 60 Trichlorofon 10

Fenitrothion 20 Triclopyr 20

Fenoprop 20 Trifluralin 500

Fensulfothion 20

Sources: NHMRC & AWRC (1987), NHMRC (1989)
Version — October 2000 page 6–1

6 Drinking water

Drinking water for Australians and New Zealanders should be safe to use and

aesthetically pleasing. Authoritative drinking water guidelines for both countries

are summarised in the sections below.

6.1 Guidelines for users in New Zealand

Guidance on what constitutes good quality drinking water is provided for New

Zealand by Drinking-water Standards for New Zealand (New Zealand Ministry of

Health 1995a) and the Guidelines for Drinking-water Quality Management (New

Zealand Ministry of Health 1995b).

6.2 Guidelines for users in Australia

In Australia guidance on what constitutes good quality drinking water is provided

by the Australian Drinking Water Guidelines (NHMRC & ARMCANZ 1996), a

companion document of the National Water Quality Management Strategy.

The Australian Drinking Water Guidelines are intended to meet the needs of

consumers and apply at the point of use; for example, at the tap. They are

applicable to any water intended for drinking irrespective of its source (municipal

supplies, rainwater tanks, bores, point-of-use treatment devices, etc.) or where it is

used (the home, restaurants, camping areas, shops, etc.).

The Guidelines provide an authoritative Australian reference on good drinking

water quality, covering a wide range of the microbiological, physical, chemical and

radiological characteristics that determine water quality. They are not intended as

guidelines for environmental water quality, nor, as the document stresses, should

they ever be seen as a licence to degrade the quality of a drinking water supply to a

guideline value.

While the individual guideline values apply at the point of use, the document deals

extensively with good system management. It points out that successful

management of water quality in a water supply system requires an understanding of

the processes and practices which can affect water quality within the system. In this

context, the term ‘system’ is defined to include everything from the point of

collection of the water, usually the catchment area, to the consumer’s tap. It

includes streams and rivers in the catchment, storage and service reservoirs,

treatment and disinfection facilities, trunk and service mains, and consumer

plumbing and appliances. Water quality can be affected at each of these points, but

all are inter-related, and integrated management is essential.

The following sections summarise the key issues contained in the Australian

Drinking Water Guidelines (NHMRC & ARMCANZ 1996).

6.2.1 Microbiological quality of drinking water

The Guidelines devote a special chapter to the microbiological quality of drinking

water because the most common and widespread health risk associated with

drinking water is contamination, either directly or indirectly, by human or animal

excreta and the micro-organisms contained in faeces. Microorganisms, including
Chapter 6 — Drinking water

page 6–2 Version — October 2000

pathogenic organisms, can enter water supplies at every stage of the collection and

distribution cycle. To ensure the microbiological safety of a water supply there

should be a wide-ranging program of protection, treatment and monitoring, with

barriers to the entry and transmission of pathogens throughout the system. The first

of these barriers should include protection of the selected source from

contamination by human or animal faeces and the maintenance of an active

catchment protection program.

The Guidelines include a general section on Catchments and Raw Water Quality

and a more specific section on Protection of the Water Catchment from Sources

of Human and Animal Faecal Matter. It is recognised that intelligent management

of land use and water resources in catchments is essential to a safe water supply.

In particular, the Guidelines emphasise the need for an active watershed

protection program, including an emergency plan for responding to major

pollution events such as spillages or contamination. Detailed advice is given on

the problems of surface and groundwater supplies, and the approaches that

should be taken for their management.

6.2.2 Chemical and radiological quality of drinking water

The same principles of catchment management are critical in dealing with issues of

chemical and radiological characteristics of drinking water. Many of these are

difficult and expensive, if not virtually impossible, to remove by treatment of the raw

source water. This applies to naturally-occurring characteristics, as well as to

contaminants introduced from human activities.

Nitrate is an important example of a chemical that occurs naturally in groundwater

supplies in some parts of inland Australia but that enters water as a result of

intensive farming or poor waste disposal practices in more densely populated

coastal settlements. The existing technologies for removing nitrate from source

waters are rarely practicable in areas where nitrate is likely to be a problem. As

nitrate is a health-related characteristic, the options may be to search for a better

water source, or to arrange an alternative supply of water for consumption by those

at risk, typically infants under three months of age.

Pesticides are an example of contaminants that can be introduced by improper

use or accidental spillage in a catchment area, and can be difficult, if not

impossible, to remove by practicable treatment processes. The Guidelines set out

the method for control of pesticide use in Australia through a national scheme of

registration, and recommend that their use in water or water catchments be

authorised only where necessary. Pesticides not authorised for such use should

not be present in drinking water.

6.2.3 Small water supplies

The Guidelines also contain a special chapter on the problems of small water

supplies, regarded as those serving less than 1000 people. For small

communities, economic constraints often mean that only untreated water can be

supplied or that treatment is limited in extent. Furthermore, monitoring may be

infrequent or absent. In such circumstances, sanitary assessment and the use of a

clean and unpolluted water source are of paramount importance. It is therefore
6.2 Guidelines for users in Australia

Version — October 2000 page 6–3

recommended that small communities carry out regular sanitary inspections of

their water supply.

Several measures can and should be taken to reduce the risk that supply to a small

community may become unsafe. A strict protocol of practices should be established

to ensure, among other things, that:

• raw water sources and storages are inspected regularly for any source of

contamination (animals, birds, drainage inflows);

• cost-effective treatment is provided where the quality of raw water is poor (e.g.

biological and pre-roughing filters).

Where problems occur, they should be thoroughly assessed. It may turn out that the

best option for a small community is to seek an alternative source of raw water.

The Guidelines give detailed advice on the way in which regular inspections should

be carried out to check for direct or potential sources of contamination. Inspection

is especially important when water is obtained from streams flowing through areas

developed for agricultural, industrial or residential purposes. The sources of

contamination of groundwater are also discussed.

The frequency of sanitary inspections of a catchment will depend on the

characteristics of each site and the source of raw water. Every catchment where

there is human habitation or free public access should be comprehensively

inspected at least once a year for potential sources of pollution.

6.2.4 Individual household supplies

Finally, consideration is given to the question of individual household supplies. For

such supplies, the emphasis should be on selecting the best quality source water

available, and on protecting its quality by the use of barrier systems and

maintenance programs. Whatever the source (ground, surface or rainwater tanks),

householders should assure themselves that the water is safe to drink. Information

on the quality of surface and groundwater may be available from state or local

governments which may monitor the particular source water as part of a state or

local water monitoring program. Alternatively, the individual should consider

having the water tested for any key health characteristics identified as being of

local concern. Where the raw water quality does not meet the relevant guidelines, a

point-of-use device may be used to treat water.

6.2.5 Guideline values

The individual guidelines cover a wide range of measurable characteristics,

compounds or constituents that can potentially be found in water and affect its

quality. They fall into the following categories:

• microorganisms, including

− bacteria

− protozoa

− toxic algae

− viruses;

• physical characteristics

− radionuclides;
Chapter 6 — Drinking water

page 6–4 Version — October 2000

• chemicals, including

– inorganic chemicals

– organic compounds

– organic disinfection by-products

– pesticides.

A health-related guideline value is the concentration or measure of a water

quality characteristic that, based on present knowledge, does not pose any

significant risk to the health of the consumer over a lifetime of consumption.

An aesthetic guideline value is the concentration or measure of a water quality

characteristic associated with good quality water.

The guideline values are intended for use in two separate but complementary ways:

• as ‘action levels’: that is, if the guideline value is exceeded, some form of action

is initiated. This will generally be short-term and immediate. For example, if the

guideline value for a health-related characteristic were exceeded, the response

should be to take immediate action to reduce the risk to consumers, and, if

necessary, to advise the health authority and consumers of the problem and the

action taken. If the characteristic were not related to health, the action might be

to advise the community of a deterioration in water quality;

• as a basis for assessing how well a water supply system meets, over time,

levels of service agreed with the community (‘performance assessment’ as

presented, for example, in an annual report). When used in this way, the data

are largely of historical rather than immediate interest, and any resulting action

to improve the quality of the supply will generally be longer-term.

In the case of pesticides, two values are provided:

• a guideline value, intended for use by regulatory authorities for surveillance

and enforcement purposes;

• a health value, intended for use by health authorities when managing health

risks associated with inadvertent exposure such as from a spill or misuse of a

pesticide.

The document emphasises that health-related guidelines define water which, based

on current knowledge, is safe to drink over a lifetime: that is, it constitutes no

significant risk to health. For most water quality characteristics covered by the

Guidelines, there is a grey area between what is clearly safe and clearly unsafe, and

the latter has often not been reliably demonstrated. Thus the guidelines always err

on the side of safety, and it follows that, for most characteristics, occasional

excursions beyond the guideline values are not necessarily an immediate threat to

health. The amount by which, and the duration for which, any health-related

guideline value can be exceeded without raising public health concern depends on

the particular circumstances. Exceedance of a guideline value should be a signal to

investigate the cause and, if appropriate, to take remedial action. If the

characteristic is health-related, the relevant health authority should be consulted.

For the individual guideline values, the reader is referred to the Australian

Drinking Water Guidelines (NHMRC & ARMCANZ 1996). This document can be

downloaded from:

http://www.nhmrc.health.gov.au/publicat/pdf/eh19.pdf
Version — October 2000 page 7.1–1

7 Monitoring and assessment

7.1 Introduction

This chapter deals with the practicalities of collecting and analysing data for the

measurement and evaluation of water quality — on the one hand, by measuring

biological indicators; on the other hand, by measuring the more traditional physical

and chemical indicators, including toxicants. Much of this chapter presupposes a

good background knowledge of the issues involved with selecting sample sites, the

timing and frequency of sampling events, and some basic principles of statistics

and the design of experiments and surveys. Much of this background is provided in

the companion document Australian Guidelines for Water Quality Monitoring and

Reporting (ANZECC & ARMCANZ 2000), the Monitoring Guidelines.

The Monitoring Guidelines lays out the framework and general principles for a

water quality monitoring program. Though the present chapter is self-contained in

terms of its coverage of monitoring and assessment, its principal aim is to

complement, not duplicate, the Monitoring Guidelines. To this end, this chapter

highlights some key issues for the users of the Water Quality Guidelines that are

either very specific to their needs, or that expand upon some of the general topics

introduced in the Monitoring Guidelines. Sections 7.1 and 7.2 generally follow the

layout of the Monitoring Guidelines while Sections 7.3 and 7.4 provide more

specialist information for monitoring using biological and physical-chemical

indicators respectively. The chapter is structured as follows:

• Section 7.1: Introduction, issues associated with integrated assessment, and the

framework for a monitoring and assessment program with reference to the

introductory steps that set the monitoring program objectives.

• Section 7.2: This section describes the remainder of the monitoring framework.

Firstly, recommendations are provided for combinations of biological and

physico-chemical indicators to apply to different situations. Then, some

generic issues that are common to both biological and physico-chemical

approaches are discussed. For example, the choice of design for a monitoring

or assessment program depends partly on whether or not there are data that predate a putative impact and on whether or not there are appropriate control sites.

Section 7.2 also recapitulates the steps needed for defining objectives and

selecting candidate indicators.

• Section 7.3: A description of issues that are specific to biological indicators.

• Section 7.4: An outline of issues for physical and chemical stressors and

toxicants in water and sediment. For many of the non-biological indicators, the

first step is to compare test data with a guideline trigger value; the procedure is

detailed in Section 7.4.4.

7.1.1 Integrated monitoring strategies

Traditionally, physical and chemical methods alone are used to assess water quality

by indirectly estimating ecological impairment. Numerical guidelines are set

according to the response of biota from different taxa to individual chemicals,

derived from single-stressor toxicity tests conducted under controlled laboratory
Chapter 7 — Monitoring and assessment

page 7.1–2 Version — October 2000

conditions. The derivation of ‘global’ guideline values, though conceptually

simple, faces a major challenge in that data derived under experimental conditions

may not be relevant to complex real world ecosystems. Nevertheless, direct

measurement of physical and chemical water quality parameters as a surrogate for

ecological health has the advantages of:

• conceptual simplicity,

• established technology,

• explicit numerical objectives,

• the ability to acquire meaningful quantities of data relatively quickly,

• comparatively low costs.

Biological indicators have a shorter history of use in monitoring in Australia and

New Zealand. Their development has been intellectually challenging and has

evoked considerable debate. This explains in part the slower acceptance of

biological indicators in environmental monitoring even though the principle is

inherently sound. Biological monitoring programs, and, to a lesser extent,

monitoring with physical and chemical parameters, can be labour intensive, prone

to quality control failures unless special care is taken, and may require data

collection over an extended period, depending on the statistical design

requirements. Environmental monitoring generally, however, has developed with

improvements in the way sampling is conducted and in application of appropriate

statistical techniques. Appendix 4, Volume 2 contains a case study that illustrates

the importance of fully optimised designs in terms of spatial and temporal controls

applied to indicators.a This case study concludes by considering the balance that

negotiating parties may be faced with in applying optimised designs to early

detection and biodiversity indicators in an essentially unmodified aquatic

ecosystem (a condition 1 ecosystem).b

As discussed in earlier chapters, these Guidelines emphasise an integrated

approach to monitoring, using an appropriate mix of indicators suited to the

primary management aims. Physico-chemical and biological indicators should be

regarded as complementary to each other. Two issues involved in this integration

are firstly, the rationale for integrated monitoring and assessment and ways to

achieve integration; and secondly consideration of ways to defray costs. These are

summarised briefly in turn.

7.1.1.1 Enhancing inferences

1. As discussed elsewhere,c it is widely acknowledged that only studies that

include the biota can define or be used to assess the overall effect of waste

waters on these organisms and the ecological health of ecosystems.

Management goals are typically biologically-based, so organisms are the

management end-point. This position holds even if the methods used for

determining global numerical guidelines, including surrogates for biological

end-points such as water chemical analytes, are acknowledged as having broad

validity. A combination of biological and physico-chemical assessment

enhances the confidence in correctly attributing causes to any observed change

in water quality: biological variables integrate effects of past and present

exposure and directly assess progress in achieving the management goals;

a See Sections

7.3 and 7.4 for

more detail;

also Chapters

3, 4 and 6 of

the Monitoring

Guidelines

b Section 3.1.3

c Sections

3.1.6 & 3.2.1.1
7.1.1 Integrated monitoring strategies

Version — October 2000 page 7.1–3

physico-chemical variables are the explanatory variables in the cause–effect

relationship.

2. Efforts should be made, wherever possible, to examine and incorporate the

results of similar types of study conducted in the region. Whether the results of

the additional studies are examined alone or are combined with those from the

study in question, inferences can be enhanced.a

3. Sometimes samples may be gathered and processed in a manner that allows the

results to be used for different purposes, each providing additional interpretative

information. An example of this is provided belowb where the advantages of

combining stream macroinvertebrate samples and data from quantitative and

rapid biological assessment studies are outlined.

4. Users need to be aware always of standard operating procedures that may be in

place at the regional scale and beyond. Comparison of results with those from

other studies is always enhanced where a common sampling and measurement

protocol is used.

Box 7.1.1 Enhancing inferences and defraying costs in environmental

monitoring programs

Whatever the indicators used in a monitoring program, savings in resources can be made in

the experimental design if data from control sites are shared amongst different bodies

conducting similar monitoring programs in the region. Apart from the advantage of cost

sharing, combined results can then be included in formal meta-analyses (analyses which

combine the results of many similar studies) and thereby allow stronger inferences to be

drawn (see also Section 7.2.5).

7.1.1.2 Defraying costs

The availability of resources is recognised as a major constraint in meeting the

level of monitoring recommended in these Guidelines. Ways to defray costs must

always be considered. Some examples to consider in this respect include:

1. As far as possible, ensure that there is a common sampling program for

collection of data on different indicators. Other than providing greater

interpretative value for the data gathered, this will reduce logistical costs

(e.g. transport etc.).

2. Share costs with similar monitoring programs being conducted in adjacent

areas.c

3. Incorporation of biological assessment in environmental monitoring programs

may lead to cost-savings for industry if ‘no-observable-effects’ in biological

responses are found, despite values for physico-chemical indicators that might

be ‘high’ or which may exceed the recommended guidelines.d Use of the

decision trees for physico-chemical indicators can also lead to cost-savings for

industry; the first of the two case studies included in the Introduction to the

Water Quality Guidelines provides such an example.

a See box 7.1.1

b Section

7.2.1.1/1

c See box 7.1.1

d Section 3.1.3.2
Chapter 7 — Monitoring and assessment

page 7.1–4 Version — October 2000

Section 7.2.1.1 recommends the type and number of indicators that should be

incorporated in a water resource monitoring and assessment program, depending

upon ecosystem condition (condition 1, 2 or 3 ecosystems).a These

recommendations need to be augmented in two special cases as described in

Section 7.2.1.2. These special cases are situations where there is inadequate

baseline datab and situations that call only for a broad-scale assessment of

ecosystem health.c

The final balance of indicators to be measured at a site will rest with local

jurisdictions and stakeholders after they have considered factors such as the nature

of contaminants, the ecosystem type, the issues of concern, level of protection,

availability of baseline data and resource constraints. While the constraints of

resources are acknowledged, local jurisdictions still have a responsibility to ensure

their water quality monitoring programs are sufficiently adequate to give

unambiguous results from which confident conclusions can be drawn.

7.1.2 Framework for a monitoring and assessment program

Although water quality monitoring with physical and chemical indicators differs in

philosophy and techniques from monitoring with biological indicators, the

approaches both rely on sound practice in environmental science, including:

• explicit written definition of the sampling site, project objectives, a hypothesis

and the sampling protocol that will support the work;

• the definition of sampling sites, sampling frequency, and spatial and temporal

variability that will permit appropriate statistical methods to be used;

• rigorous attention to field and laboratory quality control and assurance;

• incorporation of a pilot study to test the sampling protocol and determine spatial

and temporal variability.

Figure 7.1.1 outlines the basic steps involved in developing a program for

monitoring and assessing both biological and physico-chemical aspects of water

quality. This figure is consistent with the framework for the Monitoring

Guidelines, as portrayed in figure 1.1 of those Guidelines. The framework figure

shown in the Monitoring Guidelines is necessarily general in nature while figure

7.1.1 of the current Guidelines has adapted the Monitoring Guidelines framework

to incorporate aspects of the management framework outlined in Chapter 2.d

The first step of the framework, determining the primary management aims, has

been described in earlier chapters of these Guidelines.e Determining these aims will

enable stakeholders to develop an appropriate conceptual model of key ecosystem

processes and interactions. By doing this they can identify assumptions against

which monitoring outcomes can be tested, and develop appropriate working

hypotheses whose predictions can be tested using the data that the program collects

— Step 2 of the monitoring framework (figure 7.1.1). Step 2, developing a

hypothesis, is discussed earlier in these Guidelinesf

 and in Chapter 2 of the

Monitoring Guidelines. Step 1 of the Monitoring Guidelines framework (figure

1.1), ‘Monitoring Program Objectives’, combines the first two steps from the

Water Quality Guidelines framework of figure 7.1.1. The remainder of this chapter

is concerned with the other steps in figure 7.1.1. Background information that

supports the material presented here is provided in the Monitoring Guidelines.

a See Section

3.1.3

b Section

7.2.1.2/1

c Section

7.2.1.2/2

e Chapter 2 and

Section 3.1.1.1

d See Figure

2.1.1

f Section 2.2.3
7.1.2 Framework for a monitoring and assessment program

Version — October 2000 page 7.1–5

• Environmental values

• Level of protection

• Environmental concerns and issues, their extent & cause

• Natural and anthropogenic factors affecting system

• Management goals

• Provide the basis for management and information collection

• Create conceptual model of key ecosystem processes and

interactions

• Make assumption against which monitoring outcomes are

tested

• Underpins environmental goals and water quality objectives

• Select indicators

• Statistical design requirements (with decision criteria,

including effect size/guideline trigger values)

• Water quality objectives

• Determine sampling locations

• Equipment and personal inventory and preparation

• Collection protocols

• Transport and preservation

• QA/QC procedures and data quality objectives

• Chain of custody documentation

• Assess feasibility (access, resources, training, equipment)

and cost effectiveness

• Estimate of spatial and temporal variance etc

• Test and fine tune method and equipment

• Assess training needs of staff involved

• Sampling according to standard or tested protocols

• All samples should be documented: date and location; names

of staff, sampling methods, equipment used, means of

storage.

• Sample analyses according to standard or rigorously tested

methods

• Analyses should be documented: date and location; names of

analysts, methods and equipment used

• Data are adjusted to account for modifying factors (e.g. effect

of pH on chemical speciation)

• Mathematical/statistical processing

• Data are evaluated in the context of key interacting

environmental processes

• Where appropriate, refine water quality guidelines

• The report should be concise, indicate whether the

hypothesis has been supported (and management goals

met), contain recommendations for management action and

indicate refinements to the monitoring program.

• Management action will depend on outcomes, may be to

refine water quality objectives, initiate remedial action,

continue monitoring, cease monitoring, etc.

Determine primary management aims

Develop hypothesis

Study design

Pilot study

(where appropriate)

Sampling

Data analysis and

 interpretation

Evaluation/reporting

Management action

Refine monitoring program

Sample processing 

and analysis

Figure 7.1.1 Procedural framework for the monitoring and assessment of water quality (the shaded area).

(Adapted from the Monitoring & Reporting Guidelines and the framework for designing a wetland

monitoring program adopted by the Ramsar Wetland Convention

(Ramsar Convention 1996, Finlayson 1996))

Version — October 2000 page 7.2–1

7.2 Choosing a study design

This next step of the monitoring framework (figure 7.1.1) includes the selection of

indicators and requirements for experimental design, including the determination of

guideline values. General descriptions are provided in Chapter 3 of the Monitoring

Guidelines. However, the earlier chapters of the Water Quality Guidelines and its

two support volumes are the main reference sources for indicator selection and

determining guideline values, and these aspects are not discussed further. This

section recommends a balance of indicators to apply to different situations for

aquatic ecosystem protection (Section 7.2.1) and provides specific advice for

experimental design using indicators from all environmental values (Sections 7.2.2

and 7.2.3).

7.2.1 Recommendations for combinations of indicators for aquatic

ecosystems

7.2.1.1 Recommendations for each ecosystem condition

This section makes some basic recommendations for the number and mix of

indicators that should be used in integrated monitoring for each of the ecosystem

conditions.

1. Sites of high conservation value (condition 1 ecosystems)

For high conservation value sites, the goal for a water quality assessment program

should include four−six of the following aspects: (i) for contaminants other than

nutrients, ‘whole effluent’ toxicity testing to determine a safe dilution at which

effluent may be discharged; (ii) water and sediment physico-chemistry; (iii) an

‘early detection’ indicator for either water or sediment (whichever is deemed to

harbour greater risks to aquatic organisms arising from the fate and persistence of

waste substances); (iv) a quantitative biodiversity indicator; and (if applicable and

available) (v) a community metabolism indicator and (vi) a rapid biological

assessment (RBA) indicator (see rationale below).

Ideally, for early detection (item (iii) above) a biological indicator of the type

described in Section 3.2 (in particular, table 3.2.2) would be used for monitoring. It is

acknowledged, however, that such indicators have at present been developed for only

a relatively narrow range of conditions and regions. Until such indicators have been

further developed and are more widely available, it is important, nevertheless, to

adhere to the principle of early detection in monitoring and to consider alternative

approaches to meeting this important assessment objective. For example, in some

situations, adherence and responsiveness to very conservative chemical criteria and

their trends may be more protective of ecosystems than even very sensitive biological

tests. Alternatively or in addition, in Section 3.2.1.3/2 it was suggested that early

detection and predictive capabilities would be enhanced by placing additional

sampling sites for any indicator in ‘mixing zones’ — effectively measuring gradients

of spatial disturbance.

The quantitative biodiversity indicator (item (iv) above) should be selected from

Section 3.2. It would normally be expected that some species-level data would be

gathered for relevant biodiversity indicators in regions of high conservation value.

As a complement to the measurement of the quantitative biodiversity indicator,

there could be situations where it would be advisable to also collect data for an

RBA indicator. In some respects, results gathered for RBA can be better than
Chapter 7 — Monitoring and assessment

page 7.2–2 Version — October 2000

results from many quantitative approaches because they provide information about

the ecological importance of effects. As stated in Section 3.2.1.3/3, RBA programs

that have regional coverage and that encompass a full disturbance gradient can

provide regional context for the gathered data. Data gathered for RBA indicators

would not normally be expected to detect minor or subtle impacts, and for this

reason they should never be measured in isolation from quantitative indicators at

sites of high conservation value (nor, in most cases, at sites in slightly–moderately

disturbed systems).

Measurement of quantitative and RBA indicators need not add significantly to the

costs of a monitoring program. For example, replicate quantitative samples from

stream macroinvertebrate communities at a site could initially be processed as

prescribed for the AUSRIVAS RBA approach (e.g. live-sorted, see Method 3A(iii),

Appendix 3 of Volume 2) and then the residue could be preserved for later laboratory

processing in the usual (quantitative) manner. An initial pilot study could be required

to reconcile the sampling effort needed in the field to serve both RBA and

quantitative approaches. RBA data gathered from several sites would be incorporated

into, and assessed against, broader regional or state/territory AUSRIVAS models.

2 Slightly to moderately disturbed systems (condition 2 ecosystems)

For slightly–moderately disturbed sites, the recommended water quality assessment

program has the same four−six aspects prescribed in Section 7.2.1.1/1. For

measurement of biodiversity indicators, species-level data may not be necessary.

3 Highly disturbed systems (condition 3 ecosystems)

For highly disturbed sites, it is recommended that a monitoring program includes

(i) water and sediment physico-chemistry, (ii) a rapid broad-scale and/or

quantitative biodiversity indicator, depending upon the nature and degree of

contamination and level of sensitivity to impact required (selected from Section

3.2.2), and (iii) (if applicable and available) a community metabolism indicator.

