Handbook of Catchment Management

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HANDBOOK OF CATCHMENT MANAGEMENT

Also available from Wiley-Blackwell

The Lakes Handbook Volume 1 Limnology and Limnetic Ecology Volume 2 Lake Restoration and Rehabilitation Edited by P.E. O’Sullivan & C.S. Reynolds Handbook of Atmospheric Science: Principles and Applications Edited by Nick Hewitt & Andrea Jackson

Handbook of Catchment Management Edited by

Robert C. Ferrier The Macaulay Institute, Craigiebuckler, Aberdeen AB15 8QH, UK And

Alan Jenkins Centre for Ecology and Hydrology, Maclean Building, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB, UK

A John Wiley & Sons, Ltd., Publication

This edition ﬁrst published 2010, © 2010 by Blackwell Publishing Ltd Blackwell Publishing was acquired by John Wiley & Sons in February 2007. Blackwell’s publishing program has been merged with Wiley’s global Scientiﬁc, Technical and Medical business to form Wiley-Blackwell. Registered ofﬁce: John Wiley & Sons Ltd, The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK Editorial ofﬁces: 9600 Garsington Road, Oxford, OX4 2DQ, UK The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK 111 River Street, Hoboken, NJ 07030-5774, USA For details of our global editorial ofﬁces, for customer services and for information about how to apply for permission to reuse the copyright material in this book please see our website at www.wiley.com/ wiley-blackwell The right of the author to be identiﬁed as the author of this work has been asserted in accordance with the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs and Patents Act 1988, without the prior permission of the publisher. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print may not be available in electronic books. Designations used by companies to distinguish their products are often claimed as trademarks. All brand names and product names used in this book are trade names, service marks, trademarks or registered trademarks of their respective owners. The publisher is not associated with any product or vendor mentioned in this book. This publication is designed to provide accurate and authoritative information in regard to the subject matter covered. It is sold on the understanding that the publisher is not engaged in rendering professional services. If professional advice or other expert assistance is required, the services of a competent professional should be sought. Library of Congress Cataloguing-in-Publication Data Handbook of catchment management / Robert C. Ferrier and Alan Jenkins. p. cm. Includes bibliographical references and index. ISBN 978-1-4051-7122-9 (hardback : alk. paper) 1. Water quality management–Handbooks, manuals, etc. 2. Watershed management–Handbooks, manuals, etc. 3. Water resources development–Handbooks, manuals, etc. I. Ferrier, Robert C. II. Jenkins, Alan. TD365.H356 2010 333.91–dc22 2009015239 A catalogue record for this book is available from the British Library. Set in 9 on 11.5 Trump Mediaeval by SNP Best-set Typesetter Ltd., Hong Kong Printed and bound in Malaysia 1

2010

Contents

List of contributors, vii Preface, x Acknowledgments, xiii

1 TH E CATCH M E N T M AN AGE ME N T C O N CE P T, 1 Robert C. Ferrier and Alan Jenkins 2 W E TL AN D M AN AGE M E N T , 19 Mike C. Acreman and J. Owen Mountford 3 FL OOD M AN AGE M E N T , 51 Ralph M. J. Schielen 4

5

E C O L OGICAL CON S E QUE N CE S O F R IVE R CH AN N E L M A NAGE M E N T , 77 Nikolai Friberg M A NAGIN G AGR ICU L T UR A L C A TCHM E N T S T O S US T AIN P RODUCTIO N AN D W AT E R Q U AL ITY, 107 Andrew Sharpley, Marty Matlock, Louise Heathwaite and Tom Simpson

6 E FFL UE N T M AN AG E M E N T , 135 Alan Jenkins and Robert C. Ferrier 7 M A NAGIN G UR B AN R UN OF F , 155 J. Bryan Ellis

8 CAT CHMEN T T O COAST SY ST EMS – MAN AGIN G MICR OBIAL POLLUT AN T S FOR BAT HIN G AN D SHELLFISH HAR V EST IN G WAT ER S, 183 David Kay, Adrian McDonald, Carl Stapleton, Mark Wyer and John Crowther 9 IR R IGAT ION MAN AGEMEN T IN A CAT CHMEN T CON T EX T , 211 Shahbaz Khan 10 MAN AGIN G POT ABLE WAT ER SUPPLIES, 235 Bernard Barraqué 11 MAN AGIN G CAT CHMEN T S FOR HY DR OPOWER GEN ER AT ION , 253 Haakon Thaulow, Arve Tvede, Tor Simon Pedersen and Karin Seelos 12 T HE DAN UBE R IV ER – T HE MOST IN T ER N AT ION AL R IV ER BASIN , 287 Philip Weller 13 MUR R AY -DAR LIN G BASIN – IN T EGR AT ED MAN AGEMEN T IN A LAR GE, DR Y AN D T HIR ST Y BASIN , 303 Sarah Ryan

vi

Contents

14 W A TE R RE S O U R CE S IN S OUT H E A S T E N G L AN D – A DIL E MMA IN SU S TAIN AB L E D E V E L OP M EN T , 323 John C. Rodda 15 M A N AGIN G T H E CAT CHM EN T S O F THE GR E AT B AR R IE R R EE F , 351 Jane Waterhouse, Mike Grundy, Iain Gordon, Jon Brodie, Rachel Eberhard and Hugh Yorkston 16 C A T CHM E N T M AN AGE M E NT C A SE STUD Y – S E N E G AL RI V E R, 377 Mike C. Acreman 17 L A GUN A D E B AY – A T R O PICAL L A K E UN D E R P R E S S U R E , 389 Maria Victoria O. Espaldon 18 C H ESAP E AK E B AY CAT CHME N T M AN AGE M E N T – L E S S O N S L E ARN E D F R O M A C O L L AB OR AT IV E , S CIE N CE-B AS E D A P PROACH T O W AT E R QUAL IT Y RE STORAT IO N , 407 Tom Simpson 19 TH E GL AS GOW S T R AT E GIC D RA IN AGE P L AN , 427 J. Bryan Ellis

20 T HE R UHR CAT CHMEN T (GER MAN Y ) – T HE CON T R IBUT ION OF R ESER V OIR S T O IN T EGR AT ED R IV ER BASIN MAN AGEMEN T , 441 Gerd Morgenschweis 21 EV OLUT ION OF R IV ER BASIN MAN AGEMEN T IN T HE OK AV AN GO SY ST EM, SOUT HER N AFR ICA, 457 Piotr Wolski, Lars Ramberg, Lapo Magole and Dominic Mazvimavi 22 BASIN MAN AGEMEN T APPR OACHES USED IN A HIGHLAT IT UDE N OR T HER N CAT CHMEN T – T HE MACK EN Z IE R IV ER BASIN , 477 Frederick J. Wrona, Joseph M. Culp and Terry D. Prowse 23 T HE FUT UR E FOR CAT CHMEN T MAN AGEMEN T , 501 Robert C. Ferrier, Alan Jenkins and Kirsty Blackstock

Index, 516

Contributors

M I K E C . AC R E M A N Centre for Ecology and Hydrology, Maclean Building, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB, UK

M A R I A V I C T O R I A O . E S PA L D O N School of Environmental Science and Management, University of the Philippines Los Baños, Laguna 4031, Philippines

B E R N A R D B A R R AQ U E CIRED-Agroparistech, 19 avenue du Maine, Paris, France

R O B E R T C . F E R R I E R The Macaulay Institute, Craigiebuckler, Aberdeen AB15 8QH, UK

K I R S T Y B L AC K S T O C K The Macaulay Institute, Craigiebuckler, Aberdeen AB15 8QH, UK

N I KO L A I F R I B E R G The Macaulay Institute, Craigiebuckler, Aberdeen, AB15 8QH, UK; National Environmental Research Institute, Aarhus University, Silkeborg, Denmark

J O N B R O D I E Australian Centre for Tropical Freshwater Research, James Cook University, Douglas QLD 4811 and CSIRO Water for a Healthy Country Flagship, Private Mail Bag PO, Aitkenvale QLD 4814, Australia J O H N C R OW T H E R Centre for Research into Environment and Health, Lampeter, Wales, UK J O S E P H M . C U L P Canadian Rivers Institute, Environment Canada, University of New Brunswick, Fredericton, NB, Canada

I A I N G O R D O N CSIRO Sustainable Ecosystems, Private Mail Bag PO, Aitkenvale QLD 4814, Australia M I K E G R U N DY CSIRO Land and Water, 306 Carmody Road, St Lucia QLD 4067, Australia L O U I S E H E AT H WA I T E Centre for Sustainable Water Management, Lancaster University, Lancaster LA1 4YQ, UK

R AC H E L E B E R H A R D Eberhard Consulting Pty Ltd, 55 Park Road West, Dutton Park QLD 4102, Australia

A L A N J E N K I N S Centre for Ecology and Hydrology, Maclean Building, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB, UK

J . B R YA N E L L I S Urban Pollution Research Centre, Middlesex University, The Burroughs, Hendon, London NW4 4BT, UK

DAV I D K AY Centre for Research into Environment and Health, Aberystwyth, Wales, SY23 3DB, UK

viii

Contributors

S H A H B A Z K H A N UNESCO, Division of Water Sciences, 1, rue Miollis, 75732 Paris Cedex 15, France L A P O M AG O L E University of Botswana, Harry Oppenheimer Okavango Research Centre, Private Bag 285, Maun, Botswana M A R T Y M AT L O C K Biological and Agricultural Engineering Department, Division of Agriculture, University of Arkansas, Fayetteville, Arkansas, USA D O M I N I C M A Z V I M AV I University of Botswana, Harry Oppenheimer Okavango Research Centre, Private Bag 285, Maun, Botswana

S A R A H R YA N CSIRO Sustainable Ecosystems, GPO Box 284, Canberra ACT 2601, Australia R A L P H M . J . S C H I E L E N Ministry of Transport, Public Works and Water Management, Centre for Water Management, PO Box 17, 8200 AA Lelystad, The Netherlands, and University of Twente, Department of Water Engineering and Management, PO Box 217, 7500 AE Enshede, The Netherlands K A R I N S E E L O S Statkraft Energy Production, Environment and Concessions, PO Box 200, Lilleaker, 0216 Oslo, Norway

A D R I A N M c D O N A L D Faculty of Environment, School of Geography, University of Leeds, Leeds LS2 9JT, UK

A N D R E W S H A R P L E Y Department of Crop, Soil and Environmental Sciences, Division of Agriculture, University of Arkansas, Fayetteville, Arkansas, USA

GERD M O R G E N S C H W E I S Water Resources Management Department of Ruhrverband, Kronprinzenstr. 37, 45128 Essen, Germany

T O M S I M P S O N Water Stewardship, Inc., Suite 11, Bldg 7,222, Severn Avenue, Annapolis, MD 21403, USA

J . OW E N M O U N T F O R D Centre for Ecology and Hydrology, Maclean Building, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB, UK

C A R L S TA P L E T O N Centre for Research into Environment and Health, Aberystwyth, Wales, SY23 3DB, UK

T O R S I M O N P E D E R S E N Norwegian Water Resources and Energy Directorate; NVE, Licencing and Supervision Department, PO Box 5091 Majorstua, N-0301 Oslo, Norway

H A A KO N T H AU L OW Norwegian Institute for Water Research, NIVA, Gaustadalléen 21, NO0349 Oslo, Norway

T E R R Y D . P R OW S E Water and Climate Impacts Research Centre, Environment Canada, University of Victoria, Victoria, BC, Canada

A R V E T V E D E Statkraft Energy Production, Environment and Concessions, PO Box 200, Lilleaker, 0216 Oslo, Norway

L A R S R A M B E R G University of Botswana, Harry Oppenheimer Okavango Research Centre, Private Bag 285, Maun, Botswana J O H N C . R O D DA Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB and Hydro-GIS Ltd, 10 Coles Lane, Chalgrove, Oxfordshire OX44 7SY, UK

J A N E WAT E R H O U S E CSIRO Water for a Healthy Country Flagship and Reef Water Quality Partnership, Private Mail Bag PO, Aitkenvale QLD 4814, Australia P H I L I P W E L L E R International Commission for the Protection of the Danube River (ICPDR), PO Box 500, 1400 Vienna, Austria

Contributors

ix

P I O T R WO L S K I University of Botswana, Harry Oppenheimer Okavango Research Centre, Private Bag 285, Maun, Botswana

MARK W Y E R Centre for Research into Environment and Health, Aberystwyth, Wales, SY23 3DB, UK

F R E D E R I C K J . W R O N A Water and Climate Impacts Research Centre, Environment Canada, University of Victoria, Victoria, BC, Canada

H U G H YO R K S T O N Great Barrier Reef Marine Park Authority, PO Box 1379, Townsville QLD 4810, Australia

Preface

Our freshwaters and coasts are intimately connected with the land that surrounds them and the water that feeds them carries the signature of its origin as well as any modiﬁcation along its path of transport. The water trickling in a mountain headwater stream will eventually end up as part of a major river discharging into the sea. On this journey it will converge and mix with water from similar streams draining farmland, forests and cities. In many cases its course will have been altered by human activity, perhaps being dammed, canalised or its ﬂow altered by engineering. The condition of our rivers and coasts is, therefore, a direct consequence of the different landscapes water has passed through. Our waters in many ways represent a unique indicator of the quality of our environment. Whilst this all might seem fairly obvious, we still tend to manage our waters in isolation, often with little regard – or understanding – of how our land-based activities impact upon them. Our rivers have a historical legacy of carrying pollutants and nutrients into our estuaries and coasts, and so the impact of our activities reaches out beyond the river mouth itself. Climate change, industrial pollution, land management, aquaculture, urbanisation and recreation all affect our water environment and these threats, all seemingly growing by the day, result in high costs for both the environment and water consumers. To truly understand our waters, therefore, we need to study them at the level of the river basin, or catchment. A catchment is a topographically

deﬁned area of land in which water falling at any point drains downslope and downstream to a single point or outlet. The signature of how we manage our catchments, in our cities and in the countryside, is recognized in the quality of water as its moves through our landscapes. Our land and water is intimately connected and to manage waters we must understand how we manage our land. But understanding these connections between land and water is only the ﬁrst step. We also need to realize that all environmental problems start and end with people. To properly manage water, it is essential we understand how humans interact with it. It is human activities that ultimately damage our rivers, lakes, estuaries and coasts and as such it is only people – through changes in their attitudes and behaviours – that can make it better. Pollution provides a good example of this. Socalled point sources of pollution such as industrial processes or sewage discharge were identiﬁed some time ago as harmful to the aquatic environment. In many developed countries, thanks to policy initiatives and legislative changes, engineering and end-of-pipe controls were successfully put into place to mitigate or remove the negative effects. Unfortunately in many developing countries gross pollution from industrial activities continues to signiﬁcantly compromise water quality. Many pollution sources, however, are widely dispersed across a catchment area, and whilst

Preface individual sources may be small, collectively they’re very signiﬁcant. In rural areas this includes the run-off from agricultural land (containing pesticides, fertilisers, bacteria and sediment) and pollution from septic tanks, latrines, or unregulated discharge of human waste. In urban areas pollution sources, such as oils and chemicals washed off the surfaces of roads or poured down drains, can be leached into the waste water systems and in many cases end up in local streams. It is not easy, therefore, to identify and deal with a single source of pollution as often this does not exist. The term diffuse pollution is used to describe the pollution that comes from the thousands of surface water drains serving towns and cities, ﬁeld drains and streams in the countryside, and runoff from industrial yards, farmland and forests. The global ﬁnancial implications of this diffuse pollution for the water industry are daunting. Traditional, conventional legislation will not work on its own; what’s required is a set of solutions which target the hearts and minds of individuals to persuade and encourage them to change their behaviours. This sort of issue calls for a sound, informed understanding of not just the science involved, but also of people. Practical policies and enduring solutions depend on an informed appreciation of the importance of both these factors. It is not just a case of recognizing the environmental problems – we must also appreciate that only people can deliver the solutions. Some of the answer lies in imaginative policies, part of it is better education and building capacity – but mostly it requires a much greater understanding of how the individual citizens value water and how engaged they are (or feel they are) in decisions about its management. Integrated Water Resource Management (IWRM) initiatives or River Basin Planning (a strategic decision-making process introduced by legislation such as the EU Water Framework Directive) aims to integrate the management of land and water within river basin districts. For example, the EU Directive requires the preparation of a River Basin Management Plan for each

xi

River Basin District across Europe. It also speciﬁes that all interested parties must be encouraged to become actively involved in planning and management and that the process should be open, equitable and inclusive. Additionally, economic considerations play a key part in river basin planning. In particular, economic information is being used to help make judgments about which combination of measures are the most cost-effective means of improving the status of the water environment. It also ensures that set improvement targets for the status of the water environment are not disproportionately expensive to achieve; and that there be a recovery of an adequate contribution to the costs of water services. All of this demonstrates that it is no longer enough to rely on the single issue approach of the past. Rather than focussing on water quantity or quality alone, we need to consider all the water management issues across a whole catchment to ﬁnd the most environmentally sound, cost effective and socially acceptable ways of handling them. Interdisciplinary science, which tackles issues by bringing together experts from many ﬁelds, has a major role to play in all of this – through combining environmental data collection and the subsequent building of predictive models, with an understanding of how outcomes are inﬂuenced by people’s behaviour and attitudes. This social science component adds value by identifying the barriers as to why very good scientiﬁc solutions sometimes do not get implemented. There is a growing realisation that more scientists need to work alongside stakeholders to help provide practical solutions. Bringing people into the picture helps scientists understand people’s values, as well as helping people understand what the science is actually saying. By creating a shared understanding of what the issues are, shared solutions can be built. This Handbook of Catchment Management aims to elucidate our understanding of historical and current management strategies, in many cases driven by sectoral demands such as the

xii

Preface

requirement for hydropower, ﬂood management and agricultural production. The chapters presented in the ﬁrst half of the Handbook highlight examples of such management but look to rationalise the positive and negative impacts of that single issue management, and how to learn from our shared past and build vision for the future. The latter half of the Handbook presents a series of case study examples from around the world where management of water resources faces not single but multiple pressures. Additionally, we look to the challenge of understanding how future multiple drivers (such as climate change, land use, urbanisation and the requirements for energy) will place increasing pressure on our land and water resources. The case studies and issue based chapters incorporate all aquatic ecosystem types to some degree; rivers, wetlands, groundwater lakes and estuaries/shallow seas. The concept of freshwater ecosystem services and their maintenance is also a central thread to many of the chapters.

We totally depend on water as a central feature of our global heritage, so it is in the interest of everyone to work together to protect and preserve our freshwaters and to ensure equity in all aspects of the water cycle and how we interact with it. The water cycle is the most important of all earth system processes supporting life on this planet and its use by humans must acknowledge this fact. Our ever increasing demands place an enormous pressure on the freshwater environment and on water resources worldwide. With changing lifestyles, the rise in global population and the uncertainty of future climate change, we all need to look carefully at our attitude to water and the way it is used and managed. From mountain top to sea and from hydrology to human behaviour, understanding the bigger picture of how we manage our land and water together at a catchment scale is the route to successful and sustainable water management. Bob Ferrier and Alan Jenkins

Acknowledgements

This contribution to the Handbook series would not have been possible without the concerted efforts of the chapter authors, and we thank them all for the timely contributions of material, their willingness to respond to continued requests from the editors and for their patience! The substantive administrative task of coordinating this Handbook has been undertaken by two key people, Kelly Harper and Linda Moodie of the Macaulay Institute. Their sustained dedication to the project has been exceptional, and we thank them most sincerely for all their hard work in the developmental and pro-

duction phase of this book. Several people helped with the job of editing the text and we especially thank Susie Beresford and Victoria King. A special thanks are due also to Pat Carnegie of the Macaulay Institute, who had the signiﬁcant task of standardizing all the graphic material used throughout, and also to David Riley of the Macaulay Institute for providing additional photography. Finally, special thanks goes to Lorraine Robertson and Elaine Mackenzie also from the Macaulay Institute for their much appreciated effort in reviewing and amending all references throughout the chapters.

The Catchment Management Concept

1

ROBERT C. FERRIER1 AND ALAN JENKINS2 1

2

The Macaulay Institute, Craigiebuckler, Aberdeen, UK Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire, UK

1.1

Introduction

Earth systems processes such as geological, hydrological and biogeochemical have shaped our landscapes. Global and regional climatic factors have inﬂuenced the distribution of water resources (river, lakes, groundwaters, estuaries, etc.) which has inﬂuenced the distribution of ecosystems and humans. Fluvial and erosional process have determined the form of aquatic environments, inﬂuencing the development of channel characteristics, river corridors, ﬂood plains, deltas, along with instream characteristics and morphology such as reach structure, width, depth, and forms of meanders, bars and shoals. Pedological and geological processes inﬂuence the transfer of water between the saturated and unsaturated zone and the nature and extent of groundwater systems, and their interaction with surface waters. The catchment, basin, watershed or similar is basic to hydrological thinking. The catchment outlet identiﬁes the point at which all rainfall naturally drains towards or is directed to by human intervention. Natural processes result in the formation of a stable (at large scale) yet dynamic (at small scale) system bounded within physical (catchment) constraints. Undisturbed

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

catchments are in a quasi-equilibrium, but as landscape features are manipulated and changed by human activities such as land use and riparian management, natural processes are affected. This may result in downstream consequences such as movement of material (soil, water and bedload), generation of downstream ﬂoods, seasonal droughts, altered groundwater levels, increased contaminant transport, coastal sedimentation and many other impacts. A pre-requisite for sustainable resource management at a catchment scale is understanding the water cycle and its ﬂuctuations, which requires knowledge on how water moves through the environment and on the different pools and ﬂuxes that occur across spatial and temporal scales, both within catchments and across larger geographical areas. There are many hydrological processes that characterize catchment systems and their behaviour. Spatial patterns of precipitation both in terms of magnitude and intensity directly impact on the dynamics of stream ﬂow generation and groundwater recharge. Soil moisture controls the distribution and nature of vegetation which in turn links losses of water back to the atmosphere through transpiration, and also inﬁltration. The dynamic interaction between surface and groundwaters in response to ever changing conditions at seasonal, yearly and longer term climatic variation directly inﬂuences the nature of runoff in streams, rivers, lakes and to coastal environments. Emergent properties in time and space are

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robert c. ferrier and alan jenkins

Fig. 1.1 Emergent properties and uncertainty in catchment systems.

a key characteristic of catchment systems and in deﬁning process uncertainty (Fig. 1.1). Additional to understanding the water cycle is knowledge of how human activities inﬂuence both quantity and quality. Recent developments in understanding the relationships between ‘green’ and ‘blue’ water have emphasized the need for catchment scale assessment and identiﬁcation of wider ecosystem goods and services (United Nations 2005). Green water represents that consumed in plant production and evaporation from surfaces, and supports terrestrial ecosystems. Blue water, on the other hand, represents that which recharges aquifers and groundwaters and generates ﬂow in rivers and lakes. It is this latter source of water that supports aquatic ecosystems and human populations. Transfers between blue and green occur when water is

abstracted for irrigation from blue resources, and is partially consumed by green ﬂows only to return to the blue water component. Usually this transfer increases pollutant and nutrient loads into the aquatic environment (Falkenmark and Rockstrom 2006). In summary, the key principles towards developing sustainable management of catchments are ﬁrstly to understand the natural processes occurring within a catchment. In particular, to determine the physical pathways of water movement and balances, hydraulics and dynamics. To elucidate where nutrients and pollutants are generated within the landscape and how and when they may be transported and what are their downstream physical, chemical and ecological consequences. It is also essential to understand the current and future pressures on the water in the catchment and its land use to identify competing demands for the resource given its regional or global context and the history of previous management. Additionally, it is important to consider the social, ethical and political context of options for use and management, as water resources are not necessarily concomitant with administrative, institutional or country boundaries.

1.2 Historical Perspective The major pressures on catchment systems are through a historical timeline of land use and management, urbanization and industrialization driven by different cultures. Each of these pressures has direct and indirect consequences on water resources and there are many synergistic interactions between them at local, regional and global scales (Table 1.1). Land use, in particular the use of water resources for the production of biomass (food and timber), dominated the global water ﬂux. Irrigated land (which makes up 20% of the world’s cropped land but generates 40% of the world’s harvest) accounts for about 70% of water withdrawals at a global scale (Box 1.1). In many developing countries the use of water for cropping approximates

3

The Catchment Management Concept

Water facts and futures

Box 1.1

Current water use: Europe

Global Domestic

Domestic Agriculture

Industry

Agriculture Industry

According to projections made by Population Action International an increasing number of countries across the globe will face either water stress or scarcity

60 Number of countries

Consequences of global population increase:

50 40 30 20 10 0 1995

2025

2050

Year

Water availability:

10,000

The amount of water (cubic metres) available on a per capita basis is set to decline. The demand for water is expected to increase at over 60 billion cubic metres per year

Cubic meters

8,000 6,000 4,000 2,000 0 1990

2000

2025

Year

Table 1.1

Timeline of developing anthropogenic pressures on water

Timeline (years ago) 20,000 10,000 5000 4000 4000 2500–2000 2500–2000 300 200 Last 50 years Present day

Milestone Nomadic humans Move towards pastoralism Emergence of Mesopotamia and Egypt as water managers – irrigation and ﬂood control Water supply and drainage in Indus culture Evidence of water management in China Roman engineers build water supply systems throughout Europe, rise in city populations Groundwater transport and management emerges in Middle Eastern countries Agrarian revolution increases food production, increased land drainage, reclamation and water management Industrial revolution – demand for water in developed world soars World population doubled, water consumption quadrupled Fast growing Asian economies place increasing demands on regional resources Over 1 billion people still do not have access to safe drinking water One-third of the global population currently experiencing water stress, this is set to rise Uncertainties about future climate place water at the centre of a global crisis

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robert c. ferrier and alan jenkins

to 80% of the total water consumption. A compounding factor is that land use and ownership are inconsistent across the landscape and do not necessarily map to natural boundaries such as those delineated by watershed and catchment boundaries, therefore water use is spatially and temporally variable. Human population growth and economic development are placing an increasing burden on global water resources. The United Nations World Water Development Report (2006) highlighted that water withdrawals (from blue water) have increased sixfold since the 1900s, twice the rate of population growth. The increasing urbanization of the world is placing a growing burden on the world’s water resources through demands for production and human use. As large cities grow and available local surface and groundwater supplies are compromised, there is an increasing reliance on distant catchment sources to meet requirements (Showers 2002). Infrastructural integrity in relation to potable supply and appropriate wastewater treatment (if indeed they do exist) potentially struggle to keep pace with the rate of expansion of many cities, especially in developing countries. Urban populations are growing more rapidly than the capacity of governments to establish institutional structures. For example, if the Millennium Development Goal of halving the proportion of people without access to sustainable drinking water and sanitation is to be met by 2015, nearly 1000 million urban dwellers must gain access to improved access to basic functions (WHO/UNICEF 2004). Regulation (where ﬂow has been controlled or modiﬁed from its natural state) can affect the hydrology of rivers down their entire length – straightening for navigation, canalization, building of reservoirs for irrigation, storage, power generation and ﬂood management, groundwater abstraction and inter-basin transfers all generate consequences on the dynamics of ﬂow and for the quality of water environments. Most major rivers experience some form of regulation and large scale management on basins such as the MurrayDarling (Australia) and Colorado (USA) have been ongoing for decades.

Regulation for producing hydropower varies between countries, with Europe exploiting about three-quarters of the available resource whilst Africa only a few per cent. With concerns being raised about the environmental consequences of traditional energy generation, focus has moved to potential renewable energy sources at a range of scales including traditional embayment, dams and barriers through run-of-river schemes to small scale (<10 MW), mini (<2 MW) and micro (500 kW) schemes. Ecologically, freshwater biodiversity is more at risk from anthropogenic activities impacting on water quality and quantity than terrestrial systems. It is estimated that over the last thirty years or so there has been a reduction in freshwater species by approximately 50%. This ﬁgure is far higher than either land or marine ecosystems. The consequence of all of these pressures is that the current human ﬁngerprint on rivers and lakes is extensive. Presently less that 17% of the continental surface of the earth is free from direct human impact. Only in remote areas of North America, Siberia, Amazonia and the Congo Basin can near pristine rivers be found (Meybeck 2003), highlighting the importance of appropriate management of water resources globally. Major biogeochemical cycles have undergone serious perturbations as a result of human activity. The consequences of this have impacted on both the terrestrial and aquatic environment. Since the agrarian revolution, agricultural practices and the move towards intensiﬁcation of production fuelled the usage of commercial inorganic fertilizers, and the widespread availability of nutrient solutes such as nitrate in aquatic ecosystems. Increasing concentrations of nitrogen in ground and surface waters impact on the use of these resources for human consumption (current WHO guidelines highlight a maximum safe concentration of 50 mg L−1 NO3). Elevated concentrations of nitrate in coastal waters are known to drive eutrophication because phytoplankton in marine environments is nitrogen limited. The consequence of this eutrophication is widespread

The Catchment Management Concept hypoxia observed in the world’s shallow and enclosed seas, notably the Gulf of Mexico, Baltic Ocean, Black Sea, and in locations such as Chesapeake Bay and the Delta of the Mississippi. Freshwater eutrophication is primarily though not exclusively driven by the availability of phosphorus, a nutrient elevated in surface waters through runoff from soil and sediment and strongly inﬂuenced by the input of human and animal waste. Current legislation such as the EU Urban Waste Water Treatment Directive has done much to reduce the input of phosphorus from human sources in the rivers and lakes of Europe. Understanding the mode of transport of pollutants to aquatic environments is also of critical importance in the development of remediation strategies for future management plans. In recent times, classic point sources of pollution discharge such as that from sewage treatment works and industrial processes have been the principle focus for water pollution control in many countries. Despite tight regulations and controls on efﬂuent discharges, pollution of water environments continues to be a problem and the focus has now turned to assessing the impact of non-point or diffuse pollution sources. A Chartered Institution of Water and Environmental Management (CIWEM) committee discussed and agreed upon what is now a widely accepted deﬁnition of diffuse pollution published in 2000. That committee based its deliberations on published material in the standard diffuse pollution textbook of the period by Vladimir Novotny and Harvey Olem (1994). The current deﬁnition in international standard use is based on the CIWEM deﬁnition (D’Arcy et al. 2000), and reﬁned by Novotny (2003): Pollution arising from land-use activities (urban and rural) that are dispersed across a catchment or subcatchment, and do not arise as a process industrial efﬂuent, municipal sewage efﬂuent, deep mine or farm efﬂuent discharge.

Diffuse pollution therefore comprises true nonpoint source pollution together with inputs from

5

a multiplicity of minor point sources. Examples of strictly non-point sources are comparatively limited, for example nitrates seeping into groundwater or in unusual circumstances, nutrients and sediment transported in sheet erosion from farmland. Most water-driven soil erosion results in contamination at a speciﬁc point via a rill or gully formed as water traverses ﬁelds to a watercourse. Similarly runoff carrying pollutants from urban surfaces typically reaches a watercourse via a point input such as a surface water sewer outfall. And the same applies to forestry plantations and most other diffuse sources of pollution. Point or non-point is really a matter of scale; a ﬁeld of improved grassland in an upland rough grazing catchment is a nitrate point source for the underlying aquifer, just as each ﬁeld drain is a point source for understanding inputs to a ditch or small stream. In loading terms most diffuse pollution enters watercourses via pipes, channels, gullies and rills, even atmospheric deposition, since it has to be washed from the land surfaces. The important characteristics of diffuse pollution therefore are NOT whether anyone can ﬁnd the source/s, or whether a pipe is involved. Diffuse pollution is a useful concept because it allows for estimation of important loads of pollutants in waterbodies that are not from major industrial process and municipal efﬂuent discharges (that are typically well characterized, monitored and quantiﬁed). The concept is also useful because it explains features of pollution in receiving water bodies, for example why concentrations of some pollutants actually increase with ﬂow rather than are diluted, why pollution peaks are variable and difﬁcult to predict, and why impacts are often slow to develop and become evident. Novotny (2003) characterizes diffuse pollution as follows: • Diffuse discharges enter the receiving surface waters in a diffuse manner at intermittent intervals that are related mostly to the occurrence of meteorological events. • Waste generation (pollution) arises over an extensive area of land and is in transit overland

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before it reaches surface waters or inﬁltrates into shallow aquifers. • Diffuse sources are difﬁcult or impossible to be monitored at the point of origin. • Unlike traditional point sources where treatment is the most effective method of pollution control, abatement of diffuse load is focused on land and runoff management practices. • Compliance monitoring is carried out on land rather than in water. • Water quality impacts are assessed on a catchment scale. • Waste emissions and discharges cannot be measured in terms of efﬂuent limitations. • The extent of diffuse waste emissions (pollution) is related to certain uncontrollable climatic events, as well as geographical and geological conditions, and may differ greatly from place to place and from year to year. • The most important pollutants from diffuse sources subject to management and control are suspended solids, nutrients, faecal pathogens and toxic compounds. A useful way of thinking of diffuse pollution is that it is often the individually minor but collectively signiﬁcant sources in a catchment. That is the key to the control options too; measures need to be focused on the land-based activities, rather than on the point of discharge. The global demand for water is increasing. The world’s population (currently standing at approximately 6 billion), has doubled since the 1950s and is set to increase to approximately 9 billion by 2050 (United Nations 2004). Increases in demand, combined with the current status where many water resources are degraded as a consequence of historical activities, move us closer to a potential global water crisis. The uncertainties generated by climate change further increase the requirement to develop appropriate strategies for the integrated management of water resources at a catchment scale, from mountains to seas and encompassing both land and water issues. The interconnectivity between aquatic ecosystems represents a dynamic and temporally varying environment, which includes permanent and ephemeral water bodies, coastal and estua-

rine environments, and riverine connections. Compromising this continuity, compromises the inherent ecology that has evolved to exploit it.

1.3

Current Solutions

International activities related to the development of policies through which to tackle the growing need for sustainable management of water resources has gathered pace over recent decades. The United Nations Conference on Water (Mar del Plata, 1977), the International Conference on Water and the Environment (1992), the First World Water Forum (Marrakech, 1997) the Second World Water Forum (The Hague, 2000), the International conference on Freshwater (Bonn, 2001) and the World Summit on Sustainable Development (United Nations Department of Economic and Social Affairs 2002), understood the need for effective and efﬁcient management of global water resources. Principles from these international conferences evolved and strengthened the concept of ‘Integrated Water Resources Management’ (IWRM). The Global Water Partnership (2000) deﬁnes IWRM as ‘a process which promotes the co-ordinated development and management of land, water and related resources in order to maximize the resultant economic and social welfare in an equitable manner without compromising the sustainability of vital ecosystems’. Importantly the development of a basin wide approach was reinforced along with strengthened governance and participation (Box 1.2). The Dublin Principles (ICWE 1992) highlighted the importance of water as a ﬁnite resource, that development and management should be participatory, that women play a central role in global water provision especially in the poorest countries, and that water has an economic value (Box 1.2).These principles were further supported during the Second Water Forum in the Hague (WWC 2000), which reiterated that water has an economic value and that participation and empowerment are key factors in sustainable management. It further visioned that water

The Catchment Management Concept

Box 1.2

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Principles of integrated water resource management (Global Water Partnership 2000)

• Freshwater is a ﬁnite and vulnerable resource. Its management therefore requires a holistic approach, one that embraces the concept of the hydrological cycle and its interaction with natural processes and ecosystems and how human activity inﬂuences both quality and quantity. It identiﬁes that natural resource yields have limits and that regional overexploitation is a constant threat. Additionally it recognizes the important spatial component of water resources and that upstream:downstream relations must be addressed to ensure equity in common pool resource management. • Water development and management should be based on a participatory approach. This requires the creation of appropriate participatory mechanisms at a range of scales from national to local, and that capacity building is a central component. Participation in the decision-making process must be transparent, equitable, and that those who have an involvement or interest in water and its use and management must be included in the consultative process, and that consensus building should be the goal. • Women play a central part in the provision, management and safeguarding of water. This aims to promote gender awareness and highlights the important social and cultural roles of women and men in different societies. It encourages the active role of women in decision making, and at all organizational levels. • Water has an economic value in all its competing uses and should be recognized as an economic good. Clearly water has a direct economic value to users, but it also has a non-market value in relation to supporting ecosystem goods and services, and has aesthetic, cultural and religious signiﬁcance. The true value of water must embrace all of these issues. It also promotes the goal of full cost recovery.

should not be monopolized to the detriment of the citizen, and of critical signiﬁcance was that right of access to water for all is central to break down poverty barriers. It was during the International Conference on Freshwater in Bonn (ICFW 2001) that basin management principles were speciﬁcally identiﬁed as a spatial context for management. It re-iterated that ‘the key to long term harmony between nature and its neighbours is co-operative arrangements at the water basin level, including across waters that touch many shores’. This reinforced the key principles of effective IWRM in that all users of a common pool resource should be involved in its management irrespective of administrative, ﬁscal, cultural or national boundaries. Many of the world’s catchment systems are transboundary. Wolf et al. (1999) identiﬁed 261 international transboundary rivers involving 145 different countries that covered 45% of the world’s land surface, 40% of the world’s

population and 60% of the world’s freshwater resource. The World Summit on Sustainable Development (WSSD) in Johannesburg (2002), consolidated UNCED Rio Summit Principles in relation to the implementation of Agenda 21 on Sustainable Development and agreed on an important step towards sustainable water management. As part of the Plan of Implementation it challenged all countries to ‘develop integrated water resource management and water efﬁciency plans by 2005 for all major river basins of the world’. The Millennium Development Goals (MDGs) were developed to address the world’s main development challenges and are due to be achieved by 2015. They arose in the Millennium Declaration that was adopted by nearly 200 nations in September 2000. The eight MDGs break down into 18 quantiﬁable targets which are measured by a series of indicators (Box 1.3).