7.2.1.2 Combinations of indicators for two likely special cases

In addition to choosing an appropriate set of indicators for an integrated program

according to the ecosystem type, there are two situations that are likely to arise in

many applications. In the first situation,a there are insufficient baseline (i.e. ‘preimpact’) data to implement ‘before–after’ type sampling designs.b The second

situation applies where broad-scale assessment of ecosystem health is the goal of

the program.c

1 Sites where an insufficient baseline sampling period is available

If it is not possible to gather sufficient baseline data, the Guidelines recommend

additional monitoring, including a greater number of indicators and/or sites for

‘early detection’ and biodiversity measurement (i.e. the ‘multiple lines of

evidence’ conceptd). Some recent proposals to help formalise the use of ‘multiple

lines of evidence’ are described in Chapter 3 (Section 3.2.3) of the Monitoring

Guidelines.

i. For sites where development is planned, it is recommended that more extensive

biological assessment procedures be incorporated than those outlined above.e

This would include, for contaminants other than nutrients, a ‘whole effluent’

toxicity testing program for determining a safe dilution at which effluent could

be discharged. For such situations, further protocols for early detection and

biodiversity indicators will recommend the collection of data from a larger

a See Section

7.2.1.2/1

b Described in

Section 7.2.2 & in

Ch 3, Monitoring

Guidelines

c Section 7.2.1.2/2

e Section

7.2.1.1

d Section

3.2.4.1
7.2.2 Broad classes of monitoring design

Version — October 2000 page 7.2–3

number of ‘control’ and ‘to-be-disturbed’ sites than would otherwise be

gathered, so that stronger inferences may be drawn about impact by way of

disturbance gradients.a

ii. At sites where there are existing developments, adequate baseline data were

never gathered; the project approval phase pre-dated more stringent discharge

licensing conditions that have subsequently been imposed by regulators. Use

the same water quality assessment indicators as for Part (i) above, modified for

a posteriori conditions.

iii. For a posteriori monitoring of accidental discharges, use the same water

quality assessment indicators as for Part (i) above, modified for a posteriori

conditions.

2 Broad-scale assessment of ecosystem health

Applications of broad-scale monitoring procedures include assessments of biological

water quality for planning purposes, the setting of goals for remediation and

rehabilitation programs, and the monitoring and assessment of broad-scale impacts

such as diffuse pollution. For such sites, it is recommended that a monitoring

program includes (i) water and (if appropriate) sediment physico-chemistry, and (ii)

data compatible with national RBA programs (e.g. AUSRIVAS).

7.2.2 Broad classes of monitoring design

This section describes the choices of broad classes of designs of monitoring

programs which are available under different scenarios. Note that for the majority

of the physical and chemical stressors and toxicants, the initial step in assessment is

to compare data from the test waterbody or system with guideline trigger values.b

The design of a program for monitoring or assessing water quality is crucial. As

described above, this step presupposes well articulated primary management aims

and appropriate working hypotheses whose predictions can be tested using the data

that the study collects.c

However, as described below, the types of program design depend on the context

within which the investigation is taking place. The context can limit the choices

and inferential strength of the program design.

There are five broad classes of program design (figure 7.2.1; modified after Green

1979). The choice depends on whether the disturbance (putative environmental

impact) has already occurred, and whether any control sites are available for

inclusion in the program. When designing any program for monitoring and

assessment, professional statistical advice should be sought before the data are

collected. All the designs outlined in this section have assumptions, and often

involve sophisticated statistical procedures.d

Most water quality assessment and monitoring will take place relative to a definable

event, which is called a disturbance in figure 7.2.1. This will often be a potential

environmental impact (e.g. construction of a new outfall, change in land-use), but

may correspond to change in activity to improve water quality (e.g. installation of a

new treatment plant, initiation of controls on fertiliser use). If the disturbance has not

already occurred, then there is scope to collect appropriate data before the

disturbance; furthermore, if there are control areas or sites, then this leads to the

strongest class of monitoring and assessment designs, called the ‘Before–After

b Section 7.4.4

d See the

Monitoring

Guidelines,

Chapters 3, 4,

5 and 6

a See also

Sections 7.2.2

& 7.2.3 below

c Section 7.1.2
Chapter 7 — Monitoring and assessment

page 7.2–4 Version — October 2000

Control–Impact’ family of designs (BACI) (case A in figure 7.2.1). Wherever

possible these designs should be used, especially where the opportunity exists to

incorporate appropriate controls (the so-called MBACI designs of Keough &

Mapstone 1995). The logic that underpins this family of designs is described in

Section 3.2.2.1 of the Monitoring Guidelines. In general terms the MBACI design

(where multiple control sites are included) provides the strongest inferences. A

potentially important embellishment for systems with unidirectional flow is to use

matched pairs of sites (upstream and downstream) in disturbed and control locations

(MBACI-P of table 7.2.1). In this scenario it is the differences between upstream and

downstream sites that are compared.a

However, two common situations often arise. Either there are no appropriate

control sites, in which case inferences about the event need to be based on changes

through time alone (case B in figure 7.2.1); or the program has to commence after

the event, in which case inferences need to be based on spatial pattern alone (case

D in figure 7.2.1). Inferences based on spatial pattern alone will usually need to

include reference sites or sites that provide a yardstick against which to compare

the site that is being disturbed. AUSRIVAS, the rapid biological assessment

procedure based on stream macroinvertebrates, can be viewed as a special case of

this class of design.b Similarly, a variety of techniques can be used for basing

inferences on changes through time alone.

A further case, called a posteriori sampling, can arise; see case B in table 7.2.1.

Some chemicals and toxicants are so unusual that they can only come from human

activity (e.g. some specialised pesticides, some unusual isotopes). Detection of

these substances after a disturbance has occurred may be sufficient to infer

environmental impact, without the need to collect any data from before the

disturbance or from spatial control or reference sites. This is likely to be highly

unusual, and exceptional care would need to be taken to convince all stakeholders

that the substance concerned was unequivocally linked to the disturbance.c

Moreover, very good evidence would need to be assembled from auxiliary studies

to establish that concentrations of the substance below the detection level of the

laboratory analysis were ecologically harmless.

Occasionally, monitoring or assessment programs are initiated when the timing or

location of the disturbance is unknown. This leads to two further types of study

design that are not considered in any further detail in these Guidelines. Baseline

studies (case C of figure 7.2.1) refer to those carried out before an event has

occurred, where the goal is to attempt to detect unanticipated changes or trends in

the environment. Broad-scale water quality monitoring networks as well as wellplanned developments exemplify this approach. Investigative studies (case E of

figure 7.2.1) are made in response to a perception that some environmental change

has occurred; their goal is to determine the timing or nature of the change.

Examples include studies carried out after unexpected fish kills or research

programs investigating the extent and severity of acid rain.d

Finally, management for rehabilitating or restoring disturbed sites has some special

problems that need to be taken into account when designing a monitoring and

assessment program. They are outlined in box 7.2.1 below, ‘Issues for restoration

and rehabilitation’, while box 7.2.3 outlines the related procedure of ‘bioequivalence

testing’ which is appropriate for hypothesis testing in these programs.

b See the

Monitoring

Guidelines

Section 3.2

and Section

7.3.3 below

c This is

described in

more detail in

Section 7.3.3

below

d See also the

Monitoring

Guidelines

Section 3.2

a More detail

provided in

Section 7.3.2

and figure 7.2.1

for comparisons

of similarity

measures
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–5

7.2.3 Checklist of issues in refining program design

Once the broad category of design has been selected (Section 7.2.2 above), there

are a number of issues that need to be addressed, preferably in consultation with a

statistician, to refine the design and ensure that data will be collected properly for

the valid application of the chosen statistical methods. (See the Monitoring

Guidelines Chapter 6 for a discussion of the basic statistical issues, and Chapter 3

for discussion of site selection and the scope of the sampling program in space and

time.) The following sections seek to highlight the most prominent issues.

7.2.3.1 Site selection and temporal and spatial scales

To detect impacts reliably, the size and relationship of sampling areas and the

pattern of sampling in space through time need careful consideration. The

assessment objective and the nature of the disturbance also affect sampling design,

as well as site-specific and regional factors. It is difficult to be prescriptive, but

some general guidance on the issues that need to be addressed is summarised in

this section; see also discussion in the Monitoring Guidelines Chapter 3.

Independence of control and impact sites for the indicators being measured is

important for all the BACI-type and spatially-based procedures (cases A and D of

figure 7.2.1). If control and impact sites are too close, cross-contamination can

occur which can mask changes in the indicator. What constitutes too close depends

both on the nature of the indicator and dispersion of the pollutant. Where

independence cannot be ensured, there may be procedures which can take

intercorrelations between sites into account. Such procedures need expert statistical

input before the data are collected.

Information on water movements is essential for planning the extent and separation

of control and impact sites. Climatic and water velocity data can be combined with

information on discharge and morphometry in inland waters, or data on tidal

movements and oceanic circulation in marine situations, to estimate the direction

and extent of mixing and dispersion of effluents. Sometimes sophisticated

computer simulation models are available to assist in predicting these aspects of

water movement.

Spatial variation within the site(s) to be sampled can also affect the precision of

estimates of that site, which in turn can affect the outcome of any formal

significance tests. Often there are distinct habitats or strata within the sites, and

variation within the strata should be quantified in any sampling area; a single

sample unit from each stratum is inadequate. Several sample units should be taken

within the smallest scale of systematic variation, and often sites are sufficiently

large that they require several levels of successively finer spatial resolution to be

nested within each of the control and impact sites (e.g. Morrisey et al. 1992). Such

sub-sampling improves the precision of the estimates of interest, and a good pilot

study using a thorough, hierarchical design is essential for estimating which scales

of variation are important and, consequently, the most cost effective sampling

strategy likely for the final designa (theory: Sokal & Rohlf (1981), Andrew &

Mapstone (1987), McPherson (1990); examples: Morrisey et al. (1992), Downes et

al. (1993)). In addition, the behaviour of data is likely to be better at higher levels

in a sampling hierarchy: data are more likely to be normally distributed and the

influence of zeroes in the data is diminished as a result of the central limit theorem

(Keough & Mapstone 1995).

a These issues

and sampling

strategies to

deal with them

are described

in Section 3.4

of the

Monitoring

Guidelines
1. Has the

disturbance already

occurred?

3. Are there control

areas or sites?

2. Is the timing and

location of the

disturbance known?

Inference

based on:

NO

Other

study

types:

NO

(Opportunity to collect or use data

from before the disturbance)

A. BeforeAfter

ControlImpact

designs

YES

YES

B.

Temporal

changes

only

NO

C.

Baseline

studies

YES

(There are no useful data 

the disturbance

YES 

D. Spatial

pattern alone

I

Figure 7.2.1 Flow chart depicting the broad categories of designs for monitoring and assessment that apply in different cont

discussed in detail in this document. See also table 7.2.1 and the Monitoring Guidelines Sectio

page 7.2–6 Version — October 2000
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–7

Table 7.2.1 Broad categories of design from figure 7.2.1 relevant to the Guidelines listed together with the

assessment objectives that could be fulfilled by each category. Possible designs within each of the three

broad categories (A, B and D) are tabulated with a brief description and commentary together with examples

and references to other publications.

A. Inference based on the BACI (Before–After Control–Impact) family of designs

These designs are suitable for the following assessment objectives:

• early detection,

• biodiversity or ecosystem-level response.

Where comparable control sites exist and there is sufficient lead time before the disturbance, the MBACI design should be

preferred unless the prevailing situation requires one of the other BACI designs described here. The general logic of the

BACI family of designs is outlined in Section 3.2 of the Monitoring Guideines.

Possible designs Description and notes Examples and references

MBACI Before–After Control–Impact design with Multiple

control areas and (if possible) >1 impact area.

Preferred design because of increased confidence

that differences between control and impact areas

are not due to peculiarities between single control

and impact areas.

– May be modified to MBACI-P (where P stands for

pairing of sites) if indicator is best expressed in

terms of differences between paired sites.

– Short-term and long-term impacts require careful

planning of frequency of sampling.

Variation/trends amongst areas/times may be

modelled using regression, covariates, or

dynamic simulation and permutation methods.

The general principles behind this

design are outlined in Section 3.2 of

the Monitoring Guidelines (see also

figure 3.3 in that document). Keough

& Mapstone (1995; 1997) provide a

full description and discussion.

Faith et al. (1995) discuss principles

of MBACI-P designs.

‘Beyond BACI’ designs Elaboration of MBACI designs with additional nested

components in time and/or space. Appropriate

where the spatial and/or temporal scale of the

impact is unknown or where changes in the pattern

of variation of the indicator are more important than

detecting changes in the average value of the

indicator.

Underwood (1994) describes the most

elaborate models based on ANOVA;

general principles could be extended

to other statistical techniques with

more flexible assumptions (e.g.

general linear models).

BACIP (single control

site)

Modifications

Before–After Control–Impact, Paired differences.

Applicable if there is limited scope for spatial

replication (e.g. one ‘control’ and one ‘impact’ site).

Required if seasonal or other temporal changes in

response are known to occur OR if temporal

behaviour of response is unknown. Differences may

consist of multivariate dissimilarities.

– Set of differences before and after impact

compared using Student’s t-test (or equivalent)

(Appendix 4, Vol 2).

– Modelling trends or thresholds and/or inclusion of

covariates (Appendix 4, Vol 2).

Illustrated in figure 3.4 of the

Monitoring Guidelines. Described in

detail by Stewart-Oaten et al. (1986;

1992). Example provided in

Humphrey et al. (1995).

Faith et al. (1995) discuss multivariate

modification.

Simple BACI Before–After Control–Impact; only one sampling

event prior to impact. Applicable only if seasonal or

other temporal changes in the indicator have been

demonstrated to not occur.

– Of dubious value because only one sampling

event prior to the disturbance leads to a high

chance of confounding with natural changes

unrelated to the disturbance.

Described by Green (1979) and in

figure 3.2 of the Monitoring

Guidelines. This design critiqued by

Hurlbert (1984) and Stewart-Oaten et

al. (1986).
Chapter 7 — Monitoring and assessment

page 7.2–8 Version — October 2000

Table 7.2.1 (continued)

B. Inference based on temporal change alone

These designs are suitable for the following assessment objectives:

• early detection,

• biodiversity or ecosystem-level response.

These designs should be used if no comparable control sites exist. They assume that any changes in the behaviour of the

indicator after the disturbance are solely attributable to the disturbance (see Section 3.2 of the Monitoring Guidelines).

Other lines of evidence, such as would be gathered under an integrated monitoring program, would strengthen inferences

from these designs (see table 3.2 in the Monitoring Guidelines).

Possible designs Description and notes Examples and references

Intervention analysis Disturbance is regarded as an intervention and

applicable when a long time series of data has been

collected before the supposed impact which can be

used as a baseline to compare to data collected

after the disturbance. Applicable when no suitable

control sites can be found that are comparable with

the supposed impact site.

Welsh & Stewart (1989) and

Thompson et al. (1982) exemplify

intervention analysis applied to

chemical and biological indicators

respectively.

Trend analysis Objective is to describe any trend in the chosen

indicator. There are several methods that can be

used to estimate trends, including those below.

See Section 6.3 of the Monitoring

Guidelines for brief, general

descriptions and references for all the

techniques listed here.

– Time series analysis in which, if the data

sequence is long enough and sampling

sufficiently frequent, temporal autocorrelations

can be modelled and treated appropriately.

Gilbert (1987) and Galpin & Basson

(1990) provide overviews of the

complexities of applying trend

analyses to water quality data.

– Control charting and allied techniques derived

from statistical process control can be used to

measure changes of means and variance of the

values of an indicator relative to notional ‘action

thresholds’.

– GAMs (Generalised Additive Models) relatively

new, advanced group of procedures which

replace linear functions with unspecified

‘smoothers’ that are suggested by the data

themselves.

Robust smoothing is useful for displaying trends

in data with extreme values or outliers.

A posteriori sampling Applicable only if measured response (especially

chemical or biochemical marker) is unequivocally

related to the effluent (Section 3.2.3; otherwise not

elaborated upon).
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–9

Table 7.2.1 (continued)

D. Inference based on spatial pattern alone

These designs are suitable for the following assessment objectives:

• biodiversity or ecosystem-level response

• broad-scale assessment

These designs assume that disturbed sites and undisturbed sites had similar values of the indicator before the disturbance

(see Section 3.2 of the Monitoring Guidelines). Other lines of evidence, such as would be gathered under an integrated

monitoring program, would strengthen inferences from these designs; see table 3.2 in the Monitoring Guidelines.

Possible designs Description and notes Examples and references

Conventional statistical

designs (e.g. ANOVA,

ANCOVA)

Comparisons are made between disturbed and

undisturbed sites.

Discussed by Underwood (1993);

examples described by Green (1979)

– Pairing of sites upstream and downstream of

disturbance and comparison of these differences

with differences from matched pairs in

undisturbed water bodies can strengthen the

inference.

Davies & Nelson (1994) provide an

example comparing differences

between matched paired sites on

streams subjected to different forestry

operations.

– Matching disturbed site(s) with undisturbed

site(s) is essential but sometimes difficult. Use of

covariates can assist in adjusting for moderate

background differences between sites.

– For multivariate indicators (e.g. measures of

community similarity) analysis of similarity

(ANOSIM) techniques are available for some

basic designs. Future developments in

permutation and randomisation testing are likely

to expand the complexity of designs that can be

analysed.

Legendre & Legendre (1998) provide

a general overview of a variety of

multivariate techniques appropriate for

similarity data. Clarke and Green

(1988) and Clarke and Warwick

(1994) explain ANOSIM and give

some examples.

Analysis of

‘disturbance gradients’

Several sites can be identified with a range of severity

of the disturbance. Inferences are drawn from

correlation of disturbance (or surrogate disturbance)

variables with values of the indicator. A variety of

techniques can be used including those below.

– Regression relates the strength of disturbance to

the response of the indicator.

Basic description provided in Section

6.5 of the Monitoring Guidelines.

– Spatial statistical designs and methods can be

useful where the inference is based on

estimating parameters collected over a

contiguous area. The sampling intensity for such

methods is often demanding.

Cressie (1993) and Rossi et al. (1992)

detail some of the conventional spatial

statistical techniques. Thrush et al.

(1994) provide an example from

marine benthos.

– For multivariate indicators, spatial statistical

techniques are becoming available based on

permutation and randomisation tests. Again the

sampling intensity can be demanding.

Legendre & Legendre (1998) provide

a recent overview of a range of

promising techniques with copious

references to published applications.

Predictive models

based on spatial

controls only

Detection and assessment by predictive modelling

(e.g. AUSRIVAS).

– At present only AUSRIVAS for

macroinvertebrates in rivers and streams has

been developed. This method relies on a network

of reference sites against which test sites (those

thought to have been disturbed) are compared.

Test sites may not have been sampled

contemporaneously with reference sites, so this

method makes a large assumption that there is

low inter-annual variation in the family-level

composition of macroinvertebrate communities.

Because inferences are based on family-level

spatial data, this method is likely to be sensitive

to only moderate to large impacts.

AUSRIVAS is outlined in Section 7.3.3

and by Schofield & Davies (1996); it is

derived from the British RIVPACS

system, the mechanics of which are

described by Wright (1995).

The applications and limitations of

AUSRIVAS in the context of these

Guidelines are described in Sections

3.2 and 7.3.3.
Chapter 7 — Monitoring and assessment

page 7.2–10 Version — October 2000

Box 7.2.1 Issues for restoration and rehabilitation

Many of the principles identified in the accompanying sections also apply to designing programs for

monitoring and assessing the extent of biological recovery after an environmental impact has occurred. The

formal process of setting criteria for making decisions (Section 7.2.3.3) often receives little attention in

rehabilitation and restoration programs and is an area meriting further exploration (e.g. Maguire 1995).

In the majority of rehabilitation and restoration programs, there will not be reliable data collected over long

time periods before the environmental impact. Thus the main problem in setting the criteria for making

decisions for such programs lies in defining appropriate targets for the chosen indicators by which the

success of a program can be judged. If there are no pre-disturbance data at all, then the sampling program

should include appropriate undisturbed sites that can act as reference sites for the disturbed area. This, of

course, entails making assumptions about similarity in behaviour of the indicator over time in the affected

area and the control areas in the absence of the disturbance (Section 7.2), and there is a danger that the

reference sites will not represent a realistic target for the affected area (Wiens & Parker 1995). Furthermore,

there are likely to be situations where there are no appropriate reference sites, and the target reference

condition will need to be set by other means (Section 3.1.4). Setting targets in these situations is difficult, and

will often involve subjective judgements from expert panels and/or stakeholders. For example, suppose a

target value is set for an indicator and, after the prescribed time since rehabilitation, the indicator has still not

reached the target value; there are no logical grounds for determining whether the rehabilitation has failed or

the target was set too high.

In all cases, there will need to be extensive liaison between managers and stakeholders to ensure that

appropriate indicators are selected and that targets are appropriate for the constraints and context of the

impact under consideration (Maguire 1995). Within the framework provided in Chapter 7, the following four

issues need to be considered.

First, the indicators selected will need to accurately reflect the nature of the change desired. Rehabilitation

programs sometimes can concentrate on obvious, but inappropriate indicators. Norris (1986) provides a

salutary example where remediation of a disused mine site focused on obvious terrestrial and riparian works

(as indicators of remediation success) which did not result in any improvement in the biological attributes of

the river. The nature of the desired change will also depend on the time-lags between implementing a

management action and the response of the indicator. For example, changes to land use on a catchment

may take longer to result in a change in algal community composition than closing a sewage outfall; thus

sampling programs and decision criteria will need to be geared towards gradual changes in the former and

relatively abrupt changes in the latter.

Second, the context of the desired change needs to be considered in concert with the size of the effect that

needs to be detected so that timely alterations to the management of the remediation program can be made

(Section 7.2.3.3). For example, a program to assess the success of a clean-up operation after an accidental

oil spill will need tightly specified effect sizes and timelines if legal action about compensation payments

depends on the success of this operation. By contrast, the rehabilitation of a large mine area that has been a

source of serious pollution for many decades may need intermediate goals as various phases of the

rehabilitation process are implemented and their success is assessed. As a result, timelines and targets may

need to be re-set as rehabilitation proceeds.

Third, the relative risks and cost of committing a Type I or Type II error need to be considered carefully

(Section 7.2.3.3), especially in circumstances where pre-impact baseline data are limited and/or control or

reference areas are few (see box 7.2.3, ‘Application of bioequivalence testing’, for the meaning of Type I and

Type II errors under this form of hypothesis testing).

Fourth, the choice of analytical procedures and the scope of the conclusions (Sections 7.2.2 and 7.2.6) will be

limited by the availability of appropriate reference or control data. In some cases, where multiple sites are to

be rehabilitated over long time spans, MBACI designs could be implemented (Section 7.2.2), although the

use of bioequivalence testing procedures under such complex designs may need further statistical

development (McDonald & Erickson 1994). Conversely, monitoring the recovery of an indicator after an

isolated accident such as a toxic chemical spill limits the potential analytical options and strength of the

inferences (Wiens & Parker 1995, see also Section 7.2.2).
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–11

Just as spatial patterns must be considered in sampling design, so must patterns in

time. These patterns may be predictable (e.g. periodic behaviour of tides) or

episodic (e.g. floods), and range in scale from many years (e.g. El Niño–Southern

Oscillation) to diurnal or even shorter time-scales. Again, as with spatial variation,

failure to account for temporal patterns can confound impacts with natural events,

and similar sampling strategies are called for to estimate these patterns.a

7.2.3.2 The importance of good pilot data

Sampling programs can be costly, and it is important to try to optimise the

sampling program so as to address the hypotheses posed by the program (the

feedback loop in figure 7.1.1). Good pilot data collected before the monitoring

program commences are therefore highly desirable in the absence of a validated

historical database. Note that the number of samples acquired in pilot programs

should be as large as is feasible to provide accurate estimates of variation; pilot

data using small sample numbers yield unreliable information that may lead to poor

decisions in optimising the sampling program. The design of the pilot sampling

protocol must be as detailed and thoughtful as for the main project, though it

should be remembered that to refine a sampling protocol is one of the principal

objectives of a pilot study. Optimisation decisions based on a well designed pilot

study will be more soundly based and hence defensible. Another advantage of a

pilot study is that it gives field staff site-specific training, and allows anticipation of

potential hazards and logistical problems. Most practitioners recommend that a

significant fraction of total project resources should be dedicated to a pilot study;

Keith (1991) recommends 10–15%.

7.2.3.3 Setting criteria for decisions

The values of indicator variables usually respond to disturbances in a continuous

fashion (e.g. the ‘dose–response relationship’ of toxicology). As explained in

Section 3.1.7, somewhere along the continuum a value of an indicator needs to be

chosen which forms the criterion for making a decision which will precipitate some

management response.

This section outlines the procedures for setting such decision criteria, in three steps.

First it explains the use of hypothesis testing in this process; then it describes the

three stages that need to be addressed when setting decision criteria. In outline, the

first stage involves deciding what sort of change to look for in the indicators, in the

context of the environmental assessment objectives. The second stage involves

translating this change in the indicator into a quantifiable effect size. The third

stage involves assessing the risk of making a Type I error (giving false alarm) or a

Type II error (giving false sense of security)b in the light of the consequences or

costs of making either of those errors (in a purely scientific and/or social value

sense). It is important that these three interconnected stages are discussed and

iterated with the stakeholders interested in the results of any monitoring or

assessment program. The negotiations should be undertaken before implementing a

monitoring or assessment program so that effect sizes, error rates and costs are

identified explicitly. It is also best to discuss several indicators simultaneously in

this process because, inevitably, some indicators may prove to be more costeffective than others in detecting change.

a Section 3.4,

Monitoring

Guidelines,

also discusses

these issues

b See Sections

3.1.7 and 3.2.4
Chapter 7 — Monitoring and assessment

page 7.2–12 Version — October 2000

The use of hypothesis testing

These Guidelines generally adopt a statistical hypothesis testing approach to

determine whether the values of the chosen indicators have exceeded guideline

values. Users should be aware that hypothesis testing is not the only statistical

procedure that can be used in making inferences from water quality data.a (Note

that this is a separate issue to the requirement for general working hypotheses such

as those described in Section 7.1.2 above that identify key assumptions against

which monitoring outcomes can be tested.) Some background on the criticisms of

hypothesis testing and the rationale for using it in water quality monitoring are

given in box 7.2.2, ‘Hypothesis testing in environmental monitoring and

assessment’.