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Box 1.3

Delivering the millennium development goals: the importance of water

Goal 1: Eradicate extreme poverty and hunger • Access to water supply and sanitation, often used in broad deﬁnitions of ‘poverty’ • The poor carry the greatest burden of productivity-sapping disease as a result of not having access to safe water and sanitation • Water is essential to economic development, which can create productive livelihoods for the poor • Poor communities are also particularly vulnerable to ﬂoods, droughts and similar water-related disasters Goal 2: Achieve universal primary education • Children’s time is a valuable commodity in many communities especially in relation to agriculture and water provision • Water-related disease and the poor availability of adequate sanitation affects school attendance Goal 3: Promote gender equality and empower women • In many areas the fetching and storing of water is done predominantly by women Goal 4: Reduce child mortality • Children are at risk when they are without safe water to drink • In developing countries, water-related diseases are almost always amongst the most important causes of death of children under the age of ﬁve • More than 1.5 million children under ﬁve die every year from diarrhoea (more than from malaria and HIV/AIDS combined) Goal 5: Improve maternal health • The burden of fetching water and dealing with water-related disease in the family falls disproportionately on women and puts pressure on their own health. Goal 6: Combat HIV/AIDS, malaria and other diseases • Access to safe water and sanitation services can help to reduce poverty – which in turn is an important determinant in the spread of HIV/AIDS • Effective water management can reduce malaria and other diseases endemic in poor communities Goal 7: Ensure environmental sustainability (including the target of halving the number of people without access to water and sanitation) • Water is key to the sustainable utilization of land, plant and animal resources • If the water resources environment is not managed and protected, it will not be able to sustain human communities • Provision of water supply and sanitation services is reliable and sustainable • The reliability of domestic water supplies in dry seasons often depends on inﬂuencing the behaviour of other water users Goal 8: Develop a global partnership for development • Integrated water resource management is one mechanism through which such partnerships can be built, particularly where rivers and lakes are shared between more than one country

The Catchment Management Concept The importance of IWRM and catchment management to the deliver of the MDGs cannot be underestimated. Task force 7 of the Millennium Project recognized that without water resource development and greater attention to the management of water resources (namely, the water as it occurs naturally in lakes, rivers and groundwater), any gains in water services provision were unlikely to be sustained. The task force recommended that an integrated approach to land, water and ecosystem management be adopted in both policy and planning. Clearly water resources are an important component of addressing all of the MDGs. In a European context waters legislation gathered pace in the 1970s and early 1980s, followed by a second wave in the early 1990s. The majority of this water policy was use-based rather than focusing on ecological integrity. In 1995, the Commission was requested by the Council of Ministers and Environment Committee to produce a coherent water policy aiming to draw together much of the existing independent and somewhat disparate Directives (Table 1.2) to provide greater synergy, reduce points of tension

Table 1.2 Timeline of European water and associated policies Year

Milestone

1973 1975 1976 1976 1978 1979 1979 1980 1980 1991 1991 1992 1994 1995 1996 1997 2000

First Environmental Action Programme Surface Water Directive Bathing Waters Directive Dangerous Substances Directive Freshwater Fish Directive Shellﬁsh Waters Directive Birds Directive Groundwater Directive Drinking Water Directive Urban Water Treatment Directive Nitrates Directive Habitats Directive Drinking Water Directive (Revision) Bathing Waters Directive (Revision) Integrated Pollution Prevention and Control Directive Proposal for a Water Framework Directive Water Framework Directive

9

between sister Directives and promote a more integrated approach. In 1997, the Commission proposal for a Water Framework Directive (WFD) was developed and in 2000, the Directive was published (EU 2000). The requirements of some old water legislation (e.g. the Freshwater Fish Directive) have been reformulated in the Water Framework Directive to meet modern ecological thinking. After a transitional period, these old Directives will be repealed. Other pieces of legislation (e.g. the Nitrates Directive and the Urban Wastewater Treatment Directive) must be coordinated in river management plans under the WFD. An additional historical factor has been the dichotomy in the approach to pollution control at European level with some controls concentrating on what is achievable at source through the application of technology, and some dealing with the receiving environment in the form of quality objectives. Each approach has limitations. Source controls alone can allow a cumulative and environmentally detrimental pollution load, where there is a concentration of pollution sources. Quality standards can underestimate the effect of a particular substance on the ecosystem, due to the limitations in scientiﬁc knowledge regarding dose–response relationships and the mechanics of transport within the environment. The WFD addresses the need for a combined approach to pollution management. It does so as follows. On the source side, it requires that as part of the basic measures to be taken in the river basin, all existing technology-driven sourcebased controls must be implemented as a ﬁrst step. But over and above this, it also sets out a framework for developing further such controls. The framework comprises the development of a list of priority substances for action at EU level, prioritized on the basis of risk; and then the design of the most cost-effective set of measures to achieve load reduction of those substances, taking into account both product and process sources. On the effects side, it co-ordinates all the environmental objectives in existing legislation, and provides a new overall objective of good status for all waters, and requires that where the measures taken on the source side are not

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sufﬁcient to achieve these objectives, additional ones are required (WISE 2005). Main objectives of the WFD are to: • protect and where necessary to improve the quality of all inland and coastal waters, groundwater and associated wetlands and to prevent their further deterioration; • promote the sustainable use of water; • enhance protection and improvement of the aquatic environment through speciﬁc measures for the progressive reduction of discharges, emissions and losses of priority substances and the cessation or phasing-out of discharges, emissions and losses of the priority hazardous substances; • lessen the effects of ﬂooding and drought; To deliver this vision requires that: • All water bodies should be protected, including rivers, lakes, wetlands, coastal waters and groundwaters. • The basis for environmental protection to be ecological integrity (except groundwaters) and that timelined targets for achieving ‘good status’ be set. • A combined approach to the management of point and diffuse sources be developed and emissions limit values determined as appropriate. • That the fundamental unit for integrated management should be the catchment or river basin. • Economic and non-use values of water must be factored into sustainable strategies. • People must be intimately involved with the planning process (stakeholders who use or who have an interest in the water environment). • Adaptive management strategies should be adopted including continuous timelined planning cycles. The EU WFD is a complex environmental Directive, but whose principles have been developed to be simple, ﬂexible, and familiar. This is required because the WFD is to be implemented by 25 Member States, Norway, and a number of accession countries – a complex socioeconomic assemblage of cultures, traditions, languages and histories in terms of environmental management and objective setting. The WFD is a major departure from conventional environmental protection legislation in several ways. It obliges Member

States to take a holistic, inclusive, ecological approach to water management. Additionally it requires the quality ‘status’ of water bodies to be measured using ecological rather than just traditional physical and chemical parameters, with more emphasis on the biological quality of a water body. There has been much investment to date in undertaking supporting research on deﬁning the terms of reference for the implementation of the WFD in Europe from an ecological, social and economic perspective (CIS 2000). The WFD functional unit is based on river catchments or collections of river catchments (River Basin Districts), rather than traditional political divisions such as counties or regions. River basins or catchments in this context are made up of lakes, rivers, streams, groundwater and estuaries – and all the land that surrounds them, and drains into them. The WFD thus promotes a more integrated, holistic approach to water management, and one that transcends national borders. Involvement of the public is a key requirement. The WFD puts consultation, public involvement, stakeholder engagement and access to information at the heart of its development and speciﬁcally states that ‘the active involvement of all interested parties’ must be encouraged by every Member State. It is anticipated that this will enhance community, industry and stakeholder input to the management of water resources in Europe, which has previously been weak in many situations. In order to achieve the overall aims the WFD promotes an objective based approach for adaptive management based on a planning cycle composed of four main elements (Fig. 1.2), against a formalized timetable: • Characterization: including an assessment of water bodies at risk of not achieving WFD objectives as a result of man-made pressures. • Environmental Monitoring informed by the characterization. • Setting of Environmental Objectives. • Design and implementation of a Programme of Measures to achieve these environmental objectives.

The Catchment Management Concept

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Fig. 1.2 The WFD adaptive cycle.

• No further deterioration in the ecological quality of aquatic environments. • Achieving ‘good status’ for all waters by 2015. A key concept underlying the WFD is that of integration which is seen as the major focus of management of water protection with the regional planning process, and a consideration of economic aspects of water use (CIRCA 2003; WATECO 2003). These represent a mountain to seas continuum and management must embrace the complete hydrological cycle. This involves: • Integration of environmental objectives, combining quality, ecological and quantity objectives for protecting highly valuable aquatic ecosystems and ensuring a general good status of other waters. • Integration of all water resources, combining fresh surface water and groundwater bodies, wetlands, coastal water resources at the river basin scale. • Integration of all water uses, functions and values into a common policy framework, i.e. investigating water for the environment, water for health and human consumption, water for economic sectors, transport, leisure, water as a social good. • Integration of disciplines, analyses and expertise, combining hydrology, hydraulics, ecology, chemistry, soil sciences, technology, engineering and economics to assess current pressures and impacts on water resources and identify mea-

sures for achieving the environmental objectives of the Directive in the most cost-effective manner. • Integration of water legislation into a common and coherent framework. Integration of all signiﬁcant management and ecological aspects relevant to sustainable river basin planning including those which are beyond the scope of the WFD such as ﬂood protection and prevention. • Integration of a wide range of measures, including pricing and economic and ﬁnancial instruments, in a common management approach for achieving the environmental objectives of the Directive. Programmes of measures are deﬁned in River Basin Management Plans developed for each river basin district – a collection of individual catchments and water bodies speciﬁed by each Member state. • Integration of stakeholders and the civil society in decision making, by promoting transparency and information to the public, and by offering a unique opportunity for involving stakeholders in the development of river basin management plans. • Integration of different decision-making levels that inﬂuence water resources and water status, be it local, regional or national, for an effective management of all waters. • Integration of water management from different Member States, for river basins shared by

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several countries, existing and/or future Member States of the European Union. In America, the Environmental Protection Agency (US Environmental Protection Agency 1993) supported the development of a ‘watershed protection approach’ formulated around catchment management. This approach was to be an integrated, holistic, problem-solving strategy used to restore and maintain the physical, chemical and biological integrity of aquatic ecosystems, protect human health and promote economic growth. USEPA’s watershed approach has three major cornerstones. First is problem identiﬁcation, which identiﬁes the primary threats to human and ecosystem health within the watershed. Second is stakeholder involvement, which involves the people most likely to be concerned or most able to take action. And third is the integration of actions, that is, corrective efforts taken in a comprehensive, integrated manner once solutions are determined. The approach evaluates success and reﬁnes actions as necessary (US Environmental Protection Agency 1993). USEPA views this approach as placing a heavy emphasis on the many elements that affect water quality, including chemical composition (toxics and conventional pollutants), physical water quality (temperature, ﬂow and circulation), habitat quality (channel morphology, composition and health of biologic communities) and biodiversity (species number and range). The approach encompasses all waters – surface and ground, inland and coastal – and is seen as a framework for integrating existing programmes (CGER 1999). The watershed approach was also endorsed by US federal agencies through the Uniﬁed Approach to Federal Land and Resources Management (US Department of Agriculture 2000), embracing many of the concepts of IWRM. They deﬁne the term as a ‘framework to guide watershed management that; uses watershed assessments to determine existing and reference conditions, incorporates assessment results into resource management planning and, fosters collaboration with all landowners in the watershed’. The

framework considers both ground and surface water ﬂow within a hydrologically deﬁned geographical area (US Army Corps of Engineers 2000). During the WSSD in 2002, the EU announced the launching of the global EU Water Initiative. It highlighted the important role of a riverbasin-scale approach to sustainable management, particularly for transboundary catchments. Although there are a number of high proﬁle areas of regional tension water has also been an issue for developing co-operation, removing conﬂict and building joint capacity. Indeed as part of regional development negotiations water may serve as a key criteria in conﬂict prevention (Wolfe et al. 2005). The Danube has the greatest number of countries sharing its length in the world. A total of 18, over twice that of the Rhine (eight), over four times that of the Elbe (four) and six times that of the Rhone (three) (Wolf et al. 1999; ICPDR 2007). Of these, six are EU member states, three are accession countries and seven are not members of the EU. The Danube is 2800 km long, drains an area of 8000,000 km and is home to over 80 million people. The Convention on Co-operation for the Protection and Sustainable Use of the River Danube (Danube River Protection Convention, DRPC) forms the overall legal instrument for cooperation on transboundary water management in the Danube River Basin. The Convention came into force in 1998, and aims to ensure that surface waters and groundwater within the Danube River Basin are managed and used sustainably and equitably. The signatories to the DRPC have agreed to co-operate on fundamental water management issues by taking ‘all appropriate legal, administrative and technical measures to at least maintain and where possible improve the current water quality and environmental conditions of the Danube river and of the waters in its catchment area, and to prevent and reduce as far as possible adverse impacts and changes occurring or likely to be caused’. To implement the DRCP, the Danubian countries agreed when it became operational to estab-

The Catchment Management Concept lish the International Commission for the Protection of the Danube River (ICPDR) to enhance regional co-operation and the implementation of the DRCP. Following the implementation of the WFD, the ICPDR provided a platform for the development and establishment of a River Basin Management Plan for the Danube River in line with the WFD. Although not legally obligated to fulﬁl WFD requirements non-member states have embraced the concept and are working under the auspices of ICPDR under bilateral and multilateral agreements to ensure sustainable management of Europe’s second largest river. In southern Africa more than 70% of surface water resources are shared by two or more states and water is unevenly distributed in the region (Green Cross International 2000). In 1995 the member states signed a Protocol on Shared Watercourse Systems, and as part of that established River Basin Commissions and River Authorities representing the component parts of each major river basin. This recognition of catchment as a central spatial component to resource management provided a platform for co-operation between states. This Protocol is currently under review, with a view to improving political issues surrounding water management. Australia’s variability in rainfall and runoff is similar to that of southern Africa, and although it has a high per capita discharge (20× that of the Middle East) the resource is unevenly distributed both spatially and temporally. Over half the potential resource is located in the north and east of the country (approximately 20% of the land area) where only 15% of the total population reside. The greatest use of water is for irrigration agriculture, some 70% (Smith 2009). Nearly 40% of food production occurs within the MurrayDarling Basin (see Chapter 13). In 2004, an International Agreement on National Water Initiative was initiated by the Australian Governments and ratiﬁed in 2006. The overall objective of the National Water Initiative is to achieve a nationally competitive market, regulatory and planning based system of

13

managing surface and groundwater resources for rural and urban use that optimizes economic, social and environmental outcomes. Other international initiatives such as the Nile Basin Initiative (1999), La Plata River Basin Treaty (1969), the Sustainable Development of the Mekong River Basin and the establishment of the Mekong River Commission (Jacobs 2002), the Ganges Treaty (1966) and the Okavango River Basin Commision (1994) all highlight the importance of considering water resource management issues at a catchment or basin scale. The Network of Asian River Basin Organizations (NARBO), established in 2004, was speciﬁcally established to respond to the challenges of catchment management in the monsoon areas of Asia. The emerging global consensus on the necessity for an integrated approach to water resources has further embedded IWRM concepts in catchment planning. In all of these country and regional examples the general evolution from technical innovation to social engagement to moral attitudes highlights the development of thinking on water resource management from past to present to future, respectively.

1.4 Synthesizing Knowledge for the Future There are different visions on what catchment management actually represents based on sectoral perspectives even of an identical resource (Fig. 1.3). A key for the future is to look to identify synergies in those perspectives and the development of common targets and/or goals. To move this thinking forward a key issue is to identify what we can learn from the past. Global water resource management is not starting from time zero, but must work with the legacy or inheritance of historical sectoral management, legislative frameworks and scientiﬁc understanding. Primarily, production goals have strongly inﬂuenced the pattern of water resources management which we have today, which varies globally dependent upon the distribution of natural resources, climate, economics and social

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robert c. ferrier and alan jenkins

Fig. 1.3 Perspectives of catchment management.

structures. Additionally there is not a ‘one size ﬁt all’ philosophy that can deliver in all locations; rather there is a desire to build generic approaches for sustainable management. Over time, scientiﬁc understanding has begun to unravel the complexities of systems functioning and the importance of a mountain to seas holistic approach to understanding the interdependencies of physical, chemical and ecological process. There is potentially large spatiotemporal variability in the numerous physical processes and interactions within the hydrological cycle alone, before consideration of these changes on biogeochemical processes governing nutrient, carbon, sediment and, potentially, contaminant ﬂuxes within catchments. Similarly, connectivity between stores in both space and time for surface and groundwater is of critical importance and is poorly understood in many situations. In addition, the consequence of some historical actions (such as the long term chemical contamination, or overabstraction of groundwaters) is only now being realized, in some cases

decades to centuries later. Simple questions such as do how changes in land cover and its management inﬂuence soil water distribution, or what role does that soil water play in ecological processes, carbon storage or seasonal recharge still represent unknowns from an environmental science perspective, before factoring in the consequences of current and potential climatic change. Globally the quality of information on our natural environment has grown almost exponentially over recent decades, as has the computing power available to process, interpret and model such data. This, however, has not been consistent across the globe. For example, by their very nature monsoon-fed river systems contain large inherent annual and decadal variability and identifying the consequences of speciﬁc anthropogenic inﬂuences is difﬁcult to capture. Also understanding the current state of aquatic resources and the identiﬁcation of reference conditions in many developing countries is being outstripped by the rate of change in the quality

15

The Catchment Management Concept of these resources due to rapid economic growth and associated environmental degradation. This is particularly so in Asia. The relative importance of engaging in active catchment management in these contexts is of critical importance if appropriate sustainable development is to be achieved. Many emergent issues have only recently achieved enough scientiﬁc understanding to warrant debate and addressing. For example, future restoration of degraded/damaged habits has highlighted the importance of hydromorphology in ecosystem dynamics. Most major river systems have been historically engineered, resulting in a legacy of dams, impoundments, canalization, levee development and riparian management, channel manipulation, land drainage and loss of ﬂoodplain. This removes the ecological integrity and natural links between land and water, affecting overall biodiversity and potential for species migration. It has been estimated, for example, that in the Mississippi and its associated delta only 10% of the original ﬂood plain remains in a natural or semi-natural state (Gore and Sheilds 1995). Hydroecology is now a central component in the EU Water Framework Directive and other developing national legislation. The WFD requires the development of reference condition for each water body type, which equate to the ecological conditions appropriate if the waterbody was in a pristine state, undisturbed by human intervention. These include hydromorphological and physio-chemical condition, such as ﬂow regime (quantity and dynamics of ﬂow), river continuity to allow for migration of species, etc., lake levels and residence times, morphological characteristics, along with appropriate chemical parameters. Although many issues in catchment management have been focused on individual sectoral challenges there is an emerging ethos to move from single issue management to that of multiple issues and to identify potential win:win options. There is also a realization that successful management requires a balance between environmental protection and enhancement, the potential for economic growth, and an awareness of social per-

spectives, attitudes and beliefs. The catchment as a spatial unit for integration provides a focus for systems understanding, linking air, land and water. The quality and quantity of water available to both people and ecology is a barometer of the current state of the environment. What are the drivers for a catchment approach and what is the binding force which makes them operational? As highlighted earlier the catchment concept has been the cornerstone to the development of much environmental legislation especially in developed countries, but this is not necessarily the unifying force in all occasions. Collaborative catchment or watershed institutions represent a new approach to environmental governance. Sabatier et al. (2005) highlight that such collaborative institutions focused on mutually acceptable goals require participation and representation from appropriate stakeholders, a community ethos, legitimacy and wider acceptance, and a survival strategy to ensure perpetuity. Success is measured not by outputs but by outcomes which have tangible beneﬁt and acceptability within the stakeholder community. Local action and engagement, the adoption of voluntary measures, the establishment of community based water ‘champions’ are all examples of mechanisms whereby the catchment concept is being adopted as a common currency in environmental protection, whether laid down in statute books or not. The challenge for the future is to ensure that robust scientiﬁc understanding on the efﬁcacy of single and multiple actions in catchments is available to underpin the decisionmaking process. In a changing world with increasing population pressures and demands on both blue and green water, and an uncertain future driven by potential climatic changes, this has never been so pressing.

References CGER, Commission on Geosciences, Environment, and Resources (1999) New Strategies For America’s Watersheds. Committee on Watershed Management, Water Science and Technology Board, National

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Research Council. National Academy Press, Washington, DC. CIRCA, Communication and Information Resource Centre Administrator (2003) Common Implementation Strategy for the Water Framework Directive (2000/60/EC). Guidance Document No. 11: Planning process. Ofﬁce of the Ofﬁcial Publications of the European Communities, Europa. CIS (2000) Common Implementation Strategy for the Water Framework Directive. http://forum.europa. eu.int/Public/irc/env/wfd/library D’Arcy, B.J., Ellis, J.B., Ferrier, R.C., Jenkins, A. and Dils, R. (eds) (2000) Diffuse Pollution Impacts, The Environmental and Economic Impacts of Diffuse Pollution in the UK. Chartered Institution of Environmental Management (CIWEM), Lavenham Press. EU (2000) Directive 2000/60/EC of the European Parliament and the Council of 23 Oct. 2000 establishing a framework for community action in the ﬁeld of water policy. Ofﬁcial Journal of the European Communities, L327, 1–72. Falkenmark, M. and Rockstrom, J. (2006) The new blue water and green water paradigm: breaking new ground for water resources planning and management. Journal of Water Resources Planning and Management, 132, 129–208. Global Water Partnership (2000) Technical Advisory Committee of the Global Water Partnership, background paper no. 4. Integrate Water Resources Management, 2000. GWP/Sida, Stockholm. Gore, J.A. and Sheilds, F.D. (1995) Can large rivers be restored? Bioscience, 45, 145–152. Green Cross International (2000) National Sovereignty and International Watercourses. 2nd World Water Forum, The Hague, March 2000. Green Cross International. ICFW (2001) Conference Outcomes: The Bonn Keys. International Conference on Freshwater, Bonn, 3–7 December. ICPDR (2007) International Commission for the Protection of the Danube River. http://www.icpdr. org ICWE (1992) Dublin Statement. International Conference on Water and Environment, Dublin, 29–31 December. Jacobs, J.W. (2002) The Mekong River Commission: transboundary water resources planning and regional security. The Geography Journal, 168(4), 135–148. Meybeck, M. (2003) Global analysis of river systems: from Earth systems controls to Anthropocene syn-

dromes. Philosophical Transactions of the Royal Society of London B, 358, 1935–1955. Novotny, V. (2003) Water Quality: diffuse pollution and watershed management, 2nd edition. John Wiley and Sons, New York. Novotny, V. and Olem, H. (1994) Water Quality, Prevention, Identiﬁcation and Management of Diffuse Pollution. Van Nostrand Reinhold, New York. Sabatier, P.A., Focht, W., Lubell, M., Trachtenberg. Z., Vedlitz, A. and Matlock, M. (2005) Swimming Upstream – collaborative approaches to watershed management. MIT Press, Cambridge, MA. Showers, K.B. (2002) Water scarcity and urban Africa: an overview of urban–rural water linkages. World Development, 4, 621–648. Smith, D. (2009) Water resources management. In: Dovers, S. and Wild River, S. (eds), Managing Australia’s Environment. The Federation Press, NSW. United Nations, Department of Economic and Social Affairs, Division for Sustainable Development (2002) Johannesburg Plan of Implementation. UN DEAS, New York. United Nations, Department of Economic and Social Affairs, Population Division (2004) World Population to 2300. UN, New York. United Nations (2005) The Millennium Ecosystem Assessment. United Nations, New York, NY, and Island Press, Washington, DC. http://www.millenniumassessment.org/en/Index.aspx (last veriﬁed Oct 10, 2008). US Army Corps of Engineers (2000) Engineering Regulation (ER) 1105-2-100, Guidance for Conducting Civil Works Planning Studies. CEW-P, HQ-USACE, Washington, DC 20314. US Department of Agriculture, US Department of Commerce, US Department of Energy, US Department of the Interior, US Environmental Protection Agency, Tennessee Valley Authority, and the US Army Corps of Engineers (2000) Uniﬁed Federal Policy for a Watershed Approach to Federal Land and Resource Management. Federal Register 65 (2), 62565–62572. US Environmental Protection Agency (USEPA) (1993) The Watershed Protection Approach. EPA 840-S-93001. US Environmental Protection Agency, Washington, DC. WATECO (2003) Policy summary. Guidance Document No. 1 Economics and the Environment. The implementation challenge of the Water Framework

The Catchment Management Concept Directive Working Group 2.6. Ofﬁce of the Ofﬁcial Publications of the European Communities, Europa. WHO/UNICEF (2004) World Health Organization and the United Nations Childrens Fund (Joint Monitoring Programme) Meeting the MDG Drinking Water and Sanitation Target, A Mid-Term Assessment. Joint Monitoring Programme for Water Supply and Sanitation, Geneva. WISE (2005) Water Information System for Europe. Introducing the new EU Water Framework Directive.

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Wolf, A., Natharias, J., Danielson, J., Ward, B. and Pender, J. (1999) International river basins of the world. International Journal of Water Resources Development, 15, 387–427. Wolfe, A.T., Kramer, A., Carius, A. and Dabelko, D. (2005) Managing water conﬂict and co-operation. In: State of the World 2005: redeﬁning global security. The Worldwatch Institute: Washington DC, chap. 5. WWC (2000) Final Report. Second World Water Forum and Ministerial Conference. Vision to Action. World Water Council, Marseilles.

Image facing chapter title page: Courtesy of the Macaulay Institute.

2

Wetland Management

MIKE C. ACREMAN1 AND J. OWEN MOUNTFORD1 1

Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire, UK

2.1

Introduction

Wetlands occupy the transitional zones between permanently wet and generally drier areas; they share characteristics of both environments yet cannot be classiﬁed unambiguously as either fully aquatic or terrestrial. It is the presence of water for some signiﬁcant period of time which creates the soil, its micro-organisms and the plant and animal communities, such that the land functions in a different way from either aquatic or dry habitats. Mitsch and Gosselink (1993) recognized that ‘hydrology is probably the single most important determinant for the establishment and maintenance of speciﬁc types of wetlands and wetland processes when hydrologic conditions in wetlands change even slightly, the biota may respond with massive changes in species richness and ecosystem productivity’. The international Convention on Wetlands, the intergovernmental treaty established in Ramsar, Iran in 1972 (Davis 1993), provides the framework for national action and international cooperation for the conservation and wise use of wetlands and their resources. The ‘Ramsar’ Convention adopts an extremely broad approach and deﬁnes wetlands as: ‘areas of marsh, fen, peatland or water, whether natural or artiﬁcial, permanent or temporary, with water that is static Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

or ﬂowing, fresh, brackish or salt, including areas of marine water, the depth of which at low tide does not exceed six metres’. Such a deﬁnition thus includes many ecosystems from coral reefs to lakes in underground caves. The Canadian deﬁnition of wetland is more speciﬁc: ‘land that is saturated with water long enough to promote wetland or aquatic processes as indicated by poorly drained soils, hydrophytic vegetation and various kinds of biological activity which are adapted to a wet environment’ (Environment Canada 1991). In general, wetlands include a range of soils (e.g. peat in fens, alluvium in ﬂoodplains and marine clays in estuaries), vegetation communities (e.g. grasslands, forests, mangroves, reedbeds), animals (e.g. ﬁsh, reptiles, amphibians) and microbes (e.g. methane producing bacteria). Many local terms are applied to wetlands, including such general anglicized terms as ‘marsh’, ‘swamp’, ‘bog’ etc., and regionally speciﬁc terms such as aapa mires (rheotrophic mires of the boreal zone), fadamas (ﬂoodplain farmland in Nigeria) and dambos (headwater wetlands in southern Africa). There is no known means of providing a direct association between local terms and hydrological type in a fully inclusive manner. The hydrological regime and associated water quality and landscape location are the dominant processes governing the type of wetland that occurs. In coastal locations, diurnal tidal cycles and sea water salinity create the conditions for the mangrove swamps of the Sunderbans in

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mike c. acreman and j. owen mountford

Bangladesh (Hussain and Acharya 1994); seasonal inundation of the Amazon basin creates ﬂoodplain forests of the igapo (blackwater) and varzea (tidally inﬂuenced sediment-rich) types (McClain 2002); and high groundwater levels created the wetlands of the Azraq oasis in Jordan (Fariz and Hatough-Bouran 1998). Wetlands are naturally transient features of the landscape. In the short term, the dynamic nature of the hydrological regimes of wetlands leads to continuous change in habitat availability. For example, during wet periods, high water levels can create backwaters, sidearm channels, oxbow lakes and temporary pools on ﬂoodplains, which subsequently disappear. In the longer term, many wetlands naturally inﬁll with sediment and vegetation and become gradually drier as they are colonized by ﬂoating and emergent plants and successively fringing and ﬁnally terrestrial plants. As an overlay to their natural dynamics, the type of wetland is often dependent on site management. For example, grazing by horses and cattle has created and maintained wet grasslands, as on the Nene Washes in Cambridgeshire, UK (Benstead et al. 1997), where the natural vegetation would be swamp-forest. Furthermore, some wetlands have been created artiﬁcially by human action, e.g. Rutland Water is a water supply reservoir in the UK that has been designated under the Ramsar Convention as a wetland of international importance. The varied ecological character within and between wetlands results in important biological diversity (Finlayson and Moser 1991; Dugan 1993). For example, wetlands in semi-arid and arid areas, such as the inner Niger delta in Mali (Zwarts et al. 2005), are known as prime areas for biodiversity conservation and as important nursery and feeding areas for many aquatic and terrestrial migratory species. Globally, these freshwater ecosystems of the world support over 10,000 species of ﬁsh and over 4000 species of amphibians (Bergkamp et al. 1998). The varied biogeochemical processes in wetlands give rise to a range of vital functions including groundwater recharge, ﬂood storage, carbon sequestration and pollutant removal. In many

developing countries, wetlands have particular importance as major producers of food (e.g. rice and ﬁsh), fuel-wood and medicines. The functioning of wetlands has both direct economic and other ﬁnancial values that beneﬁt human welfare, health and safety. The economic valuation of wetland functions is a rapidly developing ﬁeld of study. Historically, river basin development has involved the eradication of wetlands by direct drainage and conversion to intensive agricultural land or urban development or by the construction of dams and river diversion works and exploitation of groundwater. This resulted from a view of wetlands as ‘wastelands’ (Maltby 1991). Such limited or biased perception of wetlands combined with a failure to appreciate the value of wetlands in a largely uni-sectoral planning process have been root causes of the substantial losses of wetlands and wetland functions that have resulted. Wetland destruction is still institutionalized in some countries, e.g. Turkey. In recent years, the full value of wetlands has been much more widely recognized. Since wetlands cover around 6% of the land surface of the earth (OECD 1996) and many exist in the upstream parts of river catchments, the total downstream area over which a hydrological inﬂuence may be exerted is substantial. Wetlands are now being conserved, restored and, in some cases, created with a particular focus on their wildlife and biodiversity value. By capitalizing on valuable functions, wetlands have also become central components within catchment management, such as: • groundwater recharge – Garaet Haouaria, Tunisia (SCET 1962); • ﬂood control – Charles River in Massachusetts (US Corps of Engineers 1972) and the Mississippi ﬂoodplain (Parrett et al. 1993); • shoreline stabilization – Louisiana (Costanza et al. 1989) and Brisbane Airport (Hamilton and Snedaker 1984); • sediment and toxicant removal – Pyl Brook, London (Cutbill 1993); • nutrient retention – Kattegat coast, Sweden (Eriksson 1990);

Wetland Management • waste water treatment – Calcutta (Ghosh and Sen 1987) and Kampala (Emerton et al. 1999); • ﬁsheries – mangroves; agricultural resources – Senegal Delta (Hamerlynck and Duvail 2003); and • biological diversity (McAllister et al. 1997). Conservation and restoration, particularly of peat wetlands, is being considered as a primary strategy for coping with climate change (Colston 2003). 2.1.1 The wetland ecosystem The difﬁculty in providing anything like a satisfactory comprehensive deﬁnition of ‘wetland’ arises directly from the wide diversity of wetland type, whether that understanding of type is based upon ecological or hydrological criteria. Wetlands arise as the product of complex interaction between water, soils (and geology), vegetation, fauna and microbial activity. These factors are further inﬂuenced and modiﬁed by human usage. Wetlands cover a range of environments with very different species and interactions, and even within one wetland type, the vegetation communities and faunal assemblages can vary markedly. Although water is clearly the key determining factor in the occurrence of wetlands, the diversity

21

of wetlands is inﬂuenced by many other environmental factors. Keddy (2000) discusses three principles in identifying and understanding this diversity of factors: 1 Any particular community or ecosystem is produced by multiple environmental factors acting simultaneously. 2 To understand and manage wetlands, scientists must determine the quantitative relationships between environmental factors and the properties of wetlands. 3 The multiple factors that produce a community or ecosystem will change through time. Working within these principles, some rationalization of the multitude of classiﬁcation systems for wetlands is possible, and six broad types may be proposed on the basis of the plants and animals (vegetation types and animal associations) that occur (Box 2.1). These categories may then be related to the hydrological mechanism that supplies water to the wetland. Regardless of whether the classiﬁcation is essentially biological or hydrological, there is great regional variation in the terminology used (Table 2.1). There are weaknesses in the employment of such wide categories and other approaches have been attempted, e.g. through deﬁnition of broad wetland habitats (European Topic Centre (Biodiversity) 2007) or strictly through vegetation

Table 2.1 Wetland types deﬁned by landscape location and water-supply mechanism compared with the terms applied to such wetlands in different parts of the world and different situations Landscape location and hydrological mechanism Rain-fed, ﬂat land Floodplain Surface water depression Surface water slope Groundwater depression Groundwater slope Estuarine Near-shore General

Local terms Blanket bog, domed bog, raised mire Sudd, river valley, swamp, ﬂoodplain, valley bog, marsh, cypress swamp, delta (inland), dambo, fadama, inland valley swamps and bolis, pantanal Bog, prairie pothole, slough, kettle-hole, shallow depressions Perched bog, peat, peat bog, pond, raised bog, blanket peat, muck, pocosin, saturated heath, palsa, aapa-mire, inﬁlled lakes, constructed wetland Marsh, prairie pothole, fen, pond, bog Swamp, marsh, river valley, bog, dambo, fen bogs, peat bog, peat, wetland, pocosin, cypress domes, peat fen, pakahi, fen, mire, quaking fen, pond Estuaries, mud-ﬂats, salt marshes Dune slacks, near-shore bays, rock platforms, coral reefs Swamp, bog, marsh, ‘basin storage’, ponds, hydromorphic gleys, forest wetland

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mike c. acreman and j. owen mountford

Box 2.1 Six basic types of wetland community (after Keddy 2000) 1 Swamp: dominated by trees rooted in hydric soils, but not in peat, e.g. mangroves and Taxodium swamps 2 Marsh: dominated by herbaceous plants that are (usually) emergent through water and rooted in hydric soils, but not in peat, e.g. beds dominated by Phragmites australis or Typha species 3 Bog: dominated by Sphagnum species, Cyperaceae, Ericaceous shrubs and/or evergreen rooted in deep peat, e.g. blanket bogs of boreal latitudes and ﬂoating bogs by temperate and boreal regions. The water is normally of low pH 4 Fen: dominated by Cyperaceae and Poaceae rooted in shallow peat, often with alkaline or at most mildly acidic water showing considerable movement through the peat, e.g. extensive peatlands of the Holarctic 5 Wet meadow: dominated by herbaceous plants rooted in occasionally ﬂooded soils. The temporary ﬂooding effectively excludes truly terrestrial species, whilst the dry phase excludes plant communities of moist soils. Such meadows may be created by grazing or cutting in situations that would normally support other wetland types 6 Shallow water: dominated by obligate aquatic plants growing in ≥25 cm of water, e.g. littoral zone of lakes and bays by rivers

composition (e.g. Rodwell 1991–2000). The EUR27 interpretation manual (European Topic Centre (Biodiversity) 2007) does indicate in some cases which phytosociological alliances and associations are contained within a habitat type, but the correspondence between the approaches is often inexact, with some alliances at least apparently occurring in more than one wetland habitat type. Understanding of wetland pattern and process requires examination of both hydrological type and of the zonation of associated wetland com-

munities. Hydrologically deﬁned wetlands can be subdivided into particular plant communities on the basis of, for example, soil type or grazing/ cutting regime. Wetland soils vary not only in their chemistry, texture and organic content, but in factors linked to the water regime such as speciﬁc yield and hydraulic conductivity. Movement of water through the soil (either laterally or vertically) inﬂuences the seasonal soil aeration and thus the composition of fauna and ﬂora depending on the tolerances of each species to stresses from lack of oxygen (aeration stress) or insufﬁcient water (drought stress), as demonstrated and quantiﬁed in the work of Barber et al. (2004) and Wheeler et al. (2004). More consistent systems of wetland classiﬁcation reﬂect the primacy of water – its quantity, quality, temporal and spatial distribution. Wetland function is in turn signiﬁcantly controlled by the position of the wetland in the landscape, especially with regard to river catchments and associated lakes (Maltby et al. 2005), e.g.: • Wetlands surrounding the headwaters of a river inﬂuenced by the waters reaching the river through groundwater or overland ﬂow. • The river waterbody itself and the riparian zone inﬂuenced by its ﬂow regime, e.g. through inundation of its ﬂoodplain (i.e. a hydro-morphological element of the river). • Lake water-bodies together with their littoral zone (a hydro-morphological element of the lake). • Small surface-water bodies (e.g. ponds, bogpools, rills and springs) linked to a major surface waterbody (lake or river). • A transitional waterbody (e.g. estuary) with adjacent wetlands of the inter-tidal zone (i.e. hydro-morphological element of the estuary). • The coastal waterbody (littoral zone of sea). In addition to these clearly inter-connected wetland elements in the landscape, there are also terrestrial wetlands that are directly dependent on bodies of groundwater. The development of a functional classiﬁcation of wetlands begins with the understanding that the water regime directly affects major functional characteristics of wetland ecosystems, as outlined above, i.e. carbon and nutrient cycling,

Wetland Management primary production and vegetation composition. From this basis, Wheeler (1994) advocated a classiﬁcation comprising: (i) the situation types, i.e. arrangement of the landscape (and its elements) within which the wetland occurs; and (ii) the mechanism of water supply to the wetland. This system has been further expanded through the derivation of WETMECS (Wetland Water Supply Mechanisms) and Ecological Types (Wheeler and Shaw 2001), and thence used to provide guidelines for the management and restoration of wetlands (Wheeler et al. 2004). An attempt to link ecosystem structure, hydrology and functioning was made by Brinson (1993) whose hydrogeomorphic classiﬁcation had three components: (i) the geomorphic situation; (ii) the water source and its transport; and (iii) hydrodynamics. Acreman and Miller (2007) developed a coherent classiﬁcation of wetlands based upon their position in the landscape and the water-transfer mechanism, i.e. the way that water can move into or out of a wetland (Table 2.2). The methods of water transfer used in this classiﬁcation include: • Surface water input mechanisms – precipitation, runoff, lateral inﬂow, over-bank ﬂow (from a waterbody), spring-fed and tidal inﬂow. Table 2.2 Wetland landscape location types and hydrological subtypes (after Acreman 2004) Landscape location Flat upland wetlands Slope wetlands

Valley bottom wetlands

Underground wetlands Depression wetlands

Flat lowland wetlands Coastal wetlands

Subtype based on water-transfer mechanism Upland surface water-fed Surface water-fed Surface- and groundwater-fed Groundwater-fed Surface water-fed Surface- and groundwater-fed Groundwater-fed Groundwater-fed Surface water-fed Surface- and groundwater-fed Groundwater-fed Lowland surface water-fed Surface water-fed Surface- and groundwater-fed Groundwater-fed

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• Groundwater input mechanisms – discharge and seepage. • Output mechanisms – evapo-transpiration, drainage, outﬂow, tidal outﬂow and groundwater recharge. and in addition: • Pumping (by human agency) either as an input or output mechanism. Within this classiﬁcatory framework, certain distinctive wetlands evolve in particular landscape situations and in relation to the catchment and its management. For example: 1 Peat formation may be initiated in emergent marshes (e.g. reed-swamps). As peat accumulates, the inﬂuence of mineral-rich water declines, the community evolves from a fen into a bog and the water table rises in the process of paludiﬁcation (van Breeman 1995). Where there is minimal water ﬂow (as on ﬂat uplands around watersheds) and consequently where rates of oxidation are low, the peat forms a raised mire that is controlled by climatic factors rather than local site factors. Such a mire is ombrotrophic, i.e. fed by precipitation. 2 Floodplains support a radically different type of wetland to the raised bog. Here the water regime is determined by frequent over-bank ﬂow from the river onto the ﬂoodplain and then outﬂow into the river (Duranel et al. 2007). Depending on the particular ﬂoodplain and the permeability of the underlying strata, other transfer mechanisms may include drainage, lateral ﬂow, surface runoff and groundwater discharge or recharge, as well as the transfer mechanisms of precipitation and evapotranspiration that apply to all wetland situations. Flood tolerance of biota is a primary constraint on wetlands (Keddy 2000). The seasonality and duration of the ﬂooding, coupled with the related inputs of nutrients and sediment, determine the ﬂoral and faunal assemblages that develop. In many instances, the exposed phase will correspond with the growing season for vegetation, with most species of plant and animal inactive/dormant during the ﬂooded period. 3 The alternation of inundation and exposure observed on ﬂoodplains often takes place over a seasonal cycle, with long wet and dry periods, e.g.

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mike c. acreman and j. owen mountford

the winter ﬂoods and summer dry phase typical of temperate riverine ﬂoodplains. Other wetland types have evolved where the inundation–exposure cycle is much shorter and more frequent, as in tidal situations along coasts and estuaries. Fauna and ﬂora in these saltmarsh situations must be adapted to (possibly diurnal) inundation with saline water interspersed with a period of exposure to the atmosphere. The relative duration of these phases and the frequency of the cycle will differ between the higher-lying (landward) parts of the saltmarsh and the lower (seaward) zones. The position of the wetland in the landscape is not only important in terms of individual wetland function, but also in terms of connectivity between wetlands at both the landscape and global scales. For example, ﬂyways for some migratory birds depend on sufﬁciently frequent wetlands for feeding and roosting/shelter along the route. Damage to any one of these ‘staging posts’ on the ﬂyway may result in modiﬁcation of the route or even relocation of the whole ﬂyway. Successful conservation management of wetlands for birds may therefore be necessary at the global scale, rather than simply at the landscape or catchment scale. 2.1.2 Human inﬂuence There is an overlay of human inﬂuence on the natural eco-hydrological system, a process of change that began with the earliest agricultural societies, and has become more extensive with industrialization and the growth of human populations. Human inﬂuence has included largescale destruction through systematic drainage and channelization of rivers, modiﬁcation through management and sustainable usage (e.g. Hill 1976; Mountford 1994). Within the temperate zone and especially within regions of more developed agriculture, those wetlands that survive include: (i) relict types that occur in (semi-)natural situations; (ii) modiﬁed wetlands that depend on regular human intervention to maintain their extent; and (iii) artiﬁcial wetlands created by humans to perform particular func-

tions, e.g. canals, reservoirs and ponds (both agricultural and ornamental). In the context of the developed and exploited landscape, the semi-natural and modiﬁed wetlands are especially liable to change and disappearance through both successional processes (terrestrialization) and anthropogenic inﬂuence. Such wetlands (e.g. ﬂoodplain meadows and reedbeds) illustrate how human inﬂuence has determined their occurrence and their prospects for survival. (a) Floodplain and water-meadows In Europe, lowland wet grasslands are often created by the partial reclamation of natural ﬂoodplain wetlands. Control of the ﬂooding regime was combined with removal of the swamp, marsh or fen vegetation and their replacement by wet meadows that are maintained by cutting for hay and/or grazing for agriculture (Benstead et al. 1997). This management regime prevented succession to scrub and woodland, or reversion to the original wetland type. In Britain, ﬂoodplain meadows were once highly prized as agricultural land and were carefully managed over centuries, since the nutrient-rich silt they received from river ﬂooding enabled them to sustain high hay yields (Gowing 2005). This continuity of traditional management produced a species-rich plant community (as much as 40 species m−2), but from 1940 to 1980 over 90% of their area was lost primarily due to agricultural intensiﬁcation intended to exploit their rich soils for arable crops. Since 1980, the remaining areas have been the focus of nature conservation action in the UK, with recognition of their importance at a European level (Natura 2000 habitat type 6510 Lowland hay meadows – European Topic Centre (Biodiversity) 2007). Within ﬂoodplain meadows, one distinctive type was the classic water-meadow, an engineered system of shallow surface-water channels that distributes water through the grassland. The system of channels is ﬁlled via weirs that can be used to direct water from the river to supply the meadows as and when required, to provide irrigation, protection from frost and inputs of nutrient-

Wetland Management rich sediment. In Britain, one of the most species-rich of lowland wet grasslands, the Senecioni–Brometum racemosi, is especially associated with old water-meadow systems (MG8 – Rodwell 1991–2000). (b) Conserving reedbeds (Phragmition) Although often very species-poor when judged botanically, reedbeds are nonetheless of considerable conservation importance for many animal groups, most obviously birds but also mammals, ﬁsh and several invertebrate phyla. As such, communities dominated by Phragmites australis, Schoenoplectus spp. and Typha spp., have been the subject of considerable research on the water regimes required for their conservation (Wheeler et al. 2004) or the general management programmes needed to arrest succession to fen and thence to scrub and woodland (Hawke and José 1996). Natural reedbeds occur in the littoral zone of lakes and rivers, especially in the lowlands. Reedbeds also develop in semi-natural situations as, for example, when abandoned peat-diggings are ﬂooded and succession to marsh and fen commences, e.g. within the Somerset Levels and Moors at Ham Wall (Royal Society for the Protection of Birds 2007). The conservation goal in such sites is to maintain the extent of reed, i.e. ‘freeze’ succession at one seral stage. Therefore, the main management intervention has been to remove any woody species (Alnus glutinosa, Salix spp.) that colonize the reedbed, and which threaten to accelerate the terrestrialization of the site. 2.1.3

Human inﬂuence at a global scale – climate change

Detailed reviews of the impact of climate change on wetland and aquatic ecosystems in Europe have been produced in recent years (e.g. Mitchell et al. 2006; Clément and Aidoud 2007). Most of the published material comprises broad predictions derived from models – those experimental studies on particular taxa that do exist have limited general relevance and are difﬁcult to scale up to the whole ecosystem. There remains

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uncertainty as to the impact of large-scale climatic changes, especially with respect to the key role of climate as a controlling factor in determining ecosystem attributes (Weltzin et al. 2000). Climate change affects ecosystem dynamics, community productivity and composition, and these in turn affect both the trophic structure of wetlands and their resource dynamics, with feedbacks to climate, the wetlands themselves and to associated habitats, as well as the ecotones between wetlands and other habitats (Chapin 2003). Mitchell et al. (2006) summarize those major environmental factors that affect water and wetlands and which will be altered by climate change (e.g. Mitsch and Gosselink 1993; Cannell et al. 1999; Hossell et al. 2000; Keddy 2000; Moss et al. 2003; Thompson et al. in press): • Carbon ﬂuxes – CO2 and methane. • Nitrogen mineralization and denitriﬁcation. • Precipitation patterns – amounts, seasonality and spatial distribution. • River ﬂows – quantity, timing, duration, frequency and quality, including physical quality (e.g. temperature) and chemical quality (e.g. pH, suspended sediment load). • Water supply mechanisms to wetlands, e.g. impacts on groundwater recharge, ﬂooding regimes and evaporation, timing and seasonal periodicity of water-regime ﬂuctuations. • Biological patterns of activity and the ﬂora/ faunal composition of the habitats themselves and those associated with and affecting water bodies and wetlands. • Stratiﬁcation of deeper water bodies and oxygen supply. • Primary productivity of aquatic algae in lakes. • Altered demand by human populations for water abstraction and land drainage. Although such changes at global, regional and catchment scales might overwhelm or obscure the responses of individual species (in terms of their preferred climatic amplitudes), it is clear that the attributes and performance of particular species would also be affected, e.g.: • Growth and productivity of dominant species.

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• Tolerances of individual species to dissolved materials, oxygenation and sediment loading (all inﬂuenced by altered hydrology and especially runoff regimes). • Tolerance and adaptability of individual species to changing ﬂow regimes. • Soil moisture requirements of individual species. Such changes are likely to have immediate impact in marginal and ecotonal wetland habitats, where small changes in topography and elevation may result in different soil-moisture regimes (Silvertown et al., 1999; Clément and Aidoud 2007). Mitchell et al. (2006) further summarize the direct effects of climate change on aquatic and wetland habitats, documenting (with supporting evidence) the impacts of a range of predicted climatic changes for temperate Europe, i.e. increased summer temperature, increased winter temperature, earlier onset of spring, increased summer drought, wetter winters, sea-level rise and increased ﬂooding. The impacts can be differentiated in terms of particular wetland types (peatland, ﬂoodplain wetlands, lakes, pools and rivers) and in terms of whether the climatic change is likely to affect ecosystem function or individual species survival and performance. Climate change will directly affect the functioning of rivers, lakes, pools and wetland habitats, resulting in changes in: • Phenology: mediated through both water and air temperatures and leading to changes in timings and rates of larval development and loss of synchronicity. • Distribution: will occur in response to alterations in hydrological conditions and/or temperature. • Community structure: have been observed across the full range of freshwater aquatic habitats. • Ecosystem function: may result from alteration in rates of microbial activity leading to changes in nutrient availability and possible release of greenhouse gases. The effects of climate change on the management of wetlands within catchments will also be mediated indirectly, where changes in the hydro-

logical regime result in changed (often increased) need for ﬂood-management or drought control and, in some areas, managed retreat. Flood barriers may need to be re-engineered with associated impacts on wetlands in the catchment. Under increased drought, catchment water-resource management could witness competition for water between human usage (agriculture, industry, home use and potable supplies) and the needs of wetlands. Managed re-alignment in the coastal zone may offer opportunities to compensate for loss of wetland habitat through habitat creation, but some ‘squeeze’ is likely between rising sealevels and the need for ﬂood control for housing and industry resulting in net loss of wetlands. There may also be increased visitor pressure on wetland, lakes and river habitats with impacts on water quality and the species supported and impacts on the wetland habitat and species due to disturbance (e.g. through noise and waves). 2.1.4 Setting objectives for wetlands The philosophy of nature conservation or biodiversity protection has gone through a number of changes during the century since the birth of ecology as a scientiﬁc discipline (Sheail et al. 1997). In the past 25 years, the focus has become increasingly on the ecological restoration of habitats. The reﬁnement of techniques for the construction or rehabilitation of habitats and ecosystems, however, has also elicited concern about what the goals of conservation should be. Should one seek to protect and preserve the status quo and focus all attention on the remaining natural wetlands that exist, or do artiﬁcial and reconstructed wetlands have a key role to play in successful and sustainable conservation of biodiversity? This debate extends to discussing how much intervention is ‘justiﬁable’ in achieving nature conservation goals: should one manage habitats as ‘wilderness’ minimizing the human intrusion and allowing natural processes to shape habitats, or is regular and frequent interference acceptable (even to the level of habitat ‘gardening’) if it achieves increased extent of wetland

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Box 2.2 Generic objectives for programmes and projects for the ecological conservation and restoration of wetlands (after Maltby et al. 2005 and Treweek et al. 1993) I

Deﬁnition of overall aims

II Identiﬁcation of target habitats, communities and species 1 Deﬁnition of wetland habitat types and associated species 2 Identiﬁcation of those which are rare, declining and threatened (i.e. ‘in need of conservation or restoration action’) III

Redeﬁnition of targets to take account of regional variation

1 Assessment of current range/distribution of target wetland habitats and species 2 Deﬁnition of physical circumstances in which target habitats and species naturally occur IV

Identiﬁcation of habitat requirements

1 Identiﬁcation of habitat requirements of target species and communities 2 Deﬁnition of actions necessary to satisfy those requirements V 1 2 3 4

Implementation of conservation and/or restoration strategy Formulation of conservation/restoration prescriptions Deﬁnition of criteria for assessing achievement of management objectives Deﬁnition of time-scales within which objectives are to be achieved Implementation of conservation/restoration programme

VI 1 2 3 4 5

Monitoring and reporting

Deﬁnition of appropriate indicators of wetland function and health Integration of indicators into practical monitoring plan Deﬁnition of frequency of monitoring and overall time-scale Machinery for feedback of monitoring results into management of programme/project Design and implementation of reporting procedure to all stakeholders

habitats and higher populations of wetland species (Colston 2003; Hughes et al. 2005). The problem comes into sharp focus when attempting to set realistic objectives for wetland management at a catchment scale especially where construction of roads and towns or intensiﬁcation of agriculture has fragmented the landscape. Wetlands are transient, and effective conservation should allow for their natural establishment, evolution (through succession) and decay, though this is hardly possible in developed countries. Wetland types vary in their complexity and may be provisionally ranked on the basis

of the ease with which they can be restored. Thus, complex wetlands may be impractical to restore and should have the highest priority for conservation management. Other wetlands where restoration techniques are established may be allowed to terrestrialize, at the same time as ‘replacement’ examples are re-created. It is important to assess each situation, each individual wetland and each catchment to determine priorities for conservation and restoration. Generic objectives for programmes of conservation management and/or restoration have been proposed (Box 2.2).

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mike c. acreman and j. owen mountford 2.2 Functions and Value of Wetlands 2.2.1

What makes wetlands valuable within a catchment?

The features that give wetlands their value can be divided into components, functions and attributes (Dugan 1990). Examples are given in Table 2.3. The component parts of wetlands are often used as goods or products that have direct economic values, such as ﬁsh for food, reeds for thatching, peat for fuel and horticulture, grass for cattle fodder, wildlife for tourism and water for public supply. The physical, chemical and biological processes that occur in wetlands give rise to functions that provide services to people (Table 2.4). The case for wetland conservation is often made in terms of ecosystem functioning, which

Table 2.3 Wetland components, functions and attributes Components Reeds Peat Hay Wildlife Fish Water

Table 2.4

Functions

Attributes

Groundwater discharge Flood and ﬂow control Shoreline/bank stabilisation Sediment retention Nutrient retention Archaeological preservation Carbon sequestration

Biological diversity Landscape

results in a wide range of values including groundwater recharge and discharge, ﬂood ﬂow alteration, sediment stabilization and water quality improvement (Maltby 1991). Methods of assessing the functions and services provided by wetlands have been developed in the USA (Adamus and Stockwell 1983; Novitzki et al. 1993) and Europe (Maltby et al. 1996). In particular, wetlands are said to perform ‘hydrological functions’; to ‘act like a sponge’, soaking-up water during wet periods and releasing it during dry periods (e.g. Bucher et al. 1993); these functions provide services of ﬂood reduction and water resource augmentation. However, there is a lack of detailed scientiﬁc evidence for some functions, such as the ‘sponge’ hypothesis, and not all wetlands perform all functions (Bullock and Acreman, 2003). Indeed some wetlands perform functions contrary to the normally cited list; for example some headwater wetlands are ﬂood generating areas (Hewlett and Hibbert 1967; Burt, 1995). Further attributes of wetlands arise from the combination of components and functions that produce biological diversity and landscape, which generate natural heritage, spiritual attachment and intrinsic value. In the UK alone, over 3500 species of invertebrates, 150 species of aquatic plants, 22 species of ducks and 33 species of waders have been identiﬁed living in wetlands (Merritt 1994). Detailed studies of wetland nature reserves suggest greater diversity, e.g. Wicken Fen (UK) where by 2007 over 7800 species had been recorded, the great

Wetland processes, functions and services

Processes Denitriﬁcation and other transformations of nutrients in plant cells Downward movement of water by gravity from the wetland into underlying strata Movement of water from rivers during high ﬂows to inundate a ﬂoodplain surface, ﬁll depressions and saturate the soil Production of peat from dead plant material in anaerobic conditions

Functions

Services

Uptake from water and use of nutrients by plants Recharge of aquifers Flood water storage

Improvement in water quality through reduced nutrient Augmentation of groundwater resources available for human use Reduction in downstream ﬂood risk

Carbon storage

Reduction in greenhouse gases

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Wetland Management Table 2.5 htm)

Major species groups at Wicken Fen nature reserve (from http://www.wicken.org.uk/wildlife_species. Species group Plants: algae Plants: bryophytes Plants: vascular plants Lichens Fungi Mammals Birds: breeding Birds: non-breeding Fish Amphibians Reptiles Mollusca Crustacea Araneae Myriapoda Hymenoptera: Aculeata Hymenoptera: Parasitica Hymenoptera: Symphyta Hemiptera Trichoptera Ephemeroptera Odonata Lepidoptera I Lepidoptera II Lepidoptera III Coleoptera Diptera

Vernacular name/example

Number of species

Algae and diatoms Mosses and liverworts Trees, wild-ﬂowers, grasses Lichens Brackets, toadstools Mice, deer, voles, etc. Regular and occasional Wintering, passage migrants, vagrants Rudd, roach, pike, etc. Frogs, toads, newts Lizards and snakes Snails, bivalves Woodlice, water ﬂeas Spiders Millipedes, centipedes Bees and wasps Parasitic solitary wasps Sawﬂies Bugs and aphids Caddis-ﬂies Mayﬂies Dragonﬂies, damselﬂies Butterﬂies Larger moths Smaller moths Beetles Flies

327 115 420 41 399 30 93 138 20 4 3 88 116 251 20 180 392 116 343 53 13 24 35 450 586 1473 1891

Note: Other taxonomic groups for which there are lists of species recorded, but which are not included above, are Collembola, Dermaptera, Neuroptera, Megaloptera, Mecoptera, Psocoptera, Siphonaptera, Rotifers, Triclads and Annelida.

majority of which are speciﬁc to wetland ecosystems (Table 2.5). Mitsch and Gosselink (1993) refer to wetlands both as ‘the kidneys of the landscape’, since they cleanse polluted water and play an important role in chemical cycling, and as ‘biological supermarkets’ because of the extensive food chain and rich biodiversity they support. Combinations of components, functions and attributes are also important, such as archaeological aspects of wetlands; the Sweet Track, for example, is the oldest built roadway in the world (3806–3791 bc) and has been preserved in the Somerset Levels and Moors wetland, UK (Coles 1990).

2.2.2 Wetland functions and catchment management Wetland functions provide many potentially sustainable solutions to catchment management issues, including ﬂood risk, water supply and efﬂuent disposal and thus to support and maintain economic development (Davies and Claridge 1993). The lateral extent of wetlands and their often rough vegetation has an important hydraulic effect, slowing down water movement and absorbing energy during river ﬂoods. Modelling of the River Cherwell, UK (Acreman et al. 2003) showed that removal of embankments,

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Fig. 2.1 Observed ﬂows on the River Cherwell (dashed line) compared with modelled ﬂows removing ﬂoodplain storage by embanking the river (solid line).

separating the river from its ﬂoodplain, resulted in a reduction in downstream ﬂood magnitude of 132% (Fig. 2.1). On the coast, mangroves and to a lesser extent salt marshes, can stabilize shorelines by reducing the energy of waves and currents and retaining sediment with their roots. For example, it was notable that coasts with mangroves were less damaged during the Asian tsunami in December 2004 than those where these wetlands had been removed. The World Conservation Union (IUCN) compared the death toll from two villages in Sri Lanka that were hit by the tsunami; two people died in the settlement with dense mangrove and scrub forest, while up to 6000 people died in the village without similar vegetation (Weeratunga 2005). Wetlands reduce nutrients by encouraging sedimentation (Karr and Schlosser 1978; Johnston et al. 1984), sorbing nutrients to sediments (see Khalid et al. 1977), taking up nutrients in plant biomass (Lee et al. 1975) and enhancing denitriﬁcation (Lowrance et al. 1984). Because toxicants (like pesticides) often adhere to suspended matter, sediment trapping frequently results in water quality improvements. Water quality can also be improved by the ability of wetlands to strip nutrients (nitrogen and phosphorus) from water ﬂowing through them. This has led to wetlands being managed or constructed to act as buffers or for treatment of domestic or industrial waste (Allinson et al. 2000). Fisher and Acreman (2004)

examined results of studies of 57 wetlands from around the world to assess whether wetlands affect nutrient loading of waters draining through them, and showed that the majority of wetlands reduced nutrient loading (nitrogen or phosphorus). Some wetlands however, can increase nutrient loadings for short periods by releasing soluble N and P species, thus potentially driving eutrophication in the receiving waterbody. The Nakivubo Swamp covers 5.5 km2 on the edge of Lake Victoria in Uganda. The wetland’s catchment drains 40 km2 of land dominated by industrial and urban areas of Kampala (Kansiime and Nalubega 1999). The swamp acts as a buffer for the waste water and untreated sewage from the catchment that would otherwise enter the lake. It is so effective at removing nutrients and pollutants that the intake for Kampala’s water supply is only 3 km away downstream. Thus the wetland is providing ‘free of charge’ efﬂuent and water supply treatment for Kampala and Lake Victoria. 2.2.3 Economic value of wetlands Many catchment management decisions are made on the basis of economic costs and beneﬁts of different development options. In an attempt to include wetland goods and services in this process, methods to value wetland functions and products have advanced rapidly (Bardecki 1984;

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Wetland Management Table 2.6

Economic values of wetlands

Wetland

Goods and services

Flood attenuation Groundwater recharge Nutrient cycling Carbon sequestration Fishing, grazing, recession agriculture

Turpie et al. (1999)

Fish, plants, livestock grazing, timber Recreation, amenity, existence

$8 million per year $176 million per year

Calavito (2002) Pyo (2002)

Flood attenuation

$5 million per year

Waste water treatment

$2 million per year

Emerton and Kekulandala (2002) Emerton et al. (1999)

Fishing, agriculture and fuel-wood

Indus delta mangroves, Pakistan

Fisheries

Barotseland ﬂoodplain, Zambia

Nakivubo swamp, Uganda

Reference 3

US$32 per 1000 m of water US$100 million foreign exchange in 1997 $0.4 million $5.2 million $11.3 million $27 million $2.5 million per year

Hadejia-Nguru wetlands, Nigeria

Waza-Logone ﬂoodplain, Cameroon Hail Haor wetland, Bangladesh Coastal wetlands, Youngsan, Korea Muthurajawela Marsh, Sri Lanka

Economic value

Farber and Costanza 1987; Pearce and Turner 1991; Abramovitz 1997; Costanza 1997; Barbier et al. 1997) and the cost effectiveness of wetlands in catchment management is increasingly being recognized. One of the earliest and often quoted assessments of the economic value of wetlands is of the ﬂood control function of 3800 ha of mainstream wetlands along the Charles River, USA. The US Corps of Engineers (USCE) (1972) estimated that had they been ﬁlled, as so many other wetlands in the USA have been until a change of government policy, the increase in annual ﬂood damages downstream would have been US$17 million. The study valued the retained natural wetland at $1203,000 per year, this being the difference between present annual ﬂood losses based on current land use and the projected ﬂood losses in 1990 if 30% of the wetlands had been lost. Many estimates of economic value have been undertaken in developing countries where the livelihoods of rural communities are very frequently directly linked to wetlands (Table 2.6). Barbier et al. (1991) demonstrated that the net economic beneﬁts of ﬁshing, agriculture and fuelwood from the Hadejia-Nguru wetlands in Nigeria were US$32 per 1000 m3 of water, whereas

Barbier et al. (1991) Meynell and Qureshi (1995)

Loth (2004)

returns from crops grown on the Kano river project (to which water upstream of the wetlands is diverted) were only US$0.15 per 1000 m3. The fertile ﬂoodplain wetlands of the lower Senegal River, which forms the frontier between Senegal and Mauritania, are of vital social, economic and ecological importance in this semi-arid Sahelian fringe of the Sahara desert. Natural inundation covers up to 250,000 hectares of the ﬂoodplain which itself supported up to 125,000 hectares of ﬂood recession agriculture (including maize, beans, watermelon, potatoes and millet), grazing, forests (which provide fuel-wood and construction timber), ﬁsheries and wildlife habitat. The economic value (at 1995 prices) of the ﬂoodplain has been estimated (Acreman 2003) at US$56– 136 per hectare for recession agriculture, US$140 per hectare for ﬁshing and US$70 per hectare for grazing.

2.3 The Political and Legal Setting For much of history wetlands have been perceived as unhealthy waste-lands, harbouring disease, insects and even monsters. Drainage and

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destruction of wetlands became accepted practices throughout the world as demand for agricultural and urban land increased. The battle against malaria often involved wetland drainage. However, as Montes and Bifani (1991) showed for Spain in the 1800s, and as has been repeated widely in the developing world, the simultaneous promotion of irrigation can actually exacerbate the malaria problem. Land created by wetland destruction is often called ‘reclaimed’ or ‘improved’. In much of Europe and the USA draining of wetlands was viewed as a beneﬁt to society and encouraged by laws and subsidies. Western drainage and agricultural engineers spread this perception to developing countries Historically, river basin development has involved the eradication of wetlands for a variety of reasons either directly by drainage or bypassing, or indirectly by the construction of dams and river diversion works, or the over-exploitation of groundwater. This is because far from being integrated; the development of river basins has often been uni-sectoral, concentrating on hydropower generation, irrigation or ﬂood control at the expense of, rather than mutual beneﬁt of, other sectors. The catalogue of wetland destruction is remarkable. In England the area of freshwater marsh fell by 52% between 1947 and 1982 and in the early 1980s the rate of drainage continued at between 4000 and 8000 ha per year.

2.4 Global and Regional Perspective Ramsar Convention The Convention (www. ramsar.org) provides guidance to member states that wetland must be managed within its larger surrounding ‘waterscape’ (the river basin or catchment, including the hydrological processes and functions within the basin or catchment) as well its larger surrounding landscape. Wetland management must be integrated into land use management plans and water resource management plans; speciﬁcally the business plans and operational plans of the relevant water management agencies, to ensure that wetland objectives are fully realized. The aim should be to match

water resources’ strategies with land use strategies, so that these can be implemented jointly to support the maintenance of healthy, functional wetlands that provide a full range of products and services for people (including water supply). Yet land use management and water management are generally the responsibilities of different agencies or authorities, resulting in a lack of alignment of objectives or priorities, which in turn leads to one or other of the land or water aspects of wetlands not being adequately protected or managed. Under the Convention, Contracting Parties agree to ‘formulate and implement their planning so as to promote the conservation of designated wetlands and as far as possible the wise use of wetlands in their territory’. The concept of ‘wise use’ emphasizes that human use on a sustainable basis is entirely compatible with wetland conservation. All signatories to the Convention have also committed to consultation over shared water systems )European Commission 1995). Convention on Biological Diversity At a global scale, the primary instrument for the protection of biodiversity is the Convention on Biological Diversity (CBD 2001), which has at its heart the ecosystem management approach, i.e. maintaining ecosystem functioning, an underlying theme of both Integrated Water Resources Management and Integrated River Basin Management. Natura 2000 Within the European Union, the Natura 2000 Network has become the legal cornerstone for the protection of nature. Beginning in 1973 with the ﬁrst EU Action Programme for the Environment, successive revisions strengthened the legal basis for this policy and speciﬁc ﬁnancial instruments were created for nature conservation. The ﬁfth Action Programme for the Environment speciﬁed nature conservation and the preservation of biodiversity as major priorities, and in 1992 European Union governments adopted legislation (Directive No. 92/43/CEE) designed to protect the most seriously threatened habitats and species across Europe. Known commonly as the ‘Habitats Directive’, this law complemented the ‘Birds Directive’ (No. 79/409/EEC)

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Box 2.3 The deﬁnitions of favourable conservation status according to the Habitats Directive (No. 92/43/CEE) Article 1(e) conservation status of a natural habitat means the sum of the inﬂuences acting on a natural habitat and its typical species that may affect its long-term natural distribution, structure and functions as well as the long-term survival of its typical species within the territory referred to in Article 2. The conservation status of a natural habitat will be taken as ‘favourable’ when: • its natural range and areas it covers within that range are stable or increasing, and • the speciﬁc structure and functions which are necessary for its long-term maintenance exist and are likely to continue to exist for the foreseeable future, and • the conservation status of its typical species is favourable as deﬁned in (i). Article 1(i) conservation status of a species means the sum of the inﬂuences acting on the species concerned that may affect the long-term distribution and abundance of its populations within the territory referred to in Article 2. The conservation status will be taken as ‘favourable’ when: • population dynamics data on the species concerned indicate that it is maintaining itself on a long-term basis as a viable component of its natural habitats, and • the natural range of the species is neither being reduced nor is likely to be reduced for the foreseeable future, and • there is, and will probably continue to be, a sufﬁciently large habitat to maintain its populations on a long-term basis

originally adopted in 1979 (Council of European Communities 1992). At the heart of these Directives is the creation of the network of sites called Natura 2000 comprising: (i) Special Protection Areas (SPAs) established under the Birds Directive; and (ii) Special Areas of Conservation (SACs) to be designated for other species and for habitats under the Habitats Directive. The annexes of the Directives list habitats and species that are designated for protection, and recognize some of these as of special priority for conservation. Many designated species are characteristic of aquatic or wetland habitats, and the designated habitats include wetland types, not only of standing and running freshwater, raised bogs, mires and fens, but also tidal waters, saltmarshes (coastal and inland), humid meadows and wet forests (European Topic Centre (Biodiversity) 2007). The Network is expected to cover almost 20% of the EU territory, and is not designed as a system of nature reserves that excludes people

and their livelihood but rather as an approach to planning the coexistence of nature and people at the site, landscape, national and continental scales. The Network focuses on the protection and sustainable management of species and habitats that are vulnerable in their natural range in Europe as a whole. The Directives have objectives at the national and the site level. Nationally, the aim is maintain and/or restore favourable conservation status for Natura 2000 features (i.e. species and habitats) and is deﬁned in Articles 1(e) for habitats and 1(i) for species. At the site level, management or restoration is intended to achieve favourable condition. Implicit to these measures of favourable status is that interchange between populations and sites be intact, e.g. at the catchment or national scales (Box 2.3). Following revision to the EU Common Agricultural Policy, funding of farmers and foresters has been decoupled from production and the provision of present ﬁnancial support to such stakeholders through the Rural Development Funds is reliant on good environmental status

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(and linked to Natura 2000) as well as good agricultural condition. Increasingly the focus of Natura 2000 is the management of biodiversity resources at scales above the site, including the catchment, and part of the current policy focus is to integrate these directives with the later Water Framework Directive. Water Framework Directive The Water Framework Directive of the European Union (2000/60/ EC – European Commission and Parliament 2000) requires member states to achieve ‘Good Status’ (GS) in all surface and ground waters by 2015. Good Status is a combination of Good Chemical Status and Good Ecological Status (GES). GES is deﬁned as slight deviation from the reference conditions, based on populations and communities of ﬁsh, macro-invertebrates, macrophytes and phytobenthos, and phytoplankton. It also includes morphological and physicochemical supporting elements which will affect the biological elements such as channel form, water depth and river ﬂow. Although good ecological status is not required for wetlands under the Directive per se, the following categories of wetlands are included indirectly as part of basic measures (which member states must achieve) or supplementary measures (which are at the discretion of member states): • Groundwater-fed wetlands (groundwaterdependent terrestrial ecosystems): Wetlands connected directly to groundwater bodies should achieve good ecological status since the wetland is the observable expression of the groundwaterbody and the status of the wetland is an indicator of the ecological health of the groundwater. • Floodplains: considered as part of the surface waterbody (normally a river) to which they are connected. • Any wetlands: protected areas (e.g. Natura 2000 sites) should achieve good ecological status. • Other wetlands: that are neither connected directly to groundwater bodies nor part of surface water bodies nor protected areas, such as hilltop blanket peats, are only considered under supplementary measures.

In many countries, the ingrained uni-sectoral planning process is linked to a poorly integrated legal framework for water resource management and conservation of biodiversity in wetlands and waterbodies. What legal protection does exist might be weak or misplaced, e.g. the dambos of Zimbabwe are protected against certain agricultural uses because it is believed that they reduce ﬂoods and support low ﬂows in the rivers downstream. However, McCartney (2000) undertook a detailed study of dambo hydrology and concluded the wetlands were not performing these functions. Many wetlands produce excellent agricultural land when they have been converted, at least in the years immediately following drainage when the accumulated organic soils are still relatively intact (Sheail and Wells 1983). This makes the socio-economic pressure for wetland destruction hard to resist, even in the modern era of international conservation conventions. Within more economically developed countries, the legal systems for wetland protection are more comprehensive and potentially effective, though here too there is often a gap between the requirements of the laws and their actual application. In the USA, the national policy for wetlands is that of ‘no-net loss’. In the USA damage to wetlands must be avoided but if the damage is unavoidable for some overriding reason, then that damage must be ‘mitigated’, i.e. compensatory wetlands must be created that have at least equal extent and functions to the site that has been damaged – ‘providing a habitat that is functionally equivalent to the one that was lost’ (Zedler 1996). Such functional replacement can only be achieved if the new wetland is at least hydrologically equivalent to that which was damaged or destroyed; in that way, there may be some expectation that the same biota might come to inhabit the new feature. As discussed above, wetlands vary markedly in the practicality or ease with which they can be re-created, and the potential for successful mitigation is also highly variable. Hence there may be real technical limitations to the ‘no-net loss’ approach. Attempts have been made to facilitate the mitigation process by ‘banking’ certain habitat cre-

Wetland Management ation schemes that are not directly linked to the wetland that has been damaged. Mitigation banking makes the legal approach both more practical and ﬂexible, but may partially rupture the ‘like-for-like’ philosophy that is fundamental to the American law. In applying EU directives (e.g. Natura 2000, Water Framework), Environmental Impact Assessment (EIA) is a key legal tool in planning the management of wetlands at national and catchment scales (Howe et al. 1992). This approach and the role of EIA can be illuminated through how new developments are assessed that might signiﬁcantly affect Natura 2000 designated features (European Commission 2005). Such developments are not prohibited a priori under the Birds and Habitats Directives, but must undergo an EIA and should this assessment determine that damage will be signiﬁcant, alternatives must be found to the development. However, if no alternatives to the development exist, and if the project is of overriding public interest, then the plan may still be authorized provided that certain conditions are respected. The conditions and the procedure are laid down in Article 6 of the Habitats Directive and follow the principle of sustainable development. Article 6 deﬁnes a procedure of three basic steps for considering whether plans or projects may or may not be allowed (summarized in Fig. 2.2). Decisions relating to this procedure are made by the national authorities, but the European Commission may become involved where there is an ofﬁcial complaint made against the national decision. The Commission must be informed of compensation measures that are proposed for any plans or projects that are approved in order to maintain the coherence of the Natura 2000 Network. Much early nature conservation legislation was theoretical and worded scientiﬁcally or, if the law did address practical issues, consideration was given to penalties rather than to involving stakeholders. The last 10–20 years have seen the growth of stakeholder involvement in all aspects of nature conservation, i.e. through provision of incentives and giving them ‘ownership’ of the decisions in partnership with the responsible

35

authorities. This theme underpins the ‘Conservation in Partnership’ approach which is at the core of Natura 2000 (European Commission 2005). Stakeholder involvement is not only advocated where conservation action affects private land, but also in the management planning of protected areas (Eurosite 1999). From being planned and managed reactively and without local involvement, protected areas are now increasingly run in collaboration with local stakeholders and being managed adaptively with a long-term perspective. Key lessons are: (i) involvement of stakeholders at all stages of the process; (ii) transparency of the process; and (iii) communication and awareness raising. Having set the conservation management goals of good ecological status and/or favourable conservation status, the landowner is provided with ﬁnancial incentives where, for example, site management measures to be achieved by the owner are linked to the allocation of Natura 2000 payments under the European Agricultural Fund for Rural Development (EAFRD). Other subsides and incentives for farmers have been employed at the national or regional levels through agri-environment schemes, e.g. in the UK beginning with the Countryside Stewardship and Environmentally Sensitive Area (ESA) schemes and proceeding to a more uniform Entry Level and Higher Level Stewardship system (Defra 2005). Here the landowner might be compensated for not intensifying the management of the water and the agricultural habitats (e.g. lowland wet grassland), and the level of the compensation calculated on the basis of ‘proﬁt foregone’, i.e. the extra income that might have followed through intensiﬁed agriculture (Edwards and Fraser 2001). Increasingly, however, the incentives became based not on calculated inaction but on positive management action by the landowner to maintain or enhance the biodiversity value of the site. Agri-environment schemes partly arose to address the conﬂict in rural districts where pressure to boost production was met by increasing realization of the vulnerable biodiversity that existed in these landscapes (Mercer 1990). The

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Fig. 2.2 Consideration of plans and proposals affecting Natura 2000 sites – schematic proposal illustrating the contribution of Environmental Impact Assessment. (After European Commission 2005.)

application of some prescriptions for biodiversity conservation experienced further conﬂicts as particular measures caused further unlooked for problems. Thus applying raised water-levels in English ESAs achieved its early goals of increased populations of wading birds and waterfowl, but the original regime proved unsuitable for the rich grassland plant communities for which conservation measures were also necessary (Gowing 1996, 2004). Even where raised water levels did

beneﬁt most elements of biodiversity, there was concern that this reduced the capacity of the surface drainage system and soils to accommodate excess water, therefore leaving the landscape vulnerable to ﬂoods (Acreman et al. 2007). The cost of maintaining ﬂood protection (e.g. levees, sea-walls and pumps) under rising sea levels associated with climate change also affects the management priorities in the lowland parts of river basins.