Box 7.2.2 Hypothesis testing in environmental monitoring and

assessment

There has been some argument against the use of hypothesis testing tools in environmental

assessment programs (e.g. Suter 1996). While hypothesis testing is not always either

necessary or appropriate, much of the argument about its use (or misuse) is mis-directed.

The argument is, in part, that hypothesis tests will only tell us after the event that something

‘dreadful’ has happened. However, the real issue is not whether hypothesis testing is

appropriate, but whether the criteria by which tests are made (and for which sampling

programs are designed) are sufficient or appropriate. An appropriately designed and

executed sampling program intended to detect early warning signals will provide early

warning whether analysed through hypothesis testing models or other procedures.

Therefore it is important to make a satisfactory definition of objectives and decision criteria

for each monitoring program. In real life there is a continuum or spectrum of conditions from

undisturbed to disturbed; defining statistical boundaries along this spectrum to specify

changes that are ‘acceptable’ and unacceptable to the stakeholders (as is done by the

AUSRIVAS model bands) is strongly advocated, especially where clear ‘break-points’ in the

meaning of ecological variables are not well documented.

A related issue is whether the inferences of impacts or changes should be based on

dichotomous alternatives or a continuum of conditions. Suter (1996) and Stewart-Oaten

(1996a,b) infer that hypothesis tests are constrained to test only two alternatives. However,

there is no reason why those alternatives cannot be but two of a range of conditions along a

continuum, the test being to (perhaps progressively) detect whether a response variable has

moved from one condition to the next. In this case, the dichotomous test would be used only

to test whether a particular threshold along a continuum had been crossed.

In summary, it is most important to choose appropriate ‘performance criteria’ for impact

assessments or monitoring programs. If the criteria by which a management action will be

triggered are inappropriate or insensitive or too coarse, then the issue of which tool to

choose for statistical analysis becomes irrelevant.

For hypothesis testing to be useful in making decisions, the user needs to negotiate

how much change in the indicator represents ‘background noise’. In formal terms

this means stipulating the null hypothesis (‘no change’) in terms of an effect size,

as explained in the next section. That is, the null hypothesis is best thought of as

the condition representing no important change in the value of the indicator, where

‘importance’ is determined by the context of the problem being monitored or

assessed. Similarly, Type I and Type II errors are minimised by setting a suitable

level of statistical significance when testing differences or change.

a See also the

Monitoring

Guidelines

Sections 2.4.2

and 6.4.2; the

latter touches

on alternative

procedures
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–13

Some management programs will be oriented towards restoration or rehabilitation. In

these circumstances the monitoring program will be seeking to prove that the values

of the indicators are similar to those defined by the reference conditions.a In formal

terms such programs will be trying to prove the null hypothesis (no ‘important

change’ from reference conditions), although this is formally impossible. Hypothesis

testing frameworks have been developed for such situations (they are sometimes

called ‘bioequivalence tests’ in medicine and toxicology) and these are outlined

briefly in box 7.2.3, ‘Application of bioequivalence testing’.

Box 7.2.3 Application of ‘bioequivalence testing’ for environmental

restoration

Where statistical hypothesis-testing procedures are being used to analyse the data, it may

be useful to re-cast the test using the framework of ‘bioequivalence testing’. This procedure

has been used in medical contexts (e.g. Westlake 1988, Chow & Liu 1992) and has recently

been applied to environmental restoration in the USA. It is clearly explained by McDonald

and Erickson (1994).

The problem with testing for recovery using a conventional hypothesis test is that the

investigator is attempting to ‘prove’ the null hypothesis that the selected indicator in the

disturbed site(s) has the same value as in the control or reference sites. However, failure to

reject a null hypothesis does not constitute proof.

Tests of bioequivalence solve this problem by recasting the question so that the undesirable

outcome, that the disturbed site differs substantially from the reference (i.e. the sites are not

‘bioequivalent’), becomes the ‘null hypothesis’15 and evidence is sought to reject this

hypothesis in favour of the alternative, that the impacted site is similar to the reference (i.e.

the sites are ‘bioequivalent’). Formally, the hypotheses are framed in terms of the ratio of the

values of the indicator in the disturbed site and the reference site. If recovery has been

achieved, the ratio should be sufficiently close to 1, and there should be strong evidence

against the ‘null hypothesis’ which is then rejected in favour of the alternative after

conducting the appropriate statistical test.

Under bioequivalence testing, a Type I error results in incorrectly deciding that the sites are

bioequivalent when they still differ by an important amount (i.e. inadequate recovery, a false

sense of security), whereas a Type II error results in deciding that the sites still differ when in

fact they are similar (i.e. adequate recovery, false alarm). Note that with this technique

stakeholders still must negotiate an effect size; users need to stipulate how different sites

can be before they are declared ‘non-bioequivalent’. In formal terms a critical value of the

ratio of the indicator between the sites needs to be stipulated.

Stage 1: The nature of the change and its context; the use of hypothesis testing

The criteria used for making a decision depend on the level of protection assigned.

As explained in Section 3.1.3 the level of protection depends on the condition of

the ecosystem (condition 1, condition 2 or condition 3); specific guidance on how

the level of protection affects decision criteria is given in Section 3.1.3.2 and table

3.1.2, while Section 3.1.8 elaborates on condition 3 ecosystems.

 

15 Technically, the term ‘null hypothesis’ is usually reserved for the equivalence of a test statistic

under different conditions, whereas in a bioequivalence test the investigator is quantifying the

evidence against a proposition of non-equivalence under the different conditions.

a See Section

3.1.4
Chapter 7 — Monitoring and assessment

page 7.2–14 Version — October 2000

In real life there is a continuum or spectrum of conditions from ‘undisturbed’ to

‘disturbed’. Any indicator’s response to disturbance is likely to vary according to

the strength of that disturbance, whereas the decision about whether or not an

impact has occurred is a point on that continuum. In other words, managers and

stakeholders need to determine how much change from the unimpacted or preimpact condition is acceptable.

The environmental assessment objectives (table 3.2.1) determine the point along the

continuum at which an environmental impact is deemed to have occurred. For

example, monitoring based on early detection of impact will have a different

emphasis from monitoring geared towards assessing the ecological importance of an

impact that has already happened. For early detection, a decision must be made

before the level of change becomes harmful; otherwise the change may be

irreversible. By contrast, to assess the importance of, say, an accidental ecological

impact, the monitoring team must decide whether the level of acceptable change has

been exceeded and by how much. In this situation the decision criterion is at the

point of harmful change rather than some smaller value. In general, however, the

emphasis will be on setting the decision criteria at a level that prevents harmful

effects from occurring in the first place.

Thus a very important part of setting decision criteria is knowing what management

actions will be taken if an impact is detected. The management goalsa that managers

have established provide most of this context, and some of the issues that may affect

these goals are outlined in Section 3.1.3.3. For example, if a condition 2 ecosystem is

being managed to conserve the population of a recreationally important fish, and the

threat is a persistent contaminant with the potential to reduce the fecundity of the

fish, then the decision criteria for the water quality indicators need be set at values

which are smaller than those which begin to affect the reproduction of the fish; this

will allow sufficient lead time for managers to act before the population of fish are

affected. Much scientific judgement is involved in this process, and the actual values

used as decision criteria will depend on a number of modifying factors (e.g. chemical

speciation of toxicants); such matters are covered in more detail for each of the broad

classes of indicators in Sections 3.2–3.5.

For ecosystem condition 1 (high conservation/ecological value) a criterion of ‘no

change beyond natural variability’ is prescribed for biological indicators, physical

and chemical stressors and sediments.b Operationally, this still requires users to

stipulate how much change can be expected under ‘natural’ conditions, because this

natural variation constitutes an acceptable level of change in the ecosystem.c Note

that it is still necessary to decide on an effect size (see the next sub-section) explicitly

and to ensure that sampling is intensive enough to detect effects larger than the

acceptable natural changes in the chosen indicators, and avoid Type II errors. Note

that the determination of the acceptable level of change may have both scientific

and social elements.

For those who are new to environmental assessment, defining an acceptable level

of change may seem weak, especially when management insists there must be ‘no

change’ in the indicator. A criterion of ‘no change’ cannot be used operationally

because it requires the user to prove the null hypothesis  an impossibility, as

mentioned above. However, it is possible to state some level of change in an

indicator below which it is not important to reject the null hypothesis of ‘no

change’.d This requires stakeholders to be explicit about what level of change in

a See Section

2.1.3 for

management

goals

b Section

3.1.3.1 and

table 3.1.2

c, d See also

footnote 2,

page 2–9
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–15

the indicator is regarded as harmless or acceptable. In formal terms this process

involves specifying an effect size, which is described in the next sub-section.

Stage 2: Specifying the size of the effect

The values of all the indicators used in these Guidelines vary naturally in space

and time, and estimates of their true values can only be made via samples.

Accordingly some observed changes in the indicators are likely to be ecologically

trivial. The problem for water quality monitoring is to detect non-trivial changes

in the chosen indicators soon enough to allow management to act. This means

that monitoring programs need to be sensitive enough to detect modest rather

than large changes in the indicators.

In formal terms, therefore, we need to identify the maximum amount of change in

the indicator that is tolerable before we reject the null hypothesis (no important

change) in favour of the alternative hypothesis (important or unacceptable change).

This level of ecologically important change is sometimes called the critical effect

size,

a but for brevity we refer to it as the effect size.b Some of the procedures in

these Guidelines have an implicit effect size; the relationship of guideline trigger

values to the concept of effect size is described at the end of this subsection.

Box 7.2.4 Effect sizes are implicit in some procedures

For some procedures, the effect size and error rates tend to be implicit in the methods and are

less amenable to the procedure of using scalable decision criteria described in this section.

For example, when comparing test data to a guideline trigger value, the ‘effect size’ may be

implied by the choice of percentiles used in the comparison. See Section 7.4.4 for a full

discussion of this and the trade-offs between Type I and Type II errors made in this procedure.

Similarly, in the AUSRIVAS procedure for rivers, notions of effect size and error rates are

inherent in the way the summary indices are compared with the reference conditions. See

Section 7.3.3 for more discussion.

There are two components of effect size: its form and its magnitude (Cohen 1988,

Mapstone 1995). The form of an effect is the statistical measure (e.g. mean,

variance) that is expected to differ between control and impact sites, and the pattern

of differences or trends that it is necessary to detect (Stewart-Oaten et al. 1986,

Green 1989, Underwood 1991a,b). The magnitude of an effect is the size of the

difference or change in mean, say, or variance that would be considered important.

It is difficult to be prescriptive about effect sizes in ecological assessment for two

reasons. Firstly, there is little information about the relationships between

contaminants and biological indicators in field conditions, especially in Australia

and New Zealand. Secondly, the degree of change that is important depends on the

environmental and social values that stakeholders are seeking to protect. Strategies

for setting an effect size are discussed in box 7.2.5, ‘Some suggestions for setting

effect size’. This is not an exhaustive list, and other strategies may arise as

experience in planning programs with these procedures increases.

a See box 2.3

in Section 2.2.1

b See box

7.2.4
Chapter 7 — Monitoring and assessment

page 7.2–16 Version — October 2000

Box 7.2.5 Some suggestions for setting an effect size

Where an indicator has intrinsic socio-economic value (e.g. it is a commercially or

recreationally important species), then effect sizes can be set to ensure sustainable use of

that indicator. However, many biological indicators have been selected because they are

more sensitive than commercial species or because they are thought to be ecologically

important rather than of economic value. For example, seagrass is not used directly by

humans in Australia and New Zealand, but is an important indicator because of the habitat it

provides and the number of species it supports.

Existing research, or similar impacts, preferably in comparable regions, can provide

information about the relationship between the indicator and size of potential impact,

especially if existing impacts can be found on a gradient from mild to extreme. For example,

a variety of sewage treatment plants may be present in a river basin with differing degrees of

sewage treatment. Pilot data relating indicator levels and type of treatment could be used in

stakeholder consultations to correlate stakeholders’ expectations of acceptable sewage

treatment with change in the indicator. In some cases, simulation models can use these data

to estimate how much an indicator might change under different scenarios.

For many indicators in ecosystems in Australia and New Zealand, however, such

background data are unlikely to be available. This will inevitably involve some judgement by

the planners of a program, and an arbitrary but conservative effect size will need to be

specified (e.g. Humphrey et al. 1995). This should be done explicitly, and at the beginning of

the program. Any change to the effect size later in the program must be openly and explicitly

negotiated and fully justified on scientific grounds.

Once the level of acceptable change has been negotiated, the degree of change in

the indicator may need to be set to a smaller value so that management actions can

be implemented before harmful and irreversible effects occur. When the effect size

is being set, such issues as the fate and persistence of the contaminant and timelags between a contaminant event and a measurable change in the biological

indicator should be considered. Allied to these issues are selection of appropriate

indicator(s),a and assessment of the relative costs of erroneously missing an effect

of the stipulated size (Type II error) and erroneously concluding an impact

occurred when, in fact, it did not (Type I error) (see next subsection).

For the non-biological indicators in Sections 3.3–3.5, the guideline trigger values

listed are the best currently available estimates of ecologically low-risk levels for

those indicators.b These trigger values make an implicit statement about effect

size: data from the test waterbody which are lower than the trigger value are

thought to pose little risk to the ecosystem. Depending upon the management goals,

stakeholders may need to negotiate different trigger values. There will also be

situations where trigger values are exceeded. In these cases, more complex

monitoring designs are called for,c and the steps outlined here for negotiating effect

sizes will need to be followed.

Stage 3: Specifying the error rates relative to the costs of those errors

Having stipulated an effect size, the stakeholders then need to minimise the risk of

two potential outcomes — in statistical terms, the Type I and Type II errors. These

errors can arise because the indicators we use are sampled rather than measured

completely, meaning that we are working with information which is necessarily

incomplete. The first potential error is to declare that an impact has occurred

a See Section

8.1

b Section 7.4.4

c Section 3.1.5
7.2.3 Checklist of issues in refining program design

Version — October 2000 page 7.2–17

(i.e. the effect size has been exceeded), when really there has been no actual

change that was ecologically important. The second potential error is to miss an

ecologically important change. The probabilities of each error are conventionally

denoted by the Greek letters α (for Type I) and β (for Type II).a

The challenge is to ensure that sufficient data are collected to detect the change

stipulated in the effect size while, on the other hand, not expending too many

resources on sample sizes that will detect ecologically trivial changes in the

indicator. Inevitably, resources are scarce, so all monitoring programs will need to

balance these two errors.

Conventionally, α has been set at 0.05 or smaller and few programs have stipulated β

(Toft & Shea 1983, Fairweather 1991, Mapstone 1995). Although some

recommendations for α and β are conservative default values for ecosystem

conditions 1 and 2 (for biological indicators in Section 3.2.4), it must be emphasised

that ideally these error rates should be negotiated rather than accepted uncritically.

The most important part of this negotiation is to ensure that the balance between

these two types of errors is acceptable to stakeholders in the process. To this end,

these Guidelines recommend Mapstone’s (1995, 1996) proposal that the ratio of

these two errors is negotiated as part of refining the design of a monitoring program.

This process requires iteration between stakeholders, but should be transparent,

accountable and, above all, should take place before the final monitoring or

assessment program is put in place (Mapstone 1995).

In outline, the choice of α and β involves four steps. First, establish the relative

importance or cost of the consequences of each type of error. Second, set the ratio of

the critical Type I and Type II errors relative to their costs (if there is insufficient

information to estimate the costs of the errors, Mapstone suggests they should be

weighted equally). Third, negotiate desired values of α and β with reference to the

ratio established in the previous step with the stakeholders. Fourth, design a sampling

program to meet the desirable Type II error rate, β, established in the previous step,

given the effect size which has been specified earlier; this allows the sample size and

details of the design to be finalised. Mapstone (1995) details two alternative decision

procedures that can be followed once data have been collected and analysed.

Ideally, these negotiations should include a number of potential indicators

simultaneously. In the process of balancing Type I and Type II errors, some

indicators will inevitably prove much more costly than others if the two error

rates are to be kept low. In such cases, stakeholders are faced with a choice:

either discard the costly indicators in favour of those that will detect the

stipulated effect size more cheaply, or, if the costly indicators have to be included

in the program for some reason, increase the sizes of the two errors while

maintaining the ratio between the errors. The only way to reduce the sizes of

these errors is to increase the sampling intensity. Maintaining the ratio between

the errors ensures that Type I errors are not minimised at the expense of

increasing Type II errors — i.e. that the monitoring program does not lose power

to detect an important change at the expense of being conservative about the

probability of incorrectly declaring that an important change has occurred.

In two situations in these Guidelines, this negotiation of the balance between

Type I and Type II errors is implicit; these are outlined in box 7.2.4.

a See box 2.3

in Section 2.2.1
Chapter 7 — Monitoring and assessment

page 7.2–18 Version — October 2000

The choice of the best sampling program is not a trivial issue. The strongest

evidence will come from designs that have extensive baseline data collected before

the suspected or potential impact takes place, and will involve simultaneous

monitoring in multiple control sites. The weakest evidence will result from

programs with limited or no pre-impact data. In all situations, inferences and

assessments can be strengthened by including multiple lines of evidence.a The

power of any statistical tests employed may be improved by including multiple

indicators in a multivariate analysis, depending on the pattern of responses amongst

the indicators (Green 1989).

7.2.4 Sampling protocols and documentation

From figure 7.1.1, once the sampling program has been finalised, sampling can

begin. This should take place according to standard or tested protocols. Section 8.1

and Appendix 3 of Volume 2 provide a list of protocols for biological indicators,

while Section 8.3.6 outlines sources for protocols to be used in direct toxicity tests

for toxicants. Procedures for sediment toxicity testing seem to be less well

developed, but references to and guidance through the recent literature are provided

in Section 8.4.3. Protocols for measuring physical, chemical, biological and

ecotoxicological parameters of sediments are described in general terms in Section

3.5 and Appendix 8 of Volume 2, and Chapter 4 of the Monitoring Guidelines,

with references to detailed literature.

Quality assurance and quality control (QA/QC) procedures should be part of any

sampling protocol. Quality control (QC) and quality assurance (QA) are different

but related concepts. In the context of these Guidelines, quality control means

devising and implementing safeguards to minimise the corruption of data. These

safeguards must be installed at every step of the process from project definition to

the decision on whether measured concentrations compare acceptably with the

guidelines. Quality assurance means testing the effectiveness of these safeguards.

In any QA/QC program, chain of custody documentation is essential to ensure that

errors can be traced. Chapter 4 of the Monitoring Guidelines discusses QA/QC in

some depth for key points for chemical, physical and toxicant indicators;

Section 7.4.3 below refers to that source.

7.2.5 Sample processing and analysis

Analysis here refers to the processing of sample units (e.g. field or laboratory

measurement of analytes in a water sample, counting and identifying invertebrates

in a benthic sample) rather than the statistical analysis of the resulting data. As with

sampling, standard or rigorously tested protocols should be used; many protocols

also detail methods of analysis. Because of their reliance on often complex,

rigorous laboratory procedures, more specific guidance on analytical procedures is

provided for physical and chemical stressors, toxicants and sediments.b

Again, QA/QC procedures are often described in protocols, and QA/QC is also

discussed in Chapter 5 of the Monitoring Guidelines. The monitoring team should

document at least the analytical steps and the date and location of the analyses, the

identities of the analysts, the methods used and the type and status of any

equipment used for the analysis.

a See Section

3.2.4.1

b See also

Sections 7.4

and 7.4.3
7.2.6 Data analysis, evaluation and reporting

Version — October 2000 page 7.2–19

Similar care in QA/QC should also be used during data entry and data

management. Most modern database software packages provide value-checking

routines, and clear procedures should be established to manage, track, back-up and

archive data files. Clear documentation of the features of the data (e.g. the units

that the data are entered in, codes used for missing or ‘below detection limit’ data)

need to be kept with the data files.

7.2.6 Data analysis, evaluation and reporting

The first step in evaluating the data will be the formal statistical analysis. For some

indicators, the data may need to be adjusted to account for modifying factors (e.g.

effect of pH on chemical speciation). The process of analysing the data is also

iterative, with the first step being to examine the distributions of the variables and

to check for outliers,a to see whether the data meet the assumptions of the intended

analysis. Sometimes transformation of the data can solve distributional problems.

Most statistical procedures have a second diagnostic stage after the procedure has

been applied (e.g. examination of residuals after fitting a regression or general

linear model). If these diagnostics show that the assumptions of the procedure have

been violated, alternative statistical models may need to be developed. Chapter 6 of

the Monitoring Guidelines discuss these issues, while the involvement of

professional statisticians is invaluable in ensuring the rigour of these analyses.

Once the statistical analyses have been completed, the results need to be interpreted

in the context of the key interacting environmental processes and the

environmental assessment objectives of the program. Reporting of the results needs

to clear, concise, unambiguous and timely to allow management to act on the

results. It is essential to disseminate the results to stakeholders in a form that is

readily understandable, and some general recommendations are given in Chapter 7

of the Monitoring Guidelines. Reporting will often include recommendations on

modifications to the program if it is a continuing program, thereby closing the

feedback loop in figure 7.1.1 (this chapter).

a See Chapter 6

of the Monitoring

Guidelines

Version — October 2000 page 7.3–1

7.3 Specific issues for biological indicators

This section addresses issues specific to biological indicators that need to be borne

in mind when designing monitoring and assessment programs. Section 7.3.1

outlines the issues for univariate indicators: these consist of a single response

variable such as the density or biomass of phytoplankton, measures of community

metabolism, or chemical/biochemical markers in aquatic organisms. Multivariate

indicatorsa refers to measures of community composition or structure where the

response variable is usually based on some measure of community similarity

which, in turn, is computed from the abundance (structure) or presence or absence

(composition) of many taxa within the ecosystem. Examples include measures of

the community structure of diatoms, macroalgae and invertebrates. AUSRIVAS,

the rapid biological assessment technique for Australian rivers, is also based on

multivariate community composition data, but is a special case of a design class

where inferences are based exclusively on spatial controls alone. Section 7.3.3

discusses how the outputs of AUSRIVAS relate to the issues raised in Section 7.2.

7.3.1 Issues for univariate indicators

Most of the key issues are raised in Section 7.2 and in Chapters 3 and 4 of the

Monitoring Guidelines. Univariate indicators are easily analysed using

conventional and novel statistical procedures, provided the key assumptions are

met.b However, two issues are worth emphasising.

First, many of the classical techniques of statistical analysis assume independence

of sample units through space and time. Biological indicators may violate these

assumptions temporally because of the longevity of indicators or spatially because

of dispersal or behaviour of indicators. If these phenomena are likely within the

monitoring program, then professional statistical advice should be sought to either

adjust the sampling regime or select statistical modelling tools that can

accommodate these spatial and/or temporal autocorrelations (Legendre & Legendre

1998).c

Second, data which consist of counts of organisms sometimes result in a large

number of zero values (i.e. when there are no organisms in the sampling unit). The

frequency distributions of such data are typified by a ‘spike’ at zero, then a mode at

some larger, non-zero value. Assuming that the sampling unit or device is

appropriate for the size and behaviour of the organism (most of the protocols

recommended in Chapter 8 give advice on sizes of sampling units; for a more

thorough discussion see, for example, Andrew & Mapstone 1987), such data are

usually problematic for most statistical techniques. Some recent advances have

been made in this area; as this is still an active developing area of applied statistics,

professional advice should be sought when choosing and using these techniques.

7.3.2 Issues for multivariate indicators

Multivariate data for biological indicators in these Guidelines typically consist of

either the presence or absence of taxa or their abundances across the sample units.

These data can then be summarised as similarities (or dissimilarities) between each

pair of sample units. The Bray–Curtis measure, among a few others, has been

demonstrated to be the best choice for such biological data (Faith et al. 1987), and

a See Section

7.3.2

c See also the

Monitoring

Guidelines

Sections 6.5.2,

6.6.1

b Ch 6 of the

Monitoring

Guidelines
Chapter 7 — Monitoring and assessment

page 7.3–2 Version — October 2000

sometimes transformation of the data is desirable before the similarity measure is

computed, as described in the protocols in Appendix 3 (Vol 2).

The result of these computations is a triangular matrix of similarity values. These

are not easily analysed in conventional statistical procedures such as analysis of

variance or regression. However, one situation which is amenable to the use of

similarity measures in more conventional procedures is where the ‘control’ and

‘impact’ sites can be paired, on rivers for example. In figure 7.3.1, there is a pair of

sites on each tributary river which are comparable in terms of habitat and separated

by similar distances on each river. The similarities between the upstream and

downstream sites on each river could be computed for a number of times before the

disturbance; if the similarities computed between the paired sites decreased after

the start of the impact relative to the similarities between paired sites on the control

rivers, then an impact is likely to have occurred. Examples using this design are

Faith et al. (1995) and Davies and Nelson (1994).

Figure 7.3.1 Schematic diagram of a river system with paired upstream (black diamonds)

and downstream (white diamonds) sites on each tributary.

Grey arrows indicate locations of disturbances.

Pairing of sites in this way is not always possible, however. Permutation tests that are

analogues of some of the simpler conventional techniques (Smith 1998, the

ANOSIM of Clarke & Green 1988, Clarke & Warwick 1994) have been used (e.g.

Smith 1994). Significance testing of multivariate data based on similarity measures

using permutation tests is rapidly developing (see Legendre & Legendre 1998 for an

overview; Legendre & Anderson 1999 for an attempt at analysing multifactorial

data). It is likely that methods for the analysis of similarity data in complex designs

will become available in the near future.