Wetland Management 2.5 Wetland Loss, Restoration and Management Worldwide, conversion or drainage for agricultural development has been the principal cause of inland wetland loss (Millennium Ecosystem Assessment 2005). The amount of wetland lost is difﬁcult to quantify because detailed inventories generally do not exist and deﬁnitions of ‘loss’ are subject to a wide range of interpretations; however, some estimates of wetland loss have been attempted. Many European countries have lost between 50% and 70% of their wetlands in the last century (McCartney et al. 2000). France’s largest wetland type, the freshwater meadows, bogs and woods, once covered 1.3 million ha but during the twentieth century it declined at a rate of 10,000 ha per year (Baldock 1990). In Roman times, 10% of Italy (3 million ha) was wetland; only 764,000 ha remained by 1865, and by 1972 this had diminished to only 190,000 ha. In the Castille-La Mancha region of Spain, by 1990 60% of the wetlands had been lost with three-quarters of this loss (20,000 ha) taking place in the previous 25 years (Montes and Bifani 1991). In Greece a 60% loss of wetlands, mainly lakes and marshland, took place by land drainage for agriculture after 1925 (Maltby 1986). In the same way, 28% of Tunisian wetlands had disappeared in the twentieth century (Maamouri and Hughes 1992). In Asia, about 85% of the 947 sites listed in the Directory of Asian Wetlands were under threat with 50% of the threatened sites being under serious threat (Scott and Poole 1989). By the modern era, the USA had lost some 54% of their original 87 million hectares of wetlands (Tiner 1984), primarily to drainage for agricultural production. In Asia some 67% and in Latin America and the Caribbean 50% of the major threats to wetlands were from hydrological change related to drainage for agriculture, pollution, catchment degradation or diversion of water (Dugan 1990). Large dams have brought beneﬁts to many people around the world by providing water for hydropower generation, irrigation and water supply. However, there have also been widespread criticisms about the damage done to

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upstream and downstream ecosystems by dams. In particular, the World Commission on Dams (2000) found that ﬂoodplain and deltaic wetlands downstream of dams have been degraded in many parts of the world due to lack of ﬂoods. When ﬂooded periodically, these ecosystems supply important products (e.g. arable land, ﬁsheries, livestock grazing), functions (e.g. groundwater recharge, nutrient cycling) and attributes (e.g. biodiversity), which have provided the economic, social and environmental security of rural communities for many centuries. In some cases, the operating rules for dams now allow for ﬂood releases to be made to restore and maintain downstream wetlands including the Rivers Senegal, Kafue, Tana and Phongolo in Africa (Acreman 2003). As in all ﬁelds of science, the terms used within ecological restoration are ‘theoryladen’, i.e. their usage comes with a background reﬂecting the evolution of the discipline (Bradshaw 1997). However, several terms commonly employed here are especially contentious and depend for their understanding on deﬁning what it is that the wetland restoration is intended to achieve. This controversy has been partly described above on setting objectives for wetlands, and deciding whether the goals are a natural system or whether they allow for regular intervention in a regime of wetland gardening. The word restoration implies that the purpose of the wetland management is to return the composition, productivity and processes of the site to some natural sustainable ideal. In the context of landscape fragmentation and global environmental change, it may be that genuine restoration of a perfect facsimile of the degraded ecosystem may no longer be practical. For this reason some practitioners prefer to use the words rehabilitation and/or wetland creation. Rehabilitation implies management to repair as far as possible the wetland processes, whereas creation acknowledges that a copy of the original wetland may be impossible to attain and that the best that can be built is a hydrologically equivalent ecosystem with sustainable populations of wetland species. However, in these re-created wetlands, the

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assemblage and proportions of its component species will be novel and will not ﬁt precisely with any blueprint based upon, for example, classic phytosociology. Generic objectives for wetland restoration schemes are outlined in Box 2.2. Within a realistic programme of wetland conservation and restoration, it is likely that there will be intensive efforts to manage vulnerable extant natural wetlands, combined with a targeted creation of a range of hydrologically deﬁned wetland types. A fuller discussion of wetland restoration, philosophy, objectives and techniques is provided by Keddy (2000), Hammer (1992) and Ewel (1987).

2.6

New Scientiﬁc Insights

2.6.1 Modelling wetland hydrology Assessment, restoration and management of wetlands involve altering the existing hydrological regime of a wetland. This often includes predicting conditions beyond those for which data have been measured. For example, assessment may be required of wetland conditions under future climates, or predictions may be needed for wetland conditions after re-connecting a ﬂoodplain wetland with a river. This is often achieved by constructing a computer model of the wetland, calibrating it with available data and then running it under unmeasured conditions. Few models have been developed speciﬁcally for wetlands (Restrepo et al. 1998). Furthermore, wetlands cover such diverse ecosystems that no single model could be used for all types. Instead, existing models developed for analysis of land drainage, river hydraulics and groundwater dynamics have been employed to assess wetlands. Selection of the appropriate model begins with conceptualization of how the system works (Acreman and Miller 2007), i.e. the nature of the different components of the hydrological cycle relevant to the wetland and how they interact, such as whether the water supply to the wetland is dominated by precipitation, surface runoff, groundwater or tidal ﬂows and what aspects of the wetland itself need

to be predicted. The model that best describes the conceptualization and will produce results to an acceptable accuracy is then selected. Computer spreadsheets have been widely used to model water balances of wetlands. The simplest form of these models undertakes water volume accounting by adding external inputs to the wetland, such as precipitation or inﬂow, and subtracting outputs, such as outﬂow and evaporation and do not attempt to include any internal processes, e.g. spatial variations in water table level. An example is given for Crymlyn Bog, Wales in Figure 2.3. More complex spreadsheet models include some distributional effects, for example Gasca-Tucker and Acreman (1999) describe a simple model of ditch water level and surface ﬂooding extent for application to the Pevensey Levels. Other models have focused on modelling internal behaviour, such as the relationship between ditch water level and in-ﬁeld water table levels, using land drainage and soil physics theory (Youngs et al. 1989, 1991). For example, Gowing et al. (2002) studied the water regime requirements of wetland vegetation communities using Young’s model. Two other examples of this type of model are FDRAIN (Armstrong et al. 1980) and DITCH (Armstrong and Rose 1999). Since wetland soils can be treated as a shallow aquifer, models developed originally for assessing groundwater levels have been used for studying wetlands. In particular, MODFLOW has also been used to simulate spatial and temporal variations in wetland water table levels (e.g. Reeve et al. 2000; Bradley 2002; Bradford and Acreman 2003). AQUA-3D was used to model groundwater and wetland interactions on the island of Crete, Greece (Bromley). A major drawback is that these models do not incorporate surface runoff or evaporation explicitly. However, a wetland ‘module’ has been developed for MODFLOW (Restrepo et al. 1998) that allows simple representation of ﬂow in rivers and sheet ﬂow in areas of dense vegetation (Kadlec 1990). In some wetlands, such as alluvial ﬂoodplain, the hydrology is dominated by overbank ﬂows from adjacent rivers. These can be modelled by

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Wetland Management

Fig. 2.3 Crymlyn Bog water balance. (After Acreman 2005.)

hydraulic river models, such as HEC-RAS or ISIS (Fig. 2.3). An alternative attempt is given with the much more ﬂexible spatial discretization realized by triangles in the ﬁnite element approach (Diersch and Michels 1996). An example of this model type is the FEFLOW 3D model of wetland–catchment interactions used in Germany. Additionally, the HPP-GMS ﬁnite element calculation code has been used to simulate the hydrological dynamics of the Rhine ﬂoodplain and to evaluate the role of that wetland in denitriﬁcation (Sanchez-Pérez and Trémolières 1997). However, the effort required for parameterization of such large coupled groundwater models is immense. In situations where wetland hydrology is controlled by the interaction of surface and groundwater, more complex models are required. Mansell et al. (2000) developed and applied a transient, coupled saturated–unsaturated, 2D numerical model to study the link between groundwater and cypress ponds in the Coastal Plain forest region, USA. Crowe et al. (2004) developed a 2D model for simulation of groundwater–wetland interactions. An example of coupling surface and groundwater models was given by Krause and Bronstert (2004), who coupled the surface water model WASIM-ETH-I and the groundwater model MODFLOW to simulate hydrological processes in a low lying sub-catchment of the Havel, Germany. A similar method

was employed by Thompson et al. (2004) for the analysis of wetlands. They applied the hydrological model MIKE SHE and the hydraulic model MIKE 11 together to the North Kent Marshes as part of the SHYLOC project (Al-Khudhairy et al. 2001). A major assessment of the hydro-ecology of the Niger river and its ﬂoodplains in Mali was undertaken by Zwarts et al. (2005). Hydrological modelling studies have shown that the proposed dam at Fomi in Guinea could reduce water levels in the Inner Niger delta during a ﬂood by up to 60–70 cm. Climate change may lead to additional reductions in water if the 50% reduction in rainfall predicted by the IPCC occurs. Any changes in land use upstream may also alter water availability. The implications of changes to the hydrological regime of the Niger need to be taken into account to assess the sustainability of the wetland restoration projects. 2.6.2 Hydro-ecological tools Although restoration and management may entail alteration of the hydrological regime, the success of any intervention lies in the ﬁnal impact on the wetland ecosystem. Building upon a considerable body of theory and practice in wetland restoration and eco-hydrology, there have been recent attempts to provide objective and quantitative guidelines for the maintenance

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Fig. 2.4 Water level requirements for MG8 ﬂoodplain margins. (After Wheeler et al. 2004.)

and restoration management of wetlands (Wheeler et al. 2004; Mountford et al. 2005). Within the UK, the development of such tools has been stimulated by agri-environment schemes and nature protection designations, notably in ﬂoodplain grasslands (Gowing et al. 2002) and grazing marsh, as at Tadham Moor (Mountford et al. 1997). These studies began with investigations of altered nutrient regimes and then proceeded to experimentally manipulated water levels, integrating the hydrological and nutrient aspects to produce the guidelines. Working on the basis of vegetation or habitat types, these eco-hydrological guidelines provide information on: (i) water supply mechanisms and a conceptual model; (ii) water regime; (iii) nutrient regime; (iv) management regime; (v) vulnerability; (vi) the restorability of the type; and (vii) gaps in the knowledge. For each season, the water regime parameters indicate the water depths (maximum, minimum, duration of exposure, etc.) that are favourable for the habitat, the depth ranges that the habitat can tolerate for short periods, and those that are damaging for the habitat. Figure 2.4 shows a ‘trafﬁc light-based’ water level regime zones diagram that depicts the mean water table requirements for each month of the year. The

green area shows preferred conditions. The amber area denotes conditions that are not ideal, but which the plants can withstand for short periods. The red area marks conditions which the vegetation cannot tolerate. Duranel et al. (2007) developed a method for assessing the suitability of ﬂoodplains for species-rich meadow restoration. The method was tested on the ﬂoodplain of the River Thames in central southern England and it was found that both the maximum duration of ﬂood events in autumn and winter and the depth of the groundwater table during the summer exceeded the requirements of the target species. Ecological restoration has been attempted at many wetland sites and different habitats, but in the mid-1990s, the scale of the schemes was greatly increased. The justiﬁcations for these landscape-scale restoration schemes was to: (i) safeguard existing wetland fragments through creating buffer zones around them; (ii) achieve a critical extent of wetland for species that require large areas (e.g. bigger mammals); (iii) link wetland fragments to allow dispersal of species; (iv) foster/stimulate carbon sequestration through vegetation growth and peat formation; and (v) fundamentally to meet the goals of the CBD for habitat restoration. In Europe, the ﬁrst major

Wetland Management schemes were in the Netherlands, e.g. Oostvaarders Plassen and Lauwersmeer both exceeding 5000 ha (Kampf 2000) and these were followed by a series of six schemes in the Fenland basin of eastern England forming an arc around the inland margin of this basin from Norfolk through Cambridgeshire to Lincolnshire and covering >10,000 ha in total, e.g. the Wicken Vision (Colston 2003; Hughes et al. 2005) and the Great Fen Project (Mountford et al. 2002; Great Fen Project, 2007). A further stimulus for this restoration programme was the pressure on sea-defences, and the programmes of managed retreat (Wolters et al. 2005) along the North Sea coast that not only created more saltmarsh but also threatened the viability of existing important brackish and freshwater wetlands. A ﬁnal long-term justiﬁcation for these major restoration schemes is a rationalization of water management and the potential for Integrated Water Resource Management. Both within the Netherlands and the UK, the target sites represent very extensive ﬂoodplain where several river basins join.

2.7 Integrating Wetlands within Catchment Management In the past, catchment management was dominated by the need to provide land and water resources for housing, agriculture and industry. Managing at the catchment scales has advantages for achieving these aims since, for example, water resources can be exploited in one part of the catchment and distributed to other parts where needed. However, this means that competition for resources becomes a catchment-wide issue, rather than a local issue. The Azraq oasis in Jordan, a spring-fed wetland comprising open water and seasonally ﬂooded mud ﬂats, provide an example. In the past, the wetland supported large migratory bird populations and permanent communities exceeding 4000 people, plus an estimated 40,000 nomads depended on it for water and grazing at some point in each year (Fariz and Hatough-Bouran 1998). With average annual rainfall of just

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200 mm, Amman, Jordan’s capital, relies on the aquifer underlying Azraq for its water supplies. In 1980, the (then) Amman Water Authority began pumping water to Amman directly from the Azraq springs, at an average rate of 900 m3 per hour (i.e. 75% of the springs’ discharge). In 1981 the Water Authority dug 15 artesian wells some 10 km north-west of the springs, and pumping from this well-ﬁeld replaced direct pumping from the springs in 1982. As a result of this extraction, combined with pumping from private wells, the springs’ annual discharge fell from 10.5 million cubic metres (MCM) in 1981 to 1.96 MCM in 1989 and the springs dried up completely in 1992. The drying up of the springs resulted in the virtual destruction of the ecological value of the wetland. The lake and marshes that comprised the wetland ecosystem dried up. This, in turn, led to brushﬁres, increased salinization of water sources, and a consequent decrease in agriculture and tourism. To mitigate the impacts, the Ministry of Water and Irrigation now annually pump about 1.5 MCM of water (i.e. 10% of the original input) from the well-ﬁeld into the lakes in the centre of the oasis. However, groundwater levels continue to drop at a rate of 0.5–1 m per year and so the long-term sustainability of the scheme is uncertain. A further example of implications for wetlands of catchment scale management is provided by the Indus delta. The Indus is one of the major river systems of Asia, and dominates the landscape and economy of Pakistan, providing water for the world’s largest irrigated area. The extensive delta of the Indus, stabilized by mangroves, supports an extensive ﬁshery that earned US$100 million foreign exchange in 1997 (Acreman 2003). Records from the nineteenth century suggest that freshwater ﬂows to the lower Indus were around 185 million m3 per year, with the highest ﬂow occurring in August. Since then construction of numerous dams and barrages for irrigation has reduced the ﬂows of water and sediment. During the period 1960–1971, freshwater inﬂows were only 43,000 million m3, and the Indus Water Accord only provides 12,300 million m3 per year. This occurs mainly in the

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period June–August, with little or no ﬂow in other months. Ecological studies by the Sindh Forest Department estimated that each 100 acres (40 ha) of mangrove forest requires 1 m3 s−1 (0.028 m3 s−1) during July and August to remain healthy and to support the associated ﬁsheries. For the estimated 260,000 hectares of mangroves a total volume of 27 MAF (33,300 million m3) would be needed (Meynell and Qureshi 1995). A typical ﬂood hydrograph shape would suggest a ﬂow peak of 5000 m3 s−1. The ﬂoodplain forests need to be inundated at least twice in 5 years to enable saplings to become established. However, water management is politically sensitive as more water for the delta means less for upstream irrigation. The trade-off between wetland conservation and alternative use of resources has important implications for local communities. Catchment management may lead to overall economic development for the society as a whole, but there are often unfortunate distribution impacts. In many cases, those who beneﬁt most directly from wetlands are the rural poor who use natural wetland products, farm wetland soils and graze their cattle on wetland grasses (Horowitz and SalemMurdock 1991). These people may be severely affected by wetland loss but may not beneﬁt from alternative use of resources. For example, in the Senegal valley, dam construction has led to improved electricity supply in the capital cities of Dakar, Nouakchott and Bamako, beneﬁting the urban elite (Hollis 1996). The local communities in the Senegal ﬂoodplain not only lost essential sources of fuelwood and construction timber, ﬁsheries and agricultural land, but were unable to beneﬁt from dam management as there is little rural electriﬁcation (Acreman 1996).

2.8

Conclusions

Integrated catchment management involves putting in place a series of linked actions that together deliver the goods and services required by mankind, including land for farming, housing and industry, supplies of freshwater for public

supply, irrigation and power generation, protection from ﬂoods and a healthy environment for recreation. Catchment management in the twentieth century was marked by the building of large water engineering structures, such as dams and embankments, with the aim of controlling natural processes. These developments led to massive worldwide loss of wetlands, which were considered as wastelands (Maltby 1986). The Bruntland Report, Our Common Future, and UNCED Conference in Rio in 1992 seemed to mark a turning point in modern thinking. A central principle of Agenda 21 and of Caring for the Earth (IUCN/UNEP/WWF 1991) is that the lives of people and the environment are profoundly interlinked. Ecological processes keep the planet ﬁt for living, providing our food, air to breathe, medicines and much of what we call ‘quality of life’ (Acreman 2001) and require a diversity of biological forms (McNeely et al. 1990). More recent global fora, such as the Second World Water Forum in The Hague 2000 and Earth Summit in Johannesburg, highlighted the need to ensure the integrity of ecosystems in sustainable water resources management; working with nature rather than against it. The World Commission on Dams (2000) recommended operating dams to restore and maintain ﬂoodplain and deltaic wetlands downstream. Such actions can also be very cost-effective as the full value of wetlands as providers of goods and services to people is increasingly being recognized. River basin authorities are similarly increasingly recognizing the importance of the maintenance of natural resources and functions of ecosystems, such as wetlands, as a key strategy for sustainable development (United Nations Environment Programme/Wetland International 1997). Wetland restoration and management is now a key element in catchment plans across the globe. This has been facilitated by major scientiﬁc and institutional change related to wetlands including new laws and policies, national and international organizations, scientiﬁc understanding, experts and consultants, hydrological and hydro-ecological models and methods and

Wetland Management involvement of local communities in planning and management. A constraint on realizing the beneﬁts of wetlands is that such ecosystems tend to be restricted to speciﬁc protected areas and are highly managed to maintain them in a ﬁxed ecological state, which counters their natural evolution. Realizing the full potential of wetlands would require a catchment scale approach that allows their natural establishment, evolution and decay. Within the context of European biodiversity protection, attempts to combine the creation of new wetlands with the abandonment (through succession) of other wetlands is often considered impractical and the policy focus remains on the maintenance of existing wetlands, regardless of practical difﬁculties (Tucker et al. 2008). Given the permanent nature of roads, towns and agricultural land a catchment scale approach to managing the dispersed wetlands is unlikely to be an achievable goal. Climate change in particular may lead to signiﬁcant spatial shifts in conditions that support wetlands that cannot be readily accommodated by current catchment management practices. In addition, wetlands are not a panacea for catchment management issues; whilst they can provide important goods and services at a lower cost than technological alternatives, the ability of wetlands to reduce ﬂoods or purify water is limited. Wetland should thus be treated as a highly valuable element in catchment management, but not an over-riding prerequisite.

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Wetland Management Wolters, M., Garbutt, A. and Bakker, J.P. (2005) Saltmarsh restoration: evaluating the success of deembankments in north-west Europe. Biological Conservation, 123(2), 249–268. World Commission on Dams (2000) Dams and Development: a new framework for decisionmaking. Earthscan, London. Youngs, E.G., Leeds-Harrison, P.B. and Chapman, J.M. (1989) Modelling water movement in ﬂat low-lying lands. Hydrological Proceedings, 3, 301–315. Youngs, E.G., Chapman, J.M., Leeds-Harrison, P.B. and Spoor, G. (1991) The application of a soil physics model to the management of soil water conditions in wildlife habitats. In: Hydrological Basis of

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Ecologically Sound Management of Soil and Groundwater. Proceedings of Vienna Symposium, August (1991). IAHS Publ no 202, 91–100. Zedler, J.B. (1996). Ecological issues in wetland mitigation: an introduction to the forum. Ecological Applications, 6, 33–37. Zwarts, L., van Beukering, P., Kone, B. and Wymenga, E. (eds) (2005) The Niger, A Lifeline. Effective water management in the upper Niger basin. RIZA, Lelystad/Wetlands International, Sévaré/Institute for Environmental studies (IVM), Amsterdam/A&W ecological consultants, Veenwouden, Mali/The Netherlands.

Image facing chapter title page: Courtesy of the Centre for Ecology and Hydrology.

3

Flood Management RALPH M. J. SCHIELEN1

1

Ministry of Transport, Public Works and Water Management, Centre for Water Management, Lelystad, The Netherlands and University of Twente, Department of Water Engineering and Management, Enshede, The Netherlands

3.1

Introduction

Out of the six billion people that now live on the planet, almost 50% live in the delta areas of large rivers. This number will grow to around 80% by the end of this century. Meanwhile, as a result of climate change, the sea level will rise by 18– 59 cm (according to the Intergovernmental Panel on Climate Change (IPCC)). This means that millions of people are at risk of innundation by tidal surges. How big this risk is depends on the speciﬁc geographic conditions, on the economic conditions and on the (technical) ability to reduce the risk. As risk is deﬁned as the probability (of ﬂooding) times the potential damage, reduction of risk can be achieved by either reducing the probability, or by reducing the potential damage (or both). In the Netherlands, a highly developed society has been created in which there is a subtle balance between the growing economic interest and increasing probability of extreme events (extreme discharges from rivers and storm surges at sea). This balance is kept in order by a welldeﬁned and thorough safety system and safety approach which helps defend against ﬂooding in general. The estimates of the sea level rise made by IPCC are of particular importance for the

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

Netherlands. The Royal Netherlands Meteorological Institute has interpreted this data for the Dutch situation and found a local increase of the sea level of 35–85 cm by 2100 (see Van den Hurk et al. 2006). Flood management has been invented by man. As soon as man started to live in the vicinity of rivers and their delta’s, build houses or conﬁne rivers with levees and dikes (for safety, economic or agriculture reasons), ﬂood management became an issue. The more densely populated the area, the more an issue it becomes. Indeed, it is often said that ﬂoods are natural, but hazards are human. Hence, ﬂood management in Europe, parts of America and in Southeast Asia, is a major issue, (see, for example, Wisner et al. 2005). In the Netherlands for instance, ﬂood management started around 1000 years ago. Before that, the land was nothing more than a swamp, lying in the delta of the rivers Rhine and Meuse. As agriculture developed people started to protect their new land with small levees, and over the next centuries, this became the highly sophisticated system of larger and smaller dikes and levees, together with large dams, dunes and storm surge barriers which now protect the larger part of the Netherlands against ﬂooding from the sea or from rivers. The development, which is here summarized in just one sentence, has been beautifully described in a recently published book and is still an ongoing process (Van de Ven 2004; Ten Brinke 2007).

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In Europe, the Netherlands is one of the countries most threatened by ﬂooding. Nowadays, it is often said that the threat comes from four different directions: From the west, by the sea (and hence sea-level rising), from the east, by the rivers (and the increasing discharge in wet seasons), from above (by more and more intense rainfall) and from below (due to the problems created by the groundwater and the subsidence of the soil). Threats from the sea (storms and salt intrusions) and the major rivers Rhine and Meuse are the most obvious. The Netherlands is, for the majority, located in the delta of these rivers. Many centuries ago, they were real delta’s, with alluvial fans, many branches, etc. Nowadays, they are totally conﬁned by dikes. At the point where the Rhine enters the Netherlands, it already has all the characteristics of a typical (restrained) lowland river: it is a wide, slowly ﬂowing river often with wide ﬂoodplains. Due to the system of dikes and dunes, the western part of the Netherlands resembles a big bathtub, which is kept dry by continuous pumping. As a result, the subsoil degrades at a rate of a few centimetres per year (Groenewoud et al. 1991). Heavy rainfall in the Netherlands can cause inconvenience in the low-lying western part. Life-threatening situations, however, seldom occur. The water is just pumped away in a few days. However this is not to say that there will be no damage, or annoyance for the people

involved. Therefore studies are being carried out to see whether the construction of small retention polders is possible, where the excess rain water can be temporarily stored. This may also involve a change in the sewer system: problems resulting from rain water often stems from an overloading of the sewer system. Periods of high discharge on the Dutch Rhine are almost always created by a combination of large amounts of melting water from the Alps, intense rainfall in Germany, and a subsoil in the catchment which is either saturated or frozen. Due to these circumstances, the water is rapidly collected in the main channel. The various tributaries (which also have an extreme discharge due to the same conditions) add ﬂow incessantly and the accumulated discharge then enters the Netherlands. For the Meuse, high discharges are mostly observed during winter, after a few days of intense rainfall. The Meuse basin is, compared to the Rhine, smaller, and hence the discharge reacts rapidly. A hardrock subsoil which cannot absorb large quantities of water makes the situation even worse. The Rhine and Meuse are, therefore, two rivers with different characters; however periods of high discharges on both rivers are highly correlated (Fig. 3.1) All the land around the branches of the Rhine itself and their tributaries is protected by dikes, and near the sea there also exists an extensive

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Fig. 3.1 The hydrograph of the Rhine (at Lobith) and Meuse (at Borgharen) from 1989 up to 1998. There is a clear correlation between the periods of high discharge on the Rhine during this period and the Meuse.

Flood Management system of dams, sea dikes and dunes. Almost 25% of the Netherlands lies below mean sea level, and almost two-thirds of the Netherlands is prone to ﬂooding (either from the sea or from the rivers). This ﬂood-prone area is divided into sections (Fig. 3.2) for management purposes. There is a wide variation in the spatial scale of the sections. Section 14, for instance, is 223,000 km2, contains the cities of Amsterdam, Rotterdam and The Hague and has more than 2.3 million inhabitants. Section 42 is small, only 3400 km2 and there are only 14,000 inhabitants. The biggest sections are further divided into subsections by compartments dikes, although their location is often determined by historical reasons. Note also that although the river Rhine is protected by dikes along its entire course in the Netherlands, the River Meuse is not – the ﬁrst 150 km are not protected by dikes. This has to do with the nature of the land the river is ﬂowing through. In the southern part of the Netherlands, the Meuse ﬂows through a V-shaped valley and there is no risk of inundation of large areas of land. High discharges lead to nuisance only. Nevertheless, many inhabited areas along the Meuse are now protected by a system of manmade levees (natural levees are rare, and are not part of the defense system), which created a signiﬁcant increase in the number of dike sections (up to 1995, there were 53 dike sections). In 1995, the Flood Protection Act changed, and from that time the inhabited areas along the Meuse became ofﬁcially dike sections with a prescribed safety level. Due to this change, there are now 99 dike sections. In the low-lying part of the Netherlands, where the Rhine ﬂows, the land behind the dikes is often, in contrast to the Meuse, several meters below the water surface of the rivers, and in general a collapse of a dike will have catastrophic consequences and immediate threat of loss of life (Fig. 3.3). Since 1953, when a storm surge caused devastation in the south-western part of the Netherlands (and took the lives of 1835 people) there has been a change in the approach to ﬂood protection and this has led to a re-design of the ﬂood system. Safety standards have been established for the different dike sections, and these (high) standards

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have been put in to Dutch legislation. As a result, the protection standards against ﬂooding are the highest in the world. The western part of the Netherlands has a statistical chance of ﬂooding from the sea once every 10,000 (from hereon denoted as 1 : 10,000) years. In the eastern part, where the big rivers enter the Netherlands, the probability is 1 : 1250. This difference can be explained easily by looking at the threats. Storm surges at sea are more difﬁcult to predict than high discharges in rivers, and also storms evolve more quickly than high discharges. The damage due to salt water is larger, and ﬁnally, the heart of the Dutch economy is in the western part of the country. In Ten Brinke and Bannink (2004), it is estimated that whenever dike section 14 ﬂoods, the resulting damage may cost as much as 250 billion euro’s and the lives of up to 200,000 people are at risk. Although the supply of water from the sea is actually inﬁnite, this is not really a factor. Simulations show that during low tide, signiﬁcant parts of the ﬂooded areas become dry again. In that respect, ﬂoods from rivers are actually worse, because the water keeps ﬂowing through the breach. The number of inhabitants (and hence of possible casualties), however, is generally lower in the river dike sections than in coastal areas. These protection levels may be compared with the situation in other countries. Many rivers (in Europe, and elsewhere) have protection levels with a safety of 1 : 200, 1 : 100 or even more frequent. The World Bank actually promotes a safety level of 1 : 10,000 for large cities, but these are ﬁgures which are seldom achieved. Perhaps due to the special situation in the Netherlands, ﬂood protection is part of Dutch legislation. There are no other countries where laws exist which determine the safety levels. The Flood Protection Act, in which the safety levels for the defence system are set, was introduced in 1996. The main disadvantage related to these high standards is that people increasingly believe that a ﬂood can never occur. There are fewer people who know what it is to live with the water. Up to the beginning of the twentieth century, people were adapted to living in ﬂoodplains. Their houses were built on mounds or were surrounded

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Fig. 3.2 An overview of the safety norms in the Netherlands. The areas outside the dike sections are not prone to ﬂooding from the river, because they are sufﬁciently elevated. The dike sections in the south of the Netherlands that are recently added are not on this map.

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Flood Management (a)

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Fig. 3.3 (a) A typical situation along the Rhine branches. The dikes are up to 6 metres high, and directly behind the dikes, there is development located several metres below the top of the dike. These dikes are already quite old (some of them are hundreds of years old). (b) A typical new development can be seen – a stone construction to protect a community along the Meuse. In times of high dischage, the opening on the left would be closed with concrete beams.

by small dikes. They would have tiles on the ground ﬂoor and the electricity would enter the house on the ﬁrst ﬂoor. Despite these measures, they were confronted at some time in their lives with water in their houses due to high discharges, but after the water level dropped, the house was cleaned and life continued. Now, people are less used to these situations, and hence organize their houses differently. The consequences of high discharges are obvious, but people hardly accept the fact that this may happen. They often believe it is the task of the government to make sure that they are protected. Unfortunately, people are less and less aware of what it means to live in a land which can be almost literally conquered from the sea. The defence system is so solid that the majority of the population will never experience a ﬂood. As a consequence, the awareness vanishes. The government is responsible for ﬂood management, but providing absolute safety is an illusion, as is often stated by ofﬁcials. What can be expected from the government is to provide reasonable safety, and to do everything in its ability to minimize the risks in the unfortunate case that something happens. In order to do this, public awareness is vital. Nowadays, studies are carried out to explore a possible transition from exceeding probabilities for critical water levels (upon which the current

safety standards are based, and which are only based on overtopping as the failure mechanism of dikes) to ﬂood probabilities (taking into account a number of other failure mechanisms). In this case a more realistic image of safety levels can be achieved. When this has been established, the ﬁnal step can be taken, to balance the costs of reducing the probability of ﬂooding with the protected values inside the dike section. This may lead to different safety levels for the dike sections, which may raise a societal debate about the desirability of such a situation. An argument in favour of differentiated norms, however, is given in Ten Brinke et al. (2008b), where it is argued that the current Dutch system of protection norms is actually no longer in agreement with the (spatial variation of) interests which they protect.

3.2 Historical Perspective Up to the early 1950s, ﬂood management in the Netherlands (and in most other European countries) was aimed at keeping the crests of the dikes approximately one metre above the level, corresponding to the largest discharge recorded up to that time. In the Netherlands, this peak was

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determined by the event recorded in 1926, which is still the highest discharge for the Rhine ever recorded (12,600 m3s−1 at Lobith). Before that, in the western part of the Netherlands, the last ﬂood occurred in 1703 while in 1917 there was a big ﬂood along the coast in the middle of the Netherlands (the later Lake IJssel, which at that time was still in open connection with the sea). In the years after 1926, the awareness of ﬂood risk faded. Big ﬂoods did not occur for several decades. The vast majority of the population started to believe they were entirely safe behind the (river as well as sea) dikes. All changed in the early morning of 1 February 1953. A combination of a spring tide and a severe storm surge attacked the coasts of the UK, Netherlands and Belgium. In the Netherlands, more than 500 breaches of the dikes were recorded, 200,000 ha of land was ﬂooded and 1835 people were killed. This is still an event that almost everyone is aware of. It is part of the national historical consciousness in the Netherlands. The 1953 storm is also still of huge national signiﬁcance in Scotland and England (Pollard 1978; Summers 1978). Current issues around climate change and ﬂooding of the east coast have ensured that there has been much coastal realignment to lessen the potential impact of such a surge in the future. In the Netherlands, this event was a major wake-up call. The general impression was that this should never be allowed to happen again. Clearly, the Netherlands were not safe, and that was unacceptable. It should be stated that there were concerns about the strength of the sea dikes already in the late 1930s. The Dutch Ministry of Transport, Public Works and Water Management issued a report in which it was pointed out that the dikes were in a poor condition. In the post Second World War reconstruction, however, limited attention was paid to the state of historical dikes. It is beyond the scope of this chapter to explain the details of the history of ﬂood management in the Netherlands which started more or less with the event of 1953. The interested reader is referred to Van der Ven (2004). This chapter focuses on some important elements.

After the ﬂood of 1953, the ﬁrst discussions were held regarding acceptable protection levels in relation to ‘economic’ values (including e.g. nature, cultural values and human lives) of the resources protected by the dikes. In this respect, a distinction was made between the probabilities of ﬂooding on one hand, and water levels which exceed certain critical levels with respect to dike heights on the other. It is important to understand the difference between these concepts, especially because the discussion between exceeding probabilities (of water levels) and ﬂooding probabilities (and risks) has been revisited in 2007. Flood risks are related to ﬂood probabilities. The chance that a ﬂood occurs is a combination of the chance of the water level exceeding a critical level and the chance of failure of the dike. After the ﬂood of 1953, a cost–beneﬁt analysis was made for the western part of the Netherlands, in which Amsterdam, Rotterdam and The Hague are located (for details see Van Dantzig 1956; Delta Committee 1960). The outcome was that an acceptable probability of a complete inundation for this area was found to be once in 125,000(!) years. (In the determination of this ﬁgure, the fact that there are lots of cultural and historical sites in the hinterland is taken into account in a very rough way in that the calculated probability was reduced by a factor 2.) Parallel to the cost–beneﬁt analysis, the exceeding probability of a critical water level (the so-called design water level, necessary for the design of the dikes) was determined to be once in 10,000 years which was consequently adopted as the safety level. The 1 : 125,000 probability was denoted as the average catastrophe frequency. The idea is that a dike will not immediately collapse if the design conditions are met but that in general it will be stronger due to construction properties (for details see Ten Brinke et al. 2008a). Based on the analysis for the western part of the Netherlands, the exceedance probabilities for the other dike sections (for the Rhine branches) were established. This was done in 1956, and as a result, large stretches of dikes had to be

Flood Management improved (heightened). Due to a lot of societal debate, the norms were redeﬁned a couple of times. Also the methods and models used to determine the return periods and the design water levels evolved over the years and ﬁnally reached a state that they gave sufﬁcient faith in the result. In 1992, safety standards were established at their current values, varying from 1 : 1250 for the larger part of the river area, to 1 : 2000, 1 : 4000 and 1 : 10,000 for the areas which are prone to ﬂooding from sea or lakes. Note also that the exceedance probability says little or nothing about the ‘real’ occurrence of a design discharge in terms of the next 50 or 100 years since this is too short a period. Hence, the design discharge (and the associated return period) is merely something that is needed as input for the mathematical models and methods. What we can learn from the statistical analysis is that a river discharge of 12,000 m3s−1 (as happened in 1995, for instance) may be expected on average every 60 years or so. This should immediately be followed by the statement that whenever such an event happens, there is no guarantee that it will not happen again in the coming years, as the discharge of February 1995 (only about 13 months after the 11,000 m3s−1 discharge of December 1993, see Fig. 3.4) showed. A design discharge has, therefore, little to do with reality but rather is something that is used to deﬁne a safety system with norms. 3.2.1 The Dutch organization Since 1996 safety standards and associated design water levels have been part of Dutch legislation. In the Flood Protection Act it is stated that every 5 years, the design discharge needs to be recalculated. The design discharge is based on a statistical analysis of historical discharges. As these records go back to 1901, and return periods of 1250 years are assessed for the river areas, the results have to be extrapolated signiﬁcantly. Mathematical methods are available to do this based on the theory of extreme value distribution, in which it is assumed that extreme events (of high discharge, in this case) follow a certain

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Fig. 3.4 A plot of the hydrograph for the Rhine (at Lobith) between 1 November 1993 and 28 February 1995. In this period two extreme discharges occurred: 10,940 m3 s−1 on 25 December 1993 and 11,790 m3 s−1 on 1 February 1995. According to the statistics, the return period of the 1993 event is about 30 years, and the return period of the 1995 event is about 60 years.

probability distribution. Several probability distributions are potentially applicable and it is actually not known which is the best. In practice, different distributions give different estimations for the 1 : 1250 discharge, and so the design discharge is delivered as the (rounded) average of three distributions (Gumbel, Pearson III, threeparameter log normal). There is, however, no scientiﬁc justiﬁcation for this decision. There is also a large uncertainty associated with this extrapolation, which is due to the fact that there is only data available for 106 years. The 95% conﬁdence interval of the design discharge of 16,000 m3s−1 lies roughly between 12,000 and 18,000 m3s−1. When the design discharge has been established, the associated water levels are then calculated with the latest mathematical models and a state of the art geometrical simulation of the summer and winter area of the river bed. The calculations differ for the eastern part of the Netherlands and for the western part. In the eastern part, the water levels are determined entirely by the discharge. In the western part, however, the water levels at the design conditions (i.e. with a return period of 1 : 4000 or 1 : 10,000) are mainly determined by storm surges.

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In between, there is a transition area in which (going from east to west) the inﬂuence of a high river discharge decreases while the inﬂuence of the storm condition (with respect to exceedance of critical water levels) increases. The calculations for the western part are, therefore, much more difﬁcult. There are in fact, combinations of sea-level conditions (due to storm surges), wind conditions (direction and strength) and discharges from the river side, each with their own probability, that determine the design water levels in the western part. Almost 7000 different combinations are needed for the calculation, after which a statistical processing of the results determines the actual design water levels. The design water levels are known as the ‘hydraulic boundary conditions for the Dutch water defences’, from now on denoted with the Dutch abbreviation HR1996, HR2001, etc. where the year refers to the 5-year cycle of recalculations (Ministry of Transport, Public Works and Water Management 2007). With these water levels, the Water Boards can check whether the dikes for which they are responsible are still high enough and strong enough. Testing the height is relatively simple but the strength, and hence the ability to withstand the different failure mechanisms, is much more difﬁcult to determine. In 1996 the hydraulic boundary conditions were established for the ﬁrst time. The design discharge at Lobith was found to be 15,000 m3s−1. In 2001 the results of the ﬁrst testing of the Water Boards (based on the design water levels of HR1996) became available. For 35% of the dikes, it could not be determined whether they were strong enough, mainly because of lack of data or lack of knowledge of the inner and outer slopes of the dikes. Meanwhile, the new design discharge for the hydraulic boundary conditions for 2001 was derived. Due to two periods of high discharges in 1993 and 1995 (which were not taken into account for the derivation of the hydraulic boundary conditions of 1996) the new discharge was found to be 16,000 m3s−1 (instead of 15,000 m3s−1). The immediate consequence of this increase was of course that the water levels of the hydraulic boundary conditions of 2001 increased

by up to 40 cm at some locations. Hence, a lot of dikes were assessed as being too low to comply with the standards as described in the Flood Protection Act, and action had to be taken. Accommodating the increase in design discharge of 1,000 m3s−1 was one of the main tasks of the national programme Space for the River (or Room for the River as it is often called in English texts in the Netherlands; see section 3.3.1). In 2006, the second testing round based on the water levels of HR2001 (i.e. 16,000 m3s−1) was completed. The outcome will be that most dikes will not be approved. All the plans which are currently developed in Space for the Rivers to accommodate the increase in design discharge have not been carried out, and as a result most of the dikes are too low. Clearly the testing of the dikes (every 5 years) and the completion of the Space for the River programme (to be foreseen for 2015) have different time-frames which causes some management problems.

3.3 Current Solutions In the Netherlands, since the 1990s, a process known as re-allotment has been ongoing by which agricultural land was exchanged between farmers in order to get larger, connected pieces of land which could be used more efﬁciently. Also the ideas about re-landscaping the Dutch river landscape, as a result of the reallocation of land, changed (Van Heezik 2006). The general feeling was that the role of agriculture in the area was too dominant and that more attention should be focused on nature and nature development. One of the ﬁrst expanded ideas to achieve this was issued by the Ministry of Agriculture and Fishing in 1985. In addition, the World Wildlife Foundation developed their aims to restore a more natural situation in the river area and published their ideas around 1992. They proposed plans in which nature development played as important a role in the side-channels in the ﬂoodplains as ﬂood protection. Clearly, the ‘ancient’ solution to ﬂood management (reinforcing and heightening the dikes) did not coincide with the

Flood Management idea of creating a more natural river system. Besides, it was recognized that increasing the dikes yet again (after the programmes of the previous decades) was not going to provide a robust solution to the problem of ﬂood protection. After all, higher dikes lead to higher potential damage in the event of a dike failure. These thoughts were elaborated further in the years after in some explorative studies of the Ministry of Transport, Public Works and Water Management in which it was demonstrated that there were enough possibilities (within the ﬂoodplain as well as using land outside the dikes to increase the discharge capacity) to tackle the problem. Hence, it is no surprise that the solutions for the accommodation of the increase in design discharge in 2001 were also sought in the direction of enlarging the river bed rather than dike reinforcement. This has become the formal policy of the Dutch government.

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3.3.1 Space for the river

Fig. 3.5 Schematic overviewof the KPD. Part 4, the ﬁnal presentation of the plans, has been approved by the Dutch Senate on 19 December 2006. In part 2, the KPD had to deal with almost 3000 reactions from the public. The main difference between the intention and the policy decision was the cancellation of a dike-relocation along the River Waal, and a rearrangement of measures along the River Lek. EIA, Environmental Impact Assessment.

Since complying with the standards as put forward in the Flood Protection Act is a legal responsibility, the programme Space for the River has a special status within the Dutch existing order on planning procedures. Space for the River is a so-called Key Planning Procedure (KPD). It is a joint initiative of three Ministries: the Ministry of Transport, Public Works and Water Management, the Ministry of Housing, Spatial Planning and the Environment and the Ministry of Agriculture, Nature and Food Quality. The KPD has four different phases (Fig. 3.5). It starts with an environmental impact assessment on the initial plans, and ends with the approval of the ﬁnal set of plans by the Dutch Senate. In between, there is an opportunity for the public to react to the plans and suggest improvements, after which the Dutch government has to approve the (possibly adjusted) plans. The result of this procedure is a rough sketch of the spatial ﬂood defence plans of, in this case, the river area. The location of those plans as mentioned in the KPD get a certain status with respect to future development by the local

authorities. In general, new (commercial) initiatives are hard to start, for the areas are reserved for future ﬂood management. Now, most of the plans will be located in the ﬂood plains, and these are already protected by the Flood Protection Act and the aforementioned policy. Therein it is stated that all activities which are not strictly connected to river activities are forbidden in the ﬂoodplains. This means that, for instance, building houses, siting of new industry or even recreational activities which may obstruct the discharge are not allowed. Existing industry (mainly brick factories) can be maintained, and are even allowed to expand under very strict conditions. In recent years, it was noted that economic and spatial developments in the ﬂood plains are obstructed by the policy line and, therefore, the policy line has recently been adapted. At about 15 different locations, there is room for experiments with adaptive building. These constructions have to be ﬂood-proof, and have to give an additional impulse with respect to spatial quality for the area (Fig. 3.6).

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Fig. 3.6 An artist’s impression of experimental buildings in the ﬂood plains. (Taken from a brochure issued by the Ministry of Transport, Public Works and Water Management in which the terms and conditions for building in the ﬂood plains have been summarized. Source: http://www.verkeerenwaterstaat. nl/Images/bouweninrivierbed_brochure_tcm195101301.pdf)

A consequence of the KPD is that new development in areas outside of the dikes (needed for dike relocations or side channels) is also not allowed. Even for the long term (2050 and beyond), areas have been designated where future measures might be constructed and they are thus reserved for these purposes. Hence, no large scale investments can take place. The reason for this is that the design discharge is expected to increase above 16,000 m3s−1 in the next decades. There is too little space available between the dikes to cope with this increase in design discharge. Reiterating that, the starting point of Space for the River is the calculated increase in design discharge and the result is a set of coherent spatial plans which will compensate for the increased design water level. There are some assumptions and boundary conditions which determine the scope of the programme: 1 Although possible, increasing the height of the dikes along the river branches in order to deal with increasing water levels is not the preferred solution. In the years after the high discharges of 1993 and 1995, a special Act was passed in the Dutch parliament, which enabled a rapid reinforcement of the dikes. Between 1995 and 2001, almost 600 km of dikes have been improved

(heightened and/or strengthened). In doing so, all the dikes have been brought to the level whereby they should be able to withstand the design discharge established in 1996 (which was 15,000 m3s−1, as laid down in HR1996). These dike levels were the starting point of the programme Space for the River. Hence, the increase in water level due to the increase in design discharge was to be compensated by deﬁning spatial measures (measures that increase the discharge capacity). This is, however, not to say that dike reinforcement is strictly forbidden. At places where spatial measures are almost impossible due to narrow ﬂoodplains (mainly for some branches in the western part of the Netherlands), or at locations where spatial measures are inordinately expensive, dike reinforcement is still allowed. 2 All the measures that will be selected have to be constructed before the end of 2015. 3 The total budget is +2.1 billion (based on 2002 price). As a start of the process, an inventory of all the possible plans (that increase the discharge capacity of the river, or locations to retain the water temporarily) along all the branches has been made. Over 600 different measures have been determined along the river branches. Some are small, with only a few centimetres effect; some are large with an effect of several decimetres. Some measures have a small public impact and are limited to the ﬂoodplain, whilst some have a large impact, affecting houses outside of the ﬂoodplain (e.g. dike relocations). Some measures carried the support of the people and the local communities and some were strongly opposed. The crux is to choose a set of measures that sufﬁciently reduce the water level on the one hand, while on the other hand achieve the measures within the time and budget as stated by the boundary conditions of the programme. Also, every set of measures has an impact on nature conservation, recreation and society and these aspects needed to be assessed. As there are many different sets of possible measures, choosing the most effective set of measures (taking into account all the different criteria) was the real

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Flood Management challenge of Space for the River. The programme ends with the actual construction of the plans. The timing of this stage will be different for all plans. The earliest will start in 2008, the last ones will start in 2013. In 2015, they all have to be ﬁnished. After that, the concluding stage of monitoring the results will take place and an assessment of hydraulic and morphological developments can be made. In deriving the most effective measures, it is clearly crucial to understand the impact of the measures on the rivers. A measure which increases the discharge capacity locally will have a maximum effect with respect to reduction of the water level at the location of the measure. Associated with that maximum effect is a socalled backwater curve, which gradually fades out the effect over many tens of kilometres upstream. Hence, a local measure takes care of a reduction of the water level over many kilometres in the upstream direction. Retention measures (i.e. temporary storage of excess discharge in a polder) has only a small effect in the upstream direction, but a large effect downstream. Clearly, since discharge is taken from the system downstream of the inlet of a retention area, the water level is reduced. The solution to achieving a balanced choice between all the possible measures was found in the development of a software tool: the planning kit. 3.3.2

The planning kit

This software system, which has been developed in cooperation with Delft Hydraulics, was started as essentially a database with information about the different measures in place along the river branches (see De Vriend and Dijkman 2003). A more complete inventory was then made, including many aspects of the measures, such as hydraulic data (the maximal water reducing effect, properties of the backwater curve), ﬁnancial aspects (details of the project costs, and maintenance costs), possible changes in ecotypes, consequences for agriculture, impacts on numbers of houses removed. Also, aerial views and impressions of the landscape from before and after the

introduction of a measure were included. The main application, however, was to calculate the hydraulic effects. As all of the individual hydraulic effects of the plans were calculated in advance, it was possible to get an impression of the combined effect of two or more measures, by adding the individual effects. This is, of course, an approximation: the combined effect of two measures which are in the same vicinity may be larger or smaller than the sum of the individual effects, depending on the situation. This second order effect, however, was initially neglected. The tool made it possible to assess sets of measures capable of achieving a sufﬁcient reduction of the water level at the design conditions. The results were tested by a full scale hydraulic calculation in which all the preferred measures together (and also the inﬂuence that they might have on each other) were taken into account. The combined effect on many other indicators was reported in separate ﬁles. Despite concern regarding the initial assumption that individual effects cannot just be added, the similarity between both methods was remarkable.The selection process ended in 2006 and a total of 28 plans have been deﬁned. Together, they ensure that in 75% of the main branches the calculated water levels are below those determined in HR1996 (recall that the dikes had been brought up to the standards as deﬁned by HR1996). In 25% of the branches the dikes are above those determined in HR1996. It turns out, however, that at several locations, the dikes have been historically constructed higher than was strictly necessary. This excess height is at some locations now ‘consumed’, i.e. no other measures are necessary. In 11 dike sections, by absence of alternatives, dike reinforcement was the only alternative (approximately 64 km). 3.3.3 The design table In 2007, preparation for the actual construction of the plans was started. In this phase all of the provisional plans assessed earlier are designed in detail. The provisional designs in the planning process of Space for the Rivers are the

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starting point, but the actual design may differ signiﬁcantly. This phase required a new design instrument since the planning phase was focused on making a well balanced choice between the many possible measures. To achieve the detail required in that phase, it was sufﬁcient to make a rough design, to calculate the effects in advance, and put them in the database of the planning kit. In this next phase, however, it is necessary to make many variations of a particular design in order to be able to weight all the different interests of the stakeholders. For this, it is very convenient if calculations can be made on the spot and preferably with the input of the stakeholders and combining up-to-date geographical information. Therefore, the design table has been developed. This is an interactive software tool containing the GIS data of the area of interest. The table is literally that: a table, around which people can gather (Fig. 3.7); with electronic pens, a design can be sketched using the different maps, aerial views, historical maps, etc. The beneﬁt is that all the people around the table share the same information and can immediately react. Furthermore, once everybody is satisﬁed with the sketch, the hydraulic consequences can be calculated within minutes, and if necessary, the design can be

Fig. 3.7 An impression of the design table.

adapted again. This is possible because a simpliﬁed version of the hydraulic computation software is included in the tool. In this way, the design process is condensed to just a few days or less. This is a big advance over the ‘old’ method whereby a sketch had to be implemented in a model and the calculation had to be made separately, before the results could be communicated to the stakeholders. This often took a week or more and several of these design sessions were necessary before the ﬁnal plan was approved. The ﬁrst experiences with the design table are very good. The detailed design processes are indeed very rapid, and the fact that everyone shares the same information is a big advantage. 3.3.4 The American approach to ﬂood management Nations recognized for their advanced capabilities, such as the USA, are still highly vulnerable to the various threats posed by major ﬂood and storm events. The present US system has evolved into a patchwork of structural and non-structural measures that are directed by a number of federal, state, regional and municipal level organizations. The extent of damage caused by Hurricane Katrina in 2005, however, has caused the USA to pause to reassess its emergency management and risk reduction procedures. The ﬂooding that occurs across the USA varies considerably from one region to the next due to a number of factors. Most notably among these is its large latitudinal extent, climate types that range from semi-arid to semi-tropical to cold continental, widely varying topography and the large degree of coastal exposure. Accordingly, in any given year the USA faces any combination of the following ﬂood threats: • extensive damage from ﬂash ﬂoods in the semi-arid southwest; • severe erosion from winter storms along the California coast; • extensive springtime ﬂooding in the Midwest resulting from snowmelt and heavy rains; • late summer and early fall hurricanes along the east coast and Gulf of Mexico;

Flood Management • severe summertime thunderstorms throughout the eastern half of the country. Annual ﬂood damage costs in the USA have been on the rise over the past several years, which can be attributed to a number of trends. Most notably, the US population has grown steadily and has become increasingly concentrated within highrisk areas, such as ﬂoodplains and exposed coastal zones. A second major consideration is that commercial and residential land values have increased appreciably. Also, the means by which ﬂood damages are reported continues to be reﬁned and improved. The USA is also arguably experiencing an increase in extreme hydro-meteorological events as a result of climate change. Conversely, as the ﬁnancial costs of ﬂoods have been on the rise, loss of life within the USA from ﬂooding has fallen dramatically over the years. This downward trend stems from: • An advanced forecast and warning system that is managed at the national level. • Improved satellite and land-based (e.g. doppler radar) storm tracking systems. • Improved emergency management procedures (e.g. evacuation, risk communication) that are co-ordinated between federal, state and local authorities. • A public that continues to be better informed of regional risks and personal responsibilities. • Local structural measures, such as levees, ﬂood walls and reservoirs. • The introduction of a number of non-structural ﬂood prevention measures, such as local land use planning, national ﬂood insurance, ﬂood mitigation grants, community buy-outs and ﬂoodprooﬁng practices. Within the USA, ﬂoodplain management is a responsibility that is shared between federal, state and local organizations. At the federal level, primary responsibilities for ﬂoodplain management are dispersed between the Federal Emergency Management Agency (FEMA), the National Weather Service, the US Army Corps of Engineers (USACE), the US Geological Survey (USGS) and the National Resources Conservation Service, with each operating under unique lines of authority and areas of responsibility.

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Hurricane Katrina was a catastrophic event that overwhelmed emergency services at all levels. Putting that event aside, it can be argued that with each successive major ﬂood event, conditions improve on the national scale, with the USA moving towards a more effective balance between structural and non-structural solutions. At the federal level, the National Response Framework (NRF) has been drafted to incorporate best practices and procedures from incident management disciplines – homeland security, emergency management (including ﬂoods and hurricanes), law enforcement, ﬁre-ﬁghting, public works, public health, responder and recovery worker health and safety, emergency medical services, and the private sector – and integrates them into a uniﬁed structure. The NRF describes interorganizational roles and responsibilities for the period immediately surrounding a presidential declared disaster, including ﬂood and hurricane events. The National Flood Insurance Programme, which is administered by FEMA, is an example of a non-structural approach to ﬂood management. In addition to its ﬂood insurance programme, the federal role in ﬂood mitigation relates to community buy-outs, administering the Hazard Mitigation Grant Program, the Flood Mitigation Assistance Program and the Pre-Disaster Mitigation Program – all of which are designed to empower state and local organizations. The USA maintains an extensive hydro-meteorological data collection and archive system. Near real-time and historical stream and river gauge data for most of the ofﬁcial sites are available through the USGS. Meteorological data are collected by the National Weather Service (NWS) and stored as climatic data by the National Climatic Data Center. Data for these ofﬁcial sites are available online and can be requested in either hardcopy or digital formats. Not all of the stream gauges are managed by the USGS and some issues related to interagency data standards need to be resolved. Data sharing between Mexico and Canada is open, but most of these transboundary data sets have not been standardized. The maintenance of the nation’s hydro-meteorological

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data collection network has been underfunded for the past several years. Accordingly, there has been a marked drop in the number of stream and meteorological gauges throughout the USA. Policies that guide ﬂood management can be viewed in a number of ways – regionally (basinlevel), thematically (water quality, ﬂood response, ﬂood mitigation) or by project (coastal Louisiana ﬂood protection measures). There is, however, no centralized policy that guides ﬂoodplain management within the USA because such a heavy emphasis is placed on individual states’ rights. There have been several reports and task forces formed over the years to discuss and evaluate ﬂood management policies and procedures within the country, but one central policy does not exist. Since there will be a continued emphasis on state and regional management responsibilities the national role will continue to be a supportive one, such as issuing mitigation grants, issuing clean water guidelines, providing a seamless early warning system, creating national data sets (maps, census data), and assuming leadership roles in times of ﬂood emergencies that overwhelm local management capacities. Pertinent studies and acts that help better deﬁne the US position along these lines include the Flood Control Act of 1936, the Federal Water Pollution Control Act Amendments (Clean Water Act), House Document 465 (a uniﬁed national programme for Managing Flood Losses), the National Flood Insurance Programme (NFIP), the National Environmental Policy Act (NEPA), the Coastal Zone Management Act of 1972 and the Water Resources Development Act of 1986. Related acts such as these undergo constant review and are periodically amended, on an as-needed basis. In principle, the concept of integrated water resources management (IWRM) is being embraced by all pertinent federal agencies. In practice, however, its implementation is somewhat open to interpretation due to the fragmentary nature of federal responsibilities. Possibly a more objective measure is the approach taken in multiobjective planning, which stems from the 1970 Omnibus Water Resources Act. In that act, Congress directed that multi-objective planning

procedures be implemented specifying four objectives for federal water projects: regional economic development, environmental quality, social wellbeing and national economic development. The Principles and Guidelines (P&G) were subsequently devised to guide agencies in the multiobjective approach. Multi-objective planning continues to be practised while debates ensue regarding applied methodologies and analytical tools for any given project that is undertaken. Local governments control safety standards in ﬂood-prone low-lying areas. The implementation of Floodplain Management Plans by non-Federal sponsors is a requirement for participation in federal ﬂood protection projects. For communities that decide to participate in the National Flood Insurance Programme, FEMA requires that new structures be constructed to the 100-year base ﬂood elevation. Preventive measures are put in place and enforced by local entities. The ﬂood control infrastructure that USACE manages consists of over 400 major Federal dams and reservoirs, 8500 miles of levees and dikes, and hundreds of smaller local ﬂood protection projects. Various aspects of the operations, maintenance and inspection of the collective US stock of ﬂood damage reduction infrastructure has been called into question in the aftermath of Hurricane Katrina, and there is an ever-increasing backlog of ﬂood protection projects awaiting construction. Taken as a whole, the US faces major funding shortfalls in investments needed for major rehabilitations to existing projects as well as new ones. Non-structural ﬂood protection measures are undertaken in a variety of ways, including the National Flood Insurance Program, mitigation grants, local land use planning and zoning ordnances, and ﬂood-prooﬁng education. In the case of buy-outs, for instance, FEMA and USACE have programmes available for qualiﬁed structure owners. Both agencies have speciﬁc requirements for those programmes. FEMA requirements relate to repetitive losses and they deal directly with structure owners, as well as state and local interests. The USACE buy-out programme must be economically justiﬁed and supported by a non-

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Flood Management federal sponsor (i.e. local agency). For the most part, non-structural approaches to ﬂoodplain management are shared responsibilities between all levels of government. These measures will continue to grow in importance as large-scale engineering projects become increasingly difﬁcult to justify. In terms of ﬂood forecasting and warning, these systems are operated and maintained by various agencies. The responsibility for forecasting coastal and inland ﬂooding and issuing warnings primarily lies with the NWS. The National Hurricane Center, part of NWS, issues warnings and advisories for all tropical systems that are forecast to affect the US. The River Forecast Center, also part of NWS, is responsible for monitoring, forecasting and issuing warnings of ﬂooding on all inland waterways. Local news services are used to broadcasting warnings. USACE also uses direct contact within the affected area. Local commercial businesses and organizations are also used to notify the public in the affected areas. USACE receives stage predictions from NWS and its own water control sections, which are used to predict ﬂooding and the operation of ﬂood control structures. These capabilities can be very reliable, allowing ample time for emergency action to be taken. Response to warnings varies depending on the circumstances and expected impacts, such as Category 5 hurricane versus Category 2 hurricane or 100-year ﬂood vs. 10-year ﬂood. The US public is relatively conﬁdent in its agencies’ ability to predict ﬂooding and respond appropriately. In spite of this improved ability to forecast ﬂood events, the accurate forecasting horizon is shorter than the time required for evacuation. Therefore, local decision-makers are unlikely to order a full evacuation due to the 72hour lead-time required. They do, however, often order hurricane evacuation of low-lying and unprotected areas. Flood management will most likely continue to be a shared responsibility within the USA, since there is so much resistance towards central lines of authority. The National Response Framework, which is being drafted to supersede the Federal Response Plan and the National

Response Plan, is designed to ensure better coordination between organizations during a presidential declared disaster. As a nation, the implementation of an effective ﬂood mitigation strategy remains a grand challenge. 3.3.5 Land use changes Instead of building ﬂood defences higher or executing measures to reduce excess water levels, other measures are possible, like land use planning. In the Netherlands, there are some small initiatives in this direction. Flood management is mainly restricted to the area between the dikes, while in the areas outside the ﬂoodplains no limitations are set for development. Exceptions are the areas which have been addressed in Space for the River as long-term reservations. Increasingly in new developments, however, ﬂood probabilities are taken into account leading to considerations such as ﬂood-resistant building (building on mounds, for instance). There are also some examples of ﬂoating housing projects, but up to now this is mainly in the recreational category. In other European countries (Germany, Belgium) the situation is slightly different, due to the geographical situation, often depending on whether the river has dikes or not. In the upper part of the Rhine in Germany, for instance, the river ﬂows through a natural valley and there are no dikes to protect the surroundings. Here the possibility of removing the smaller dikes, which protect agricultural land, could be considered to provide extra conveyance capacity to the river in times of high discharge. A method has been developed to assess the effects of climate and land use change on hydrological conditions and species composition in Dutch riverine (limited to small streams) grasslands (see Runhaar et al. 2002). The study concluded that climate effects are limited compared to land use effects. Land use effects, in this case modelled as realization of an ecological network, an additional buffer zone and the allowance of free meandering streams, will increase the area of grasslands. Increase of grassland areas may lead to increased roughness of the ﬂoodplains and

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hence to higher water levels and more intense ﬂooding. This may have an additional effect (positive or negative) in the contribution of ﬂoods on the main rivers to which they contribute. In Pﬁster et al. (2004), it is concluded that for small basins, land use change can have signiﬁcant negative local effects on ﬂooding (this holds in particular for the combination of urbanization and heavy local rainfall). No evidence is available, however, that peak ﬂows in the greater Rhine and Meuse catchment will be affected by changes in land use. This is despite the change in the character of the Rhine from a combined rain-fed/ meltwater river into a mainly rain-fed river as a result of climate change.

3.4

New Scientiﬁc Insight

Flood management in the Netherlands (as in many other countries) is currently still based on the chance that the water level in the rivers exceeds a critical level. This critical level is related to the statistical return period of a certain discharge: the return period is 1 : 1250 and the design discharge is 16,000 m3s−1 at Lobith. There is of course a relationship between the design water levels and the actual dike height. In the design of the dikes, the wave run up (due to wind) is taken into account. The height needed for this run up (with a minimum of 50 cm, and calculated deterministically) is added to the design water level as a freeboard. Furthermore, some additional height for setting the properties of the dikes is included as a safety factor. If a dike has been designed in accordance with the guidelines, then according to the theory and models, the dike would withstand a discharge of 16,000 m3s−1 but would collapse at a discharge of 16,001 m3s−1, because of overtopping, subsequent erosion of the outer side of the dike and ﬁnally dike failure. This is of course not the reality. On the one hand, dikes are often robustly designed, such that they can withstand more than precisely the design discharge. On the other hand, there are a lot of uncertainties which may lead to a failure of the dike even under conditions which are

below design levels. In general, apart from the wave run up, the freeboard also accounts for some of the uncertainties. In recent (2007) guidelines for dike development, this element has been given a more formal status. In every new dike, a height of 30 cm has to be incorporated to account for uncertainties. This means that the freeboard for new dikes has to be at least 80 cm. Because of this rather large freeboard there are conditions possible (for instance no wind and very internally strong dikes) such that the maximum discharge is actually larger than 16,000 m3s−1. In these circumstances, the freeboard is not needed for wave run up but may help add strength to withstand excess discharges. It is clear that a lot of uncertainties remain that are either not included in the model or are modelled in too simpliﬁed a way, because our knowledge of these processes is limited. There are two main categories of uncertainty (Kok et al. 2003); one is natural variability and the second relates to lack of knowledge. Natural variability denotes the intrinsic unpredictability of modelling physical processes. Modelling wind, hydraulic roughness, the discharge ratio between summer bed status and winter bed status, all belong to this category, as well as uncertainties in the peak discharge and mechanisms like seeping and piping. The discharge ratio at bifurcation points is another source of (natural) uncertainty, which turns out to be very important. In the Flood Protection Act, a certain discharge ratio is deﬁned which has to be maintained. Deviations may cause unexpectedly high water levels along the branches downstream of the bifurcation points, which might be above the design water levels. Therefore, nowadays there are some initial thoughts to make some constructions near these bifurcations to control the discharge ratio whenever necessary (see Schielen et al. 2007, 2008). The second category of uncertainty is related to lack of knowledge. This comes back to the model itself (how does the ﬂow behave around structures for instance, and is this mimicked by the model?). There are also uncertainties related to probability distributions of discharge series (statistical uncertainties). All these factors mean

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Fig. 3.8 The relation between the discharge and the return period. The design discharge is deﬁned as the discharge with the return period of 1250 years (see dashed lines). From a statistical analysis of historical discharges, it turns out that the relations between discharge and return period is given as Q = a + ln(T) = b, with a = 1316.44 and b = 6612.61 (see Van den Langemheen and Berger 2001).

that results from the computer model must be accompanied by uncertainty analysis to put the results in an appropriate perspective. Studies of the Ministry of Transport, Public Works and Water Management (only published as internal memos) show an absolute uncertainty with respect to predicted design water levels up up to almost 20 cm for the river IJssel (and up to 15 cm for the other branches) as a result of natural variability and lack of knowledge. When the uncertainties are taken into account, a transition from exceedance probabilities (i.e. the exceedance of a critical water level) to ﬂooding probability (i.e. the probability that a dike fails due to overtopping, failure of slopes, or other) can be made. In Stijnen et al. (2002), this analysis is done in a systematic way, although in Vrijling (2001) the ideas were already pointed out. The starting point is the situation in which the design discharge at Lobith (16,000 m3 s−1) leads to a design water level (say x m + MSL), and that the dikes at those locations have an (assumed) height of x + y, where y is the freeboard. If the dike height is exactly equal to the design water level, and only the discharge is included as a variable, the ﬂooding probability is exactly 1 : 1250. When we assume that the discharge is the only variable, and also assume that the freeboard is strong enough to withstand the water, it is clear that more discharge can be transported. The exact ﬁgure can be determined by a simple cal-

culation using an increased discharge. The relationship between the discharge and the return period (Fig. 3.8) then gives the ‘actual’ ﬂooding probability. The result is that this decreases signiﬁcantly. In addition, apart from the discharge, other sources of uncertainty can be introduced and, just as described above, the probability of failure can be repeatedly calculated. To incorporate the uncertainties due to the peak discharge under the condition that dikes can also fail due to overtopping of waves, the design conditions for wind are incorporated in the calculations. Note that in the guiding principles for dike design which are usually used in the Netherlands, the wind conditions are prescribed in a deterministic way. If we take these deterministic conditions for wind into account, and look at wave overtopping as a failure mechanism (which is a stronger condition than ‘just’ overﬂowing due to high water levels as a result of a high discharge), the probabilities of ﬂooding decrease (at Lobith) up to 1 : 3907 (with a discharge of 17,500 m3 s−1). Compare this to the design discharge of 10,000 m3 s−1 at 1 : 1250. These steps can be repeated for the other uncertainties (e.g. wind direction, wind strength, water level), and also the combination of all those uncertainties. Wind direction and wind strength, for instance, lead to a decrease of probability to 1 : 1903 and 1 : 1841, respectively.

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Once the probabilities have been determined, the risk of ﬂooding can be established. Note the difference between ﬂood probability and ﬂood risk. With ﬂood probability, the probability of failure of the protection structures leading to hinterland ﬂooding (levees, dikes, sluices) is denoted. Flood risk is deﬁned as ﬂood probability times the potential damage caused by the ﬂood. Necessary for the estimation of ﬂood risk, therefore, is an investigation of the economic value of the land that is protected by the dikes. The value of real estate can quite easily be determined, but historical and cultural values, nature values and human lives are difﬁcult (if not impossible) to express in monetary terms, and may even incorporate an ethical discussion. This subject has been extensively dealt with by Jonkman (2007). In order to reduce the ﬂood risk, there are two possibilities: either reduce the probability, or reduce the potential damage (Box 3.1). The probability can be reduced by reducing the water levels, reinforcing the dikes or by trying to reduce the failure mechanisms. Reduction of the potential damage can be achieved by dividing the dike section into separate compartments, or by reducing the economic value through industrial relocation, for example. Making new compartments is not straightforward. Smaller compartments may

Box 3.1

ﬁll up more quickly, and may hence lead to more damage. The location for compartment dikes must therefore be assessed carefully. All these actions cost money, and in order to see whether this is a sensible investment, the reduction of risk (the difference between the risk in a reference situation in a certain year, and the risk in that same year, with risk-reducing measures taken) must be considered against the investments. Note that both quantities are expressed in (/yr. If the reduction in risk is larger than the investments, the investment is sensible. The interesting point is that in this way, the optimal probability for every dike section can also be determined. The investments and reduction of risk must be in balance, otherwise the investment is useless from an economic point of view. A risk analysis can also reveal that some sections are already ‘too safe’. This is to say that no investment balances a reduction in risk. Increasing the risk, however (for instance by lowering the dike), is of course not considered. Also, future economic developments play an important role in the risk analysis. New investments change the risk and hence also the optimal probability. This means that signiﬁcant new investments in a dike section would actually mean that the risk analysis must be repeated and appropriate measures might have to be taken.

Flood probabilities – the FLORIS Study

The approach of looking at ﬂooding probabilities rather than exceeding probabilities has been followed in several studies. In the FLORIS Study (Van der Most and Wehrung 2005), the ﬂooding probabilities are calculated for all the different dike sections, with a more thorough modelling of the failure mechanisms of the dikes. Failure mechanisms, like overtopping, overﬂowing, instability of inner or outer slope, piping and erosion of outer slope, are taken into account. The piping mechanism (at discharges signiﬁcantly below design conditions!) turns out to have a very large inﬂuence on the probabilities, as is shown in the ﬁrst phase of the FLORIS Study. For some sections, the analysis results in larger ﬂooding probabilities than exceeding probabilities. However, it is generally agreed that the physical aspects of the mechanism are extremely difﬁcult to model and as a consequence, the results are very sensitive. Therefore, in the next phase of FLORIS, special attention is paid to the subject of piping. In practice, piping does occur, but seldom results in the collapse of a dike. If piping is detected in time (as a result of an intensive inspection during periods of high discharge), appropriate measures can be taken to control the phenomenon. The FLORIS Study is not ﬁnished yet. Preliminary results can be seen on www.projectvnk.nl; in Dutch.