The more conventional methods of analysing dissimilarity data have focused on

displays of the data via such techniques as multidimensional scaling, principal

components analysis and correspondence analysis (review: Legendre & Legendre

1998; brief description of principal components analysis and multidimensional

scaling in Sections 6.5.4 and 6.6.3 of the Monitoring Guidelines). Inferences have

been made purely on the basis of striking patterns in such displays, and Green

(1979) argued that obvious patterns in such graphs were likely to correspond to

large impacts. Such a procedure obviously lacks the sensitivity required for some

assessment objectives. However, these displays remain an important tool for

interpretation and communication after a formal hypothesis test via a

randomisation or permutation procedure (Clarke & Warwick 1994).
7.3.3 Use of AUSRIVAS

Version — October 2000 page 7.3–3

7.3.3 Use of AUSRIVAS

7.3.3.1 Outline of AUSRIVAS

AUSRIVAS is a rapid biological assessment procedure developed for rivers and

streams (Schofield & Davies 1996) similar to the British RIVPACS system

(Wright et al. 1993). It currently uses macroinvertebrate data, but the use of other

taxa (e.g. diatoms and fish) is being researched. Its applicability to wetlands and to

New Zealand rivers is also being investigated.

AUSRIVAS is a specialised example of a monitoring design that relies on spatial

information alone to infer whether a disturbance has caused an impact. In general

terms the problem can be stated thus: to judge whether a particular site has been

disturbed by some activity or event, other, apparently undisturbed, sites that are

similar in their environmental attributes must be found to act as a standard or

control for comparison. The site suffering the supposed disturbance is designated

the test site, while the sites acting as controls are called reference sites.

In AUSRIVAS a large number of reference sites with as high an environmental

quality as possible have been identified across a wide variety of river types and

ecosystems, sampled for their macroinvertebrates and had their habitats

characterised by a standard set of physical and chemical variables that are largely

unrelated to likely pollutants. This set of reference sites has then been classified

according to their biota to produce groups of sites containing similar fauna. A

numerical analysis has then been used to identify the environmental attributes

which describe each group of reference sites. Now, any test site requiring

assessment has its environmental attributes compared with those of the reference

sites to determine which group or groups of reference sites it resembles most

closely. The fauna of these corresponding reference sites is then compared with the

test site: if the test site supports fewer taxa than are predicted by the reference sites,

it is judged to be disturbed.

7.3.3.2 Sampling protocol and issues about effect size and sensitivity

The AUSRIVAS protocols modify much of the advice given in Sections 7.2.3 to

7.2.6 above. Site selection, the methods of stratifying habitats within sampling

sites, the timing of sampling and the analytical methods and outputs are all

specified in protocols for each state and territory in Australia. The summary indices

and recommendations for reporting procedures are also standardised, and the

sampling, sorting and identification steps are subject to QA/QC programs. The

software for analysing the data is maintained and developed at a central location

accessible via the AUSRIVAS home page.

Decisions about effect size are implicit in the procedure. The degree of impact

upon a site is judged by the values of summary indices relative to a stipulated

percentile of the reference sites that act as spatial controls. If a site scores a value

on these indices that is smaller than 90% of the values recorded for reference sites,

the fauna is deemed to be lacking some of the families of invertebrates that could

be expected at that site. Although designation of such a percentile threshold

expresses an effect size in terms of how deviant a site is from reference conditions,

it is analogous to but not exactly equivalent of the process of setting Type I and II

error rates for conventional statistical procedures.
Chapter 7 — Monitoring and assessment

page 7.3–4 Version — October 2000

Nevertheless, the issue of how far a potentially-disturbed site can deviate from

reference or control conditions before an impact is deemed to have occurred must

still be resolved with stakeholders. Strategies similar to those described above for

procedures based on statistical methods can be employed: i.e. use of existing

information, examining the response of the indicator variable to known impacts of

different degrees of severity, and use of pilot data in simulation modelling. Note

that indicators used in rapid, broad-scale methods are often quite coarse (e.g. use of

family-level rather than species-level identifications). Thus the threshold value at

which the decision is made that an impact has occurred should take account of the

harmfulness of the potential impact, its reversibility and the time-lags between an

event and the implementation of actions to prevent harm. The threshold value may

need to be set at a more conservative value than that deemed acceptable by

stakeholders so that management has time to react and prevent irreversible harm.

7.3.3.3 Application and cautions

AUSRIVAS has been promoted in these Guidelines as ideally suited for the rapid,

cost-effective, first-pass determination of the extent of a problem or potential

problem, e.g. as applied to broad-scale land-use issues. Earlier, a note of caution

was provided for use of the method in applications other than these,a particularly

for detecting impacts of a minor nature and for site-specific assessments where the

method requires additional testing and the addition of more data. A perspective on

some of the limitations of the approach is provided below, together with comments

on ongoing data collection and proposed research and development aimed at

improving the sensitivity and broadening the application of this procedure.

An important aspect of AUSRIVAS is the availability and selection of suitable

reference sites. In some regions of Australia it is easy to find reference sites on

rivers and streams draining relatively intact catchments. Unfortunately, large

regions of Australia have been subject to broad-scale impacts and there are no

‘near-pristine’ sites from which to select biogeographically relevant reference sites

(e.g. wheat belt of Western Australia, lowland reaches of the Murray-Darling

Basin). Thus, in AUSRIVAS, the least impacted sites of such regions have been

targeted to provide reference sites for setting targets for rehabilitation of the more

degraded sites; however, this does not solve the problem of assessing the degree of

degradation of the reference sites themselves. Without pre-impact data, this task is

outside the ambit of routine prescriptive procedures and would require a variety of

situation-specific case studies to arrive at some assessment. The issue is being

addressed as part of the current Australia Wide Assessment of River Health

(AWARH), which aims to report on the ecological condition of around 4000

Australia river sites by the year 2000 using AUSRIVAS.

A related but more tractable problem results when a test site has no close

environmental equivalents in the reference database. Therefore, an important initial

step in evaluating a test site is a statistical comparison between its environmental

attributes and those of all the reference sites: if it has no sites with similar attributes

in the reference set, no further assessment can be made, i.e. there are insufficient

sites in the database that can be regarded as a ‘control’ (Furse et al. 1987). The

current AUSRIVAS software (available from the AUSRIVAS homepage) contains

a testing routine to assess whether test sites fall within the ‘domain’ of the existing

reference site set on which the bioassessment models are based. It then must be

a See box

3.2.1, Section

3.2.2.1/3
7.3.3 Use of AUSRIVAS

Version — October 2000 page 7.3–5

decided whether the test site warrants the added expense of adding sufficient

comparable reference sites to the database to enable an assessment to be made.

A potential drawback is the relatively large number of reference sites that must be

sampled to build reliable models for predicting the presence or absence of the

target organisms. This is particularly relevant to site-specific assessments, where

adequate characterisation of the local reference condition is critical. The

development of AUSRIVAS is predicated on the collection of a large amount of

reference site data nationally, and it is anticipated that the geographic spread, as

well as the spatial density of sites, will gradually increase to improve the

applicability of the predictive models.

Another important aspect for some Australian streams is the natural inconstancy of

animal community composition amongst years. Thus, ‘high’ temporal variability of

macroinvertebrate communities over large parts of Australia, particularly semi-arid

and northern regions prone to drought and/or cyclonic disturbance (Humphrey et

al. 2000), may reduce to some (as yet unknown) extent the sensitivity of ‘static’

models derived for these locations. To this end, it is recommended that test site

assessment using the AUSRIVAS protocol should be done in parallel with

reference (‘control’) site assessment to assess the degree of natural temporal

change in macroinvertebrate community composition and compare it with the

summary index value for the test site.

In some regions of Australia it is clear that some reference sites are naturally

depauperate; that is, the number of macroinvertebrate taxa is low. For procedures

such as AUSRIVAS, where the final reporting indices are based on the ratio of

the number of taxa observed to the number of taxa expected, this poses potential

problems for the sensitivity and robustness of the final assessment, even at

species level.

Finally, AUSRIVAS and related procedures (Reynoldson et al. 1995) are rapid

assessment tools and will only detect impacts that are severe enough to eliminate

taxonomic groups of organisms. The formal hypothesis testing associated with

conventional statistical methods has no clear analogue here. This procedure uses

a suite of reference sites to predict the expected composition of families of

invertebrates at a test site; if the test site has fewer families than expected based

on the distribution of reference site values, it is deemed to be disturbed.

Nevertheless, several basic considerations of survey design (sample and site

replication, etc.) still apply to assessments or surveys conducted with

AUSRIVAS. These considerations become particularly important at small spatial

scales (i.e. a specific activity, development, point-source disturbance, within a

catchment) where stronger inference and greater sensitivity to impact may be

required. If AUSRIVAS is to be adopted in these situations, it must be conducted

in a design framework that has adequate sample and site replication to enable the

study objectives to be met. If necessary, aspects of the rapid biological

assessment protocol may need to be adapted or modified so that the data gathered

are amenable to both AUSRIVAS and quantitative assessment.a

a

Complementary

roles for

quantitative

and rapid

assessment in

monitoring

programs are

recommended

in Section

7.2.1.1 above

Version — October 2000 page 7.4–1

7.4 Specific issues for physical and chemical indicators

(including toxicants) of water and sediment

This section outlines issues specific to physical and chemical indicators (including

toxicants) of water and sediment that need to be borne in mind when designing

monitoring and assessment programs. The issues in Sections 7.4.1 and 7.4.3 are

comprehensively discussed in Chapters 3–6 of the Monitoring Guidelines — see

table 7.4.1 for a checklist of these issues and appropriate cross-reference to the

Monitoring Guidelines.

Table 7.4.1 Checklist for sampling and analysis of physical and chemical indicators with

cross-reference to details provided in the Monitoring Guidelines.

Issue Chapter or section from the Monitoring

Guidelines

Representative sampling Chapter 3, Chapter 4

spatial boundaries 3.3.1

scale 3.3.2

duration 3.3.3

patterns of sampling 3.4.1

selection of sites 3.4.2

frequency of sampling 3.4.3

numbers of samples 3.4.4, A5.1.10

Surface water sampling 4.3.1, 4.3.2

hydrology, flow variations, runoff 3.4.3, 3.4.3.2, 4.3.1

stratification 3.4.1.2, 3.4.2.1

human effects on contaminant loads and timing 3.4.3.2

automatic samplers 3.4.3.2, 4.3.2

time of day 3.4.3, 3.4.3.2

Sediment sampling and sediment sample handling 4.3.1, 4.3.5, 3.4.2.1, 5.5.8

potential for contamination 4.3.5, 4.3.1

suspended sediments 4.3.5

Sample storage and handling 4.5, 4.6, 4.3

Chemical speciation 5.5.8.2, Tables 4.5 & 5.2

Bioavailable concentration vs total concentration 3.5

Quality assurance/Quality control in the field 4.6 and subsections

chain of custody 4.6.1

training of staff 4.7.2

quality assurance samples: blanks 4.6.3.1

quality assurance samples: replicates 4.6.3.2

quality assurance samples: spiked samples 4.6.3.3

pilot trial 3.4, 3.4.2, 3.4.4

equipment 4.6, 4.6.2, 4.6.3.1

sample transport 4.6.1, 4.6.2, 4.6.3.1

site access 3.4.2, 4.2, 4.7.1, 4.7.3

occupational health and safety 4.7 and subsections

analytes 5.3

cleaning and calibration 4.3.1, 4.3.2.1, 4.3.6, 4.6.1, 4.6.2

protocols 4.3.7, 4.6.2

Quality assurance/Quality control in the laboratory 5.5 and subsections

chain of custody 5.4.1.2

occupational health and safety 5.6 and subsections

training of staff 5.6.3

quality assurance program 5.5.5 and subsections

quality assurance samples subsections of 5.5.5

matrix compatibility 5.5.5.1

accurate recording of data 5.4, 6.2
Chapter 7 — Monitoring and assessment

page 7.4–2 Version — October 2000

In addition to the cross-references provided to sampling and analysis of sediments

in table 7.4.1, a protocol describing key aspects of collection and laboratory

analysis of sediment samples is provided in Appendix 8 of Volume 2, while advice

on comparing sediment ‘test’ data with default guideline values is provided in

Section 7.4.4.4 below.

These Guidelines emphasise the use of guideline trigger values for assessing the

environmental significance of physical and chemical indicators. The statistical

procedure for comparing test data and a trigger value is described in Section 7.4.4.a

The generic considerations for sampling design given in Section 7.2 also apply to

physical and chemical indicators.

7.4.1 Hydrology and representative sampling

Sampling of waters and sediments must be representative.b The challenge is to

sample in enough detail to outline a picture of the natural variations in time and

space and to reliably detect deviations from this natural ‘background’ variation.

Natural variations in surface waters and groundwaters, whether flowing or

standing, can affect the values of physical and chemical indicators. For example,

all water bodies can form vertical or horizontal layers of differing temperature or

salinity that may or may not need to be sampled separately, according to the

sampling plan. Currents and the lateral and vertical movements of different water

masses also need to be considered during the planning of field sampling, analyses

and study design. Natural periodicity, and the timing of industrial discharges into

water bodies, and the considerable effects of runoff in inland waters can make large

differences to the loads and concentrations of physical and chemical indicators, and

must also be planned for.

In sediments,c the sampling plan and study design must consider the effects of

natural layering, mixing, and variations in particle size and porosity on the

indicator being sampled. The likelihood of disturbance and cross-contamination

during sampling must not be forgotten. Suspended sediments need to be collected

in a representative manner (Batley 1989), as do sediment pore waters.

For all samples, precautions must be planned and taken to prevent the values of the

indicators changing during storage and transport.

7.4.2 Chemical speciation in water samples

The issue of the chemical form of physical and chemical indicators (that is, the

compound(s) of the indicator present in the sample) are relevant regardless of the

use envisaged for the water. Speciation (the form of the chemical) assumes critical

importance where the environmental value concerns ecosystem protection or

human health. The form of the indicators needs to be determined and those

chemical species that are likely to affect the environmental value must be

identified. In the past, total (i.e. unfiltered) concentrations were measured and

compared with guideline values on the understanding that this approach probably

overestimates the amount of deleterious form(s) of the indicator. This approach to

protection may be overconservative. A refinement is to measure and compare total

filtered concentrations. This, too, is a conservative approach (though less so)

because the diversity of chemical forms of a physical and chemical indicator in the

solution may have different effects on an environmental value.

c See Vol 2,

App 8 and

Monitoring

Guidelines

Sections 4.3,

4.3.5

a See also the

Monitoring

Guidelines

Section 6.4.3

b See

Monitoring

Guidelines

Chapters 3 & 4
7.4.3 Quality Assurance and Quality Control (QA/QC)

Version — October 2000 page 7.4–3

There are at least two ways to resolve the speciation problem.

• Determine the indicator using an analytical method that is specific to the

chemical species. While this is an improvement on using total filtered

concentrations, it requires the species or range of species detected by the

method to be arbitrarily defined as a surrogate for the species affecting (usually

detrimentally) the environmental value. An example of this approach is the use

of anodic stripping voltametry in the determination of copper. The fraction

determined under operationally defined conditions is identified as labile forms

which, in turn, are believed to be the forms in which copper is most

bioavailable.

• Use thermodynamic speciation modelling. One requirement of this mathematical

tool is that all aqueous chemical species that may be important to the chemical

form of indicators be measured. This usually increases the analytical

requirements because of the inclusion of chemical species that would not

otherwise be determined. The technique requires that the system measured is in

equilibrium, and that the equilibrium is the same as that existing at the time of

sampling. This has implications for preservation, transport and storage of

samples. The specification and interpretation of thermodynamic speciation

models is complex and requires considerable facility in the use of computers, and

in the interpretation of chemical data. A more detailed discussion of speciation

modelling is beyond the scope of these Guidelines.

7.4.3 Quality Assurance and Quality Control (QA/QC)

Quality control and quality assurance were defined generically in Section 7.2.4. A

specific formal statement of quality control for physical and chemical indicators is

this:

The overall objective of quality control in the measurement of physical and chemical

variables is the determination of the exact indicator concentration that existed at a

specifically defined location at the time the sample was taken. In most cases this

requirement extends to the chemical speciation of the indicator.

Neglect of QA/QC is probably the most important reason for the unreliability of

most historical chemical data.

Protocols for field and laboratory aspects of sampling must be followed carefully,

as discussed in the Monitoring Guidelines.a QA/QC begins with the choice and

training of competent field and laboratory staff; it includes the choice and

maintenance of field and laboratory equipment and vehicles. It extends to the

checking of analytical methods and analytical performance, the tracking of each

sample throughout sampling and analysis, and the accurate recording of data in the

final database.

7.4.4 Comparing test data with guideline trigger values

7.4.4.1 Physical and chemical stressors

This section provides a summary of the approach recommended for comparing

results from a test site with a guideline trigger value. Details of the method are

contained in Appendix 7 of Volume 2; Section 6.4.3 of the Monitoring Guidelines

touches on it also. There are a number of common statistical methods that are

a See

Monitoring

Guidelines

Chapters 4, 5

and 6
Chapter 7 — Monitoring and assessment

page 7.4–4 Version — October 2000

potentially applicable for this purpose, although experience suggests that the

assumptions underpinning many ‘conventional’16 statistical tests are often violated

by water quality data (and sometimes quite seriously so). Section 6.3.4 of the

Monitoring Guidelines recommends transformations to correct specific problems,

although the action required will depend very much on the characteristics of the

data at hand. This lack of consistency in the way site-specific data may be

processed and interpreted is an impediment to the development of a simple,

straightforward trigger rule.

Compounding this difficulty is the usual requirement to specify the magnitude of

change in a particular statistical parameter (e.g. mean, variance, percentile) that is

deemed to be ‘significant’ — either ecologically or statistically or both. The

quantification of a minimum effect size that can be claimed to be ecologically

important is difficult. With respect to the trigger rule outlined in this section, this

issue of ecological importance is discussed further below and more generally in

Section 7.2.3.3. The important observation to note at this stage, however, is that

exceedances of the trigger values are an ‘early warning’ mechanism to alert

managers of a potential problem. They are not intended to be an instrument to

assess ‘compliance’ and should not be used in this capacity.

During the development of a suitable trigger mechanism, considerable attention

was given to the following design requirements:

• explicit recognition of the inherent (and usually large) variability of natural

systems;

• robustness under a wide range of operating conditions and environments;

• no, or only weak, distributional assumptions about the population of values from

which the test and reference data are obtained;

• known statistical properties, consistent with and supporting the monitoring

objectives of this document;

• ease of implementation and interpretation;

• suitability for visual display and analysis;

• intuitive appeal.

The recommended trigger-based approach for physico-chemical stressors may be

stated as follows.

A trigger for further investigation will be deemed to have occurred when the median

concentration of n independent samples taken at a test site exceeds the eightieth percentile

of the same indicator at a suitably chosen reference site. Where suitable reference site data

do not exist, the comparison should be with the relevant guideline value published in this

document.

This rule satisfies the first dot point above since it is statistically-based and

acknowledges natural background variation by comparison to a reference site. Its

robustness derives from the fact that it accommodates site-specific anomalies and

uses a robust statistical measure as the basis for triggering. No assumptions are

 

16 In this context, the term conventional is used to denote statistical procedures based on the general

linear statistical model having normally distributed errors.
7.4.4 Comparing test data with guideline trigger values

Version — October 2000 page 7.4–5

required to be made about the distributional properties of the data obtained from

either the test or reference sites. The computational requirements of the approach

are minimal and can be performed without the need for statistical tables, formulae,

or computer software. As demonstrated later in this section, the temporal sequence

of trigger events is readily captured in a simple plot.

It should be understood that the trigger protocol is responsive to shifts in the

location (i.e. ‘average’) of the distribution of values at the test site. While

differences in shape of the reference and test distribution may be important in some

instances, this is a secondary consideration that is not specifically addressed by this

protocol. It is also important to note that the role of the 80th percentile at the

reference site is to simply quantify the notion of a ‘measurable perturbation’ at the

test site. The protocol is not a statistical test of the equivalence of the 50th and 80th

percentiles per se. The advantages of using a percentile of the reference

distribution are 1) it avoids the need to specify an absolute quantity and 2) because

the reference site is being monitored over time, the trigger criterion is being

constantly updated to reflect temporal trends and the effects of extraneous factors

(e.g. climate variability, seasonality).

Implementation of the trigger criterion is both flexible and adaptive. For example,

the user can identify a level of routine sampling (through the specification of the

sample size n) that provides an acceptable balance between cost of sampling and

analysis and the risk of false triggering. The method also encourages the

establishment and maintenance of long-term reference monitoring as an alternative

to comparisons with the default guideline values provided in Section 3.3 that do not

account for site-specific anomalies.

The remainder of this section addresses sampling issues, data requirements,

computational procedures and statistical properties associated with the proposed

method. The mathematical detail associated with computation of Type I and Type

II errors may be found in the Annex of Appendix 7 of Volume 2. Worked examples

of the computations and performance aspects of the trigger rule are provided in

Appendix 7 (Volume 2).

1 Data requirements at the reference sites

Prior to implementing the trigger rule, the user will need to have addressed some

data collection issues.

• Reference site selection: selection of (a) suitable reference site(s) has been

addressed in Section 3.1.4.

• Minimum data requirements at the reference site: a minimum of two years of

contiguous monthly data at the reference site is required before a valid trigger

value can be established. Until this minimum data requirement has been

established, comparison of the test site median should be made with reference

to the default guideline values identified in Section 3.3 of this document.

2 Computation of the 80th percentile at the reference site

The computation of the 80th percentile at the reference site is always based on the

most recent 24 monthly observations. The procedure is as follows:

(i) arrange the 24 data values in ascending (i.e. lowest to highest) order,

(ii) take the simple average (mean) of the 19th and 20th observations in this

ordered set.
Chapter 7 — Monitoring and assessment

page 7.4–6 Version — October 2000

3 Updating the reference site data and 80th percentile

Each month, a new reading at the reference (and test) site is obtained. The

reference site observation is appended to the end of the original (i.e. unsorted) time

sequence. Steps (i) and (ii) from 2 above are applied to the most recent 24 data

values. Note, even though only the most recent two years of data is used in the

computations, no data should be discarded.

Maintenance of the complete data record will allow longer-term statistics to be

computed. For example, after five years of monthly monitoring, all sixty

observations could be used to compute the overall 80th percentile. This could serve

as a useful benchmark against which the ‘rolling’ monthly percentiles could be

compared for evidence of trends.

4 Data requirements at the test site

A feature of the method is the flexibility it provides the user for the allocation of

resources to the sampling effort. As previously mentioned, there is no fixed

requirement to monitor at a reference location (i.e. the default guideline values can

be applied). Similarly, the choice of sample size at the test site is arbitrary,

although there are implications for the rate of false triggering. For example, a

minimum resource allocation would set n=1 for the number of samples to be

collected each month from the test site. It is clear that the chance of a single

observation from the test site exceeding the 80th percentile of a reference

distribution which is identical to the test distribution is precisely 20%. Thus the

Type I error in this case is 20%. This figure can be reduced by increasing n. For

example, when n=5 the Type I error rate is approximately 0.05. The concomitant

advantage of larger sample sizes is the reduction in Type II error (the probability of

a false no-trigger). So-called ‘power curves’ are provided in Appendix 7

(Volume 2) to assist in understanding the consequences upon error rates of a

particular sampling strategy at the test location.

5 Computation of the median at the test site.

The median is defined to be the ‘middle’ value in a set of data such that half of the

observations have values numerically greater than the median and half have values

numerically less than the median. For small data sets, the sample median is

obtained as either the single middle value after sorting in ascending order when n is

odd, or the average of the two middle observations when n is even.

6 Ecological importance

The proposed trigger rule does not purport to define or represent an ecologically

important change. As previously explained, the trigger approach is an early

warning mechanism to alert the resource manager of a potential or emerging

change that should be followed up. Whether or not the actual change in condition

at the test site has biological and/or ecological ramifications can only be

ascertained by a much more comprehensive investigation and analysis. To make

this distinction clear, the concept of a measurable perturbation is introduced. Our

de facto definition of a measurable perturbation is that it is the magnitude of the

shift between the 50th and 80th percentiles at a reference site. While the definition

is arbitrary, it does have broad acceptance and intuitive appeal among experts. It

should also be noted that the statistical significance associated with a change in

condition equal to or greater than a measurable perturbation would require a

separate analysis.
7.4.4 Comparing test data with guideline trigger values

Version — October 2000 page 7.4–7

7 Performance characteristics

It is important that the statistical performance characteristics of any test or

decision-making rule are documented and understood to avoid unduly conservative

or liberal triggering.

The foregoing discussion makes no assumptions regarding the shape of the reference

and test distributions. Without this knowledge, a formal calculation of Type I and

Type II errors is not possible. However, as a general principle, increasing the

frequency of collection of independent samples will reduce the magnitude of both

errors. A more complete discussion of the performance characteristics of the

recommended approach is provided in Appendix 7 of Volume 2.

8 On-going monitoring — the control chart

The foregoing has been provided to assist with the month-by-month comparisons.

It is suggested that these monthly results be plotted in a manner indicated in

figure 7.4.1 below. This provides a visual inspection of all results and helps

identify trends, anomalies, periodicities and other phenomena. The methods in

Chapter 6 of the Monitoring Guidelines can be used to model trends and other

data behaviour if required.

Month

1 2 3 4 5 6 7 8 9 10 11 12

concentration units

0

1

2

3

next level of investigation triggered

no action required

warning - investigation may be necessary

Test site median

Reference site P80

Figure 7.4.1 Control chart showing physical and chemical data (Y axis) for test and

reference sites plotted against time, and recommended actions

9 Comparing test data against single guideline (default values)

In the absence of suitable reference site data (as defined in step 1 above), the

median of the test site data is to be compared with the default guideline value

identified in Section 3.3.2.5 of this document. This guideline value has been

computed as the 80th percentile of the amalgamation of a number of historical data

sets across broad geographical regions. Unlike the comparison with a locallyderived 80th percentile, the guideline value is static and will not reflect any local

spatial and/or temporal anomalies. Reference site monitoring is strongly advocated
Chapter 7 — Monitoring and assessment

page 7.4–8 Version — October 2000

if these effects are considered to represent a significant source of departure from

the guideline value.

Figure 7.4.2 below illustrates the difference in control charting procedures when

the guideline value is used in place of a trigger obtained using the 80th percentile

from reference site monitoring.