Flood Management These elements have been studied intensively by the Ministry of Transport, Public Works and Water Management (2006). The starting point of the study was the situation after the completion of Space for the Rivers, and after the completion of an additional programme which ensures that the existing dikes are sufﬁciently strong and high. Hence, the defence system is up to date. In this situation, ﬁve different options have been studied in order to assess an increase of the design discharge of up to 1000 m3 s−1. The ﬁve options were: 1 International cooperation – In this option, measures implemented in Germany that might lead to a reduction of discharge in the Netherlands were assesed. It turned out that this was hardly an option, since the safety levels in Germany are lower than in the Netherlands. 2 Organizational measures – This includes reinforcing dikes with sandbags when necessary and reducing piping problems through the construction of boxes around the place where the water up-wells (Fig. 3.9). 3 Emergency retention polders – The location of several areas in which the excess water could temporarily be stored was studied. This is a costly measure, since all kinds of protection measures have to be taken to keep the water in the designated areas. 4 Dividing large sections into smaller sections by compartments – This involved a study of the potential location for the compartment dikes and a system of smaller dikes and levees to guide the water. 5 Increase the norm to a design discharge to 17,000 m3 s−1 – In this option, both additional spatial measures in addition to Space for the River have been investigated, as well as raising the dikes. Interestingly, only option 2 is found to be successful and provides for a relatively small investment which leads to a very signiﬁcant reduction of the risk. Whether construction of compartment dikes is successful depends on the speciﬁc locations. The same holds for emergency retention polders. Other options are just not proﬁtable and although there is obviously a reduction in

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Fig. 3.9 A typical example of piping. Water wells at the land side of the dike. (Photograph courtesy of Deltares.)

risk, the investments needed to achieve that reduction do not balance. It should be mentioned that these results hold for the Rhine, but the options for the Meuse have also been studied. One of the results is that applying organizational measures is proﬁtable for the Meuse, but to a signiﬁcantly lesser extent. Altogether, it is clear that if a straightforward risk analysis is undertaken, it is very likely that different dike sections end up with different risks. The question is whether this is a preferable situation. The current system, based on the exceeding probabilities, produces a difference within the Netherlands (1 : 1250 in the eastern part, 1 : 10,000 in the western part). Whether it is fair to differentiate risk levels between the different dike sections in the Netherlands even further is the subject of ongoing discussion. There is one important reason to take into account lower probabilities than those which come out of a risk approach. Although the probability might be very small, the result of a ﬂood can be potentially devastating for a total society. The ﬂooding of cities like The Hague, Amsterdam and Rotterdam might bankrupt the Netherlands. Putting an economic value on human lives is almost impossible, so including them in the assessment of ‘potential damage’ is questionable. The loss of cultural elements is equally unacceptable. All of these arguments can mean that in

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cases where the reduction of the risk does not balance the investment, the investments are justiﬁed from the point of view of these noneconomic arguments. Besides, ﬂood protection, or keeping the water outside, is ﬁrmly ﬁxed within the Dutch culture. Hence, all initiatives to return land to the rivers or the sea is almost always strongly opposed by the public Other ﬂood protection measures are generally accepted in the Netherlands, although as soon as measures are constructed in the direct neighbourhood of private properties, there is often a clear and obvious opposition to the plans. It is interesting to raise the question of what norms should be applied for an appropriate ﬂood management strategy: exceedance probabilities, ﬂooding probabilities or risk numbers. Every option has advantages and disadvantages. Exceedance norms are the easiest to test from a governmental point of view. They are difﬁcult to explain to the general public, however. Norms based on actual ﬂooding probabilities are perhaps easier to explain to the general public, but uncertainty remains about the nature and the modelling of the failure mechanisms and hence the interpretation of the outcome. Beside, these ‘norms’ do not take into account economic activity or nature and cultural values in the protected areas. From the latter point of view, a system of risk numbers might be the best, although this raises the question of what to do in case of a future large scale investment. This increases the risk, which must be balanced in order not to exceed the ‘norm’. Besides, in an area with little economic development there might still be an acceptable risk number, but in combination with an unacceptably high ﬂooding probability. So this system should be accompanied by something like a maximum ﬂooding probability. 3.4.1

The ‘safety-chain’

In many disciplines related to risks and crises, it is good practice to use the so-called ‘safety chain’ to structure and evaluate the potential for disruption of society. The safety chain has ﬁve elements:

• pro-action: excluding risks; • prevention: minimizing risks; • preparation: preparation to suppress a crisis; • repression: actual suppressing of the crisis; • recovery: returning to a normal situation. Sometimes, a sixth one is added: learning from the event. In the Netherlands, up to the time of and including the Space for the River programme, attention was focussed on the ﬁrst two links of the chain. In doing this, there was little attention to recovery (for this was seldom needed). However, nowadays there is a clear tendency to pay attention to the links between preparation and repression (Ten Brinke et al. 2008a). In case a crisis threatens, the Dutch authorities (on different levels, from ministries to local communities) have an adequate information system, evacuation plans and crisis-management teams. From time to time, signiﬁcant exercises involving cooperation between ministries, local communities and emergency services (ﬁre brigades and ambulances) are held. These procedures are constantly evaluated and improved. Preparation is strongly connected to awareness. It becomes increasingly clear that if people are not aware of the nature of the land they are living in, it is hard to get them to think about ﬂood management in the ﬁrst place. This goes along with the impression that the majority of the population has an almost absolute faith in the protecting capacity of the dikes and dunes. Partly, this is because the ﬂood defence system, with the concepts of return periods, probabilities, failure mechanisms, etc., is hard to understand. Partly, also, it is because the memory of people is short: events that took place in 1993 and 1995 (let alone the big ﬂood of 1953) are almost forgotten. Finally, it is partly because people think that the government will protect them in all cases, a thought which to some extent is motivated by the high standards in the Netherlands. There is, however, a growing need for increased awareness. The Dutch government has public campaigns to draw attention to climate change and the consequences for water, for instance, and on internet sites people can get information about the prob-

Flood Management ability of all kinds of risks, including ﬂooding risks. This is based on postal codes and hence, people can get information about the elevation of their private properties with respect to high water levels. The thoughts around safety, risks and probabilities of ﬂoods change over time. New and more detailed data become available as well as new scientiﬁc insight. The method used in the Netherlands to assess ﬂood risk dates from the 1950s. As already proposed in Ten Brinke and Bannink (2004), it is worthwhile evaluating the safety approach itself every 25 years or so. This is exactly what is happening now and there is a tendency towards a transition: going from exceeding probabilities to ﬂood risk. Meanwhile, the legal standards are still valid, and have to be met. This is why Space for the River is implemented. The thinking about a new approach continues, however. When Space for the River is completed in 2015, it will be an appropriate time to switch to another approach, for the river system is then ready for the next decades of potentially high discharges.

3.5

Synthesis

On 4 October 1986, the Dutch queen opened the Delta Works following completion of the storm surge barrier in the Eastern Scheld. The queen stated that ‘The storm surge barrier is closed. The Delta Works are completed. (The province of) Zeeland is safe.’ In 1997, she opened the storm surge barrier in the ‘Nieuwe Waterweg’ near Rotterdam. This was actually the ﬁnal construction of the Delta Works. Again it was stated that the Netherlands were safe. Safety, however, is a relative concept and absolute safety does not exist. A statement like ‘reasonably safe’ would be more appropiate, but this is something that the public might ﬁnd confusing. Statistics are a useful and necessary tool on which to found ﬂood management approaches. Statistics, however, do not predict when the next period of (extremely) high discharge will come, and neither can it be used to ‘predict’

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that after a high discharge there will not be another one for a number of years. Therefore, it is vital that any country not only has its ﬂood protection measures up to date, but also has crisis management up to date. So whenever, despite all the measures taken, a crisis occurs, the government and the people themselves are fully prepared. For this, another way of dealing with ﬂoods is needed. In Schielen and Roovers (2008) some initial thoughts on this topic are discussed. For this ﬂood awareness is vital, as is keeping evacuation plans and crisis control tools up to date. This cannot prevent casualties in all cases, but will certainly help to minimize it. In the very long run, say in the next few centuries and beyond, the question remains whether a country like the Netherlands can continue the current strategy of ﬂood management. Climate change will most probably play a very important role: the sea level will rise and the rivers will show more extremes in their discharge regime. The general impression is that a sea-level rise of 1–2 m can still be dealt with using sand supplementation to reinforce the coastal line. The real problem will be in the rivers, where the sea-level rise will cause higher water levels in the downstream part of rivers, more difﬁculties in transporting the water to the sea and more salt intrusion. There are, therefore, opinions which question the Dutch approach (e.g. see Enserink and Bohannon 2005). For the next 200 years it is likely that sea-level rise can be taken into account in building ﬂood protection. A large industrial extension of the harbour of Rotterdam for instance is built at 5 m+ MSL which guarantees sufﬁcient safety. It would be wise to take similar measures for every large investment project in the western part of the Netherlands. One way or another, to protect the Netherlands from ﬂooding in the future will cost more and more money and there is an urge for innovative solutions and perhaps also a paradigm shift in our behaviour towards the water. Only then, the words of the famous Dutch engineer Johan van der Veen, who stated half a century ago that there will be a moment in the (far) future when the Dutch, with

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a sigh of relief, will give their land back to the sea, will not come true.

3.6

Conclusions

As a result of natural variation in meteorological circumstances which cause the wide range of discharge and water levels, ﬂooding from rivers will always be a potential threat. Many centuries ago, people started to restrain rivers by building dikes to protect themselves from ﬂooding and ﬂood protection has been the goal. Nowadays, ﬂood management is the trend. The protection levels throughout the world differ. In the Netherlands, the protection levels are very high as a result of the geographical location of the country. The sea and the rivers Rhine and Meuse pose a constant threat, and in combination with the high population density and economic development, there is a need to maintain high standards. Apart from that, there are new developments in the approaches adopted to ﬂood management, for instance a risk approach, where the potential damage is taken into account. Another develop-

Box 3.2

ment is a study towards a differentiated protection level in which less developed areas, from an economic point of view, or less populated areas, might have a lower protection level. This requires that the evacuation possibilities or other measures, such that loss of lives and goods are minimized, are in place. Which ‘norm’ is accepted, based on exceedance or ﬂooding probabilities, or risk numbers, is the subject of discussion and is dependent on the speciﬁc situation in a country. Throughout Europe, similar developments have taken place. In Hungary, for example, a large programme has been undertaken (the Vásárhely Programme) to increase the protection levels along the Tisza River up to a return period of 100 years (Box 3.2). This is done through the construction of side channels, dike relocation and dike reinforcement. Large retention polders are constructed such that discharges with a return period of 1000 years do not cause any problems. Local communities are protected by concrete levees instead of dikes, just as is done on parts of the River Meuse in the Netherlands. Note that the return period of 1000 years is close to the safety norm which is used in the Netherlands.

The New Vásárhelyi Plan

The four extreme ﬂoods of 1998, 1999, 2000 and 2001 made it clear that the ﬂood protection of the Tisza River needed to be improved. Flood protection on the Tisza River dates back to the middle of the nineteenth century, when the engineer Pál Vásárhelyi drafted the original ﬂood protection plan and although he died before the constructions started, the name ‘Vásárhelyi Plan’ was given to the project in his memory. After a thorough but quick preparation, the Government of Hungary in its decree of 1022/2003 (III. 27) accepted the conceptual plan of the improvement of ﬂood protection of the Tisza River and launched phase one of the New Vásárhelyi Plan. The basic concept of the New Vásárhelyi Plan is to convey the 1 in 100 year ﬂoods between the dikes and to protect the land against higher ﬂoods by storing the water in excess of the 1 in 100 year ﬂood in ﬂood retention reservoirs. The use of these reservoirs is expected to reduce the level of 1 in 1000 year ﬂood by about 1 m along the Hungarian Tisza. To achieve this goal: 1 The existing dikes have to be built up to the present design conditions (e.g. the level of 1 in 100 year ﬂood plus 1 m freeboard). At present about 40% of the 2822 km long dikes system (including the tributaries of the Tisza River too in Hungary) do not meet the design conditions. 2 The ﬂood conveyance capacity of the ﬂoodplains has to be restored by removing obstacles from the ﬂoodplains. Because the ﬂoodplains of the Tisza River are wide (e.g. 1–5 km), in most of the cases the ‘removal of obstacles’ means opening a 300–600 m wide shortcut on the ﬂoodplain (‘hydraulic corridor’).

Flood Management

73

3 The extreme ﬂoods have to be reduced by storing some of the ﬂood volumes in ﬂood retention reservoirs. 4 The Plan has to be implemented in co-ordination with ecological requirements, land use planning and agricultural programmes. During the feasibility phase 30 potential reservoir sites were investigated. The selection process considered geological, technical, ecological, economic and social conditions. Priority was given to the ecological requirements and to social acceptance. During the preparatory phase more than 500 meetings were organized to present the plan to the local population and to seek their advice. Finally, 10–11 sites were selected for construction with the potential to store 1.5 billion m3 of water. The reservoirs can be ﬁlled and drained by hydraulic structures (mainly gates, although ﬁxed weirs are used where the lead time is short). The structures can also be used to inundate the reservoirs during smaller ﬂoods to provide water for agriculture and/or wetlands. During the public debates requests to establish wetlands, to provide water for agricultural use and even to create standing water surfaces were received. These requirements were incorporated into the Plan. Phase 1 of the Plan includes construction of six ﬂood retention reservoirs, opening ‘hydraulic corridors’ at several locations and relocating the dikes at one location. Construction of two reservoirs was completed in 2008 while the preparation of the construction of the others is ongoing. The New Vásárhelyi Plan is expected to be ﬁnished in 25 years.

There are numerous further similarities between the Vásárhely Programme in Hungary, and the Space for the River Programme in the Netherlands. Due to European regulation, nature conservation and nature development (Natura 2000) must be taken into account, which demands a thorough communication between the different authorities. Also the fact that the rivers ﬂow through historical cities means that for both countries, there must be a subtle engineering of measures to spare characteristic buildings along the rivers. Awareness of the ﬂood management in both countries is also strikingly similar. In the Netherlands, there are campaigns on national television to raise ﬂood awareness. These campaigns are necessary because of the high level of protection, and only very occasionally does high water itself cause a problem. In Hungary, the situation is different, since in the last decade there have been ﬂoods almost every year and, hence, public awareness is high. This is also clear from a recent poll, in which 98% of the population in the Tisza region responded that they were well aware of the potential threats of the river. A few years previously this was only 2%! There are also, however, some differences. Constructing retention polders is relatively easy

in Hungary, because there are only very few people living in those areas. The land around the rivers is mainly agricultural and farmers can be compensated if the retention polder is used. Retention polders in the Netherlands are, due to the large population density, very expensive. There are no empty areas left, all the villages need to be protected by levees and public opposition is strong. Also in Belgium (Meuse, Scheldt), Germany (Rhine, Elbe) and Romania (Bega, Timis), the governments make and execute plans to manage ﬂoods. The solutions are often quite similar: dike reinforcement in combination with ﬂood plain plans to increase structural safety on one the hand, and retention polders where possible on the other hand. The plans are almost always dealt with in an integral way: the interests of agriculture, nature and recreation are taken into account and often determine how the plans are ﬁnally engineered. It is encouraging to see that throughout Europe, increasing efforts are being taken to manage ﬂoods. The associated projects try to couple ﬂood management with many other issues and so improve the spatial quality of the river area. As many European countries share the same

74

Box 3.3

ralph m. j. schielen

The Belgian perspective

In Flandres, the struggle against ﬂooding was based on the concept that as little land as possible was allowed to be inundated. As a result, high dikes were built close to the river. More recently, Belgium has adopted an alternative approach which looks to work with nature, instead of against nature. It is now realized that ﬂooding cannot be conquered, but merely managed. Flooding is no longer prevented at all costs. Managing damage is now equally important. Space is given back to the river, in the sense that natural ﬂood areas are restored. In Wallony, the regional government has taken an initiative called ‘Plan Pluie’ (Rain Plan or ‘Prévention et Lutte contre les Inondations et leurs Effets sur les Sinistrés’ which is translated as ‘Prevention and Battle against Inundation and Related Effects in Case of Floods’). The ﬁve objectives of the plan are: • Improve the knowledge of risks related to high water levels and ﬂoods. • Diminish and slow down water ﬂows from tributaries and basins. • Restore ﬂoodplains. • Reduce vulnerability in ﬂood zones. • Improve crisis management in case of disasters. This text is based on the information found on the ofﬁcial website of the Flemish Government (http:// www.lin.vlaanderen.be/awz/html/overstrom.htm; in Dutch) and the ofﬁcial website of the Walloon Government (http://environnement.wallonie.be/de/dcenn/plan_pluies/index.htm; in French) problems with their rivers, international cooperation is vital. Exchange of knowledge is one way of cooperation. Another way is ﬁnancing river projects in other countries. In Germany, some retention polders have been constructed with Dutch ﬁnancial backing. The construction in Germany is much simpler, since the population density is less. The beneﬁt for the Netherlands is clear, since the retention polders create lower water levels in periods of high discharge. Many countries struggle with the issue of climate change. It is generally thought that more measures need to be taken to provide continued ﬂood protection into the future, but the magnitude is often not clear due to the wide variation in predictions of sea-level rise or increase in precipitation. At this time, it is meaningless to make realistic plans to abandon the densely populated deltas, like the western part of the Netherlands. It would be wise, however, to take precautionary measures for new, large-scale projects to anticipate the consequences of climate change. For instance, a series of new islands in front of the present shore creating a second shoreline, ﬂoating houses in the ﬂoodplains, or very wide dikes upon which housing or recreation is facilitated, have been suggested. Initiatives like these are worthy

of study. They create awareness, and help to create a sustainable society for the future.

Acknowledgements The author is grateful for the constructive comments of his former colleagues Wilfried ten Brinke and Marcel de Wit. The text of section 3.3.4 (the American approach to ﬂood management) was kindly provided by Paul Bourget, Institute for Water Resources United States Army Corps of Engineers, USA. The text in the box on the New Vásárhelyi Plan was kindly provided by Péter Bakonyi, National Hydrologic Forecasting Center VITUKI, Hungary.

References Delta Committee (1960) Report of the Delta Committee. The Hague, The Netherlands (in Dutch). De Vriend, H.J. and Dijkman, J.P.M. (2003) A new method of decision support to river ﬂood management. In: Proceedings of the First International River Forum, October 2003, Zhengzhou, China, pp. 63–71. Enserink, M. and Bohannon, J. (2005) Hurricane Katrina: questioning the Dutch solution. Science, 309, 1809.

Flood Management Groenewoud, W., Lorentz, G.K., Brouwer, F.J.J. and Molendijk, R. (1991) Geodetic determination of recent land subsidence in the Netherlands. In: Land Subsidence. Proceedings of the 4th International Symposium on Land Subsidence, May 1991. IAHS Publication 200. Balkema, Rotterdam, pp. 463–471. Jonkman, S.N. (2007) Loss of Life Estimation in Flood Risk Assessment: theory and applications. PhD Thesis, Delft University of Technology, The Netherlands. Kok, M., Stijnen, J.W. and Silva, W. (2003) Uncertainty Analysis of River Flood Management in the Netherlands, Proceedings of ESREL 2003 (European Safety and Reliability Conference 2003), 15–18 June 2003, Maastricht, The Netherlands. Ministry of Transport, Public Works and Water Management (2006) Syntheses – report of the research programme crisis management strategies for Rhine and Meuse. The Netherlands (in Dutch). Ministry of Transport, Public Works and Water Management (2007) Hydraulic Boundary Conditions for the Primary Water Defences for the Third Assessment 2006–2011. The Netherlands (in Dutch). Pﬁster, L., Kwadijk, J., Musy, A. Bronstert, A. and Hoffmann, L. (2004) Climate change, land use change and runoff prediction in the Rhine-Meuse basins. River Research and Applications, 20, 229–241. Pollard, M. (1978) North Sea Surge: the story of the East Coast ﬂoods of 1953. Terence Dalton, Lavenham. Runhaar, J., Van Walsum, P.E.V. and Prins, A.H. (2002) Effects of climate and land use change on hydrological conditions and species composition in Dutch riverine grasslands (Calthion, Junco-Molinion). Ecohydrology & Hydrobiology, 2, 219–226. Schielen, R.M.J. and Roovers, G. (2008) Adaptation as a Way of Flood Management, Proceedings of the 4th International Symposium on Flood Defense, Toronto, Ontario, Canada. May 4–6, 2008. Schielen, R., Jesse, P. and Bolwidt, L. (2007) On the use of ﬂexible spilllways to control the discharge distribution of the Rhine in the Netherlands: hydraulic and morphological observations. Netherlands Journal of Geosciences-Geologie en Mijnbouw, 86, 319–330. Schielen, R.M.J., Havinga, H. and Lemans, M. (2008) Dynamic control of the discharge distributions of the Rhine River in the Netherlands. In: Proceedings of Riverﬂow 2008, 3–5 September, 2008, Izmir, Turkey. Stijnen, J.W., Kok, M. and Duits, T.M. (2002) Uncertainty Analysis Flood Defence Rhine Branches: sources of uncertainty and consequences of mea-

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sures. PR464, HKV_lijn in water, The Netherlands (in Dutch). Summers, D.H. (1978) The East Coast Floods. David and Charles, Newton Abbott. Ten Brinke, W.B.M. (2007) Land in the Sea, The Water History of The Netherlands. Veen Magazines B.V., Diemen, The Netherlands (in Dutch). Ten Brinke, W.B.M. and Bannink B.A. (2004) Dutch Dikes and Risk Hikes. A thematic policy evaluation of risks of ﬂooding in the Netherlands. RIVM rapport 500799002. Rijksinstituut voor Volksgezondheid en Milieu RIVM, Bilthoven, The Netherlands (in Dutch). Ten Brinke, W.B.M., Bannink, B.A. and Ligtvoet, W. (2008a) The evaluation of ﬂood risk policy in the Netherlands. Proceedings of the Institution of Civil Engineers – Water Management, 161, 181–188. Ten Brinke, W.B.M., Saeijs, G., Helsloot, I. and Van Alphen, J. (2008b) Safety chain approach in ﬂood risk management. Proceedings of the Institution of Civil Engineers – Municipal Engineer, 161, 93–102. Van Dantzig, D. (1956) Economic decision problems for ﬂood protection. Econometrica, 24, 276–287. Van de Ven, G.P. (ed.) (2004) Man-made Lowlands. History of water management and land reclamation in the Netherlands. Matrijs, Utrecht, The Netherlands. Van den Hurk, B., Klein Tank, A., Lenderink, G. et al. (2006) KNMI Climate Change Scenarios 2006 for the Netherlands. KNMI Scientiﬁc Report WR 2006-01. Available: http://www.knmi.nl/klimaatscenarios/ knmi06/achtergrond/WR23mei2006.pdf (accessed 9 October 2008) Van den Langemheen, W. and Berger, H.E.J. (2001) Hydraulic Boundary Conditions 2001: design water levels for Rhine and Meuse. RIZA report 2002.014. Ministry of Transport, Public Works and Water Management, The Netherlands. Van der Most, H. and Wehrung, M (2005) Dealing with uncertainty in ﬂood risk: assessment of dike rings in the Netherlands. Natural Hazards, 36, 191–206. Van Heezik, A. (2006) Battle for the Rivers, 200 Years of River Policy in the Netherlands. HNT Historische Producties, Den Haag (in Dutch). Vrijling, J.K. (2001) Probabilistic design of water defense systems in the Netherlands. Reliability Engineering and System Safety, 74, 337–344. Wisner, B., Blaikie, P., Cannon, T. and Davis, I. (2005) At Risk. Natural hazards, people’s vulnerability and disasters, 2nd edn. Routledge, London.

Image facing chapter title page: Courtesy of the Centre for Ecology and Hydrology.

4 Ecological Consequences of River Channel Management NIKOLAI FRIBERG1 1

The Macaulay Institute, Craigiebuckler, Aberdeen, UK and National Environmental Research Institute, Aarhus University, Silkeborg, Denmark

4.1

Introduction

Streams and rivers create a dense network of freshwater throughout the landscape and they are intimately related to their valley, constituting an ecological entity. On a larger scale, processes and persistence in rivers are dependent on the catchment from which they derive their water. Compared to other ecosystems, streams and rivers are among the most species rich as the close linkage with the terrestrial ecosystem across the ecotone between land and water allows co-existence of numerous species of plants and animals. Flow and habitat features vary considerably in space and time, in and among river systems, which further increases diversity. Due to their unidirectional ﬂow, rivers are vital in transporting nutrients and organic matter from the terrestrial environment to the sea with an important function of the ecosystem being transformation of coarse organic matter to ﬁne particles and continuous mineralization of nutrients. In many parts of the world, landscapes have been changed by human activities for thousands of years. The intrinsic properties of running waters, the dense channel network and the unidirectional ﬂow provided a unique solution to infrastructure and a source of energy for human

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

settlements. Early changes to streams and rivers typically involved damming, minor diversions of water for irrigation and regulation of channels to allow navigation. Rivers were used for water mills and for transporting large quantities of goods. Concurrent landscape changes, often involving considerable deforestation, have additionally altered the ﬂuxes of carbon and nutrients between the terrestrial ecosystems and rivers. After the industrial revolution, pressure on river ecosystems from human activities has increased substantially and growing population densities have increased demand for ecosystem services such as water for consumption and waste disposal. Rivers have been straightened and channelized to ensure drainage of surrounding agricultural areas, to facilitate navigation and as ﬂood protection. Furthermore, many river reaches are continuously maintained by dredging of bottom sediments and removal of woody debris and weeds. This has resulted in a loss of habitats and ecosystem complexity and the interaction with the river valley has been disconnected. Furthermore, the ecosystem capacity to resist other pressures such as organic pollution is reduced because surface area and re-oxidation are decreased. The current physical structure of rivers and the diversity of biological communities are closely linked with past and present human activities in the catchment. Human activities inﬂuence stream ecosystems on multiple scales, ranging from direct manipulation of the in-stream

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Longitudinal processes: • Large scale dispersal of biotic components • Landscape fragmentation • Loss of longitudinal connectivity • Changes in flow regime In-stream processes: • In-stream habitat complexity • Habitat modification (dredging, weed cutting) - simplification

Interaction processes: • Small scale dispersal • Energy transfer • Riparian habitat structure

• Channelisation - truncation of natural gradients • Land-use (afforestation, agriculture etc.)

environment (channelization, removal of large woody debris, etc.) on the stream reach to altering the landscape and land use in the catchment, thereby inﬂuencing the hydrological pathways and morphological structure (Fitzpatrick et al. 2001; Allan 2004). Past and present disturbances impact the stream ecosystem elements and it can be difﬁcult to distinguish the exact disturbance from the individual stressors on the biotic communities (Harding et al. 1998; Allan 2004). Channel management has a multitude of impacts on river ecosystem at different spatial and temporal scales (Fig. 4.1). Rivers play a key role in modern society and their primary role is in providing ecosystem services related to consumption, irrigation and transportation. These services have been utilized without considering sustainability and throughout the world destruction of river ecosystems has been substantial. By destroying ecosystem processes and properties several other economically valuable services and long-term beneﬁts to society are lost. It is vital, therefore, that catchment management is put into an ecological context to ensure that river ecosystem services are used in a sustainable way in the future.

4.2

Fig. 4.1 Large-scale changes include fragmentation of the riparian habitat, loss of longitudinal connectivity by, for example, the instalment of weirs as part of channelization and changes in ﬂow regime due to drainage, etc. Localized in-stream habitats are being simpliﬁed by removal of coarse substrates and biological material such as plants and dead organic matter. Channelization and land use will further truncate the natural interaction between the stream and the riparian zone which is essential in maintaining exchange of energy over the ecotone and for sustaining the high species diversity.

River Channel Morphology and Management

Natural river morphology depends on catchment scale structural controls, reach scale channel pattern differences and micro-scale variations in channel bedform, all of which vary over different time-scales (Frissell et al. 1986). This knowledge can be used in hierarchical classiﬁcation of different types of natural morphology based on descriptions of features such as dominant bed type, entrenchment ratio, sinuosity, width:depth ratio and water surface slope (Rosgen 1994). Naturally, therefore, there is a variety of morphologies that will respond differently to management (Downs 1995). In an ecological context, however, the important issue relates to spatial and temporal heterogeneity provided by natural stream channels which creates a range of habitats for the biota. At the reach scale, current velocity, depth and stream bed substratum vary in a predictable manner in natural stream channels with the distance between two neighbouring rifﬂes in most cases being 5–7 times channel width (Fig. 4.2). Rifﬂes and pools are distinctively different biotopes (Box 4.1) with respect to substratum,

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79

Fig. 4.2 Current velocity, depth and streambed substratum vary in a predictable manner in natural stream channels.

depth and current velocities which enable the co-existence of a large number of species. Rifﬂes are gravel accumulation characterized by high water velocity, shallow depth and coarse substrates that often provide a three-dimensional sub-surface habitat. Constant ﬂow through the gravel ensures a ﬂux of oxygen and matter which will support invertebrate communities living in the interstitional spaces or secure the survival of

Box 4.1

ﬁsh eggs. A natural stream channel will, furthermore, contain well deﬁned areas of both erosion and deposition which induces a high degree of habitat heterogeneity. In addition, accumulation of dead organic matter in depositional areas on the stream bed provides the major source of carbon to higher trophic levels in running waters and in-stream channel retention of organic matter is essential in sustaining ecosystem function.

Where organisms live – the habitat concept

In understanding how the biota in running waters respond to the physical environment and its degradation it is essential to make observations at a scale relevant to the organisms. The most common operational unit used is ‘habitat’ which in its original deﬁnition is an ecological or environmental area that is inhabited by a particular species. (Townsend et al. 2008). However, the term is often used more broadly to include assemblages of species living together in the same space and this inconsistency with the original deﬁnition has recently prompted the increased use of the term ‘biotope’ in studies on physical–biological linkages in running waters. A biotope is deﬁned as an area of uniform environmental conditions providing a living place for a speciﬁc assemblage of plant and animals and is, therefore, a more appropriate and precise term than habitat. In studies from running waters the terms ‘functional habitat’ (Kemp et al. 1999) and ‘mesohabitat’ (Armitage et al. 1995) have all been used more or less synonymously for the word biotope, i.e. as homogeneous areas in terms of bottom substrate and current velocity range that are especially important for explaining species distribution. The underlying concept is that number of biotopes will naturally reﬂect river types but also be directly affected by anthropogenic intervention thus establishing a link between channel simpliﬁcation and overall macroinvertebrate composition at the reach (river channel) scale.

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Fig. 4.3 Channelized rivers have uniform cross-section and varied in-stream habitat conditions compared with natural channels.

In channelized rivers and streams, uniform cross-sections with steep banks have replaced natural irregular cross-sections and more gently sloping banks. Furthermore, coarse substrate has frequently been removed by dredging as part of stream maintenance. The uniform conditions have reduced numbers and spatial variability of stream habitats (Fig. 4.3). Channelized streams are dominated by large spatial cover of a few habitat types that are very similar with regard to substrate conditions and ﬂow, often characterized as glides or runs. Natural rifﬂe-pool sequence is often broken or missing altogether and the extent of near bank deposition areas is greatly reduced. River channel management can be subdivided into predominantly two elements that differ in intensity and temporal resolution. Often these two components are combined but not always. One element is when the channel form of the river is changed by large-scale engineering works.

This could be channelization of meandering rivers, creating single channels in braided systems, conﬁning river channels using artiﬁcial embankment structures or removing large, natural, in-stream structures such as boulders. These changes to channel form are undertaken once within a well deﬁned time frame (days to years depending on scale) but have a large-scale impact and they are in many cases historical. These changes to a river channel will signiﬁcantly impair natural functions and processes and create a suite of undesired side effects. The other element of channel management is through maintenance which is characterized by repeated, but lower intensity, interventions. Examples are dredging of accumulated sediments, removal of weeds and large woody debris and localized bank stabilization including cutting down riparian vegetation. Maintenance has often been made necessary by large-scale changes to channel form because of undesired side effects such as excess

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Ecological Consequences of River Channel Management Table 4.1

List of common types of channel management

Type Channelization

Boulder removal Dredging

Removing and managing bank vegetation In-stream weed cutting

Removal of (woody) debris Instalment of artiﬁcial structures

Method

Purpose

Straightening and deepening of river channels or conﬁning (braided) multi-channels into single (straight) channels Blasting using explosives Mechanical removal Manual digging in small streams or using a backhoe excavator either on the bank or on a boat

Reduced ﬂooding by increasing bank height Secure drainage of surrounding land by lowering of the water table Reclaiming land by decreasing stream length Boulders and other large physical structures removed to allow navigation Removal of unwanted obstacles to ﬂow and navigation (including timber) to increase in-channel ﬂow capacity To counteract increasing mobilization, transport and accumulation of sediments in channelized streams To increase channel ﬂow To ensure that machinery used for dredging can access river banks Prevent interference with leisure activities such as sports ﬁshing and canoeing

Trees and shrubs are removed and the grass/herbal vegetation is subsequently often managed in summer by cutting Weeds are removed employing a range of methods ranging from hand cutting in shallow streams using a scythe to heavy machinery on boats or using herbicides Cut up manually or removed using different types of machinery Instalment of rip-rap structures, concrete, wood, gabions, deﬂectors and croys

sediment accumulation, increased erosion, increased growth of in-stream plants due to an improved light environment and more ﬁne substrates, etc. (Table 4.1). Processes in river channel management cannot be separated from their riparian zones and, hence, the interaction with the terrestrial ecosystems. Riparian zones provide a substantial heterogeneity to river channels in addition to being donors of organic matter which provide the main source of energy to most stream ecosystems (Allan 1995). River regulation has had a substantial impact not only on channel plan form but also on the riparian zones and the entire ﬂoodplain as exempliﬁed by the River Skjern, Denmark. The regulation of the lower 20 km of the river, which took six years in the 1960s, straightened the main channel and moved its course on the ﬂoodplain (Fig. 4.4a). River length and channel complexity

To secure drainage large woody debris are often actively removed Bank reinforcement of river channels in areas with increased erosion risk Localized habitat improvement for especially ﬁsh in modiﬁed channels

were reduced by the removal of most meanders and a number of lakes. Flooding was prevented by large dikes and the groundwater table was lowered by installing pumping stations and drainage ditches and 40 km2 of arable farm land could be claimed from what were previously meadows and wetlands. This had a signiﬁcant impact on the extent of wetlands in the ﬂoodplain when compared to the situation prior to the regulations (Fig. 4.4b) and the area covered by wetlands totalled only approx. 4.3 km2 at the end of 1990s (Svendsen and Hansen 1997). At a local scale, riparian vegetation has a direct impact on river bank form through its inﬂuence as a roughness element reducing marginal ﬂow velocities and by roots adding tensile strength to bank sediments (Thorne 1990). In addition to the direct loss of habitats, channelization can have negative impacts on water

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Fig. 4.4 (a) The course of the lower 20 km of the River Skjern, Denmark, as mapped in 1871 and after regulation in 1987. (b) Extent of meadows and wetlands along the River Skjern, Denmark, as mapped in 1871 and after regulation in 1987.

Ecological Consequences of River Channel Management

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Fig. 4.5 Annual transport of total and dissolved iron in the southern drainage channel (Søndre Parallelkanal) which drained soils in the central part of the regulated area of the lower River Skjern, Denmark.

chemistry. The channelization of the lower parts of Skjern River resulted in large amounts of soluble iron (Fe) being released when the pyriterich soils in the ﬂoodplain were drained. Soluble Fe was formed when the former anoxic soils were aerated as the water table was lowered. Consequently, the transport of total and dissolved Fe was substantial in the period from 1966 (4 years into the drainage project) to 1995 (Fig. 4.5). It is estimated that 85,000 tonnes of total Fe were transported in the River Skjern during this 30-year period and, compared to the Fe transport in the unregulated part of the tributary River Omme, this corresponds to an approximately 40% post-channelization increase. Concentrations of total Fe decreased signiﬁcantly, however, during the 30-year period reﬂecting that pyrite stores were being used up and that soil depth was reduced by up to 1.5 m as organic matter was decomposed and further compressed by heavy farming machinery which made soils

wetter and more anoxic. Parallel with the decrease in Fe, sulphate (SO4) concentrations decreased signiﬁcantly and this was probably responsible for a concurrent increase in pH. Relationships between large woody debris (LWD), river form and process, and in-stream ecology are complex and operate over a wide range of temporal and spatial scales (Montgomery et al. 2003). LWD provide immediate impacts on stream ﬂow, sediment transport and channel morphology at the local scale (Bufﬁngton et al. 2002) that in some cases will cascade over broader spatial and temporal scales inﬂuencing river and valley form and available aquatic habitats (Collins et al. 2002). At local scales, wood increases physical heterogeneity by enhancing scour of pools, deposition of bars, and creating spatial variation of channel hydraulics that promotes hyporheic exchange and the development of sedimentary patches of different substrate sizes (Mutz et al. 2007).

84 4.3

nikolai friberg Adaptations to a Life in Running Waters

Habitat characteristics, in particular those related to disturbance regime (i.e. frequency, harshness or predictability of events) and habitat diversity (e.g. the availability of refugia), are strong determinants of population distributions and community assemblage by providing evolutionary conditions from which life history attributes and community properties are derived (Southwood 1977, 1988; Townsend 1989). Organisms living in running waters consequently show a range of adaptations that increase both their resistance and resilience to disturbance and the physical impact of ﬂow. Resistance refers to the ability of a system to withstand displacement by a distur-

Box 4.2

bance whereas resilience refers to the rate at which a system recovers following a perturbation. Organisms experience drag forces which are dependent on the mass density of water, water velocity and area of the organism (Box 4.2) Most running water ecosystems contain benthic algae and macrophytes (rooted vascular plants), and their distribution throughout the river continuum reﬂects a multitude of environmental factors including water chemistry, irradiance levels and substrate conditions (Vannote et al. 1980). In the context of channel management macrophytes are of key importance in creating a diverse physical environment and have, through this role as ‘ecological engineers’, an

Flow and drag forces

Reynold’s number is a dimensionless ratio between inertial forces of moving water creating turbulence and viscous forces where water moves unidirectionally in a laminar manner. Reynold’s number can be calculated either for a stream channel or an aquatic organism using: Re = UD v

(1)

where U is water velocity, D is channel depth or length of an organism and ν is the kinematic viscosity. Main ﬂow in streams is almost always turbulent and most organisms encounter turbulent ﬂow although at variable intensity. However, close to surfaces of inorganic substrates and dead organic matter, in plant patches and close to banks, ﬂow can be laminar as well. Small organisms as microorganisms (bacteria, fungi and algae) and small invertebrates will encounter laminar ﬂow. Flow generates important dynamic forces and can be calculated using: F = 0.5ρU 2 AC

(2)

where ρ is the mass density of water (temperature dependent), U is water velocity, A is the organisms area and C is dimensionless constant. The unit is Newton (kg m−2). In running water three forces are relevant: form drag, friction drag and lift. Form drag is the most important because of the turbulent nature of most streams and here A relates to frontal area of the organism. In friction drag A is the entire surface area while regarding lift, A is the projected area as this force acts perpendicular to the unidirectional ﬂow. The relationships between the force exerted on the organisms, water velocity and the relevant area are not straightforward as could be assumed from equation (1) and will show differences depending on morphology of the organism and water velocity which is reﬂected by the coefﬁcient C. As an example form drag in plant patches will increase at higher velocities with an exponent between 1 and 1.75 (Sand-Jensen 2006).

Ecological Consequences of River Channel Management enormous potential in the rehabilitation of physically degraded river systems. By contrast, although algal communities clearly respond to channel management, as exempliﬁed by dense mats of ﬁlamentous green algae often covering the stream bed after dredging, they do generally only interact with the physical environment on a smaller, near-bed scale. Many attached algae are so small that they live within the boundary layers dominated by laminar ﬂow (<0.1 mm) and depth of water, gradient and bed roughness in the channel all play a role in determining algal microhabitats (Reynolds 1996). Algae are found on stones (epiliths), macrophytes (epihytes), on ﬁne grained substrates (episamms/epipels) and suspended in the water column. With regard to the presence of suspended algae in running waters, it is important that dead zones with little or no ﬂow exist or that inocula from upstream lentic waters occur as they will otherwise be ﬂushed out of the system (Reynolds 1996). Algae in running waters are of key importance for the food web as they provide a high quality source of energy for higher trophic levels as well as a source of dissolved organic matter (Allan 1995). Furthermore, dense assemblages of periphyton attenuate near bed ﬂows and this is essential for the provision of microhabitats for other organisms as well as for the movement of solutes (Dodds and Biggs 2002). Macrophytes in running water form patches as an adaptation to ﬂow as hydraulic resistance does not increase proportionally with the size of the patch. Flume experiments have shown a 4–7 times increase in resistance for a 10-fold increase in biomass (Sand-Jensen 2006; Fig. 4.6). Furthermore, plant species have variable resistance to ﬂow when they grow at low densities, in contrast to when they are forming patches. In general, single shoots of stiff, ramifying species are more resistant to ﬂow than ﬂexible, streamlined species. When growing in patches these individual differences relating to plant morphology are less pronounced because of mutual protection and shelter of shoots. Stiff, ramifying shoots are subjected to considerable drag forces when solitary but their resistance to ﬂow creates more

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benign conditions for downstream shoots than more ﬂexible plant species would. As a consequence of this, resistance of species with different morphologies tends to converge towards a common value at high velocities and plant biomasses. When drag forces reach a critical level, however, macrophytes will be uprooted and washed away. The shear stress necessary for uproot will be species speciﬁc and dependent on growth form, but submerged macrophytes are the most ecologically resistant whereas amphibic or terrestrial species are more liable to uprooting. Under natural conditions most streams will lose their macrophytes in autumn when high discharge events occur and overwintering populations are mainly found in groundwaterfed streams characterized by relatively constant ﬂows over the year. Macrophytes recover from surviving roots and rhizomes buried in the stream bed or grow from propagules (seeds and drifting shoots). It is critical for recolonization that populations upstream are available and that seeds and roots are able to settle on the stream bed. When settled, however, most macrophytes can, within a few months, expand from a single shoot to a well-established patch of several square metres. Patch dynamics are dependent on substrate composition, and hence stability, with more well deﬁned patches occurring on ﬁne grained substrates, reﬂecting that the likelihood of a shoot settling outside a patch is low on easily disturbed, ﬁne grained sediments. Macrophytes create a physically very heterogeneous environment through the way they form patches. Water velocity is accelerated above and around patches whereas ﬂow within and behind the macrophytes can be almost zero (Sand-Jensen and Mebus 1998; Sand-Jensen and Pedersen 1999). Fine sediments and organic matter will consequently be deposited in the low ﬂowing areas and erosion of ﬁner particles between patches will result in localized coarse substrate habitats (Sand-Jensen 1998). Overall, however, macrophytes will decrease ﬂow and at the reach scale they will increase water height for a given

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Fig. 4.6 The drag on plants in ﬂowing water increases with increasing current velocity and plant biomass. α, slope of the log–log relationship; B, biomass; U, mean velocity. Black circles, 4 shoots; white circles, 64 shoots. (From Sand-Jensen 2006.)

discharge thereby increasing the likelihood of over bank ﬂooding (Fig. 4.7). Macroinvertebrates exhibit numerous adaptations to a life in ﬂow and this is reﬂected in their morphology, behaviour, feeding, physiology and life cycle strategies. Body shapes such as dorsoventrally ﬂattened or streamlined are common among stream macroinvertebrates which will reduce the forces of ﬂow, and attributes such as claws or suckers help in maintaining position in high energy environments (Vogel 1994). These types of morphologies are, however, not exclusive to running waters but are also found in organisms inhabiting stagnant waters.

Fig. 4.7 The water stage in streams increases with discharge but also with in-channel macrophyte biomass. (From Sand-Jensen 2006.)

Ecological Consequences of River Channel Management Furthermore, despite many macroinvertbrates being capable of withstanding high water velocities (>1 m s−1) there will be a threshold above which they cannot resist the imposed drag and lift forces. More importantly, however, is how macroinvertebrate behaviour interacts with the physical environment in streams to increase both their resistance and resilience to ﬂow disturbances. 4.3.1 Flow refugia: physical–biological coupling Natural streams include a number of patches which remain stable and provide areas with reduced ﬂow even during spates. A pre-requisite for macroinvertebrates to remain in position is that the substrate is stable and not mobilized by ﬂow. In addition to being resistant to ﬂow, large structural elements create a mosaic of patches on the stream bed with different levels of shear stress even when discharge is increasing: certain patches remain in areas of low shear stress, or dead zones, when external forces increase, while other patches show a proportional response (Fig. 4.8). Macroinvertebrates have been found both in

Fig. 4.8 Certain patches (dead zones) on the stream bed sustain low shear stress even when external forces increase while other patches will respond with increased shear stress. (Modiﬁed from Lancaster and Hildrew 1993a.)

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observational and experimental studies to use these dead zones as ﬂow refugia, with numbers increasing in permanent low shear patches when discharge increases (Lancaster and Hildrew 1993a,b; Winterbottom et al. 1997a). The underlying mechanism with regard to how macroinvertebrates move into ﬂow refugia remains unresolved as this process could be either passive or active or a combination of both (Fig. 4.9). Macroinvertebrates are effectively dispersed by the unidirectional ﬂow and there is an overall relationship with discharge which suggests a strong passive element (Winterbottom et al. 1997b) (Fig. 4.10). Numerous other studies have shown, however, that macroinvertebrates have a strong behavioural component so the ﬁndings are likely to reﬂect a combination of active and passive processes (Allan 1995). Furthermore, aggregation of macroinvertebrates in patches, with permanently low shear stress, might also be a response to increased levels of food resources such as dead organic matter which is deposited in the patches (Lancaster and Hildrew 1993b) and the benign environmental conditions are likely to increase species interactions such as predation and competition which inﬂuence size

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Fig. 4.9 Colonization rate (total macroinvertebrate assemblage and the stoneﬂy Leuctra nigra) of artiﬁcially provided ﬂow refugia increased with increasing discharge. (Modiﬁed from Winterbottom et al. 1997a.)

Fig. 4.10 Mobility of the total macroinvertebrate assemblage and the stoneﬂy Leuctra nigra show a positive linear relationship with discharge indicating a strong positive element to macroinvertebrate dispersal. (Modiﬁed from Winterbottom et al. 1997a.)

distribution and species composition. Another very important provider of ﬂow refugia for macroinvertebrates in running waters are interstitial spaces in coarse substrates such as gravel beds (Giller and Malmquist 1998). The scientiﬁc literature reports many linkages between individual parameters (substratum, current velocity, etc.) describing the in-stream physical environment and different attributes of the macroinvertebrate community (Allan 1995). On a ﬁne scale, using high resolution measurements without other confounding factors, the importance of physical–biological coupling in streams has been shown although many aspects are still not fully understood (Hart and Finelli 1999). One reason is that numerous physical

factors interact across different temporal and spatial scales, resulting in very different biotopes (rifﬂes, pools, etc.) as well as inducing variation within these biotopes. One example of the latter (Fig. 4.11) is the response of two macroinvertebrate metrics (total abundance and EPT* taxa abundance) to mean particle size which varied considerably between two adjacent rifﬂe biotypes in the same stream. The majority of studies have shown that diversity and abundance of macroinvertebrates increase with particle size as this increases habitat complexity and volume of * EPT = species belonging to the three insect orders Ephemeroptera, Plecoptera and Trichoptera, which is a metric sensitive to perturbations.

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Fig. 4.11 Two adjacent rifﬂes (1 and 2) in a lowland river showed very different relationships between particle size and abundance of all macroinvertebrates (A,B) and EPT taxa (C,D). (From Pedersen and Friberg 2007.)

interstitial spaces (Giller and Malmquist 1998). In one rifﬂe (B and D) there was the expected positive relationship between particle size and macroinvertebrate metrics whereas there was no relationship in the other rifﬂe (A and C), the reason being that it was consolidated by ﬁne sediments thereby reducing habitat availability. This study illustrates that even within the same biotype that appears completely similar when visually assessed, there can be considerable differences in the physical–biological coupling. The importance of macrophytes as a direct food resource for macroinvertebrates in temperate streams is limited although signiﬁcant effects of grazing have been reported in certain circumstances (Jacobsen and Sand-Jensen 1994). Indirectly, however, they play a key role for macroinvertebrate communities. Up to 400,000 indi-

viduals of macroinvertebrates per m2 of stream bed have been found when including those living on macrophytes which constituted 95% of the total density (Iversen et al. 1985). Architectural complexity of macrophytes has been found to increase taxa richness (Taniguchi et al. 2003) and in some cases densities (Rooke 1986; Jeffries 1993). In addition, Rooke (1986) found that different macrophyte taxa did not support the same proportion of macroinvertebrate feeding guilds, indicating that they provided different food resources. Moreover, besides plant morphology, growth form and local site characteristics have been shown to inﬂuence interactions between macroinvertebrates and macrophytes. Harrison et al. (2005) found differences in macroinvertebrate characteristics in macrophyte patches positioned differently in the stream

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channel, and Gregg and Rose (1985) observed that the response of taxa richness to macrophyte complexity varied depending on physical site characteristics with increased macroinvertebrate taxa richness only found in high complexity Ranunculus over a certain current velocity threshold. Fish are a prominent component of running water ecosystems and of very high commercial importance in many places of the world primarily for recreational ﬁshing. They inhabit the entire river continuum and exhibit a succession of species from source to estuary reﬂecting a shift in environmental conditions. Fish living in high energy environments are adapted to drag forces with stream-lined fusiform body shapes and common examples are the salmonids (Vogel 1994). Common salmonids in the upper parts of river systems throughout Europe are stationary populations of brown trout (Salmo trutta fario) which in addition to body shape also exhibit behavioural adaptations to a life in running water. They utilize areas of reduced water velocity behind stones and woody debris from which they feed on drifting macroinvertebrates. Migratory individuals of the same species, S. trutta, as well as the Atlantic salmon (Salmo salar) have an anadromous life cycle and use rivers as vectors between the feeding areas at sea and spawning grounds in the upper part of river systems. Throughout the world, ﬁsh species show similar life cycle adaptations with anadromous or catadromous migratory movements using the longitudinal connectivity of rivers. Recent research has demonstrated that ﬁsh species which stay in freshwater during their entire life cycle show a much more pronounced migratory behaviour in connection with, for example, spawning than previously believed (Hladik and Kubecka 2003; McMahon and Matter 2006). Depending on habitat of a given species, or where it occurs in the river continuum, alternative body shapes occur as exempliﬁed by, for example, bullheads that live on the bed of stony upland streams in which roughness caused by the coarse substrate has reduced ﬂow or further downstream where a number of more lentic

species with less streamlined body shapes occur as velocities are reduced. Upland streams have steep channel slopes that discourage the accumulation of gravel-sized material and the development of a heterogeneous morphology which potentially can limit habitat diversity. It has been demonstrated theoretically that the presence of wood in upland streams increases channel roughness sufﬁciently to permit signiﬁcant accumulations of gravel-sized material in channels that would otherwise be too energetic to maintain a substrate within this size range, thereby providing an important habitat (especially spawning) for salmonids (Bufﬁngton et al. 2004). Further downstream wood creates pools and holding areas for both migratory and more stationary ﬁsh species (Bilby and Bisson 1998).

4.4

Impacts of River Channel Management

By comparing the changes that management inﬂicts on stream channel properties with the adaptations and biotope demands of running water organisms, it is evident that there are strong negative impacts on the ecology. Most types of channel management will negatively impact both ecosystem resistance and resilience, primarily through the loss of spatial heterogeneity and by impairing recovery routes of the biota. 4.4.1 Direct impacts of channel management Macrophyte diversity and composition can be impacted by channel regulation (Table 4.2). A summary of both unregulated and regulated streams which were subjected to contemporary management in the form of weed cutting twice a year showed that macrophyte coverage was independent of channel management history whereas species richness and diversity were signiﬁcantly higher in unregulated than in regulated stream reaches. In addition, spatial distribution was signiﬁcantly more heterogeneous in the unregulated

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Table 4.2 Vegetation data from unregulated and regulated stream reaches. Values marked with an asterisk (*) are signiﬁcantly different (from Baattrup-Pedersen & Riis 1999) Unregulated (n = 7)

Species coverage (%) Species richness Shannon diversity index Plant heterogeneity

Regulated (n = 7)

Mean

Range

Mean

Range

86 22* 0.47* 0.51*

44–100 13–32 0.25–00.77 0.37–0.66

75 15* 0.27* 0.38*

25–100 7–21 0.08–0.68 0.24–0.53

stream reaches. These results indicate that there is management legacy and changes to channel form will have persistent impacts on macrophyte communities. Another common management practice is removal of macrophytes either through weed cutting or dredging. Long-term use of weed cutting changes the composition and structural complexity of macrophyte communities making them less species rich and spatially more homogenous. Substantial changes in composition patterns can occur with increasing dominance of fast growing species with high dispersal capacity (Baattrup-Pedersen et al. 2003). Different macroinvertebrates exhibit a range of adaptations to life in running waters and overall diversity, as well as other community attributes, and to a large extent rely on the heterogeneity of the environment in which they dwell. Consequently, channel management has the potential of decreasing diversity as well as abundance, biomass and energy transfer of macroinvertebrate communities by creating less varied biotopes. Changes in the composition of the macroinvertebrate community, however, can be subtle and difﬁcult to assess using standard assessment methods. Armitage and Pardo (1995) investigated the impact of river regulation on macroinvertebrates and found that conventional assessment techniques were unable to demonstrate regulation effects. By contrast, the proportions of meso-habitats and their faunal community were found to be a sensitive indicator of regulation. In a study of 75 independent river reaches, Feld and Hering (2007) found that land use and hydromorphology were the two main variables in explaining macroinvertebrate communites at dif-

ferent spatial scales. At the local scale, the presence of artiﬁcial structures used for bank enforcement was related to degraded macroinvertebrate communities. Recovery of a river system after the termination of channel management (weed cutting and bankside vegetation control) in a long-term study (20 years) was found to be minor with only a slight increase in macroinvertebrate family richness which could be attributed to a change in marginal vegetation (Wright et al. 2003). The study did show, however, an increase in leaf eating macroinvertebrates at the expense of algae grazers, reﬂecting the larger input and retention of organic matter. Studies have shown that ﬁsh density and biomass can be negatively impacted by channelization (Hermansen and Krog 1984; Oscoz et al. 2005), whereas impacts on species richness appear less strong, at least on the reach scale (Steinberger and Wohl 2003; Oscoz et al. 2005). A study by Aarts et al. (2004), however, found that despite the marked improvement in chemical water quality in recent years in large European rivers, the ﬁsh fauna is still impoverished when compared with historical records. The main reason is loss of habitats with rheophilic species highly adapted to riverine conditions declining far more than generalist species. Fish habitats in main and secondary channels have suffered most from channel regulation as they have been severely degraded, fragmented or completely destroyed. The immediate effect of removing macrophytes is the loss of in-stream biotopes and habitats for macroinvertebrates and ﬁsh, which is the combined result of removing plants and lowering

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of the water level (Kaenel et al. 1998). The longterm impact of weed cutting is a reduced diversity and structural complexity of macrophyte communities which can negatively affect macroinvertebrates and ﬁsh. This probably relates to a lower spatial and temporal physical heterogeneity, i.e. less varied substrate conditions and a more narrow range of ﬂow conditions with decreasing structural diversity of the macrophyte community (Garner and Bass 1996). An investigation of two reaches in a Danish stream indicated the importance of the interaction between macrophytes, physical features and macroinvertebrates in a management context (Friberg 2006). One stream reach had not been exposed to channel management, mainly weed cutting, for over 20 years whereas weed cutting was performed twice per year in the other reach. Channel dimension and shape were only slightly modiﬁed in both reaches. Species composition and structural variation of the macroinvertebrate community reﬂected management practices with un-branched bur-reed (Sparganium emersum) completely dominating (>50% coverage) the frequently weed cut reach. By contrast, macrophyte species were much more evenly distributed in

the reach that had not been managed with broadleaved pondweed (Potamogeton natans) covering approximately 25% of the stream bed and at a smaller scale macrophyte patches were more complex containing several species. These differences in the macrophyte community were also reﬂected in the macroinvertebrates: the species richness was signiﬁcantly higher in samples taken between plants in the reach without management compared to the reach with frequent weed cutting (Fig. 4.12). A similar tendency was found with regard to macroinvertebrates sampled on substrates directly under macrophytes whereas no clear pattern emerged when comparing macroinvertebrates found on plants between the two reaches. One exception, however, was the mixed macrophyte patches (three species) on the unmanaged reach which had a higher diversity of macroinvertebrates compared to single macrophyte species. Results from this investigation clearly suggest that management of macrophytes can have negative impacts on other parts of the biota either indirectly or directly. As weed cutting in the study reported here had not occurred in several months prior to sampling it is most likely that differences found reﬂected macrophyte com-

Fig. 4.12 The greatest macroinvertebrate diversity was found on the reach where channel management (weed cutting) had not been undertaken for more than 22 years between plants and on mixed stands. (From Friberg 2006.)

Ecological Consequences of River Channel Management munity structure and this seems to be mediated primarily through changes to the physical environment. 4.4.2 Management impacts mediated through changes of riparian zones Narrowing of stream channels, when riparian forest is removed, is caused by bank encroachment of herbaceous plants that otherwise would have been shaded out by the trees (Allan 1995; Hession et al. 2003a; Sweeny et al. 2004). Forested stream reaches exhibit slower channel migration and lower ﬂoodplain accretion rates of sediments which render them more stable than channels with modiﬁed riparian vegetation (Hession et al. 2003b). Changes to the riparian vegetation will have an impact on in-stream channel characteristics which will affect ecological process rates as demonstrated by Sweeny et al. (2004). Forested streams channels were wider with higher bed roughness than adjacent deforestated channels (Fig. 4.13a). Uptake rates of ammonium (NH4) were substantially higher in reaches with an intact riparian zone as were net daily metabolic rates measured by unit stream length (Fig.

Fig. 4.13 The ratio of stream width (a), ammonium uptake (b), net daily metabolism (c) and number of macroinvertebrates (d) obtained from adjacent paired reaches with and without riparian deforestation of 16 temperate streams in eastern North America. Deforested stream reaches were short (<1.5 km) and not impacted by other pressures such as watering of cattle, arable farm practices or urbanization. Stream size ranged from ﬁrst to ﬁfth order (0.1–123 km2 in catchment area). (Modiﬁed from Sweeny et al. 2004.)

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4.13b,c). The high capacity of nitrogen (N) removal in the form of NH4 was likely to reﬂect the increase in surface area and increased interaction with the bed promoting both abiotic sorption and microbial uptake. Consequently, the transport of N to downstream recipients will be retarded and the likelihood of a permanent loss through denitriﬁcation, and through trophic linkages with the terrestrial ecosystem, will increase. The forested stream reaches were more heterotrophic compared with the deforested reaches expressed by more negative net daily metabolic rates (up to 500% more negative) which indicate that reaches with an undisturbed riparian zone process signiﬁcantly greater amounts of organic matter per unit stream length than reaches without a natural riparian zone. In addition, Sweeny et al. (2004) found more macroinvertebrates in the forested reaches (Fig. 4.13d) and that stream narrowing, when removing the riparian vegetation, nulliﬁed any potential advantages regarding ﬁsh, quality of dissolved organic matter and pesticide degradation. Riparian zones in managed rivers will generally get dryer and the vegetation will change either directly by anthropogenic intervention

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Fig. 4.14 Number of riparian plant species (i.e. species preferring moist conditions) and their frequency were signiﬁcantly higher along natural stream channels (Nat) compared with channelized streams (Cha; within 10 m from the stream). (Modiﬁed from Baattrup-Pedersen et al. 2005.)

or by the altered environmental conditions. Lowering of the water table will further increase decomposition and mineralization of organic soils resulting in subsidence and release of excessive N, phosphorous (P) and in some cases Fe (Fig. 4.5). A study in which plant species were divided into functional groups using the Ellenberger indicator value of moisture showed that diversity and relative frequency of riparian plant species (i.e. species preferring moist conditions) were signiﬁcantly higher within 5 m of an unregulated stream compared to a regulated stream (Fig. 4.14). Furthermore, analysis of species data revealed an intricate relationship between nutrient tolerance and distance to the stream along the natural streams, indicating how a natural ﬂood regime, through the additions of nutrients, can create complex habitats. This study suggests that channel management, by truncating the interaction between the stream and its surroundings, can have indirect negative impacts on plant biodiversity in the riparian zone. Leyer (2005) found similarly that grassland species composition and spatial distribution in ﬂoodplains along the regulated River Elbe reﬂected soil moisture and water level ﬂuctuations. The riparian zone is not only important for plants, however, as 70–80% of the macroinvertebrates in streams are insects with an adult terrestrial life stage. Negative impacts

on adult aquatic insects could be mediated through changes to the vegetation as studies have shown that the dispersal of adults is affected by the type of riparian vegetation (Petersen et al. 1999). Furthermore, there is a number of terrestrial arthropods that are linked to riparian zones and they are likely to be inﬂuenced by riparian management as well (Burdon and Harding 2008).

4.5 Channel Management and Legislation Historically, streams and rivers were considered as merely conveyors of water and legislation was put in place to secure the effectiveness of the process through a number of drainage acts. With increasing pollution of freshwater ecosystems during the last century, most countries developed water quality legislation with environmental quality standards for nutrient concentrations and BOD (Biological Oxygen Demand) levels which did not regard either habitat quality or water quantity. This created a dichotomy in water legislation where one part of the ecosystem was protected, namely water quality, whereas issues related to discharge and habitats were regulated solely to provide speciﬁc ecosystem services. More recently, however, legislation has evolved

Ecological Consequences of River Channel Management to become more integrated as exempliﬁed in Denmark where a new watercourse act was adopted in 1982 (revised in 1993) which has had signiﬁcant implications for mitigation and consequently overall stream quality. Its primary aim was to ensure that watercourses could be used for the discharge of water (surface, waste and drain water) but implementation of measures to secure this aim should be performed with due consideration for the environmental demands laid down in other legislation. Speciﬁcally the act introduced restoration of the physical environment as a way of achieving quality objectives relating to macroinvertebrates and ﬁsh. This has resulted to date in more than 1000 restoration projects ranging from laying out spawning gravel to comprehensive projects re-meandering streams and improving connectivity with the ﬂoodplain (Hansen and Baattrup-Pedersen 2006). Similarly, it was reasoned in a court case in 1994 that the deﬁnition of the Clean Water Act in the USA encompasses the physical and biological integrity of water, as well as chemical integrity (Masonis and Bodi 2001) and the Water Resource Act of 1991 covering England and Wales also exposed a more uniﬁed approach to water regulation (Howard 1992). More recently, the Council of the European Communities adopted in 2000 the European Union (EU) Water Framework Directive (WFD) which requires Member States of the EU to assess, monitor and, where necessary, improve the ecological quality of surface waters including rivers. The WFD is the most substantial piece of EU water legislation to date and provides an integrated, catchment based approach to waterbody management. The WFD sets out normative deﬁnitions describing biological and physico-chemical standards expected of ‘high’, ‘good’ and ‘moderate’ quality of streams and rivers for four biological quality elements (phytoplankton, macrophytes and benthic algae, macroinvertebrates and ﬁsh) and a variety of hydromorphological and physiochemical quality elements. In fact, the term ‘hydromorphology’ is introduced by the WFD and signiﬁes how important physical features are considered in determining the ecological status of freshwa-

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ters. In relation to channel management, the focus on hydromorphology as a quality element in its own right is very important. The hydromorphological quality of a river in ‘high’ status class is deﬁned as follows: ‘Channel patterns, width and depth variations, ﬂow velocities, substrate conditions and both structure and condition of the riparian zones correspond to totally or nearly totally to undisturbed conditions’. With regard to the other status classes, hydromorphology is considered a supporting element and there are no normative deﬁnitions. The WFD states that hydromorphological assessment should be undertaken at 6-year intervals as part of operational monitoring programmes in each Member State. Consequently, the WFD integrates hydromorphology in the overall quality assessment of ecological status in a way that is likely to reduce the negative consequences of channel management. The WFD, however, also acknowledges that some streams and rivers are so impacted by hydromorphological alteration while providing important ecosystem services that costs to society of restoring them back to a more natural condition would be disproportionately high and in which case they would be classiﬁed as ‘Heavily Modiﬁed Waterbodies’. The WFD and other relevant water acts that integrate both physical and chemical variables are a vital step forward in reducing the negative impact of channel management.

4.6 Monitoring the Impacts of Habitat Degradation The impacts of habitat degradation can be monitored either directly by using hydromorphological assessment systems, by combining habitat characteristics with speciﬁc species preferences or by assessing the ecological impact. In assessing the overall ecological status of rivers it is important to characterize them according to their morphology. Ideally, this morphological characterization should be targeted to be a causal link between the anthropogenic pressure on river morphology and the

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resulting impacts on the biota. In Europe a range of different methods are currently used: • The Austrian Habitat Survey (Muhar et al. 1996) in Austria. • The Danish Stream Habitat Index (Pedersen and Baattrup-Pedersen 2003) in Denmark. • The SEQ Physique (Agences de l’Eau and Ministere de l’Environment 1998) in France. • The Ecomorphological Survey for large rivers (Fleischhacker and Kern 2002) in Germany. • The River Habitat Survey (RHS) (Raven et al. 1998) in the UK. The assessment systems use a variable number of parameters relating to discharge classes, channel features, structural bank properties and ﬂoodplain land cover and a scoring system to evaluate the hydromorphological status of rivers. High status class is in most systems identiﬁed on the basis of a top percentage of sites according to their score and not by using reference conditions. Habitat models describe the relationship between water discharge and the biota in a river reach. Most habitat models use preference indices, which determine the suitability of a given physical element (velocity, water depth, substrate) for certain species and their individual development/life history stages. These models then combine the results from hydraulic models with the preference indices to produce values of river area weighted by habitat quality (weighted usable area, WUA) as a function of ﬂow for a given species and life stage. Currently a number of different software systems exist which are capable of calculating WUA: • PHABSIM (Milhous et al. 1989) and RHABSIM used in the USA; • RHYHABSIM (Jowett 1989) used in New Zealand; • EVHA (Pouilly et al. 1995) used in France; • CASMIR used in Germany (Jorde 1997); • RSS (Killingtveit and Harby 1994) used in Norway; • HABITAT (Duel et al. 2003) used in the Netherlands. Habitat suitability indices (HSIs) are a similar approach but without the modelling component

which allows the relation of habitat preferences for a given species and life stage to hydraulic parameters (again typically velocity, depth and substrate). Most HSIs have been deﬁned as univariate response functions but more complex multivariate relationships also exist (Parasiewicz and Dunbar 2001). The relationships established by HSIs are useful for decision-making on the protection of relevant habitats or their restoration. Evidence of the linkage of reach scale physical parameters and biotic samples on sites which have only been disturbed by physical alterations is scarce. There has been moderate success in linking biota with the currently used habitat assessment systems, which is partly due to the mixed nature of pressures acting on a river reach and the habitat surveys using parameters of at least two spatial scales. The latter reﬂects that methods for sampling macroinvertebrates often are on a different scale to that of the hydromorphological assessment and that the sampling strategy was developed to target mainly organic pollution. The impact of channel management has been assessed using a standard biotic index (DSFI) in six ﬁrst and second order streams. Each stream contained an upper reach that was modiﬁed by channelization, wood removal and dredging in earlier years and a lower reach (<1 km) that had remained relatively undisturbed. There were no dispersal barriers between the two reaches and no differences in water chemistry. The reaches showed clear differences in number of available habitats and this was mirrored in the ecological assessment class which in the DSFI-system ranged between 1 (severe pollution) to 7 (unpolluted). The ecological assessment class was higher in the unmodiﬁed than in modiﬁed reaches in all streams (Table 4.3) which could only be attributed to differences in habitats. Speciﬁcally focussing on the diversity of taxa that are regarded as intolerant towards pollution, the relationship is even more pronounced. These ﬁndings reﬂect that existing assessment systems originally developed to target organic pollution show sensitivity toward habitat degradation but also that

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Table 4.3 Outcome of an ecological assessment system based on macroinvertebrates (DSFI, Skriver et al. 2000) used in 6 different streams with reaches that were either unmodiﬁed or modiﬁed by channel management. Positive diversity groups consist of macroinvertebrate taxa that are pollution intolerant and negative diversity groups of taxa that are tolerant towards pollution. Number of negative groups is subtracted from positive groups (last column) to calculate DSFI Stream no. 1-unmodiﬁed 1-modiﬁed 2-unmodiﬁed 2-modiﬁed 3-unmodiﬁed 3-modiﬁed 4-unmodiﬁed 4-modiﬁed 5-unmodiﬁed 5-modiﬁed 6-unmodiﬁed 6-modiﬁed

Ecological assessment

Positive diversity groups

Negative diversity groups

Sum of diversity groups

5 4 7 4 6 5 5 4 5 5 7 4

12 6 10 4 11 9 9 2 10 7 13 3

0 4 0 3 0 2 0 3 1 3 0 2

12 2 10 1 11 7 9 –1 9 4 13 1

this sensitivity is likely to be improved by the development of speciﬁc tools. At present, a few assessment systems that target the impacts of low ﬂow (Extence et al. 1999) and degraded hydromorphology (Barbour et al. 1996, Lorenz et al. 2004) do exist but they are not generally applicable and could be further developed. One key issue when assessing the inﬂuence of hydromorphology is that interpretation of results is often confounded by multiple stressors inﬂuencing freshwater communities (Matthaei et al. 2006). Typically, streams that have undergone a high degree of habitat degradation or alterations of ﬂow regimes will be situated in areas with multiple anthropogenic pressures. In addition, stressors can interact in a synergistic manner and increased concentrations of easily degradable organic matter will have a more detrimental impact in a habitat degraded stream because the number of reactive surfaces and the re-aeration capacity are reduced (Andersen 1994). Development of more indicator systems sensitive to hydromorphological degradation will be a key issue in the future and the use of species traits instead of the more traditional approach of taxonomic lists is a promising way forward.

There has been increased focus recently on using species traits of macroinvertebrates in biomonitoring in general (Dolédec et al. 1999; Statzner et al. 2005) and speciﬁcally in relation to habitat disturbance (Dolédec et al. 2006) and at the mesohabitat scale (Lamouroux et al. 2004). Species trait combinations are assumed to be elicited, modiﬁed or selected depending on the environmental characteristics of systems or habitats. Townsend and Hildrew (1994) predicted that stream species assemblages would exhibit traits that are well suited to the spatial and temporal heterogeneity of their habitat and considered that similar environmental drivers should induce similar trait composition in stream benthic communities. As a result, species traits have the potential to indicate the mechanisms, including human-induced disturbances, which structure communities. Traits reﬂect the life history of taxa (e.g. body size, reproduction cycles per year, life cycle duration); the resistance and resilience potential (e.g. substrate relation, dispersal mechanisms, body shapes) and general behavioural and physiological features of organisms (e.g. feeding habits, food, respiration, reproduction technique).

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nikolai friberg 4.7 Restoration and Rehabilitation

During the last decade there has been an increase in river restoration projects globally to counteract the deterioration caused by river channel management (Ormerod 2004: Palmer et al. 2005). Restoration projects, however, should also be considered as river management as they often involve considerable ‘hard’ engineering and changes to the existing stream channels. Projects have been undertaken for a number of reasons and on very different scales, ranging from instalment of single structures, such as reﬂectors or removal of small weirs, to large-scale restoration of entire ﬂoodplains. Restoration projects can be divided into three types depending on the aim and scale (Hansen and Baattrup-Pedersen 2006): where restoration has been carried out locally on single reaches (type 1; Fig. 4.15), where restoration re-establishes the longitudinal connectivity (type 2; Fig. 4.15); and where the restoration includes both river and riparian zones/ﬂoodplains (type 3; Fig. 4.15 ) The majority of restoration projects carried out to date have used some sort of ‘hard’ engineering approach. Type 1 projects are typically the re-introduction of spawning gravel for ﬁsh, adding structures such as gabions or deﬂectors and/or re-meandering the channel by excavating a new course, often stabilized by rip-rap structures, etc. (Friberg et al. 1994). Type 2 projects mainly include direct removal of obstacles to migration such as weirs and small dams or installing ﬁsh ladders/bypass streams to enable passage of barriers that cannot be removed. Type 3 projects are, due to their scale, the most uncommon restoration type and more process orientated than the other two types. An example of

the latter is the restoration of the River Skjern in Denmark which is one the most comprehensive restoration projects conducted to date (Pedersen et al. 2007). In this project the lower part of the river (average discharge of 35 m3 s−1) was re-meandered by increasing the river length from 19 to 26 km and the groundwater table was raised in an area of 22 km2 in the surrounding river valley where land use was changed from arable farming to wetlands. Frequency of ﬂooding was increased by narrowing the over-wide stream channel and creating a natural proﬁle which also increased the range of biotopes present (Fig. 4.16). Only a minority of restoration projects, however, allow natural processes to operate and are undertaken without sufﬁcient knowledge about the reference conditions. Walter and Merritts (2008) dated and mapped deposits along mid-Atlantic streams in North America and consulted historical records to ﬁnd that their reference conditions before European settlement were small branching channels with extensive vegetated wetlands. In contrast to their reference conditions, these streams had been restored by re-meandering as it was assumed from contemporary sediment conditions that they were natural gravel bed streams in self-formed, ﬁne-grained ﬂoodplains. In reality the presence of ﬁne sediments could be attributed to tens of thousands of seventeenth to ninetenth century mill-dams behind which sediments had accumulated and buried the natural stream features. Compared with the large number of restoration projects undertaken, relatively few monitoring studies have assessed the ecological outcome of these often very expensive projects (Ormerod 2004; Palmer et al. 2005). The restoration of a 1.3-km straightened and channelized reach of

Fig. 4.15 Schematic outline of the three main types of river restoration. See text for explanation. (From Hansen and Baattrup-Pedersen 2006.)

Ecological Consequences of River Channel Management

Fig. 4.16 Example of the changes to the crosssectional proﬁles in River Skjern. The cross-sectional area has generally decreased by approximately 30%. The morphology of the proﬁles has changed from the constructed rectangular shape to a more natural physical appearance. (From Pedersen et al. 2007.)

the lowland River Gelsaa, Denmark to a 1.9-km meandering course in 1989 was one of the earliest restoration projects of this type and a positive impact on the macroinvertebrate community was found after 2 years (Friberg et al. 1994). Compared with samples from before the restoration and from an upstream reach that remained channelized, higher abundances of stone-dwelling species occurred including the mayﬂy Heptagenia sulphurea which was not previously found. This ecological response reﬂects the large number of stones and gravel introduced as part of the restoration and is similar to ﬁndings when gravel is introduced alone as a restoration measure (Merz and Chan 2005). By contrast, other studies have found no effect of introducing gravel on macroinvertebrate communities (Harrison et al. 2004) or ﬁsh (Pretty et al. 2003), and the conﬂicting results can probably be attributed to a number of reasons relating to comparability of the restored system with reference conditions to which species are adapted, dimensions/scale issues relating to project as well as dispersal barriers for organisms within the riverine landscape (Ormerod 2004).