Month

1 2 3 4 5 6 7 8 9 10 11 12

concentration units 0

1

2

3

next level of investigation triggered

no action required

warning - investigation may be necessary

Guideline trigger

Test site median

Figure 7.4.2 Control chart showing physical and chemical data (Y axis) for test site plotted

against default trigger value, time and recommended actions

7.4.4.2 Toxicants

This section describes the general needs for comparing toxicant test data with

guideline trigger values. Conceptually, toxicants and ‘physical and chemical

stressors’ are subcategories of the same class of potentially hazardous indicators,

being properties or (usually) constituents of the aquatic environment. However, the

treatment of these groups for guideline purposes is different. Specifically, toxicants

are usually compared with a single default trigger value, less commonly with a

background or reference distribution. The default values are prepared by analysis

of a comprehensive set of available ecotoxicological data. Physical and chemical

stressors at a test site are usually compared with those at a reference site. The latter

reference-comparison approach, however, has its parallels in measurement

programs for toxicants, as described in 1 below.

1 Background data that may supplant guideline default trigger values

Some surface waters will contain concentrations of toxicants that may naturally

exceed the default guideline trigger values tabulated in Section 3.4. Where this is

the case and as recommended in Section 3.4.3.2, new trigger values should be

based on background (or baseline) data. (Note that ‘background’ in this case, refers

to natural toxicant concentrations that are unrelated to human disturbance.) As a

matter of course, gathering of background data is always recommended, at least in
7.4.4 Comparing test data with guideline trigger values

Version — October 2000 page 7.4–9

the initial stages of a water quality management program, to establish whether or

not concentrations of toxicants are naturally high.

Toxicant concentrations may vary seasonally. Because of this and the need to be

confident about the best estimate of background concentrations, it is recommended

that background data be gathered on a monthly basis for at least two years. In all

respects, data requirements and collection are the same as for physical and

chemical stressors, as described above.a Until this minimum data requirement has

been established, comparison of the test site median should be made with reference

to the default guidelines identified in Section 3.4.3 of this document.

For those months, seasons or flow periods that constitute logical time intervals or

events to consider and derive background data, the 80th percentile of background

data (from a minimum of 10 observations) should be compared with the default

guideline value. This 80th percentile value is used as the new trigger value for this

period if it exceeds the default guideline value provided in Section 3.4.3 of this

document. Test data are compared with the new trigger values using the same

principles as outlined in steps 2–8 above for physical and chemical stressors.

Where background toxicant values fall consistently below default trigger values,

sampling intensity at these sites could be reduced after a suitable period (e.g. two

years).

2 Comparing test data with default guideline values

In practical terms, the method for comparing toxicant test data with default

guideline values should be similar to the approach recommended in step 9 above

for physical and chemical stressors.b However, it is recommended that a more

conservative approach should apply to the comparison of toxicant test data with

default guideline values. Specifically, it is recommended that action is triggered if

the 95th percentile of the test distribution exceeds the default value (or stated

differently, no action is triggered if 95% of the values fall below the guideline

value). The more stringent approach is recommended here because, unlike physical

and chemical stressors, toxicant default values are based upon actual biological

effects data and so by implication, exceedance of the value indicates the potential

for ecological harm. Note that because the proportion of values required to be less

than the default trigger value is very high (95%), a single observation greater than

the trigger value would be legitimate grounds for action in most cases, even early

in a sampling program.

7.4.4.3 Physical and chemical (including toxicant) data gathered from surface waters

‘upstream’ of the test site

In many situations, particularly where additional human use activities are present

‘upstream’ of the test site of interest, the regular collection of data from upstream

of the test site will be necessary. These data will be compared with the test data of

interest to assist in determining the source and cause of any possible elevated

toxicant concentrations found at the test site. Where there are multiple sources of

toxicants along a water-course, catchment managers will need to establish

appropriate data analysis and assessment procedures to apply.

a See step 1,

Section 7.4.4.1

b Section

7.4.4.1
Chapter 7 — Monitoring and assessment

page 7.4–10 Version — October 2000

7.4.4.4 Sediments

The application of the decision treea reverts to reference or background site

concentrations if these exceed the trigger values. The selection of an acceptable

reference site for water quality studies has been discussed earlier.b Basically the

same considerations apply to sediments, with the additional option, that a reference

or background condition can also be established from measurements at depths in

sediment cores below observed concentration excursions.

While temporal variability is used to characterise water quality parameters at a

reference site, this is clearly inappropriate for sediments where the accumulation

rates are typically below 1 cm/y. It is more appropriate to use spatial variability,

either based on depth profiles at a test site or an appropriate number of surface

sediment samples, to characterise a site. Sites will typically contain a range of grain

sizes, and determining median concentrations and 80th or 95th percentile values may

distort any comparison. It is important that in comparing test and reference sites,

samples with a similar grain-size distribution are used. Normalising to a fine grain

size (e.g. <63 μm) is inappropriate, as the normalised value will have less of an

impact on biota when diluted with coarser sediments that usually contain lower

contaminant concentrations.

The spatial scale over which the reference and test site measurements are taken is a

matter for decision by stakeholders, based on sound scientific judgement. The

heterogeneity of sediment samples with respect to contaminants largely mirrors the

differences in grain size. Defining the size of the test site will be a regulatory

responsibility, in terms of the spatial extent of contaminated sediment that is

acceptable in the region of interest. As a guide, the spatial extent of a test site may be

a geographical feature, for example, a delta or an embayment within a harbour.

Alternatively, a test site may comprise a recognised ecological habitat, for instance a

riffle zone in a stream or a defined area of fine sediment in a lake. In a large water

body the test site might be larger than in a narrow river or creek, where biota might

have difficulty in avoiding the contamination. The area of any reference site should

be comparable to that of the test site, and the grain size must be similar.

Because of the poor reliability of the sediment trigger values it is difficult to be

prescriptive about how these can be compared with test values. The same applies to

the comparison of reference site values with test sites, where comparisons of

reference median or 80th percentile with the test site median may be equally

appropriate in giving an estimate of the relative concentrations, which is really all

that is required in the case of sediments. However, where sediment samples within

a test site clearly exceed trigger values, or are reasonably inferred to be

ecologically hazardous, these Guidelines recommend additional sampling to more

precisely delineate contaminated zones within the site.

a See Section

3.5

b Section 3.1.4
Version — October 2000 page R1–1

References

Chapter 1 Introduction

ANZECC 1992. Australian water quality guidelines for fresh and marine waters. National Water

Quality Management Strategy Paper No 4, Australian and New Zealand Environment and

Conservation Council, Canberra.

ANZECC & ARMCANZ 1994. Policies and principles: A reference document. National Water

Quality Management Strategy Paper No 2, Australian and New Zealand Environment and

Conservation Council & Agriculture and Resource Management Council of Australia and New

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Version — October 2000 page A–1

Appendix 1 Acronyms and glossary of terms

Acronyms

ACC Acceptable Contaminant Concentration

ACR Acute-to-chronic ratios

AE Alcohol ethoxylated surfactants

AES Alcohol ethoxyolated sulfate surfactants

AGPS Australian Government Publishing Service

ANCA Australian Nature Conservation Agency

ANZECC Australian and New Zealand Environment and Conservation Council

ANZFA Australia New Zealand Food Authority

AQUIRE Aquatic Toxicity Information Retrieval Database

ARMCANZ Agricultural and Resource Management Council of Australia and New

Zealand

ASSAC Australian Shellfish Sanitation Advisory Committee

ASSCP Australian Shellfish Sanitation Control Program

ASQAP Australian Shellfish Quality Assurance Program

ASTM American Society for Testing and Materials Designation

AUSRIVAS Australian River Assessment Scheme

AVS Acid volatile sulfide

BACI Before– After, Control–Impact

BACIP Before–After, Control–Impact Paired

BCF Bioconcentration factor

BEDS Biological effects database

BOD Biological oxygen demand

BOM Biodegradable organic matter

CCL Cumulative Contaminant Loading Limit

CCME Canadian Council for Ministers of the Environment

CCREM Canadian Council for Resource and Environment Ministers

CEC Cation exchange capacity

CFU Colony forming units

COAG Council of Australian Governments

COD Chemical oxygen demand

CSIRO Commonwealth Scientific & Industrial Research Organisation

DASET Department of Arts, Sport, Environment and Territories

DCC Desirable contaminant concentration

DEST Department of Environment, Sport and Territories

DISR Department of Industry, Science and Resources

DO Dissolved oxygen

DOC Dissolved organic carbon

DTA Direct toxicity assessment

DWAF Department of Water Affairs and Forestry

DUAP Department of Urban Affairs and Planning
Appendix 1 Acronyms and glossary of terms

page A–2 Version — October 2000

EC Electrical conductivity

ECLs Environmental concern levels

EEZ Exclusive Economic Zone

EIA Environmental Impact Assessment

EIS Environmental Impact Statement

ENSO El Nino Southern Oscillation

ERIN Environmental Resource Information Network

eriss Environmental Research Institute of the Supervising Scientist

ES Effect size

ESD Ecologically sustainable development

ESP Exchangeable sodium percentage

EV Environmental value

FNARH First National Assessment of River Health

GESAMP Joint Group of Experts on the Scientific Aspects of Marine Pollution

ICM Integrated catchment management

ICPMS Inductively coupled plasma mass spectrometry

IMO International Maritime Organisation

ISQG Interim sediment quality guideline

LAS Linear alkylbenzene sulfonates

LOEC Lowest observed effect concentration

LWRRDC Land and Water Resources Research and Development Corporation

MATC Maximum acceptable toxicant concentration

MBACI Multiple Before–After, Control–Impact

MBACIP Multiple Before–After, Control–Impact, Paired

MDBC Murray Darling Basin Commission

MHSPE Ministry for Housing, Spatial Planning and the Environment

MPC Maximum permitted concentration

NATA National Association of Testing Authorities of Australia

NHMRC National Health and Medical Research Council

NICNAS National Industrial Chemicals Notification and Assessment Scheme

NIWA National Institute of Water and Atmospheric Research

NOAA US National Oceanic and Atmospheric Administration

NOEC No observable effect concentration

NPDES National Pollutant Discharge Elimination System

NRC National Research Council

NRHP National River Health Program

NSSP US National Shellfish Sanitation Program

NSWDWR NSW Department of Water Resources

NSWEPA NSW Environmental Protection Authority

NWQMS National Water Quality Management Strategy

OECD Organisation for Economic Co-operation and Development

PAR Photosynthetically available radiation

PCBs Polychlorinated biphenyls
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–3

PQL Practical quantitation limit

QA/QC Quality assurance/quality control

RBA Rapid biological assessment

RIVPACS Riverine Invertebrate Prediction and Classification System

SACC State Algal Coordinating Committee

SAR Sodium adsorption ratio

SCARM ARMCANZ Standing Committee for Agricultural and Resource Management

SCEP ANZECC Standing Committee on Environmental Protection

SEM Simultaneously extracted metals

SoE State of the Environment

SPM Suspended particulate matter

TAN Total ammonia nitrogen

TCM Total catchment management

TDS Total dissolved solids

TIE Toxicity identification & evaluation

TTM Total toxicity of mixtures

UNESCO United Nations Education Scientific and Cultural Organization

USEPA United States Environmental Protection Agency

VEPA Victoria Environment Protection Authority

WADEP Western Australian Department of Environmental Protection

WAEPA WA Environmental Protection Authority

WAWA WA Water Authority (now split between the Water Corporation Western

Australia and Waters and Rivers Commission (WA)

WET Whole effluent toxicity

WHO World Health Organization

WQG Water quality guideline

WWW World Wide Web
Appendix 1 Acronyms and glossary of terms

page A–4 Version — October 2000

Glossary

Term Definition

Abalone/paua Haliotis spp. of shellfish

Abiotic The non-living components of a system (see biota)

Absorption In chemistry: Penetration of one substance into the body of

another

In biology: The act of absorbing (i.e. to take in as fluids or gases

through a cell membrane). To take a substance (e.g. water,

nutrients) into the body through the skin or mucous membranes or,

in plants, through root hairs.

Acceptable Contaminant

Concentration (ACC)

The ACC is the maximum concentration (mg/L) of contaminant in

irrigation water which can be tolerated for a shorter period of time

(20 years) assuming the same maximum annual irrigation loading

as DCC

Acclimation Short-term adaptation of individual organisms to specific

environmental conditions

Acid-soluble metal The concentration of the metal that passes through a 0.45 μm

membrane filter after the sample is acidified to pH 1.5–2.0 with

nitric acid

Acidic Having a high hydrogen ion concentration (low pH)

Acid volatile sulfides (AVS) Sulfides in sediments that liberate hydrogen sulfide on reaction

with cold dilute acid (mainly FeS or MnS in sediments)

Acute toxicity Rapid adverse effect (e.g. death) caused by a substance in a living

organism. Can be used to define either the exposure or the

response to an exposure (effect).

Acute–chronic ratio The species mean acute value divided by the chronic value for the

same species

Additive toxicity The toxicity of a mixture of chemicals that is approximately

equivalent to that expected from a simple summation of the known

toxicities of the individual chemicals present in the mixture (i.e.

algebraic summation of effects).

Adsorption The taking up of one substance at the surface of another

Aeration Any process where a substance becomes permeated with air or

another gas. The term is usually applied to aqueous liquids being

brought into intimate contact with air by spraying, bubbling or

agitating the liquid.

Aerobic Of organisms requiring oxygen for respiration or conditions where

oxygen is available

Aesthetic Aspects of, say, a water body, that can be considered beautiful or

pleasant to the senses

‘Aggressive’ carbon dioxide The amount of dissolved carbon dioxide in excess of that required

to stabilise the bicarbonate ion present in water

Algae Comparatively simple chlorophyll-bearing plants, most of which are

aquatic and microscopic in size

Alkalinity The quantitative capacity of aqueous media to react with hydroxyl

ions. The equivalent sum of the bases that are titratable with

strong acid. Alkalinity is a capacity factor that represents the acidneutralising capacity of an aqueous system.

Alkaloids Nitrogenous organic bases found in plants

Allochthonous Organic material that is developed or derived outside a particular

waterbody

Ambient waters All surrounding waters, generally of largely natural occurrence
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–5

Amphipods Invertebrates belonging to the order Crustacea

Anaerobic Conditions where oxygen is lacking; organisms not requiring

oxygen for respiration

Analytes The physical and chemical species (indicators) to be determined

Anion Negatively charged ion

Anionic Characteristic behaviour or property of an ion that has a negative

charge. Anions move to the anode in electrolysis.

Anode The electrode where oxidation occurs

Antagonism A phenomenon in which the effect or toxicity of a mixture of

chemicals is less than that which would be expected from a simple

summation of the effects or toxicities of the individual chemicals

present in the mixture (i.e. algebraic subtraction of effects)

Anthropogenic Produced or caused by humans

A posteriori Identifying causes by inductive reasoning based on actual effects,

consequences or facts (i.e. from observation, experience or

experiment)

A priori Predicting effects by deductive reasoning based on causes rather

than actual observation, experience or experiment

Aquaculture Commonly termed fish farming, but it broadly refers to the

commercial growing of marine (mariculture) or freshwater animals

and aquatic plants

Aquatic ecosystem Any watery environment from small to large, from pond to ocean,

in which plants and animals interact with the chemical and physical

features of the environment

Aquifer An underground layer of permeable rock, sand or gravel that

absorbs water and allows it free passage through pore spaces

Assessment factors A unitless number applied to the lowest toxicity figure for a

chemical to derive a concentration that should not cause adverse

environmental effects; also called ‘application factor’ or ‘safety

factor’, the size of the AF varies with the type of data (section

8.3.3.2)

Assimilation The incorporation of absorbed substances into cellular material

Assimilative capacity The maximum loading rate of a particular pollutant that can be

tolerated or processed by the receiving environment without

causing significant degradation to the quality of the ecosystem and

hence the environmental values it supports

Ataxia Inability to coordinate voluntary movement

Autochthonous Organic material that is developed or produced within a particular

waterbody

Autotrophy The nutrition of organisms that produce their own organic

constituents from inorganic compounds, using energy from

sunlight or oxidation processes (e.g. most plants and some

bacteria)

Avoidance threshold The lowest concentration of a substance that causes a fish to

actively move away from the source

Barramundi Lates calcarifer

Baseline data (studies) Also called pre-operational data (studies); collected (undertaken)

before a development begins

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

habitats (lakes, rivers, ponds, etc.)

Benthos The sum total of organisms living in, or on, the sediments of

aquatic habitats
Appendix 1 Acronyms and glossary of terms

page A–6 Version — October 2000

Binding sites Sites on a substrate where chemical interaction with an indicator

(qv) may occur. The interaction may be electrostatic, polar,

hydrogen bonding or covalent bonding.

Bioaccumulation General term describing a process by which chemical substances

are accumulated by aquatic organisms from water, either directly

or through consumption of food containing the chemicals

Bioassay A test that exposes living organisms to several levels of a

substance that is under investigation, and evaluates the

organisms’ responses

Bioavailable The fraction of the total of a chemical in the surrounding

environment that can be taken up by organisms. The environment

may include water, sediment, soil, suspended particles, and food

items.

Biochemical (or biological)

oxygen demand

The decrease in oxygen content in mg/L of a sample of water in

the dark at a certain temperature over a certain of period of time

which is brought about by the bacterial breakdown of organic

matter

Usually the decomposition has proceeded so far after 20 days that

no further change occurs. The oxygen demand is measured after 5

days (BOD5), at which time 70% of the final value has usually been

reached.

Bioclogging Clogging of irrigation infrastructure due to excessive algae or

microbial growth

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

Bioconcentration factor (BCF) A unitless value describing the degree to which a chemical can be

concentrated in the tissues of an organism in the aquatic

environment

At apparent equilibrium during the uptake phase of a

bioconcentration test, the BCF is the concentration of a chemical in

one or more tissues of the aquatic organisms divided by the

average exposure concentration in the test.

Biocorrosion Corrosion caused by microorganisms through formation of biofilms

on the metal surface

Biodiversity (biological diversity) The variety of life forms, including the plants, animals and microorganisms, the genes they contain and the ecosystems and

ecological processes of which they are a part

Biofilm Layer of materials created by microorganisms on an underwater

surface

Biological assessment Use and measurement of the biota to monitor and assess the

ecological health of an ecosystem

Biological community An assemblage of organisms characterised by a distinctive

combination of species occupying a common environment and

interacting with one another

Biomagnification 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 residue concentrations increase systematically

from one trophic level to the next.

Biomass The living weight of a plant or animal population, usually

expressed on a unit area basis

Biosolids Sewage sludge, organic residuals remaining after domestic

sewage treatment
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–7

Biota The sum total of the living organisms of any designated area

Biotoxins A toxin (a poison) which originates from a living thing (a plant,

animal, fungi, bacteria, etc.)

Bioturbation The physical disturbance of sediments by burrowing and other

activities of organisms

Bivalve A mollusc with a hinged double shell

Black bream Acanthopagrus butcheri

Black tiger prawn Penaeus monodon

Bloom An unusually large number of organisms per unit of water, usually

algae, made up of one or a few species

Blue mussel Mytilus edulis

Buffer A solution containing a weak acid and its conjugate weak base, the

pH of which changes only slightly on the addition of acid or alkali

Buffering capacity A measure of the relative sensitivity of a solution to pH changes on

addition of acids or base

°C Degrees Celsius

Carcinogen A substance that induces cancer in a living organism

Catchment The total area draining into a river, reservoir, or other body of

water

Cathode The electrode where reduction occurs

Cation Positively charged ion

Cation exchange capacity (CEC) A measure of a soil’s ability to retain cations

Cationic The characteristic behaviour or property of an ion with a positive

charge. Cations move to the cathode in electrolysis.

Chelate The type of co-ordination compound in which a central metal ion is

attached by co-ordinate links to two or more non-metal atoms in

the same molecule, called ligands

Chemical oxygen demand The amount of oxygen required to oxidise all organic matter that is

susceptible to oxidation by a strong chemical oxidant

Chlorination 1) The process of introducing one or more chlorine atoms into a

compound

2) The application of chlorine to water, sewage or industrial wastes

for disinfection

Chronic Lingering or continuing for a long time; often for periods from

several weeks to years. Can be used to define either the exposure

of an aquatic species or its response to an exposure (effect).

Chronic exposure typically includes a biological response of

relatively slow progress and long continuance, often affecting a life

stage.

Chronic value The geometric mean of the lower and upper limits obtained from

an acceptable chronic test or by analysing chronic data using a

regression analysis

A lower chronic limit is the highest tested concentration that did not

cause an unacceptable amount of adverse effect on any of the

specified biological measurements, and below which no tested

concentration caused unacceptable effect

An upper chronic limit is the lowest tested concentration that did

cause an unacceptable amount of adverse effect on one or more

biological measurements and above which all tested

concentrations also caused such an effect

Cladoceran Water flea; zooplankton belonging to the fourth order of the

Branchiopoda, the Cladocera
Appendix 1 Acronyms and glossary of terms

page A–8 Version — October 2000

Colloid Material in solution typically 1 nm–100 nm in diameter. Colloidal

particles do not settle out of solution through the force of gravity.

Organic colloidal matter is considered especially important in the

transport of inorganic substances such as P through the soil

profile.

Community An assemblage of organisms characterised by a distinctive

combination of species occupying a common environment and

interacting with one another

Community composition All the types of taxa present in a community

Community metabolism The biological movement of carbon in an ecosystem, involving two

processes, production (via photosynthesis) and respiration

Community structure All the types of taxa present in a community and their relative

abundances

Complexation The formation of a compound by the union of a metal ion with a

non-metallic ion or molecule called a ligand or complexing agent

Compliance Action in accordance with upholding a ‘standard’ (water quality)

Concentration The quantifiable amount of chemical in, say, water, food or

sediment

Condition indicators or targets Indicators of the condition or state of the ecosystem. They are

normally biological indicators (e.g. species composition, species

abundance), but may also be physical or chemical indicators (e.g.

dissolved oxygen concentration, temperature, flow duration).

These often represent the targets, or water quality objectives, that

need to be met in order to actually achieve the desired level of

ecosystem protection.

Contaminant Biological (e.g. bacterial and viral pathogens) and chemical (see

Toxicants) introductions capable of producing an adverse

response (effect) in a biological system, seriously injuring structure

or function or producing death

Control That part of an experimental procedure which is like the treated

part in every respect except that it is not subjected to the test

conditions. The control is used as a standard of comparison, to

check that the outcome of the experiment is a reflection of the test

conditions and not of some unknown factor.

Corrosion Deterioration of surfaces through erosion processes such as the

conversion of metals to oxides and carbonates

Cresylic Acidic commercial mixture of phenolic materials boiling above the

cresol range (greater than 240°C)

Criteria (water quality) Scientific data evaluated to derive the recommended quality of

water for different uses

Crop quality With regard to inorganic contaminants, increased concentration of

contaminant in plant tissue that while not phytotoxic, reduces the

economic value of the crop due to increased residues

Cumulative Resulting from successive additions at different times or in

different ways

Cumulative Contaminant

Loading Limit (CCL)

The CCL is the maximum contaminant loading in soil defined in

gravimetric units (kg/ha); it indicates the cumulative amount of

contaminant added, above which site-specific risk assessment is

recommended if irrigation and contaminant addition is continued.

Cyanobacteria A division of photosynthetic bacteria, formerly known as bluegreen algae, that can produce strong toxins

Cyanosis A blueness in the appearance of surficial tissues, generally due to

a deficiency of oxygen bound to haemoglobin

Cytotoxic Having an adverse impact on cells

Decision criteria Criteria by which decisions will be made as a result of monitoring

for potential impacts. (See also effect size, Type I error, Type II

error)
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–9

Decision framework or decision

tree

A series of steps for tailoring guideline trigger values to a specific

site or region and for assessing water quality by considering the

local or regional environmental factors that will modify the effect of

the particular water quality parameter

The decision frameworks or trees begin with the simplest steps

and finish with the most difficult and expensive.

Depuration Process that uses a controlled aquatic environment to reduce the

level of pathogenic organisms that may be present in live shellfish

Desirable Contaminant

Concentration (DCC)

The DCC is the maximum concentration (mg/L) of contaminant in

irrigation water which can be tolerated assuming 100 years of

irrigation based on stated irrigation loading assumptions

Detection limit The smallest concentration or amount of a substance that can be

reported as present with a specified degree of certainty by definite

complete analytical procedures

Detritus Unconsolidated sediments composed of both inorganic and dead

and decaying organic material

Dinoflagellates Major class of marine algae that move by flagella. They are often

red in color, and can produce strong toxins that can kill many fish

and other marine organisms.

Direct toxicity assessment (DTA) The use of toxicity tests to determine the acute and/or chronic

toxicity of waste water discharges or total pollutant loads in

receiving waters. (Assesses the toxicity of mixtures of chemicals

rather than individual chemicals.)

Diuresis Increased discharge of urine

Diurnal Daily

Divalent Having a valence (combining power at atomic level) of two (e.g.

calcium, Ca2+)

Dose The quantifiable amount of a material introduced into an animal

Dysphagia Difficulty in swallowing

Early detection Measurable biological, physical or chemical response in relation to

a particular stress, prior to significant adverse affects occurring on

the system of interest.

Early life-stage test 28-day to 32-day exposures (60-day post-hatch for salmonids) of

the early life stages of a species of fish from shortly after

fertilisation through embryonic, larval and early juvenile

development. Data are obtained on survival and growth.

EC50 (median effective

concentration)

The concentration of material in water that is estimated to be

effective in producing some lethal response in 50% of the test

organisms. The LC50 is usually expressed as a time-dependent

value (e.g. 24-hour or 96-hour LC50).

ECse The electrical conductivity of the soil saturation extract

ECs The electrical conductivity of the soil solution at maximum field

water content

EC1:5 The electrical conductivity of a 1:5 soil:water extract

Ecological integrity (health) The ‘health’ or ‘condition’ of an ecosystem

The ability of an ecosystem to support and maintain key ecological

processes and organisms so that their species compositions,

diversity and functional organisations are as comparable as

possible to those occurring in natural habitats within a region

Ecologically sustainable

development

Development that improves the total quality of life, both now and in

the future, in a way that maintains the ecological processes on

which life depends

Ecosystem condition Current or desired status of health of an ecosystem, as affected by

human disturbance

Ecosystem health In this document synonymous with ‘ecological integrity’ (qv)
Appendix 1 Acronyms and glossary of terms

page A–10 Version — October 2000

Ecosystem-specific modifying

factor

A factor that can influence (mostly reduce) the biological effects

caused by a Particular toxicant or stressor

Effect size The size of impact that would cause concern (or constitute an early

warning). Often defined as a level of (ecological) change that is

acceptable in comparison to a defined reference.

Effluent A complex waste material (e.g. liquid industrial discharge or

sewage) that may be discharged into the environment

Electrical conductivity The ability of water or soil solution to conduct an electric current

Encrustation Accumulation of material on surfaces through chemical or

biological processes

End-points Measured attainment response, typically applied to ecotoxicity or

management goals

Endemic, endemism Confined in occurrence to a local region

Enterococci Any streptococcal bacteria normally found in the human intestinal

tract; usually nonpathogenic

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 pollution, waste

discharges and deposits. Several environmental values may be

designated for a specific waterbody.

Epilimnion The uppermost layer of water in a lake, characterised by an

essentially uniform temperature that is generally warmer than

elsewhere in the lake, and by relatively uniform mixing by wind

and wave action

Epilithon Organisms attached to rocks, such as algae and lichens

Epiphyte A plant that grows on the outside of another plant, using it for

support only and not obtaining food from it

ESP The exchangeable sodium content of a soil expressed as a

percentage of the cation exchange capacity

Eukaryotes An organism characterised by the presence of membrane-bound

organelles (see prokaryote)

Euphotic Of surface waters to a depth of approximately 80–100 m; the lit

region that extends virtually from the water surface to the level at

which photosynthesis fails to occur because of reduced light

penetration

Euryhaline Describes organisms that are capable of osmo-regulating over a

wide range of salinities

Eutrophic Abundant in nutrients and having high rates of productivity

frequently resulting in oxygen depletion below the surface layer of

a waterbody

Eutrophication Enrichment of waters with nutrients, primarily phosphorus, causing

abundant aquatic plant growth and often leading to seasonal

deficiencies in dissolved oxygen

Evapotranspiration The combined loss of water from a given area during a specified

period of time by evaporation from the soil or water surface and by

transpiration from plants

Exchangeable sodium

percentage (ESP)

The sodium adsorbed onto clay mineral surfaces as a proportion of

the total cation exchange capacity of those surfaces

Exposure The amount of physical or chemical agent that reaches a target or

receptor

Fate Disposition of a material in various environmental compartments

(e.g. soil or sediment, water, air, biota) as a result of transport,

transformation and degradation

Field capacity The greatest amount of water that it is possible for a soil to hold in

its pore spaces after excess water has drained away
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–11

Flocculation (1) The process by which suspended colloidal or very fine particles

coalesce and agglomerate into well-defined hydrated floccules of

sufficient size to settle rapidly

(2) The stirring of water after coagulant chemicals have been

added to promote the formation of particles that will settle

Flounder Rhombosolea spp.

Flow-through system An exposure system for aquatic toxicity tests in which the test

material solutions and control water flow into and out of test

chambers on a once-through basis either intermittently or

continuously

Fluorosis Chronic poisoning by fluorine

Fouling Accumulation of material through chemical, physical or biological

processes

Free carbon dioxide The amount of dissolved carbon dioxide in excess of that required

to stabilise the bicarbonate ion present in water

Freshwater shrimp A decapod crustacean, including the genus Macrobrachium spp.

Gastropod A mollusc of the Class Gastropoda, with a locomotive organ

placed ventrally (e.g. snail and limpet)

Gilvin The coloured component of dissolved organic matter in water. It is

composed mainly of humic, fulvic and tannic compounds.

Green shell mussel Perna canaliculus

Gross alpha (activity) A measure of the concentration of alpha-particle emitting

radionuclides in water. This is determined by standard techniques

involving the evaporation of a water sample and measurement of

the alpha activity of the residue.

Gross beta (activity) A measure of the concentration of beta-particle emitting

radionuclides in water. This is determined by standard techniques

involving the evaporation of a water sample and measurement of

the beta activity of the residue.

Groundwater Water stored underground in rock crevices and in the pores of

geologic materials that make up the earth's crust; water that

supplies springs and wells

Guideline package Decision trees that are applied to physical and chemical stressors

and/or associated issues for aquatic ecosystems

Guideline trigger values These are the concentrations (or loads) of the key performance

indicators measured for the ecosystem, below which there exists a

low risk that adverse biological (ecological) effects will occur. They

indicate a risk of impact if exceeded and should ‘trigger’ some

action, either further ecosystem specific investigations or

implementation of management/remedial actions.

Guideline (water quality) Numerical concentration limit or narrative statement recommended

to support and maintain a designated water use

Habitat The place where a population (e.g. human, animal, plant,

microorganism) lives and its surroundings, both living and nonliving

Half-life Time required to reduce by one-half the concentration of a material

in a medium (e.g. soil or water) or organism (e.g. fish tissue) by

transport, degradation, transformation or depuration

Hardness The concentration of all metallic cations, except those of the alkali

metals, present in water. In general, hardness is a measure of the

concentration of calcium and magnesium ions in water and is

frequently expressed as mg/L calcium carbonate equivalent.

Hazard The potential or capacity of a known or potential environmental

contaminant to cause adverse ecological effects
Appendix 1 Acronyms and glossary of terms

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Helminth Helminths are worms; the helminths discussed in this document

are human and animal parasites

Hepatotoxin Toxic substances which adversely affect the liver

Heterotrophy The nutrition of plants and animals that are dependent on organic

matter for food

High reliability guideline trigger

values

Trigger values that have a higher degree of confidence because

they are derived from an adequate set of chronic toxicity data

(section 8.3.4) and hence require less extrapolation from the data

to protect ecosystems

Humic substances Organic substances only partially broken down that occur in water

mainly in a colloidal state. Humic acids are large-molecule organic

acids that dissolve in water.

Hydrogeology Study of subsurface waters and with related geologic aspects of

surface water

Hydrograph Graphical representation of river or stream discharge or of waterlevel fluctuations in a well

Hydrolysis (1) The formation of an acid and a base from a salt by the ionic

dissociation of water

(2) The decomposition of organic compounds by interaction with

water

Hydrophilic Having an affinity for water, readily absorbs water

Hydrophobic Having little or no affinity for water, repels or does not absorb

water

Hypolimnion The region of a waterbody that extends from below the thermocline

to the bottom of the lake; it is thus removed from much of the

surface influence

Hypothesis Supposition made from known facts as a starting-point for further

investigation

Hypoxia Deficiency of oxygen in tissues or in blood; anoxia

Incipient LC50 The concentration of a chemical that is lethal to 50% of the test

organisms as a result of exposure for periods sufficiently long that

acute lethal action has essentially ceased. The asymptote (part of

the toxicity curve parallel to the time axis) of the toxicity curve

indicates the value of the incipient LC50 approximately.

Indicator A parameter that can be used to provide a measure of the quality

of water or the condition of an ecosystem

Ingestion The swallowing or taking in of food material

Inorganic carbon Generally, simple ions and molecules that contain carbon bonded

only to inorganic atoms. Carbonates are the most common group,

although the cyanide ion is also considered to be inorganic.

Interstitial Occurring in interstices or spaces; applied to water and to flora and

fauna living between sand grains and soil particles

Invertebrates Animals lacking a dorsal column of vertebrae or a notochord

In vitro Outside the intact organism; generally applied to experiments

involving biochemical events occurring in tissue fragments or

fractions in a laboratory

Ion An electrically charged atom

Kuruma prawn Penaeus japonicus

Langelier Saturation Index (SI) An index based on the tendency of water to deposit or dissolve

calcium carbonate. It relates the actual pH of water with the pH at

which water is saturated with calcium carbonate (SI = pH - pHs).

LC100 Lowest concentration of a toxicant that kills all the test organisms
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–13

LC50 (median lethal

concentration)

The concentration of material in water that is estimated to be lethal

to 50% of the test organisms. The LC50 is usually expressed as a

time-dependent value, e.g. 24-hour or 96-hour LC50, the

concentration estimated to be lethal to 50% of the test organisms

after 24 or 96 hours of exposure.

LD50 (median lethal dose) The dose of material that is estimated to be lethal to 50% of the

test organisms. Appropriate for use with test animals such as rats,

mice and dogs, it is rarely applicable to aquatic organisms

because it indicates the quantity of a material introduced directly

into the body by injection or ingestion rather than the concentration

of the material in water in which aquatic organisms are exposed

during toxicity tests.

Leachate Water that has passed through a soil and that contains soluble

material removed from that soil

Leaching Where referred to in the salinity and sodicity section of Chapter 4,

the downward movement of water and solutes below the root zone

Leaching fraction (LF) The proportion of water applied to the surface of a soil (e.g. as

irrigation or rainfall) that drains below the root zone in the soil

profile

Lentic Standing body of water such as a lake or pond

Lethal Causing death by direct action. Death of aquatic organisms is the

cessation of all visible signs of biological activity.

Level of protection A level of quality desired by stakeholders and implied by the selected

management goals and water quality objectives for the water

resource

Life-cycle study A chronic (or full chronic) study in which all the significant life

stages of an organism are exposed to a test material. Generally, a

life-cycle test involves an entire reproductive cycle of the organism.

A partial life-cycle toxicity test includes the part of the life cycle

observed to be especially sensitive to chemical exposure.

Ligand A molecule, ion or atom that is attached to the central atom of a

co-ordination compound, a chelate or other complex. May also be

called complexing agent.

Liveweight Weight of the living animal

LOEC (Lowest observed effect

concentration)

The lowest concentration of a material used in a toxicity test that

has a statistically significant adverse effect on the exposed

population of test organisms as compared with the controls. When

derived from a life-cycle or partial life-cycle test, it is numerically

the same as the upper limit of the MATC.

LOEL (Lowest observed effect

level)

The lowest concentration that produces an observable effect in a

test species. Below this concentration there are no observed

effects in the test species.

Long-term trigger value (LTV) The maximum concentration of contaminant in irrigation water

which can be tolerated assuming 100 years of irrigation, based on

key irrigation loading assumptions

Lotic Flowing waters (e.g. rivers and streams)

Low reliability guideline trigger

values

Trigger values that have a low degree of confidence because they

are derived from an incomplete data set (section 8.3.4.1). They

are derived using either assessment factors or from modelled data

using the statistical method. They should only be used as interim

indicative working levels.

Macrophyte A member of the macroscopic plant life of an area, especially of a

body of water; large aquatic plant
Appendix 1 Acronyms and glossary of terms

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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.

Management goals will mostly be narrative statements focusing

management on the relevant water quality objectives.

Marron Cherax tenuimanus

MATC (Maximum acceptable

toxicant concentration)

The maximum concentration of a toxic substance that a receiving

water may contain without causing significant harm to its

productivity or uses as determined by chronic toxicity tests

Maximum tolerable daily level

(MTDL)

The dietary level that when fed for a limited period, will not impair

animal performance and should not produce unsafe residues in

produce for human consumption

Median Middle value in a sequence of numbers

Mesotrophic Water bodies or organisms which are intermediate between

nutrient-rich and nutrient-poor

Metabolite Any product of metabolism

Methylation The introduction of methyl (CH3) groups into organic and inorganic

compounds

Methyl mercury The most common form is the cation CH3Hg+ although (CH3)2Hg

also occurs. Both are extremely potent toxicants and can lead to

secondary poisoning through biomagnification. They are usually

formed in anoxic sediments.

Mixing zones An explicitly defined area around an effluent discharge where

effluent concentrations may exceed guideline values and therefore

result in certain environmental values not being protected. The

size of the mixing zone is site specific.

Moderate reliability guideline

trigger values

Trigger values that have a moderate degree of confidence

because they are derived from an adequate set of acute toxicity

data (section 8.3.4) and hence require more extrapolation than

high reliability trigger values, including an acute-to-chronic

conversion

Monomeric A chemical compound comprising single molecules

Morphometry The form, shape and dimensions of an entity, e.g. waterbody or

animal

Multiple lines of evidence Weight of the evidence based on different types of information

from a variety of different sources and studies

Munsell Scale A means of expressing the colour of a soil by matching it against a

colour chart

Necrotic Localised dying tissue

Neurotoxin Toxic substances which adversely affect the nervous system

NOEC (No observed effect

concentration)

The highest concentration of a toxicant at which no statistically

significant effect is observable, compared to the controls; the

statistical significance is measured at the 95% confidence level

Not detectable Below the limit of detection of a specified method of analysis

Nutrient solution Plant growth medium providing all essential elements for plant

growth in the absence of soil or other support media. Also referred

to as solution culture.

Octanol:water partition

coefficient (Pow)

The ratio of a chemical's solubilities in n-octanol and water at

equilibrium. The logarithm of Pow is used as an indication of a

chemical's propensity for bioconcentration by aquatic organisms.

Off-flavour Result of the accumulation of certain pollutants such as petroleum

hydrocarbons in fish or shellfish to a level that affects the flavour,

making the product undesirable for human consumption; also

known as tainting
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–15

Oligotrophic Waters with a small supply of nutrients

Organic carbon Generally carbon which is chemically bonded to other carbon

atoms, although methane (one carbon atom only) and its

derivatives are considered organic

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

processes

Osmoregulation The biological process of maintaining the proper salt concentration

in body tissues to support life

Osmosis Diffusion of a solvent through a semi-permeable membrane into a

more concentrated solution, tending to equalise the concentrations

on both sides of the membrane

Oxidation The combination of oxygen with a substance, or the removal of

hydrogen from it or, more generally, any reaction in which an atom

loses electrons

Oxygenation The process of adding dissolved oxygen to a solution

Pacific oyster Crassostrea gigas

PAH Polycyclic aromatic hydrocarbons

Parameter A measurable or quantifiable characteristic or feature

Partition coefficient A ratio of the equilibrium concentration of the chemical between a

non-polar and polar solvent

Pathogen An organism capable of eliciting disease symptoms in another

organism

Pelagic Term applied to organisms of the plankton and nekton which

inhabit the open water of a sea or lake

Percentile Division of a frequency distribution into one hundredths

Performance indicators These are the indicators used to assess the risk that a particular

issue will occur (they are used in the guideline packages to

compare against the trigger values). They are generally median (or

mean) concentrations in the ambient water, and may be stressor

and/or condition indicators.

Periphyton The organisms attached to submerged plants

Pesticide A substance or mixture of substances used to kill unwanted

species of plants or animals

pH Value that represents the acidity or alkalinity of an aqueous

solution. It is defined as the negative logarithm of the hydrogen ion

concentration of the solution.

pH (CaCl2) Measurement of soil pH in a 1:2.5 solution of soil:0.01M CaCl2 .

The CaCl2 solution is used because it has an ionic strength similar

to that of soil water.

Phenols Phenol is a benzene ring with one -OH radical replacing hydrogen.

Phenols are compounds which contain additional chemical groups

bound to this basic structure (each replacing hydrogen).

Photodegradation Breakdown of a substance by exposure to light; the process

whereby ultra-violet radiation in sunlight attacks a chemical bond

or link in a chemical structure

Photolysis The decomposition of a compound into simpler units as a result of

the absorption of one or more quanta of radiation

Photosynthesis The conversion of carbon dioxide to carbohydrates in the presence

of chlorophyll using light energy

Physico-chemical Refers to the physical (e.g. temperature, electrical conductivity)

and chemical (e.g. concentrations of nitrate, mercury)

characteristics of water

Physiology The study of the functioning of organisms and their parts

Phytoplankton Small (often microscopic) aquatic plants suspended in water
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Phytotoxicity Toxicity of contaminants to plants

Pilot program A field investigation similar in design to a sampling program, but

less ambitious in scope. It is used to assess preliminary indicator

values, spatial and temporal variability and logistic issues before

definitive sampling.

Plankton Plants (phytoplankton) and animals (zooplankton), usually

microscopic, floating in aquatic systems

Pollution The introduction of unwanted components into waters, air or soil,

usually as result of human activity; e.g. hot water in rivers, sewage

in the sea, oil on land

Polychlorinated biphenyls These are highly toxic and persistent compounds derived from the

replacement by Cl radicals of numerous H radicals on biphenyl,

which consists of two benzene rings joined by a covalent bond,

with the elimination of two H radicals (C12H10).

Potable water Water suitable, on the basis of both health and aesthetic

considerations, for drinking or culinary purposes

Practical Quantitation Limit

(PQL)

The Practical Quantitation Limit (PQL) is the lowest level

achievable among laboratories within specified limits during routine

laboratory operations. The PQL represents a practical and

routinely achievable detection level with a relatively good certainty

that any reported value is reliable (Clesceri et al. 1998). The PQL

is often around 5 times the method detection limit.

Precipitation (1) The formation of solid particles in a solution; generally, the

settling out of small particles

(2) The settling-out of water from cloud, in the form of rain, hail,

snow, etc.

Primary production The production of organic matter from inorganic materials

Producers Organisms that are able to build up their body substance from

inorganic materials

Prokaryotes Organisms characterised by the absence of membrane-bound

organelles (opposite to eukaryotes)

Prolarvae Newly hatched larvae during the first few days when they feed on

their supply of embryonic yolk

Protocol A formally agreed method and procedure for measuring an

indicator; it defines the sampling, sample handling procedures and

sample analysis

Protozoans Single-celled, animal-like organisms of the kingdom Protista

Quality assurance (QA) The implementation of checks on the success of quality control

(e.g. replicate samples, analysis of samples of known

concentration)

Quality control (QC) The implementation of procedures to maximise the integrity of

monitoring data (e.g. cleaning procedures, contamination

avoidance, sample preservation methods)

Rainbow trout Oncorhynchus mykiss

Radiological Pertaining to nuclear radiation

Rapid biological assessment A form of biological assessment, best developed using stream

macroinvertebrate communities, that uses standardised, costeffective protocols to provide rapid sample processing, data

analysis, reporting and management response. The results from

such assessments are likely to be reliable to detect large impacts

but not small or minor impacts.

Recruitment In these Guidelines, the replenishment or addition of individuals of

an animal or plant population through reproduction, dispersion and

migration

Red claw Cherax quadricarinatus
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Redox potential An expression of the oxidising or reducing power of a solution

relative to a reference potential. This potential is dependent on the

nature of the substances dissolved in the water, as well as on the

proportion of their oxidised and reduced components.

Reference condition An environmental quality or condition that is defined from as many

similar systems as possible 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

Relaying Transfer of shellfish from restricted areas and conditional approved

areas (when closed due to harvesting criteria being exceeded) to

approved or conditional approved areas for natural biological

cleansing using the ambient environment as a treatment system

Risk A statistical concept defined as the expected likelihood or

probability of undesirable effects resulting from a specified

exposure to known or potential environmental concentrations of a

material. 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.

Ryznar (Stability) index Index relating the pH of water (pH) to the pH of water just

saturated with calcium carborate (pHs)

Salinity The presence of soluble salts in or on soils or in water

Sediment Unconsolidated mineral and organic particulate material that

settles to the bottom of aquatic environment

Sediment pore waters Water that occupies the space between particles in a sediment, as

distinct from overlying water which is the water above the sediment

layer

Sewage fungus A thick filamentous growth that develops in water contaminated

with sewage. The filamentous material is composed predominately

of the bacterium Sphaerotilus natans.

Short-term trigger value (STV) The maximum concentration of contaminant in irrigation water

which can be tolerated for a shorter period of time (20 years)

assuming the same maximum annual irrigation loading to soil as

for the long-term trigger value (qv)

Silver perch Bidyanus bidyanus

Simultaneously extracted metals The sum of the molar concentrations of heavy metals (excluding

iron and manganese) that are solubilised with cold dilute acid

(usually measured simultaneously with the measurement of AVS).

Snapper Pagrus auratus

Sodicity The presence of a high proportion of sodium ions relative to other

cations in a soil

Sodium adsorption ratio (SAR) The concentration of sodium relative to calcium and magnesium in

the soil solution

Solution concentration Concentration of solutes in the soil water phase. The solutes,

which may be contaminants, in the soil water are generally

regarded as being highly available to organisms.

Sorption Process whereby contaminants in soils adhere to the inorganic and

organic soil particles

Speciation The intimate chemical environment of the indicator (qv), that is the

compound or ion of which it forms a part

Species 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 produce viable offspring if bred with

members of another group
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Species richness The number of species present (generally applied to a sample or

community)

SPM — suspended particulate

matter

This is insoluble material which resides in the water column, or is

dispersed in a sample upon agitation

Stakeholder A person or group (e.g. an industry, a government jurisdiction, a

community group, the public, etc.) who have an interest or concern

in something

Standard (water quality) An objective that is recognised in enforceable environmental

control laws of a level of government

Standing crop The weight of organic material that can be sampled or harvested

by normal methods at any one time from a given area

Static system An exposure system of aquatic toxicity tests in which the test

chambers contain solutions of the test material or control water

that are not usually changed during the test. Depending upon

conditions, a static system may or may not be in equilibrium.

Steady state or dynamic

equilibrium

The state at which the competing rates of uptake and elimination of

a chemical within an organism or tissue are equal. An apparent

steady state is reached when the concentration of a chemical in

tissue remains essentially constant during a continuous exposure.

Stressors The physical, chemical or biological factors that can cause an

adverse effect in an aquatic ecosystem as measured by the

condition indicators (see Section 3.3.2)

Sub-lethal Involving a stimulus below the level that causes death

Supersaturation Refers to a solution containing more solute than equilibrium

conditions will allow

Survival time The time interval between death and the initial exposure of an

aquatic organism to a harmful substance

Suspension A system in which very small particles (solid, semi-solid, or liquid)

are more or less uniformly dispersed in a liquid or gaseous

medium.

If the particles are small enough to pass through filter membranes,

the system is termed a colloidal suspension. If the particles are of

larger than colloidal dimensions they will tend to precipitate, if

heavier than the suspending medium, or to agglomerate and rise

to the surface, if lighter.

Sydney rock oyster Saccostrea commercialis

Synergism A phenomenon in which the effect or toxicity of a mixture of

chemicals is greater than that to be expected from a simple

summation of the effects or toxicities of the individual chemicals

present in the mixture

Tainting See ‘Off-flavour’

Taxa richness Number of taxa present

Taxon (Taxa) Any group of organisms considered to be sufficiently distinct from

other such groups to be treated as a separate unit (e.g. species,

genera, families)

Taxonomic (group, resolution) An organism’s location in the biological classification system used

to identify and group organisms with similar physical, chemical

and/or structural composition.

Teratogen An agent that increases the incidence of congenital malformations

Thermodynamic equilibrium Property of a system which is in mechanical, chemical and thermal

equilibrium

Thermotolerant coliform Also known as faecal coliforms. In tropical and sub-tropical areas,

thermotolerant coliforms may on some occasions include

microorganisms of environmental rather than faecal origin.
Appendix 1 Acronyms and glossary of terms

Version — October 2000 page A–19

Threshold concentration A concentration above which some effect (or response) will be

produced and below which it will not

Tolerance The ability of an organism to withstand adverse or other

environmental conditions for an indefinitely long exposure without

dying

Total dissolved solids (TDS) A measure of the inorganic salts (and organic compounds)

dissolved in water

Total metal The concentration of a metal in an unfiltered sample that is

digested in strong nitric acid

Toxicant A chemical capable of producing an adverse response (effect) in a

biological system at concentrations that might be encountered in

the environment, seriously injuring structure or function or

producing death. Examples include pesticides, heavy metals and

biotoxins (i.e. domoic acid, ciguatoxin and saxitoxins).

Toxicity The inherent potential or capacity of a material to cause adverse

effects in a living organism

Toxicity identification and

evaluation (TIE)

Toxicity characterisation procedures involving use of selective

chemical manipulations or separations and analyses coupled with

toxicity testing to identify specific classes of chemicals and

ultimately individual chemicals that are responsible for the toxicity

observed in a particular sample

Toxicity test The means by which the toxicity of a chemical or other test

material is determined. A toxicity test is used to measure the

degree of response produced by exposure to a specific level of

stimulus (or concentration of chemical).

Trigger values These are the concentrations (or loads) of the key performance

indicators measured for the ecosystem, below which there exists a

low risk that adverse biological (ecological) effects will occur. They

indicate a risk of impact if exceeded and should ‘trigger’ some

action, either further ecosystem specific investigations or

implementation of management/remedial actions.

Trochus Trochus niloticus

True colour The colour of water resulting from substances that are totally in

solution; not to be mistaken for apparent colour resulting from

colloidal or suspended matter

Turbulence Unorganised movement in liquids and gases resulting from eddy

formation

Type I error Probability of concluding that an impact has occurred when, in fact,

an impact has not occurred

Type II error Probability of concluding that an impact has not occurred when, in

fact, an impact has occurred

Univariate Statistical analysis concerned with data collected on one

dimension of the same organism

Uptake A process by which materials are absorbed and incorporated into a

living organism

Value judgements A decision involving basic issues of fairness, reasonableness,

justice, or morality

Volatile Having a low boiling or subliming pressure (a high vapour

pressure)

Water quality criteria Scientific data evaluated to derive the recommended quality of

water for various uses

Water quality guideline See ‘Guideline (water quality)’

Water quality objective A numerical concentration limit or narrative statement that has

been established to support and protect the designated uses of

water at a specified site. It is based on scientific criteria or water

quality guidelines but may be modified by other inputs such as

social or political constraints.
Appendix 1 Acronyms and glossary of terms

page A–20 Version — October 2000

Watertable The level of groundwater; the upper surface of the zone of

saturation for underground water.

Whiting Sillago spp. of marine fish

Whole effluent toxicity testing The use of toxicity tests to determine the acute and/or chronic

toxicity of effluents

Xenobiotic A foreign chemical or material not produced in nature and not

normally considered a constituent of a specified biological system.

This term is usually applied to manufactured chemicals.

Yabby Cherax destructor

Zooplankton The animal portion of the plankton
Version — October 2000 page A–21

Appendix 2 The National Water Quality Management Strategy

The Australian and New Zealand Environment and Conservation Council

(ANZECC) and the Agriculture and Resource Management Council of Australia

and New Zealand (ARMCANZ) are working together to develop a National Water

Quality Management Strategy (NWQMS).

The guiding principles for the Strategy are set out in National Water Quality

Management Strategy: Policies and Principles — A Reference Document

(NWQMS Paper 2, ANZECC & ARMCANZ, 1994) which emphasises the

importance of:

• ecologically sustainable development

• integrated (or total) catchment management

• best management practices, including the use of acceptable modern technology,

and waste minimisation and utilisation

• the role of economic measures, including user pays and polluter pays.

The process of implementing the National Water Quality Management Strategy

involves the community working in concert with government in setting and

achieving local environmental values, which are designed to maintain good water

quality and to progressively improve poor water quality. It involves development

of a plan for each catchment and aquifer, which takes account of all existing and

proposed activities and developments, and which contains the agreed

environmental values and feasible management options.

Figure A1 National Water Quality Management Strategy
Appendix 2 The National Water Quality Management Strategy

page A–22 Version — October 2000

Documents of the National Water Quality

Management Strategy

Paper No. Title

Policies and Process for Water Quality Management

1 Water Quality Management — An Outline of the Policies

2 Policies and Principles — A Reference Document

3 Implementation Guidelines

Water Quality Benchmarks

4 Australian and New Zealand Water Quality Guidelines for Fresh and

Marine Waters

5 Australian Drinking Water Guidelines — Summary

6 Australian Drinking Water Guidelines

7 Australian Guidelines for Water Quality Monitoring and Reporting

Groundwater Management

8 Guidelines for Groundwater Protection

Guidelines for Diffuse and Point Sources*

9 Rural Land Uses and Water Quality

10 Guidelines for Urban Stormwater Management

11 Guidelines for Sewerage Systems — Effluent Management

12 Guidelines for Sewerage Systems — Acceptance of Trade Waste

(Industrial Waste)

13 Guidelines for Sewerage Systems — Sludge (Biosolids) Management

14 Guidelines for Sewerage Systems — Use of Reclaimed Water

15 Guidelines for Sewerage Systems — Sewerage System Overflows

16a Effluent Management Guidelines for Dairy Sheds

16b Effluent Management Guidelines for Dairy Processing Plants

17 Effluent Management Guidelines for Intensive Piggeries

18 Effluent Management Guidelines for Aqueous Wool Scouring

and Carbonising

19 Effluent Management Guidelines for Tanning and Related Industries in

Australia

20 Effluent Management Guidelines for Australian Wineries and Distilleries

* The guidelines for diffuse and point sources are national guidelines which aim to ensure high levels of environmental protection that

are broadly consistent across Australia.
Version — October 2000 page A–23

Appendix 3 Recent water quality documents of the NZ Ministry for

the Environment

• Flow Guidelines for Instream Values (NZ Ministry for the Environment 1995)

• New Zealand Drinking Water Guidelines (NZ Ministry of Health 1995).

• Water Quality Guidelines No. 1: Biological Growths (NZ Ministry for the

Environment 1992)

• Water Quality Guidelines No. 2: Colour and Clarity (NZ Ministry for the

Environment 1994)

• Periphyton Guidelines (NZ Ministry for the Environment, in press)

• Recreational Water Quality Guidelines (NZ Ministry for the Environment

1999)

• Monitoring the Trophic Status of New Zealand’s Lakes (NZ Ministry for the

Environment, in press)

• Managing Waterways on Farms (NZ Ministry for the Environment, in press)

• A discussion on reasonable mixing in water quality management, Resource

Management Ideas No. 10 (NZ Ministry for the Environment 1994)

• Reducing the Impacts of Agricultural Runoff on Water Quality: A discussion of

policy approaches (NZ Ministry for the Environment 1997)

Version — October 2000 page A–25

Appendix 4 Development of the revised guidelines

This Appendix outlines the revision process for the Guidelines, and the various

instrumentalities involved.

The revision program

In March 1993 the ANZECC Standing Committee on Environmental Protection

(SCEP) made a decision to review the Australian Water Quality Guidelines for

Fresh and Marine Waters, periodically, with the following two primary objectives:

• to incorporate current scientific, international and national information which is

appropriate to Australian and New Zealand conditions; and

• to produce a document that is sufficiently clear and understandable for the

relevant authorities to use for consultation.

The Environmental Research Institute of the Supervising Scientist (eriss) was

given the responsibility for managing the review of the Guidelines on behalf of

ANZECC.

The review strategy was endorsed by the ANZECC SCEP at its meeting on 15 May

1996 which involved key government and non-government groups in recognition of

the need for an open and transparent process with broad community consultation. As

part of the strategy, a project committee was established and given very broad

representation to oversee and facilitate the process. Membership included state

environment agencies, a number of agencies and representatives of the Agricultural

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

and representatives of the National Health and Medical Research Council

(NHMRC), the National Farmers Federation, the Australian Seafood Industry

Council, and peak conservation organisations (full membership listed below).

The outcomes from a preliminary workshop, together with the issues raised in early

public submissions on the Guidelines, were used to assist in identifying and

defining the scope of the tasks necessary for reviewing the 1992 report. Expert

groups from Australia and New Zealand were then commissioned to review the

technical aspects of the report and a draft of the revised guidelines was compiled

by eriss.

After consideration by the project committee, the draft document was referred

firstly to the ARMCANZ/ANZECC Contact Group (listed below), and then to the

ARMCANZ subcommittee — the Sustainable Land & Water Resource

Management Committee (SLWRMC) — for feedback and subsequent endorsement

prior to its release. The Guidelines were released for a three-month public

comment period in July 1999 by the Australian Commonwealth Minister for the

Environment & Heritage, without the endorsement of SCEP and the ARMCANZ

Standing Committee for Agricultural and Resource Management (SCARM).

eriss received and collated 96 public submissions. On the basis of these

comments, the Guidelines were redrafted in close consultation with the Contact

Group and its working parties (listed below). The revised Guidelines were

endorsed by the Contact Group in May 2000 and by SCEP in June 2000, with the

final document then referred to ANZECC who approved the Guidelines for

publication under the National Water Quality Management Strategy in July 2000.
Appendix 4 Development of the revised guidelines

page A–26 Version — October 2000

The Project Committee

Name Organisation Name Organisation

Graeme Batley ARMCANZ — CSIRO Energy

Technology

Kevin McAlpine

(Secretary)

Com. Environmental Research Institute

of the Supervising Scientist

John Chapman NSW Environmental Protection

Authority — Centre for Ecotoxicology

Bill Maher University of Canberra

John Cugley SA Department of Environment and

Natural Resources

Scott Markich Australian Nuclear Science and

Technology Organisation

Lisa Dixon Victorian Environmental Protection

Authority

Greg Miller peak conservation organisations

Barry Hart Monash University — Water Studies

Centre

Andrew Moss Queensland Department of Environment

Chris Humphrey Com. Environmental Research

Institute of the Supervising Scientist

Barry Noller Northern Territory Department of Mines

and Energy

Heather Hunter Queensland Department of Natural

Resources

Eric Pyle New Zealand Ministry for the

Environment

Arthur Johnston

(Chairman)

Com. Environmental Research

Institute of the Supervising Scientist

Nigel Scullion Australian Seafood Industry Council

Warren Jones Tasmanian Department of

Environment and Land Management

Graham Skyring ARMCANZ — Skyring Environment

Enterprises

David Klessa Com. Environmental Research

Institute of the Supervising Scientist

Victor Talbot WA Department of Environmental

Protection

Mike Lawton Northern Territory Department of Land

Planning and Environment

Alan Thomas Com. Environment Protection Group

Chris leGras Com. Environmental Research

Institute of the Supervising Scientist

Pam Waudby National Farmers Federation

Richard Lugg National Health and Medical Research

Council

Rosalyn

Vulcano

Northern Territory Power and Water

Authority

Contributing members of the Contact Group and/or proxies (period 1996–2000)

Name Organisation Name Organisation

Paul Bainton &

Alan Thomas

Com. Environment Australia,

Environment Protection Group

Rachel

Gregson, Ross

Dalton, Dennis

Alyliffe, David

Lambert,

Stephen Clark

Agriculture, Fisheries and Forestry —

Australia

Barbara

Richardson &

Carolyn Davies

NSW Environment Protection Authority Bruce Cooper New South Wales Department of Land

and Water Conservation

Chris Bell Victorian Environment Protection

Authority

Anne Woolley &

Peter Thompson

Queensland Department of Natural

Resources

John Cugley SA Environment Protection Authority Peter Scott Melbourne Water Corporation

Stephen Fisher

(EPA) & Ian

Eskdale (DE)

Queensland Environment Protection

Agency & Queensland Department of

Environment

Alan Maus &

Barry Sanders

Water Corporation of Western Australia

Victor Talbot Western Australian Department of

Environmental Protection

Michael Lawton Northern Territory Department of Lands,

Planning & Environment

Greg Dowson &

Warren Jones

Tasmanian Department of Environment

and Land Management

Robert Neil Environment ACT

Bob Zuur & Eric

Pyle

New Zealand Ministry of the

Environment

Philip Callan National Health and Medical Research

Council
Appendix 4 Development of the revised guidelines

Version — October 2000 page A–27

Working Parties to the Contact Group

1. Aquaculture: David Cunliffe, Pauline Semple, Victor Talbot, Rob Cordover,

Christine Cowie, Kerry Jackson and Michelle Burford

2. Agriculture: Liz Rogers, Greg Dowson, Karen Benn and Karina Watkins

3. Physico-chemical stressors: Klaus Koop, Greg Dowson, John Cugley, Peter

Scott, Andrew Moss and Bob Humphries

4. Toxicants and sediments: Peter Thompson, Bob Humphries, John Cugley,

Bruce Cooper, Munro Mortimer, Victor Talbot, Karina Watkins and Gus

Fabris

Appendix 5 Basis of the proposed guidelines for recreational water quality and aesthetics in Australia

Version — October 2000 page A–29

Appendix 5 Basis of the proposed guidelines for recreational

water quality and aesthetics in Australia

The draft World Health Organization (WHO) Guidelines for Safe Recreationalwater Environments: Coastal and Fresh-waters (WHO 1998), in referring to the

different types of recreational usage of water, give the following examples:

• no contact, where enjoyment is of aesthetic beauty of the water environment;

• limited contact, e.g. boating, rowing, fishing;

• meaningful direct contact that involves a negligible risk of swallowing water,

e.g. wading;

• extensive direct contact with full body immersion and a meaningful risk of

swallowing water, e.g. swimming.

The WHO Health-based Monitoring of Recreational Waters: The Feasibility of a

New Approach (The ‘Annapolis’ Protocol) (WHO 1999) considers the adequacy

and effectiveness of present approaches to the monitoring and assessment of

recreational water, particularly where the monitoring is linked to the effective

management of microbiological hazards in coastal and freshwater areas.

A number of types of hazards that can be encountered in recreational water are

dealt with in the WHO Guidelines; they include:

• poisoning and toxicoses, including stings of poisonous and venomous animals,

ingestion or inhalation of, or contact with, chemically contaminated water or

blooms of toxic cyanobacteria or dinoflagellates;

• physiological effects, including chilling, thermal shock;

• exposure to pathogenic bacteria, viruses, fungi or parasites;

• aesthetic quality including visual clarity, colour, odour, surface scum.

The WHO Guidelines also include guidance on assessment and control measures,

public health advice and intervention requirements when guideline values are

exceeded.

References

WHO 1998. Guidelines for safe recreational-water environments: Coastal and fresh-waters. Draft for

Consultation, EOS/DRAFT/98.14, World Health Organization, Geneva.

WHO 1999. Health-based monitoring of recreational waters: The feasibility of a new approach (The

‘Annapolis Protocol’), WHO/SDEW/WSH/99.1, World Health Organization, Geneva.

Version — October 2000 index 1

Index

Note: Page entries beginning with ‘A’ refer to appendices

a posteriori 3.2–19, 7.2–4, A–5

a priori 3.2–13, 3.2–14, A–5

abalone 4.4–4, A–4

abiotic factors 3.1–2, A–4

absorption A–4

absorptive capacity 2–18

acceptable contaminant concentration (ACC)

A–4

acceptable ecological change, measurement

3.1–20 to 3.1–23, 3.3–19 to 3.3–20, 7.2–15

to 7.2–16

acclimation A–4

acid volatile sulfides (AVS) 3.5–8, A–4

acid-soluble metal A–4

acidic A–4

acronyms A–1 to A–3

acute-chronic ratio A–4

acute toxicity 3.4–2, A–4

additive toxicity A–4

aeration 4.2–5, A–4

aerobic A–4

aesthetic guideline value 6–4

aesthetics A–4

see also recreational water quality and

aesthetics

‘aggressive’ carbon dioxide A–4

agricultural chemical preparation 4.2–16

Agriculture and Resource Management

Council of Australia and New Zealand

(ARMCANZ) A–21

Agriculture and Resource Management

Council of Australia and New Zealand

(ARMCANZ/ANZECC Contact Group

A–25, A–26

Working Parties A–27

algae A–4

in irrigation waters 4.2–2 to 4.2–3

algal biomass 3.3–20

algal blooms see cyanobacteria

alkalinity A–4

alkaloids A–4

allochthonous A–4

ambient waters 2–17, A–4

ammonia 3.4–3

ammonium 3.3–10, 3.3–12, 3.3–14, 3.3–16,

3.3–17

amphipods A–5

anaerobic A–5

analytes A–5

animal pathogens, in irrigation waters 4.2–3 to

4.2–4

anion A–5

anionic A–5

anode A–5

antagonism A–5

anthropogenic A–5

aquaculture A–5

aquaculture guidelines 4.1–1, 4.4–1

for protection of cultured fish, molluscs

and crustaceans 4.4–3 to 4.4–12

philosophy 4.4–2

physico-chemical stressors 4.4–7, 4.4–9 to

4.4–10

precautionary comments 4.4–18 to 4.4–19

research and development priorities 4.4–19

scope 4.4–2 to 4.4–3

toxicants 4.4–8, 4.4–10 to 4.4–11

aquatic ecosystems 1–3 to 1–4, A–5

biological assessment 2–16

classification 3.1–7 to 3.1–9

features affecting water quality assessment

and ecosystem protection 3.1–7

philosophy and steps to applying the

guidelines 3.1–1 to 3.1–7

aquatic plants, nuisance growth of 3.3–22 to

3.3–25

aquifer A–5

area classification approach 4.4–17 to 4.4–18

assessment factors 3.4–2, A–5

assessment objectives, for ecosystem

protection 3.2–4 to 3.2–6

assimilation A–5

assimilative capacity 2–18, A–5

ataxia A–5

AUSRIVAS rapid biological assessment

3.2–4, 3.2–8, 3.2–20, 7.2–4, 7.3–3 to 7.3–5

applications and cautions 7.3–4 to 7.3–5

cautionary note 3.2–9

division of O/E indices into bands 3.2–15

to 3.2–17

outline 7.3–3

sampling protocol and issues about effect

size and sensitivity 7.3–3 to 7.3–4

Australia, ecosystem features 3.1–7, 3.1–8

Australian and New Zealand Environment and

Conservation Council (ANZECC) A–21

Standing Committee on Environmental

Protection (SCEP) A–25
Index

index 2 Version — October 2000

Australian National Water Quality

Management Strategy see National Water

Quality Management Strategy

autochthonous A–5

autotrophy A–5

avoidance threshold A–5

BACI (Before-After Control-Impact) family of

designs 3.2–12, 7.2–5, 7.2–7, 7.2–3 to 7.2–4

background data 3.1–14

bacterial infections, impact on human health

4.4–13 to 4.4–14

barramundi A–5

baseline data collection 3.2–10, 3.2–17 to

3.2–18

baseline data (studies) 7.2–4, A–5

benthos A–5

bicarbonate, in irrigation waters 4.2–9

binding sites A–6

bioaccumulation 3.4–17, A–6

bioassay A–6

bioassessment see biological assessment

bioavailable fraction 3.1–22, A–6

biochemical oxygen demand 3.3–8, 4.4–7,

A–6

bioclogging A–6

bioconcentration A–6

bioconcentration factor (BCF) 4.4–17, A–6

biocorrosion A–6

biodiversity 3.1–3, A–6

biodiversity assessment 3.2–5, 3.2–7, 3.2–8 to

3.2–9

biodiversity assessment objective 3.2–8

biodiversity indicators 3.2–5, 3.2–8 to 3.2–9,

7.2–1 to 7.2–2

bioequivalence testing, for environmental

restoration 7.2–13

biofilm A–6

biological assessment 3.2–1, A–6

definition 3.2–1

framework 3.2–2 to 3.2–3

philosophy and approach behind

bioindicators 3.2–1 to 3.2–2

biological assessment objectives, for

ecosystem protection 3.2–4

biological community A–6

biological contaminants

effect on aquaculture species 4.4–11 to

4.4–12

impact on human health 4.4–13 to 4.4–15

biological diversity see biodiversity

biological indicators 3.1–3, 3.1–7, 3.2–2,

7.1–2

multivariate indicators 7.3–1 to 7.3–2

univariate indicators 7.3–1

biological oxygen demand A–6

biomagnification 3.4–17, A–6

biomass A–6

biosolids A–6

biota A–7

biotoxins 4.4–15, A–7

bioturbation A–7

biting insects 5–5

bivalve A–7

black bream A–7

black tiger prawn A–7

bloom A–7

blue mussel A–7

blue-green algae see cyanobacteria

broad-scale assessment of ecosystem health

3.2–4, 3.2–6, 3.2–7, 3.2–20, 7.2–3

buffer A–7

buffering capacity A–7

°C (degrees Celsius) A–7

cadmium, in soil, interaction with chloride in

irrigation water 4.2–10

calcium, in livestock drinking water 4.3–2

carcinogen A–7

carrying capacity 2–18

catchment A–7

catchment level responsibility 2–3

catchment management plans 2–14

cathode A–7

cation A–7

cation exchange capacity (CEC) A–7

cationic A–7

chelate A–7

chemical oxygen demand 4.4–7, A–7

chemical speciation in water samples 7.4–2 to

7.4–3

chemical stressors see physico-chemical

stressors

chemicals

toxicity assessment 2–19 to 2–20

see also toxicants

chloride, in irrigation waters 4.2–9 to 4.2–10

chlorination A–7

chlorophyll 3.3–10, 3.3–12, 3.3–14, 3.3–16,

3.3–17, 3.3–25

chronic 3.4–1, A–7

chronic value A–7

cladoceran A–7

climate variability 3.1–7

colloid A–8

community A–8

community composition 7.3–1, A–8

community metabolism A–8
Index

Version — October 2000 index 3

complexation A–8

compliance 2–13, A–8

concentration A–8

conceptual models 2–18 to 2–19, 3.1–19 to

3.1–20

condition 1 ecosystems see high

conservation/ecological value ecosystems

condition 2 ecosystems see slightly to

moderately disturbed ecosystems

condition 3 ecosystems see highly disturbed

ecosystems

condition indicators or targets A–8

contaminant bioavailability 3.5–7, 4.2–12

contaminants A–8

see also biological contaminants

continual improvement 2–16

control charting 3.2–14, 3.3–19, 7.4–7

controls 3.1–15 to 3.1–16

definition A–8, 3.1–15

cooperative best management 2–13 to 2–14,

3.1–22, 3.2–13 to 3.2–14

corrosion 4.2–15 to 4.2–16, A–8

cresylic A–8

criteria (water quality) A–8

crop quality A–8

cultural importance of water 2–6 to 2–7

cultural issues 1–4

cultured fish, molluscs and crustaceans

protection of 4.4–2

water quality guidelines 4.4–3 to 4.4–12

cumulative A–8

cumulative contaminant loading limit (CCL)

4.2–11, 4.2–12, A–8

cyanobacteria A–8

in irrigation waters 4.2–3

in livestock drinking water 4.3–1 to 4.3–2

in recreational waters 5–6

cyanobacterial blooms, risk assessment 3.3–28

cyanosis A–8

cytotoxic A–8

data analysis, evaluation and reporting 7.2–19

decision criteria A–8

and trigger values 3.1–20 to 3.1–22, 7.2–11

to 7.2–18

decision trees/frameworks 2–10, 3.1–4, 3.1–6,

3.1–7, A–9

for applying guideline trigger values for

toxicants 3.4–13 to 3.4–20

for assessing physico-chemical stressors in

ambient waters 3.3–1 to 3.3–2

for assessing test site data and deriving sitespecific water quality guidelines 3.1–17

to 3.1–19

for assessment of contaminated sediments

3.5–5 to 3.5–10

for biological assessment of water quality

3.2–2 to 3.2–3

for metal speciation guidelines 3.4–19

for water quality for protection of

aquaculture species 4.4–4 to 4.4–5, 4.4–6

defraying costs, in environmental monitoring

programs 7.1–3 to 7.1–4

depuration A–9

desirable contaminant concentration (DCC)

A–9

detection limit 3.4–15, A–9

detritus A–9

dinoflagellates A–9

direct effect stressors 3.3–3 to 3.3–4

direct toxicity assessment (DTA) 2–19 to

2–20, 3.4–16, 3.4–20, A–9

discharge controls 2–17

dissolved oxygen 3.3–10, 3.3–12, 3.3–14,

3.3–16, 3.3–17

low 3.3–25 to 3.3–27

disturbance intensity, and indicator response

3.1–20

disturbances 3.1–15, 3.2–12, 7.2–3

detection and assessment 3.2–18 to 3.2–19

diuresis A–9

diurnal A–9

divalent A–9

documentation 7.2–18

dose A–9

drinking water 1–4, 6–1

chemical and radiological quality 6–2

guideline values 6–3 to 6–4

guidelines for users

in Australia 6–1

in New Zealand 6–1

individual household supplies 6–3

microbiological quality 6–1 to 6–2

small water supplies 6–2 to 6–3

dynamic equilibrium A–18

early detection A–9

short- or longer-term changes 3.2–4 to

3.2–5, 3.2–6 to 3.2–8

early life-stage test A–9

EC1:5 A–9

EC50 (median effective concentration) A–9

ECi 4.2–6, 4.2–7, 4.2–8

ECs A–9

ECse 4.2–6, 4.2–7, 4.2–8, A–9

ecological effects data 3.1–14, 3.3–7

ecological integrity (health) 3.1–1 to 3.1–2,

A–9
Index

index 4 Version — October 2000

ecologically sustainable development (ESD)

1–5, A–9

ecosystem-based approach 2–18 to 2–19

ecosystem classification 3.1–4 to 3.1–5, 3.1–7

to 3.1–9

ecosystem condition 3.4–14, A–9

and level of protection 3.1–9 to 3.1–12

ecosystem health A–9

broad-scale assessment 3.2–4, 3.2–6,

3.2–7, 3.2–30, 7.2–3

see also ecological integrity

ecosystem protection, biological assessment

objectives for 3.2–4 to 3.2–6

ecosystem-specific modifying factor A–10

effect size 3.1–21, 3.2–13, 7.2–15, 7.3–3,

A–10

effluent A–10

effluent discharge 1–3

effluent outfalls 2–17

El Niño-Southern Oscillation (ENSO) 3.1–7

electrical conductivity 3.3–11, 3.3–13, 3.3–15,

3.3–16, 4.3–4, A–10

encrustation A–10

end-points A–10

endemic/endemism A–10

enhancing inferences, in environmental

monitoring programs 7.1–2 to 7.1–3

enterococci 5–4, A–10

environmental capacity 2–18

environmental concerns 3.1–5

environmental restoration 7.2–10, 7.2–13

environmental values 1–2, 2–1, 2–6 to 2–9,

2–15, 2–16, A–10

epilimnion A–10

epilithon A–10

epiphyte A–10

ESP see exchangeable sodium percentage

eukaryotes A–10

euphotic A–10

euryhaline A–10

eutrophic A–10

eutrophication 3.1–2, A–10

evapotranspiration A–10

exchangeable sodium percentage (ESP) A–10

experimental design and analysis procedures,

for generic protocols 3.2–10 to 3.2–13,

7.2–3 to 7.2–17

exposure A–10

faecal coliforms 5–4

farmstead water supplies 4.1–2

fate A–10

field capacity A–10

fish resources 4.4–1

flexible decision-making 3.2–13 to 3.2–14

flocculation A–11

flounder A–11

flow-through system A–11

fluorosis A–11

foliar injury

from chloride 4.2–9 to 4.2–10

from sodium 4.2–10

fouling 4.2–16, A–11

free carbon dioxide A–11

freshwater shrimp A–11

gastropod A–11

general water uses 4.2–1

agricultural chemical preparation 4.2–16

corrosion 4.2–15 to 4.2–16

fouling 4.2–16

pH 4.2–15

gilvin A–11

green shell mussel 4.4–1, A–11

gross alpha (activity) 4.2–15, 4.3–6, A–11

gross beta (activity) 4.2–15, 4.3–6, A–11

groundwater 4.1–1, A–11

Guideline applicability 1–2

guideline (water quality) A–11

guideline packages 3.1–4, 3.1–6, A–11

for applying the guideline trigger values to

sites 3.3–21 to 3.3–29

guideline trigger values 1–2, 2–10, 3.1–3,

3.1–4, 3.1–6, 3.1–7, 3.1–17 to 3.1–19, A–11

and decision criteria 3.1–20 to 3.1–22

determining for selected indicators 3.1–5 to

3.1–6

drinking water 6–3 to 6–4

for toxicants 3.4–5 to 3.4–11, 7.4–8 to

7.4–9

levels of refinement 3.1–18

local reference data 3.1–14

physico-chemical stressors 3.3–5 to 3.3–7,

7.4–3 to 7.4–8

Guidelines 1–1

application to groundwater 1–2

applications for water quality management

2–11 to 2–19

background 1–4 to 1–5

features 1–1

framework 2–1 to 2–20

guiding principles 1–5

not standards 2–16 to 2–17

not to be used as mandatory standards 1–1

to 1–3

objectives 1–6

philosophical approach to applying 2–12 to

2–17
Index

Version — October 2000 index 5

revision program A–25 to A–26

targeted for local conditions 2–15

habitat 2–16, A–11

half-life A–11

hardness 3.4–18, 3.4–21, A–11

hazard A–11

health-related guideline value 6–4

heavy metals

effect on aquatic ecosystems 3.4–5

effect on sediments 3.5–3

effect on aquaculture species 4.4–8, 4.4–10

in irrigation water 4.2–11 to 4.2–12

in livestock drinking water 4.3–4 to 4.3–5

in recreational waters 5–9

helminth A–12

hepatotoxin A–12

in aquatic ecosystems 3.4–9 to 3.4–10

herbicides, in irrigation water 4.2–14

heterotrophy A–12

high conservation/ecological value systems

2–9, 3.1–10, 3.1–11

monitoring and assessment indicators

7.2–1 to 7.2–2

situation-dependent guidelines 3.2–17 to

3.2–19

high reliability guideline trigger values 3.4–2

to 3.4–3, A–12

highly disturbed ecosystems 2–9, 2–10,

3.1–10, 3.1–12, 3.2–19

guidelines 3.1–22 to 3.1–24

lack of suitable reference sites or data

3.1–23

low-risk trigger values 3.3–23 to 3.3–24,

3.3–26

monitoring and assessment indicators

7.2–2

urban regions, guidelines 3.1–23 to 3.1–24

water quality objectives 3.1–22 to 3.1–23

historical data 3.1–14

human consumers of aquatic foods 4.4–1

precautionary comments 4.4–18 to 4.4–19

preventative and management approaches

4.4–17 to 4.4–18

protection of 4.4–2

water quality guidelines for protection of

4.4–12 to 4.4–18

human factors

affecting aquatic ecosystems 3.1–1 to

3.1–2

affecting ecosystems 3.1–5

human pathogens, in irrigation waters 4.2–3 to

4.2–4

humic substances A–12

hydrogeology A–12

hydrograph A–12

hydrological sampling 7.4–2

hydrolysis A–12

hydrophilic A–12

hydrophobic A–12

hypolimnion A–12

hypothesis 3.1–19, A–12

hypothesis testing 7.2–11, 7.2–12 to 7.2–13

nature of change and its context 7.2–13 to

7.2–15

specifying error rates relative to costs of

those errors 7.2–16 to 7.2–18

specifying the size of the effect 7.2–15 to

7.2–16

hypoxia A–12

in vitro A–12

incipient LC50 A–12

indicator response, and disturbance intensity

3.1–20

indicators 3.1–3, 3.1–5, A–12

broad classes and desired attributes 3.2–6

to 3.2–9

matching to problems 3.2–9 to 3.2–10

see also biodiversity indicators; biological

indicators; non-biological indicators

Indigenous Australians, cultural importance of

water 2–7

indirect effect stressors 3.3–4

individual household supplies 6–3

industrial water 1–4

ingestion A–12

inorganic carbon A–12

inorganic toxicants

aquaculture species 4.4–8, 4.4–10

impact on human health 4.4–13

in recreational waters 5–9

integrated catchment management 2–3

integrated water quality assessment 2–16,

7.1–1 to 7.1–4

management goals for 3.1–19 to 3.1–20

interstitial A–12

invertebrates 3.1–23, 3.2–4, A–12

investigative studies 7.2–4

ion A–12

irrigation water

biological parameters 4.2–2 to 4.2–4

effect on soil, plants and water resources

4.2–1 to 4.2–2

heavy metals and metalloids 4.2–11 to

4.2–12

ions of concern for water quality 4.2–9 to

4.2–11

nitrogen and phosphorus 4.2–12 to 4.2–13
Index

index 6 Version — October 2000

pesticides 4.2–13

radiological quality 4.2–13, 4.2–15

salinity and sodicity 4.2–4 to 4.2–9

water quality guidelines 4.1–1, 4.2–1 to

4.2–15

philosophy 4.2–1

Kuruma prawn A–12

lack of dissolved oxygen 3.3–25

key indicators 3.3–25 to 3.3–26

low-risk trigger values 3.3–26

use of guideline package 3.3–27

Langelier Saturation Index (SI) A–12

LC50 (median lethal concentration) A–12

LC100 A–12

LD50 (median lethal dose) A–13

leachate A–13

leaching A–13

leaching fraction (LF) 4.2–6, A–13

lentic A–13

lethal A–13

levels of protection 3.1–5, A–13

alternative levels 3.1–12

and ecosystem condition 3.1–9 to 3.1–11

assignment 3.1–9 to 3.1–13

framework for assigning 3.1–11 to 3.1–12

recommended 3.1–13

life-cycle study A–13

ligand A–13

livestock drinking water quality 4.1–1, 4.3–1

biological parameters 4.3–1 to 4.3–2

derivation and use of guidelines 4.3–1

heavy metals and metalloids 4.3–4 to 4.3–5

major ions of concern for 4.3–2 to 4.3–4

pesticides 4.3–5

radiological quality 4.3–6

liveweight A–13

load-based guidelines 3.3–20

local jurisdictions, site-specific guideline

values for highly disturbed ecosystems

3.1–23

LOEC (lowest observed effect concentration)

A–13

LOEL (lowest observed effect level) A–13

long-term trigger values (LTV) A–13

heavy metals and metalloids in irrigation

water 4.2–11 to 4.2–12

nitrogen and phosphorus 4.2–12 to 4.2–13

longer-term changes, early detection 3.2–4 to

3.2–5, 3.2–6 to 3.2–8

lotic A–13

low reliability guideline trigger values 3.4–2

to 3.4–3, A–13

low-risk guideline trigger values

comparison with 3.3–19

default approach to deriving 3.3–8 to 3.3–9

definition 3.3–5 to 3.3–6

ecosystem conditions 3.3–6 to 3.3–7

preferred approaches to deriving 3.3–7 to

3.3–8

sources of information for 3.3–6

macroinvertebrate communities 3.1–23, 3.2–4

macrophyte A–13

magnesium, in livestock drinking water 4.3–3

management focus on issues not guidelines

2–15

management framework for applying the

guidelines 2–1 to 2–11

management goals 2–2, 2–8, 3.1–5, A–14

to integrate water quality assessment

3.1–19 to 3.1–20

management responses 2–2, 2–3

management strategy 1–3, 2–3

Maori, cultural importance of water 2–7

marine biotoxins, impact on human health

4.4–14 to 4.4–15

marron A–14

MATC (maximum acceptable toxicant

concentration) A–14

maximum tolerable daily level (MTDL) A–14

measurable perturbation 7.4–6

median 7.4–6, A–14

mercury 3.4–2

mesotrophic A–14

metabolite A–14

metalloids

guideline values 3.5–3, 3.5–4

in irrigation waters 4.2–11 to 4.2–12

in livestock drinking water 4.3–4 to 4.3–5

metals

guideline values 3.5–3, 3.5–4

speciation considerations 3.5–7 to 3.5–8

speciation guidelines, decision trees 3.4–19

see also heavy metals

methyl mercury A–14

methylation A–14

microalgal toxins 4.4–15

microbiological characteristics

recreational waters 5–4 to 5–5

see also pathogens

microbiological quality, drinking water 6–1 to

6–2

mixing zones 2–17, A–14

adjacent to effluent outfalls 2–17

moderate reliability guideline trigger values

3.4–2 to 3.4–3, A–14
Index

Version — October 2000 index 7

monitoring and assessment program 1–4, 2–2,

2–3, 2–14, 7.1–1

biological indicator issues 7.3–1 to 7.3–5

broad classes of monitoring design 7.2–3 to

7.2–4

checklist of issues in refining program

design 7.2–5

importance of good pilot data 7.2–11

setting criteria for decisions 7.2–11 to

7.2–18

site selection and temporal and spatial

scales 7.2–5 to 7.2–11

choosing a study design 7.2–1

combinations of indicators for two likely

special cases 7.2–2 to 7.2–3

recommended indicators for each

ecosystem condition 7.2–1 to 7.2–2

data analysis, evaluation and reporting

7.2–19

framework 7.1–4 to 7.1–5

indicator selection 3.1–6, 7.1–2

integrated strategies 7.1–1 to 7.1–2

defraying costs 7.1–3 to 7.1–4

enhancing inferences 7.1–2 to 7.1–3

issues for restoration and rehabilitation

7.2–10

physico-chemical indicator issues 7.4–1 to

7.4–10

sample processing and analysis 7.2–18 to

7.2–19

sampling protocols and documentation

7.2–18

monomeric A–14

morphometry A–14

multiple lines of evidence 3.2–15, 3.4–13,

7.2–2, 7.2–18, A–14

multivariate indicators 7.3–1 to 7.3–2

multivariate studies 3.2–10, 3.2–12

Munsell Scale 5–3, A–14

national level responsibility 2–3

National Water Quality Management Strategy

1–3, A–21

documents A–22

objectives 1–6

principles 1–5

responsibilities 2–3 to 2–4

natural factors affecting ecosystems 3.1–5

necrotic A–14

neurotoxin A–14

New Zealand

default trigger values, physico-chemical

stressors 3.3–17 to 3.3–18

ecosystem features 3.1–7, 3.1–8

New Zealand Ministry for the Environment,

water quality documents 1–3, A–23

New Zealand’s National Agenda for

Sustainable Water Management (NASWM)

1–3

nitrate and nitrite

in aquatic ecosystems 3.3–10 to 3.3–17,

3.4–5, 3.5–3

in drinking water 6–2

in livestock drinking water 4.3–3

nitrogen 3.3–10, 3.3–12, 3.3–14, 3.3–16,

3.3–17

in irrigation water 4.2–12 to 4.2–13

NOEC (no observed effect concentration)

A–14

non-biological indicators, guideline trigger

values 3.1–21

non-biting insects 5–5

non-toxic direct effect stressors 3.3–3 to 3.3–4

not detectable 2–9 (‘no change’), A–14

nuisance growth of aquatic plants 3.3–22

key indicators 3.3–23

low-risk trigger values 3.3–23 to 3.3–24

recreational waters 5–5 to 5–6

suitable nutrient loads 3.3–24 to 3.3–25

use of guideline package 3.3–24

nuisance organisms, recreational waters 5–5 to

5–6

nutrient loadings, in ecosystems 2–18 to 2–19

nutrient solution A–14

nutrients, in sediments 3.5–3, 3.5–5

octanol:water partition coefficient (Pow) A–14

off-flavour compounds 4.4–15 to 4.4–17,

A–14

oil 5–8

oligotrophic A–15

organic carbon A–15

organic compounds, guideline values 3.4–4 to

3.4–10 (in water), 3.5–3 to 3.5–4 (in

sediments)

organic toxicants

effect on aquaculture species 4.4–8, 4.4–10

to 4.4–11

impact on human health 4.4–13

in recreational waters 5–9

organism A–15

organochlorine pesticides 3.4–2

organometallic compounds, guideline values

3.4–5 (in water), 3.5–3, 3.5–4 (in sediments)

osmoregulation A–15

osmosis A–15

oxidation A–15
Index

index 8 Version — October 2000

oxides of nitrogen 3.3–10, 3.3–12, 3.3–14,

3.3–16, 3.3–17

oxygenation A–15

Pacific oyster 4.4–1, A–15

PAH A–15

parameter A–15

parasites

impact on human health 4.4–14

in livestock drinking water 4.3–2

partition coefficient A–15

pathogens A–15

effect on aquaculture species 4.4–11 to

4.4–12

in irrigation waters 4.2–3 to 4.2–4

in livestock drinking water 4.3–2

in recreational waters 5–4 to 5–5

paua A–4

percentile 3.3–7 to 3.3–8, A–15

performance indicators A–15

periphyton 5–6, A–15

pesticides 3.4–2, A–15

in aquatic ecosystems 3.4–9 to 3.4–10

(water), 3.4–4 (sediment)

in drinking water 6–2

in irrigation water 4.2–13, 4.2–14

in livestock drinking water 4.3–5

in recreational waters 5–10

petrochemicals 5–8

pH 3.3–10, 3.3–12, 3.3–14, 3.3–16, 3.3–17,

A–15

of recreational waters 5–7

of water 4.2–15

pH (CaCl2) A–15

phenols A–15

philosophical approach to applying guidelines

2–12 to 2–17

phosphate 3.3–10, 3.3–12, 3.3–14, 3.3–16,

3.3–17, 3.3–25

phosphorus

in irrigation water 4.2–12 to 4.2–13

in sediments 3.5–5

photodegradation A–15

photolysis A–15

photosynthesis A–15

physical and chemical indicators

checklist for 7.4–1 to 7.4–2

chemical speciation in water samples 7.4–2

to 7.4–3

comparing test data with guideline trigger

values

physical and chemical stressors 7.4–3 to

7.4–8

sediments 7.4–10

toxicants 7.4–8 to 7.4–9

‘upstream’ from test sites 7.4–9

hydrology and representative sampling

7.4–2

quality assurance and quality control 7.4–3

physico-chemical A–15

physico-chemical stressors 3.3–1 to 3.3–2

affecting aquatic ecosystems 3.3–4 to

3.3–5

aquaculture species 4.4–7, 4.4–9 to 4.4–10

comparing test data with guideline trigger

values 7.4–3 to 7.4–8

comparison with low-risk guideline trigger

values 3.3–19

default trigger values 3.3–8 to 3.3–18

defining low-risk guideline trigger values

3.3–5 to 3.3–7

load-based guidelines 3.3–20

measuring acceptable ecological change

3.3–19 to 3.3–20

preferred approaches to deriving low-risk

guideline trigger values 3.3–7 to 3.3–8

types of 3.3–3 to 3.3–4

physiology A–15

phytoplankton A–15

phytoplankton monitoring 4.4–18

phytotoxicity A–16

pilot program 3.1–3, 7.2–11, A–16

plankton A–16

plant pathogens, in irrigation waters 4.2–4

plants, salinity tolerance 4.2–8

pollution A–16

polychlorinated biphenyls 3.4–2, A–16

potable water A–16

Practical Quantitation Limit (PQL) 3.4–15 to

3.4–16, A–16

precipitation A–16

primary contact activities 5–3

microbiological characteristics 5–4

primary industries 1–4, 4.1–1 to 4.1–2

primary management aims 3.1–4 to 3.1–5

primary production A–16

Project Committee A–26

prokaryotes A–16

prolarvae A–16

protocols A–16

early detection of sediment toxicity 3.2–9

to 3.2–10, 3.2–11

experimental design and analysis

procedures 3.2–10 to 3.2–13

protozoans A–16

quality assurance (QA) 7.2–18, 7.4–3, A–16

quality control (QC) 7.2–18, 7.4–3, A–16
Index

Version — October 2000 index 9

radiological A–16

radiological quality

irrigation water 4.2–13, 4.2–15

livestock drinking water 4.3–6

recreational waters 5–9

radionuclides, impact on human health 4.4–13

rainbow trout A–16

rapid biological assessment (RBA) 3.2–4,

3.2–8, 3.2–9, A–16

see also AUSRIVAS rapid biological

assessment

receiving capacity 2–18

recreational water quality and aesthetics 1–4,

5–1

basis of proposed guidelines A–29

guidelines for users

in Australia 5–1 to 5–2

in New Zealand 5–1

microbiological characteristics 5–4 to 5–5

nuisance organisms 5–5 to 5–6

physical and chemical characteristics 5–6

to 5–10

recreational categories 5–3 to 5–4

recruitment A–16

red claw A–17

redox potential A–17

reference conditions 3.1–14, 3.1–16, A–17

sources of information 3.1–14 to 3.1–15

reference data 3.3–7 to 3.3–8

reference sites 3.1–16, 7.4–5 to 7.4–6

references R1–1 to R7–3

regional level responsibility 2–3

rehabilitation 3.2–20, 7.2–10

relaying A–17

remedial actions, assessing success of 3.2–20,

7.2–10

representative sampling 7.4–2

Resource Management Act (1991) (NZ) 1–5,

2–3 to 2–4

responsibilities, for water quality management

2–3 to 2–4

restoration 7.2–10, 7.2–13

risk A–17

risk assessment 2–11, 2–14

cyanobacterial blooms in a lowland river

3.3–28

risk-based application of the guidelines 3.1–6

to 3.1–7

risk-based guideline packages 3.3–21 to

3.3–22

lack of dissolved oxygen 3.3–25 to 3.3–27

nuisance growth of aquatic plants 3.3–22 to

3.3–25

Ryznar (stability) index 4.2–15, 4.2–16, A–17

salinity 3.3–11, 3.3–13, 3.3–15, 3.3–16, A–17

and plant tolerance 4.2–8

in irrigation water 4.2–4 to 4.2–9

sample processing and analysis 7.2–18 to

7.2–19

sampling protocols 3.2–9 to 3.2–10, 7.2–18

secondary contact activities 5–3

microbiological characteristics 5–4 to 5–5

sediment contaminants 3.5–3

sediment pore water 3.1–22, 3.5–8, A–17

sediment quality 2–15, 2–16

sediment quality guidelines 3.5–1

application

chemical testing 3.5–5 to 3.5–10

sediment sampling 3.5–5

approach and methodology used in trigger

value derivation 3.5–2 to 3.5–3

comparing test data with guideline trigger

values 7.4–10

recommended guideline values

absence of guidelines 3.5–5

ammonia, sulfide, nutrients and other

sediment contaminants 3.5–3

metals, metalloids, organometallic and

organic compounds 3.5–3

underlying philosophy 3.5–1 to 3.5–2

sediment sampling 3.5–5

sediment speciation 3.5–7 to 3.5–8

sediment toxicity, early detection protocols

3.2–10, 3.2–11

sediments A–17

sewage fungus A–17

shellfish-growing areas, area classification

approach 4.4–17 to 4.4–18

short-term changes, early detection 3.2–4 to

3.2–5, 3.2–6 to 3.2–8

short-term trigger values (STV) A–17

heavy metals and metalloids in irrigation

water 4.2–11 to 4.2–12

nitrogen and phosphorus 4.2–12 to 4.2–13

silver perch A–17

simultaneously extracted metals A–17

site selection and temporal and spatial scales

7.2–5 to 7.2–11

site-specific guideline levels 3.1–22, 3.1–23

sites, definition 3.1–16

sites where an insufficient baseline sampling

period is available 3.2–19, 7.2–2 to 7.2–3

situation-dependent guidelines 3.2–17 to

3.2–20

slightly to moderately disturbed ecosystems

2–9, 3.1–11 to 3.1–12, 3.2–19

biodiversity indicators 7.2–2

low-risk trigger values 3.3–23, 3.3–26

small water supplies 6–2 to 6–3
Index

index 10 Version — October 2000

snapper A–17

sodicity A–17

in irrigation water 4.2–4 to 4.2–9

sodium, in irrigation water 4.2–10 to 4.2–11

sodium adsorption ratio (SAR) 4.2–7, 4.2–11,

A–17

soil structure degradation, caused by irrigation

water quality 4.2–7

solution concentration A–17

sorption A–17

south central Australia, low rainfall areas,

default trigger values, physico-chemical

stressors 3.3–16

south-east Australia, default trigger values,

physico-chemical stressors 3.3–10 to 3.3–11

south-west Australia, default trigger values,

physico-chemical stressors 3.3–14 to 3.3–15

spatial control 3.1–15

spatial data 3.1–14

spatial scales 7.2–9

spatial variation (within sites) 7.2–5

speciation 7.4–1, A–17

see also sediment speciation

species A–17

species richness A–18

SPM (suspended particulate matter) 3.3–11,

3.3–13, 3.3–15, 3.3–16, A–18

stakeholder involvement 2–4 to 2–6

stakeholders 2–10, 2–14, 3.1–21, A–18

selection of level of protection 3.1–10 to

3.1–11

standard (water quality) 1–1, A–18

standing crop A–18

state level responsibility 2–3

static system A–18

statistical control 3.1–16

statistical decision making, alternative

approach 2–14

statistical performance characteristics 7.4–7

statistical performance criteria 2–2

steady state A–18

stormwater 1–3, 2–17

stressors 3.2–9 to 3.2–10, A–18

effect on biological diversity 3.1–17

see also physico-chemical stressors

sub-lethal A–18

sub-lethal organism responses 3.2–8

sulfate, in livestock drinking water 4.3–3

sulfide 3.4–3

supersaturation A–18

surface films, recreational waters 5–8

survival time A–18

suspended particulate matter see SPM

suspension A–18

sustainable nutrient loading 3.3–24

sustainable organic matter loads for standing

waterbodies 3.3–29

sustainable use 2–12

Sydney rock oyster A–18

synergism A–18

tainting A–18

chemical compounds causing 4.4–15 to

4.4–17

taxa richness A–18

taxon (taxa) A–18

taxonomic (group, resolution) A–18

temperature

of aquatic ecosystems 3.3–3 to 3.3–5,

3.3–8 to 3.3–9, 3.3–21

of waters for aquaculture 4.4–7

of recreational waters 5–8

effects on toxicants 3.4–13, 3.4–17

temporal change 7.2–8

temporal control 3.1–15, 7.1–2

teratogen A–18

territory level responsibility 2–3

test sites

physical and chemical stressors 7.4–3 to

7.4–8

sediments 7.4–10

toxicants 7.4–8 to 7.4–9

‘upstream’ surface waters from 7.4–9

thermodynamic equilibrium A–18

thermotolerant coliforms A–18

in irrigation waters 4.2–3 to 4.2–4

in livestock drinking water 4.3–2

three-phased screening approach 4.4–18

threshold concentration A–19

threshold value of indicator 3.1–20, 3.1–21

tolerance A–19

of plants to salinity 4.2–8

total catchment management 2–3

total dissolved solids (TDS) A–19

in livestock drinking water 4.3–3 to 4.3–4

total metal A–19

total organic carbon (TOC) 3.5–8

toxic stressors 3.3–3

toxicants 3.4–1

applying guideline trigger values to sites

3.4–11 to 3.4–12

comparing monitoring data with trigger

values 3.4–21 to 3.4–22

decision tree 3.4–13 to 3.4–20

underlying philosophy 3.4–12 to 3.4–13

aquaculture species 4.4–8, 4.4–10 to

4.4–11

background data 3.1–14

background levels 3.1–22
Index

Version — October 2000 index 11

comparing test data with guideline trigger

values 7.4–8 to 7.4–9

definition 3.4–1, A–19

effect on ecological integrity 3.1–17

guideline development 3.4–1

altering the level of protection for

different ecosystem conditions 3.4–3

to 3.4–4, 3.4–11

extrapolating from laboratory data 3.4–2

procedures for deriving trigger values

3.4–2 to 3.4–3

toxicity data for deriving guideline

trigger values 3.4–1 to 3.4–2

impact on human health 4.4–13

recreational waters 5–8, 5–9 to 5–10

trigger values for alternative levels of

protection 3.4–5 to 3.4–11

see also heavy metals; inorganic toxicants;

organic toxicants; pesticides

toxicity 3.4–17 to 3.4–18 A–19

toxicity assessment 2–19 to 2–20

see also Direct toxicity assessment

toxicity identification and evaluation (TIE)

3.4–16, 3.5–10 A–19

toxicity testing 3.5–9 to 3.5–10 A–19

treatments 3.1–15, 3.1–16

trigger values see guideline trigger values

trochus A–19

tropical Australia, default trigger values,

physico-chemical stressors 3.3–12 to 3.3–13

tropical ecosystems 3.3–20 to 3.3–21

true colour A–19

turbidity 3.3–11, 3.3–13, 3.3–15, 3.3–16,

3.3–18

turbulence A–19

Type I errors 2–14, 3.1–21, 7.2–11, 7.2–12,

7.2–16 to 7.2–17 A–19

Type II errors 2–14, 3.1–21, 7.2–11, 7.2–12,

7.2–16 to 7.2–17 A–19

unacceptable level of change

data assessed against bands of AUSRIVAS

predictive models 3.2–15 to 3.2–17

guidelines for determining

compliance and legal framework 3.2–14

flexible decision-making 3.2–13 to

3.2–14

‘weight-of-evidence’ approach 3.2–15

situation-dependent guidelines 3.2–17

broad-scale assessment of ecosystem

health 3.2–20

highly disturbed systems 3.2–19

sites of high conservation value 3.2–17

to 3.2–19

sites where an insufficient baseline

sampling period is available to meet

key default guideline decision criteria

3.2–19

slightly to moderately disturbed systems

3.2–19

univariate analysis 3.2–10, 3.2–12, 7.3–1,

A–19

univariate indicators 7.3–1

‘upstream’ from test site data 7.4–9

uptake A–19

urban regions

highly disturbed ecosystems guidelines

3.1–23 to 3.1–24

water quality assessment programs 3.1–24

value judgements A–19

viral infections, impact on human health

4.4–14

visual use 5–4

volatile A–19

water body 3.1–4

water clarity 3.3–18, 5–6 to 5–7

water colour 5–6 to 5–7

water quality 2–15 to 2–16

water quality criteria A–19

water quality guidelines 2–2, 2–9 to 2–11

definition 2–9, 2–11 A–19

for biological indicators 3.2–1 to 3.2–20

for drinking water 6–1 to 6–4

for irrigation and general water use 4.2–1

to 4.2–16

for livestock drinking water 4.3–1 to 4.3–6

for physical and chemical stressors 3.3–1

to 3.3–30

for protection of cultured fish, molluscs and

crustaceans 4.4–3 to 4.4–12

for protection of human consumers of

aquatic foods 4.4–12 to 4.4–18

for recreational water activities and

aesthetics 5–1 to 5–10

for sediments 3.5–1 to 3.5–10

for toxicants 3.4–1 to 3.4–22

highly-disturbed ecosystems 3.1–22 to

3.1–23

used to trigger action 1–2

water quality management, application of

guidelines for 2–11 to 2–19

water quality management framework 2–1

broad strategy 2–1 to 2–3

responsibilities 2–3 to 2–4

environmental values 2–6 to 2–9
Index

index 12 Version — October 2000

stakeholder involvement 2–4 to 2–6

water quality objectives 2–2, 2–11, 3.1–22

A–19

water quality prediction models 2–18 to 2–19

watertable A–20

whiting A–20

whole effluent toxicity testing A–20

see also Toxicity assessment

within-site variation 3.1–16

World Health Organization, guidelines A–29

xenobiotic A–20

yabby A–20

zooplankton A–20