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Another important issue is temporal resolution when evaluating restoration projects. In the case of the River Gelsaa, the macroinvertebrate community at the restored reach was similar in density and diversity to that of the upstream channelized reach 4 and 6 years after restoration (Friberg et al. 1998) but signiﬁcantly lower after 8 years (Friberg et al. 2000). Density of trout (Salmo trutta) was also signiﬁcantly lower in the restored reach after 8 years (Carl 2000) showing that timing of the ecological assessment of restoration success could provide entirely different answers. In the River Gelsaa, the survey showed markedly less macrophyte coverage on the restored reach after 8 years compared with the upstream channelized reach and this is the likely explanation of the observed macroinvertebrate and ﬁsh patterns (Friberg et al. 2000). Interestingly, both reaches were not managed by weed cutting since the restoration and the recovery of macrophytes in the channelized reach appear to be more beneﬁcial for the overall biota than the active restoration in which meanders and stones were added. There has recently been more focus on restoring ecosystem processes and especially on enhanced retention of organic matter. Loss of moss cover when restoring substrate heterogeneity has been shown to decrease leaf litter retention (Muotka and Laasonen 2002) whereas the introduction of boulders increased detritus retentiveness which increased species richness and abundance of leaf eating macroinvertebrates (Lepori et al. 2005). Both studies exemplify how restoration projects can beneﬁt from an ecosystem perspective which goes beyond measuring structural elements. Wood can be introduced as a restoration measure to rivers in previously forested areas or where wood is removed actively from the stream channel as part of a management strategy. This restoration technique has been undertaken in both North America (Roni et al. 2002) and Europe (Kail et al. 2007), and has been justiﬁed by studies showing that LWD has a strong inﬂuence on channel morphology (Montgomery et al. 2003) and salmonid habitat (Bryant et al. 2005). A review of experiences in using wood as a

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restoration measure in Germany and Austria found an increase in density of brown trout (Salmo trutta fario) in three out of ﬁve studies from lower mountain and alpine streams. There were additional positive effects for other ﬁsh species such as, for example, bullhead (Cottus gobio) and chub (Leuciscus cephalus) (Kail et al. 2007). The ultimate restoration using wood should include natural recruitment from replanted riparian woodlands providing a natural and sustainable supply of wood to the channel. This approach, however, is a long-term solution which might not be feasible everywhere as replanting of riparian zones is likely to evoke a number of issues relating to competing and conﬂicting interests in connection with commercial forestry, agriculture and sport ﬁshing.

4.8

Future Directions

Most managed rivers are ecosystems completely out of balance and they can only retain their services to society through a continuous effort to keep this imbalance in place. This has a very high ecological cost with a signiﬁcant loss of both structural attributes, such as species diversity, as well as ecosystem function. Impairment of hydromorphology in rivers as a consequence of channel management will furthermore signiﬁcantly reduce resistance to other pressures on the ecosystem, thus making it more vulnerable to overall degradation. At the same time, legislation such as the WFD has ambitious targets for restoring the ecological integrity of rivers which can only be achieved by releasing pressure on some of the currently heavily managed systems. There is a need for a more holistic management approach, better tools to detect and model impacts of channel management and an evidence-based restoration strategy if this dichotomy in societal demands on rivers is to be bridged. The majority of management decisions in relation to stream channels are still taken on a local scale largely ignoring the larger, process orientated controls of geomorphology and important ecological drivers such as species dispersal,

habitat needs, etc. The use of existing hydromorphological classiﬁcation and assessment schemes is not ideal as they are static in nature and introduce aspects of subjectivity, reducing the usefulness of this approach in predicting effects of river channel management. Furthermore, linkages between river channel hydromorphological classiﬁcations and the biota are weak making predictions about how channel management will inﬂuence ecological status very uncertain. Similarly, the impact of hydromorphological degradation on the biota is currently assessed using a range of techniques that are suboptimal as they lack appropriate sensitivity as well as the ability to quantify the importance of individual pressures. Management also needs to be on a catchment scale when single reaches are considered. Furthermore, management strategies have to acknowledge that catchments are nested within climatic and biogeographical regions that will inﬂuence the outcome of management interventions. To aid managers, new and reﬁned tools need to be developed which will provide understanding of the linkages between hydromorphology and biota across the relevant spatial and temporal scales. In addition, although channel management primarily inﬂuences the physical environment, future strategies need to incorporate the fact that multiple stressors are operating in most cases. A better understanding of how stressors interact is needed, therefore, together with the development of tools to quantify the importance of individual ecosystems stressors. Restoration is an obvious way forward to mitigate the impacts of channel management and strategies should be stratiﬁed according to the possibility of recovery. Less intensive restoration with a larger spatial cover is likely to have more beneﬁcial effects on ecosystem processes than localized and highly intensive projects. Reclaiming riparian areas, with the aim of reinstating natural processes where possible, could be a way forward rather than re-instating speciﬁc features (meanders, rifﬂes) in a conﬁned channel. Soft engineering using carbon neutral or positive

Ecological Consequences of River Channel Management restoration approaches should be promoted using both riparian and aquatic vegetation as well as the introduction of wood. Being a management tool, restoration should be prioritized at the catchment level and incorporate knowledge of reference conditions. References Aarts, B.G., van der Brink, F.W.B and Nienhuis, P.H. (2004) Habitat loss as the main cause of the slow recovery of ﬁsh faunas of regulated large rivers in Europe: the transversal ﬂoodplain gradient. River Research and Applications, 20, 3–23. Agences de l’Eau and Ministere de l’Environment (1998) SEQ Physique: a system for the evaluation of the physical quality of water courses. Paris. Allan, J.D. (1995) Stream Ecology. Chapman and Hall, London. Allan, J.D. (2004) Landscapes and riverscapes: the inﬂuence of land use on stream ecosystems. Annual Review of Ecology Evolution and Systematics, 35, 257–284. Andersen J.M. (1994) Water quality management in the River Gudenaa, a Danish lake–stream–estuary system. Hydrobiology, 275/276, 499–507. Armitage, P.D. and Pardo, I. (1995) Impact assessment of regulation at the reach level using macroinvertebrate information from mesohabitats. Regulated Rivers: Research and Management, 10, 147–158. Armitage P.D., Pardo I. and Brown A. (1995) Temporal constancy of faunal assemblages in ‘mesohabitats’. Application to management? Archiv für Hydrobiologie, 133, 367–387. Baattrup-Pedersen, A. and Riis, T. (1999) Macrophyte diversity and composition in relation to substratum characteristics in regulated and unregulated Danish streams. Freshwater Biology, 42, 375–385. Baattrup-Pedersen, A., Larsen, S.E. and Riis, T. (2003) Composition and richness of macrophyte communities in small Danish streams – inﬂuence of environmental factors and weed cutting. Hydrobiologia, 495, 171–179. Baattrup-Pedersen, A., Friberg, N., Larsen, S.E. and Riis, T. (2005) The inﬂuence of channelization on riparian plant assemblages. Freshwater Biology, 50, 1248–1261. Barbour, M.T., Diamond, J. and Yoder, C. (1996) Biological assessment strategies: applications and limitations. In: Grothe, D., Dickson, K. and Reed, D.

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Lepori, F., Palm, D. and Malmqvist, B. (2005) Effects of stream restoration on ecosystem functioning: detritus retentiveness and decomposition. Journal of Applied Ecology, 42, 228–238. Leyer, I. (2005) Predicting plant species’ responses to river regulation: the role of water level ﬂuctuations. Journal of Applied Ecology, 42, 239–250 Lorenz, A., Hering, D., Feld, C.K. and Rolauffs, P. (2004) A new method for assessing the impact of hydromorphological degradation on the macroinvertebrate fauna of ﬁve German stream types. Hydrobiologia, 516, 107–127. Masonis, R.J. and Bodi, F.L. (2001) River law. In: Naiman, R.J. and Bilby, R.E. (eds), River Ecology and Management: lessons from the Paciﬁc Coastal Ecoregion. Springer-Verlag, New York. Matthaei, C.D., Weller, F., Kelly, D.W. and Townsend, C.R. (2006) Impacts of ﬁne sediment addition to tussock, pasture, dairy and deer farming streams in New Zealand. Freshwater Biology, 51, 2154–2172. McMahon, T.E. and Matter, M.J. (2006) Linking habitat selection, emigration and population dynamics of freshwater ﬁshes: a synthesis of ideas and approaches. Ecology of Freshwater Fish, 15, 200–210. Merz, J.E. and Chan, L.K.O. (2005) Effects of gravel augmentation on macroinvertebrate assemblages in a regulated California river. River Research and Applications, 21, 61–74. Milhous, R.T., Updike, M.A. and Schneider, D.M. (1989) Physical Habitat Simulation System Reference Manual Version II. Instream ﬂow information paper 26. United States Fish and Wildlife Service, Fort Collins, Colorado. Montgomery, D.R., Collins, B.D., Bufﬁngton, J.M. and Abbe, T.B. (2003) Geomorphic effects of wood in rivers. In: Gregory, S., Boyer K. and Gurnell, A.M. (eds), The Ecology and Management of Wood in World Rivers. American Fisheries Society, Bethesda, MD, pp. 21–47. Muhar, S., Kainz, M., Kaufmann, M. and Schwarz, M. (1996) Ausweisungfusstypespeziﬁsch erhaltener Fliesswasserabschnitte in Österreich – Österreichische Bundesgewasser. BMLF, Wasserwirtschaftskataster, Vienna (in German). Mutz, M., Kalbus, E. and Meinecke, S. (2007) Effect of instream wood on vertical water ﬂux in low-energy sand bed ﬂume experiments. Water Resources Research, 43, W10424. Muotka, T. and Laasonen, P. (2002) Ecosystem recovery in restored headwater streams: the role of enhanced

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Reynolds, C.S. (1996) Algae. In: Petts, G. and Calow, P. (eds), River Biota – selected extracts from the Rivers Handbook. Blackwell Science Ltd., Oxford. Roni, P., Beechie, T.J., Bilby, R.E., Leonetti, F.E., Pollock, M.M. and Pess, G.R. (2002) A review of stream restoration techniques and a hierarchical strategy in prioritizing restoration in the Paciﬁc Northwest watersheds. North American Journal of Fisheries Management, 22, 1–20. Rooke, J.B. (1986) Macroinvertebrates associated with macrophytes and plastic imitations in the Ermaosa River, Ontario, Canada. Archiv für Hydrobiologie, 106, 307–325. Rosgen, D.L. (1994) A classiﬁcation of natural rivers. Catena, 22, 169–199. Sand-Jensen, K. (1998) Inﬂuence of submerged macrophytes on sediment composition and near bed ﬂow in lowland streams. Freshwater Biology, 39, 663–679. Sand-Jensen, K. (2006) Water ﬂow at all scales. In: SandJensen, K., Friberg, N. and Murphy, J. (eds), Running Waters. National Environmental Research Institute, Silkeborg, Denmark. Sand-Jensen, K. and Mebus, J.R. (1998) Fine-scale patterns of water velocity within macrophyte patches in Danish lowland streams. Oikos, 76, 169–180. Sand-Jensen, K. and Pedersen, O. (1999) Velocity gradients and turbulence around macrophyte patches in streams. Freshwater Biology, 42, 315–328. Skriver, J., Friberg, N. and Kirkegaard, J. (2000) Biological assessment of running waters in Denmark: introduction of the Danish Stream Fauna Index (DSFI). Verhandlungen internationale Vereinigung für theoretische und angewandte Limnologie, 27, 1822–1830. Southwood, T.R.E. (1977) Habitat, the templet for ecological strategies. Journal of Animal Ecology, 46, 337–365. Southwood, T.R.E. (1988) Tactics, strategies and templets. Oikos, 52, 3–18. Statzner, B., Bady, P., Dolédec, S. and Schöll, F. (2005) Invertebrate traits for the biomonitoring of large European rivers: an initial assessment of trait patterns in least impacted river reaches. Freshwater Biology, 50, 2136–2161. Steinberger, N. and Wohl, E. (2003) Impacts of water quality and ﬁsh habitat associated with maintaining natural channels for ﬂood control. Environmental Management, 31, 724–740. Svendsen, L.M. and Hansen, H.O. (1997) Skjern Aa. Sammenfatning af den eksisterende viden om de

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Image facing chapter title page: Courtesy of Bent Lauge Madsen.

Managing Agricultural Catchments to Sustain Production and Water Quality 5

A N D R E W S H A R P L E Y 1, M A R T Y M A T L O C K 2, LOUISE HEATHWAITE3 AND TOM SIMPSON4 1

Department of Crop, Soil and Environmental Sciences, Division of Agriculture, University of Arkansas, Fayetteville, Arkansas, USA 2 Biological and Agricultural Engineering Department, Division of Agriculture, University of Arkansas, Fayetteville, Arkansas, USA 3 Centre for Sustainable Water Management, Lancaster University, Lancaster, UK 4 Water Stewardship, Inc., Annapolis, Maryland, USA

5.1

Introduction

Improvements in water quality at local and regional levels have been documented with implementation of conservation measures or Best Management Practices (BMPs) that were targeted to speciﬁc and identiﬁable nutrient sources (Richards et al. 2002; Jokela et al. 2004). However, at the scale of major impact (e.g. Gulf of Mexico, Chesapeake Bay, Baltic Sea), management strategies become much more complex and the ability to assign a change in water quality to altered management becomes more difﬁcult. Further, when political borders are crossed, at either state, county, regional or national levels, then conﬂicting political strategies often compete and further complicate solutions, which can override local water quality issues and interests. This is partly due to the extremely diverse nature of the users of land and water resources ‘encouraging’ different and often conﬂicting constraints be placed on these resources.

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

Nitrogen (N) and phosphorus (P) have long been recognized as essential inputs for crop and animal production. Dramatic improvements in the economic efﬁciency of grain and animal production over the last 50 years have been coupled with increasing surface and ground water quality problems (Boesch et al. 2001; Rabalais et al. 2001). Point sources of nutrients, such as wastewater treatment and industrial plants, have been reduced because they are generally easy to identify and regulate. In a 1996 US Environmental Protection Agency (USEPA) report, non-pointsource nutrients were the primary source of concern in 40% of rivers, 50% of lakes and 60% of estuaries surveyed and listed as impaired (US Environmental Protection Agency 1996). As in the USA, the European Union Water Framework Directive (EU WFD) (Council of European Communities 2000) now requires widespread control of N and P inputs to rivers speciﬁcally to improve riverine ecology (Hilton et al. 2006). Related research in the UK, shows eutrophication risk in rivers is primarily linked to dissolved P concentrations during periods of ecological sensitivity (Reynolds and Davies 2001; Neal and Jarvie 2005; Jarvie et al. 2006), rather than gross annual mass P loads. In fact, many of the agricultural streams

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monitored for the Phosphorus and Sediment Yield Characterisation in Catchments (Mainstone et al. 2008) project had average soluble reactive P (SRP) concentrations above the guideline concentration of 0.10 mg P L−1, beyond which running waters are classiﬁed as being mesotropic or eutrophic (Environment Agency 1998) and which is being proposed as a boundary between moderate and good ecological status for lowland calcareous rivers (UKTAG 2006). Nutrient-related water quality issues include groundwater contamination with nitrate and accelerated eutrophication of surface water. Nitrogen, in the form of nitrate, is a groundwater quality concern because it has been linked to methemoglobinemia in infants and toxicity in adults. The USEPA has established a maximum contaminant level for nitrate-N in drinking water of 10 mg L−1 (45 mg nitrate L−1), while the European Union (EU) has established nitrate vulnerability zones to limit nitrate contamination of groundwater resources (Department for Environment, Food and Rural Affairs 2002a; US Environmental Protection Agency 2002).

Box 5.1

Research has identiﬁed agricultural land use practices which are of highest risk for nutrient loss at ﬁeld and farm scales, and forms in which N and P are exported (Withers and Lord 2002; Sharpley et al. 2006). Their impact on stream water quality at a catchment scale is less well deﬁned because ﬁeld and farm inputs are modiﬁed by catchment hydrology, in-stream processing, and contributions from a diverse range of other rural nutrient sources, including farmyard and septic tanks runoff and rural sewage treatment works (Gburek and Sharpley 1998; Jarvie et al. 2006). In response, new requirements to improve the ecological status of rivers under the EU WFD (EC, 2000) focus attention on identifying nutrient sources and their ecological impacts in rivers. This chapter examines how agriculture can be managed at catchment scale within the conﬁnes of production and environmental pressures, the potential measures and mechanisms by which this might be achieved, and how agricultural management can be rapidly and effectively adapted to work within a catchment strategy or framework.

Eutrophication

Eutrophication is the organic enrichment of a water body, accelerated by greater nutrient inputs. In most cases, eutrophication restricts water use for ﬁsheries, recreation, and industry because of the increased growth of undesirable algae and aquatic weeds and oxygen shortages caused by their death and decomposition. An increasing number of surface waters also experience periodic and massive harmful algal blooms (e.g. cyanobacteria and Pﬁesteria) that contribute to summer ﬁsh kills, unpalatable drinking water, formation of carcinogens during water chlorination and links to neurological impairment in humans (Kotak et al. 1993; Burkholder and Glasgow 1997). While eutrophication of most fresh water is accelerated by increased inputs of P (Sharpley 2000), coastal waters tend to be limited by N and seasonally by P (Howarth et al. 2000). The economic ramiﬁcations of nutrient enrichment of ground and surface water are profound, as are the effects on catchment management planning strategies. For example, harmful algal bloom outbreaks have severely affected both the ﬁshing and tourism industries in many areas of the world. In the USA, for example, the economic loss to eastern coastal states with impaired waters is estimated in excess of $1 billion since 1990 (Goodman 1999; Howarth et al. 2000). Further, it has become more cost-effective to decrease nutrient losses than to treat the negative impacts. For instance, New York City decided in the early 1990s that identifying, targeting and remediating the sources of N and P in its water supply catchments would be more cost-effective than building a new $8-billion water treatment facility. As a result, the state has invested $10 million to decrease nutrient sources in its water supply catchments

Managing Agricultural Catchments 5.2 Historical Background Post Second World War improvements in agriculture have dramatically increased grain and protein production in a very cost-effective manner. The specialization and fragmentation of crop and animal production systems, however, has brought new pressures to bear on agricultural management within catchments. Catchments generally had a sustainable nutrient balance, while nutrients are now moved, either as inputs (fertilizer and feed products) or produce on a global scale. As a result, nutrient and land management in agricultural catchments has moved from a rural concern to regional and in some cases national economic security issue. Agricultural catchments now operate in a global market place with new pressures, challenges and therefore solutions. For instance, increased grain and animal production in Brazil is making inroads into traditional US markets and US producers supply a large percentage of the Japanese meat as water quality constraints in Japan limit costeffective production there. Also, the recent fervor for biofuel production is likely to have a dramatic and lasting impact on agriculture and catchment management and will be discussed later in this chapter in terms of trying to avoid previous infrastructural shortcomings that could adversely affect water quality. As a consequence of the spatial separation of crop and animal production systems, fertilizers are produced (i.e. N) or imported (i.e. P) to areas of grain production. The grain and harvested N and P are then transported to areas of animal production, where inefﬁcient animal utilization of nutrients in feed (<30% utilized) are excreted as manure. This has led to a large-scale, one-way transfer of nutrients from grain- to animal-producing areas that crosses catchment and even national boundaries and has dramatically broadened the emphasis of catchment management strategies. The rapid growth of the animal industry in certain areas of the USA, which is recorded

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on a smaller scale in the UK, especially in the west and south-west and in, for example, South Island New Zealand, has been coupled with an intensiﬁcation of operations. Over the last 10 years, for example, the number of dairies has decreased by 40%, but herd size has increased by 50%, while the number of hogs has increased 18% on 72% fewer farms (Gardner 1998). Similarly, 97% of poultry production comes from farms with more than 100,000 birds and over a third of beef production from <2% of the feedlots. As a result of this regional intensiﬁcation, modern farming systems have become fragmented with increasing separation of crop and livestock production, even across regional boundaries (Lanyon 2005). Similar trends in historical changes have been noted in Europe (Withers et al. 2002). In general, catchments dominated by animal feeding operations (AFOs) have become net sinks for nutrients imported in fertilizer to apply to local crops or in animal feed, as animals utilize <30% of the nutrients they are fed (Poulsen 2000; Valk et al. 2000). Consequently, catchments that include AFOs determine the magnitude of nutrient surpluses at farm and catchment scales depending on the type and intensity of livestock operations and the land area available for spreading of the manure produced. For example, the potential for N and P surpluses on farms with AFOs can be much greater than in cropping systems where nutrient inputs become dominated by livestock feed rather than fertilizer purchased explicitly for application to crops (Table 5.1). With a greater reliance on imported feeds, only 30% of N and 29% of P in purchased feed for a 1280-hog operation on a 30-ha farm in Pennsylvania could be accounted for in farm outputs (Table 5.1). As the intensity of animal production within a catchment increases, more P must be recycled, the P farm surplus (inputoutput) becomes greater, soil P levels increase and the overall risk of nutrient loss tends to increase (Haygarth et al. 1998; Sharpley 2000; Withers et al. 2003a).

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Farming system and nutrient budget (data adapted from Lanyon 2000, and Bacon et al. 1990) Nutrient input (kg ha−1 yr−1) in:

Farming system

Animal unitsa

Feed

Fertilizer

Output in produce

Surplus

Nutrient utilization

Nitrogen budget Cash cropb Dairyc Hogd Poultrye

– 123 142 296

– 155 390 5800

95 40 10 –

92 75 120 1990

3 120 280 3810

97 38 30 34

Phosphorus budget Cash cropb Dairyc Hogd Poultrye

– 123 142 296

– 30 105 1560

22 11 – –

20 15 30 440

2 26 75 1120

91 37 29 28

a

One animal unit (AU) is 454 kg live weight and is equivalent to 0.74 dairy cow, 1.14 dry heifer, 9 slaughter hogs and 250 poultry layers (US Department of Agriculture National Agricultural Statistics Service 2007). b 30-hectare cash crop farm growing corn and alfalfa. c 40-hectare farm with 65 dairy Holsteins averaging 6600 kg milk cow−1 yr−1, 5 dry cows and 35 heifers. Crops were corn for silage and grain, alfalfa and rye for forage. d 30-hectare farm with 1280 hogs; surplus includes 36 kg P and 140 kg N ha−1 yr−1 manure exported from the farm. e 12-hectare farm with 74,000 poultry layers; surplus includes 180 kg P and 720 kg N ha−1 yr−1 manure exported from the farm.

5.3 Current Consequences In order to establish guidelines for management change within agricultural catchments, it is ﬁrst necessary to deﬁne and quantify what actually constitutes impairment along with baseline or acceptable conditions. In response, the USEPA established the National Regional Nutrient Criteria Program in the Ofﬁce of Water (Table 5.2; US Environmental Protection Agency 2001). Similarly, in Europe, the restoration, protection and maintenance of ‘good’ water quality is a key goal in managing agricultural catchments. The EU WFD requires the setting of ‘reference conditions’ for different water body types in order to deﬁne how far they have become impaired (Council of European Communities 2000). Reference condition is deﬁned as ‘no or only very minor alterations’ resulting from human activities to the water body’s physiochemical, biological and hydrological properties. Clearly, nutrient status is only one of many factors inﬂuencing the health of aquatic ecosystems.

In the USA, baseline or reference conditions of pristine waters are deﬁned by monitoring total P, total N, chlorophyll a and clarity, in waters draining areas with minimal human impact (Gibson et al. 2000). Because it can be argued that most, if not all, lakes have been impacted by human activity to some degree, reference conditions represent the least impacted conditions or what is considered to be most attainable. The following guidelines are used to deﬁne minimally impacted lakes: catchments with <1% urban land use, >65% of lakeshore has at least 10 m of natural bank-side vegetation, no point source discharge inputs and no control of water-level ﬂuctuations (Gibson et al. 2000). Baseline concentrations of total N and P for freshwater systems in the continental USA are shown in Table 5.2. The difference between baseline and current N and P levels is used to determine and target reductions required in a catchment to alleviate water quality impairment. In the EU, countries have governmental powers to implement the WFD and water quality goals are achieved through river basin management planning.

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Managing Agricultural Catchments Table 5.2 Baseline N and P concentrations for each of the aggregated nutrient Ecoregions in the USA for freshwater systems (adapted from US Environmental Protection Agency 2001) Total N (mg L−1) Aggregated ecoregion Region I II III IV V VI VII VIII IX X XI XII XIII XIV a

Total P (μg L−1)

Description

Rivers and streams

Lakes and reservoirs

Rivers and streams

Lakes and reservoirs

Willamette and Central Valleys Western Forested Mountains Xeric West Great Plain Grass and Shrub Lands South Central Cultivated Great Plains Corn Belt and Northern Great Plains Mostly Glaciated Dairy Region Nutrient Poor Largely Glaciated Upper Midwest and Northeast Southeastern Temperate Forested Plains and Hills Texas–Louisiana Coastal and Mississippi Alluvial Plain Central and Eastern Forested Uplands Southern Coastal Plains Southern Florida Coastal Plains Eastern Coastal Plains

0.31 0.12 0.38 0.56 0.88 2.18 0.54 0.38 0.69 0.76 0.31 0.91 – 0.71

– 0.10 0.40 0.44 0.56 0.78 0.66 0.24 0.36 – 0.46 0.52 1.27 0.32

47 10 22 23 67 76 33 10 37 128a 10 40 – 31

– 9 17 20 33 38 15 8 20 – 8 10 18 8

This high value may be either a statistical anomaly or reﬂects a unique condition.

Even so, water quality concerns associated with agriculture have arisen at a catchment rather than farm scale, reﬂecting the cumulative impacts of individual nutrient and land management activities on farms. In many areas, agriculture’s contribution to water quality impairment has been, or is, masked by non-agricultural nonpoint and point sources to varying degrees. This raises issues of scale (both spatial and temporal), difﬁculties attributing cause and effect to ensure adaptive management, and the need to consider interactions with other nutrient sources, which must all be taken into account when developing catchment strategies for agricultural management that sustain productive viability and maintain water quality. 5.3.1 Scale of management As catchment area increases, so does the complexity of management, in terms of source identiﬁcation, remedial efforts and assessing which

strategies/practices have actually contributed to export reductions. The scale and coverage of catchment area to be targeted for BMPs requires an understanding of the processes controlling N and P loss. Nitrogen losses are generally less scale-dependent and more management-related, occurring from a large area of a catchment (Heathwaite et al. 2000). For P, however, BMP measures are targeted at critical source areas based on catchment research showing that the majority of the loss originates from only a small proportion of the catchment, the 80 : 20 rule. These are essentially P hotspots with active hydrological connectivity by overland ﬂow (Pionke et al. 1996). In the UK, a twin-track approach to controlling nutrient loss from agricultural catchments is being considered (Department for Environment, Food and Rural Affairs 2003). The ﬁrst track seeks to proactively implement speciﬁc solutions in high priority ‘sensitive’ areas based on detailed problem and monitored assessments (Fig. 5.1).

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Fig. 5.1 The twin-track policy for control of non-point source impairment in agricultural catchments. (Adapted from Department of Environment, Food and Rural Affairs 2003; Sharpley et al. 2005.)

Targeting of speciﬁc solutions to problem areas has a number of advantages; it helps secure practical action where it is most urgently needed, provides a basis for assessing the impact of control options through monitoring and helps to meet national legislative commitments. The second track seeks countrywide adoption of basic good nutrient and land management practice to prevent accelerating nutrient loss and further deterioration in water quality in less vulnerable areas and in the longer term. It provides a mechanism for improving the general level of environmental management on farms and provides a lever for adoption of those targeted solutions that can be easily and widely adopted at minimal cost (Fig. 5.1). 5.3.2

Adaptive management

There are several land management practices and system options available, which in and of themselves can lead to nutrient loss reductions. To be

effective in decreasing nutrient loads, where there is a wide range in production systems, however, there must be careful selection and targeting of conservation practices and management strategies. These practices vary in effectiveness among catchments and there will be synergistic effects on nutrient loss reductions, where combinations of these practices can further enhance overall reduction, when appropriately targeted to areas of high nutrient source availability and those which are hydrologically active. This is the basis for adaptive management of these practices, such that if nutrient loss reductions are not achieved by implementing nutrient efﬁciency, edge-of-ﬁeld buffers or offsite wetlands, then one or all are reassessed and modiﬁed. An adaptive management approach provides an appropriate way for decision makers to deal with the uncertainties inherent in the environmental repercussions of prescribed actions and their inﬂuences on water quality. Adaptive management can be applied at multiple scales.

Managing Agricultural Catchments At a basin scale, adaptive management requires measurement of both nutrient loadings and the extent and duration of speciﬁc water quality impairment. Although it will not be possible to relate these changes to speciﬁc changes in a basin, these data will provide better understanding of the relationships between nutrients and water quality impairment. On smaller scales, management actions can be treated as experiments that test hypotheses, answer questions and thus provide future management guidance. This approach requires that conceptual models are developed and used and relevant data is collected and analysed to improve understanding of the implications of alternative practices. Research driven by adaptive management is conducted in a framework where the testing of hypotheses and the new knowledge gained is then used to drive management adaptations, new hypotheses and new data gathering on endpoints. Unlike the traditional model of hypothesisdriven research, adaptive management implies co-ordination with stakeholders and consideration of the economic and technological limitations on management. Unlike traditional demonstration projects, adaptive management implies an understanding that complex problems will require iterative solutions that will only be possible through generation of new knowledge as successive approximations to problem solving are attempted. A basin-level response to practices cannot be expected to be observed for some time. We need a better understanding of the spatial and temporal aspects of basin-level responses, but must also focus on other scales at which response can occur in a more timely fashion. This would likely be smaller sub-catchment scales, where local water quality and quantity beneﬁts may become evident more quickly, and which enhance practice adoption. 5.3.3 Changing the paradigm of catchment management The accumulation of nutrients within terrestrial and aquatic environments is such that even if N

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or P was no longer added to agricultural systems, there would be a considerable time-lag (years or decades) before improvements in water quality, or regeneration of diverse habitats, might become apparent. Thus, the emphasis of catchment management should be on preventing further deterioration and taking strategic and sustainable actions sooner rather than later, otherwise we are simply and literally storing up more severe problems for future generations to confront. There has been a fundamental shift from current general guidance on Good Agricultural Practice (e.g. Department for Environment, Food and Rural Affairs 2002b) to more proactive implementation of cost-effective and targeted Best Management Practices (BMPs) (Department for Environment, Food and Rural Affairs 2003), and more recently the Whole Farm Approach (http://www.defra. gov.uk/farm/wholefarm/index.htm) with mutual farmer-regulator agreement of local solutions to local problems. In turn, this will require provision for additional farmer awareness, training and advisory support, involve a commitment to better record keeping and farm planning, and incur variable levels of cost including capital grant support (Withers et al. 2003b). It is recognized that a combination of BMPs, involving not just better management of nutrient inputs but also better management of land, are required, and that these must be implemented at a sufﬁcient intensity across the catchment to achieve desired goals (Sharpley and Rekolainen 1997; Withers and Jarvis 1998; Sims and Kleinman 2005). It is still, however, unclear to what extent N and P loss from existing land uses can be overcome by more sensitive management rather then the alternative of widespread land use change or restrictive farming. Under proposed reforms of the EU Common Agricultural Policy (CAP), the role of agriculture in the rural economy is being re-evaluated. Signiﬁcant areas of agricultural land may come out of production in an attempt to strive towards a more sustainable balance between a viable agriculture, a diverse range of habitats and good water quality. In the USA, the Conservation Reserve Program (CRP) has proved successful in improving wildlife habitats and

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water quality through establishing perennial ground cover.

Box 5.2 Socioeconomic barriers to conservation adoption A good example of this can be seen from the Cannonsville Watershed (1180 km2), which is a drinking water supply for New York City and claims a 93% participation in volunteer conservation programmes (Watershed Agricultural Council 2004). A survey of CRP participants showed they were generally older and more likely to obtain information from extension agents, consultants, and catchment council personnel than non-participants, but there was no difference in educational level or farming status (full or part time) (James 2005). Overall, negative attitudes toward voluntary adoption of BMPs focused on the loss of productive land and loss of being able to decide independently what to do on their own land. This survey illustrates the complexities of adopting BMPs among farmers in any given catchment, complexities that are related not only to the transfer of new BMP technology but also to socioeconomic pressures.

5.4 New Scientiﬁc Insights 5.4.1 Managing connectivity and complexity The interconnected character of landscape features within catchments is only recently being quantiﬁed (Bis et al. 2000; Lane et al. 2006; US Environmental Protection Agency 2006). Catchments are the minimum unit of ecosystem management for water quality (Matlock et al. 1994; Verdonschot 2000); certainly some ecological processes require larger geographical areas of management, but for ecosystem processes to be considered, no smaller unit of management is reasonable. Beginning with the River Continuum

Concept in 1980, stream ecologists recognized that the ecological integrity of the stream, and thus water quality, were the product of all the activities in the catchment. This concept is a simple statement of connectivity: From headwaters to mouth, the physical variables within a river system present a continuous gradient of physical conditions. This gradient should elicit a series of responses within constituent populations resulting in a continuum of biotic adjustments and consistent with patterns of loading, transport, utilization, and storage of organic matter along the length of the river. (Vannote et al. 1980).

This complexity is illustrated by measuring the impact of agricultural or municipal enrichment of nitrogen and phosphorus in water bodies. Simple measurements of water chemistry, or estimates of load, do not provide adequate information to assess the potential impact of nutrient enrichment (Matlock et al. 1999). Rather, the biotic and ecosystem responses to nutrient enrichment are often muted or even masked by other ecosystem processes or characteristics (Fig. 5.2). Even this very simple model of connectivity, when quantiﬁed, yields very complex interactions with indiscrete thresholds involving multiple variables. Thus, predicting some indication of algal growth (chlorophyll a production, for example) under multiple regimes of N and P enrichment is not possible without also understanding light regime (itself a function of canopy, stream bank entrenchment, channel orientation, season and suspended solids), substrate stability (a function of size, hydrological ﬂow regime, land use characteristics) and grazing (a function of benthic macroinvertebrate and ﬁsh community structures). 5.4.2 Assessing catchment processes and condition There are numerous biotic and chemical indices for assessing the condition of a waterbody within a catchment context (Verdonschot 2000; Sharpley et al. 2005). Of the numerous biotic indicators of

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Fig. 5.2 Interactions of factors potentially affecting algal growth in streams (Ludwig 2006)

catchment processes that have been evaluated, the benthic macroinvertebrate community is most commonly measured (Yagow et al. 2006). In 2005, the USEPA conducted the Wadeable Streams Assessment (WSA) to provide a statistically defensible analysis of the status of the streams and small rivers (wadeable) across the USA (US Environmental Protection Agency 2006). This was the ﬁrst analysis of this scale in the USA, and included chemical, physical, biological, habitat, geomorphological and catchment reach data from 1392 randomly selected perennial stream sites across the USA. Results indicated that 42% of the streams measured were in poor biological condition when compared with an ecoregion-based least-disturbed site. Only 28% of the streams were in good biological condition (Fig. 5.3). The most common stressors affecting stream biota were nitrogen, phosphorus, riparian disturbance and streambed sediments (Fig. 5.4). The sources of these stressors were not speciﬁcally identiﬁed. These stressors are characteristic of pollutants from agricultural catchments. The

relative risk to biological condition was determined to be greatest from streambed sediments, P and N (Fig. 5.4). Streams in the Interior Highlands had almost twice the percentage of stream lengths with relatively high phosphorus concentrations and half again more reaches with high nitrogen concentrations (Figs 5.5, 5.6). Over 51% of the wadeable streams in the interior highlands region of the USA were rated as having poor biota (Fig. 5.7). However, 26% of both interior highlands and plains/lowlands streams showed poor condition with respect to streambed sediments (Fig. 5.8). Stream condition was affected by catchment characteristics, especially location within the drainage network. 5.4.3 Managing complex ecosystem processes in catchments Managing complex ecosystem processes requires explicit understanding and mapping of these connections. The USEPA developed the Causal Analysis/Diagnosis Decision Information System (CADDIS) using a formal stressor identiﬁcation

Fig. 5.3 Biological condition of the wadeable streams of the USA, measured in 2005 at 1392 randomly selected perennials and streams and small rivers across the USA (US Environmental Protection Agency 2006)

Fig. 5.4 Extent of stressors and their relative risks to the biological conditions of US streams, as measured by the Wadeable Stream Assessment (USEPA 2006)

Fig. 5.5 Percent stream with low, medium and high P concentrations determined as part of the Wadeable Stream Assessment for streams across the USA and by three major geographical regions. Streams with low stressors or in good conditions were comparable to reference conditions within the appropriate ecoregion. Streams with medium stressors or in fair condition were greater than 75% of the reference condition, and high stressors or poor condition were worse than the 95th percentile of the reference condition. (Adapted and compiled from USEPA 2006.)

Fig. 5.6 Percent stream with low, medium and high N concentrations determined as part of the Wadeable Stream Assessment for streams across the USA and by three major geographical regions. Streams with low stressors or in good conditions were comparable to FLerence conditions within the appropriate ecoregion. Streams with medium stressors or in fair condition were greater than 75% of the reference condition, and high stressors or poor condition were worse than the 95th percentile of the reference condition. (Adapted and compiled from USEPA 2006.)

Fig. 5.7 Percent stream with low, medium and high streambedsediment concentrations determined as part of the Wadeable Stream Assessment for streams across the USA and by three major geographical regions. Streams with low stressors or in good conditions were comparable to reference conditions within the appropriate ecoregion. Streams with medium stressors or in fair condition were greater than 75% of the reference condition, and high stressors or poor condition were worse than the 95th percentile of the reference condition. (Adapted and compiled from USEPA 2006.)

Fig. 5.8 Percent stream with low, medium and high total biological control determined as part of the Wadeable Stream Assessment for streams across the USA and by three major geographical regions. Streams with low stressors or in good conditions were comparable to reference conditions within the appropriate ecoregion. Streams with medium stressors or in fair condition were greater than 75% of the reference condition, and high stressors or poor condition were worse than the 95th percentile of the reference condition. (Adapted and compiled from USEPA 2006.)

Managing Agricultural Catchments process that incorporates conceptual models using a stepwise process of elimination and riskbased assessment (US Environmental Protection Agency 2007). Based upon the USEPA Stressor Identiﬁcation Guidance Document (US Environmental Protection Agency 2000), CADDIS is a major deviation from the traditional methods of managing ecosystems because it does not utilize a reference or baseline dataset for comparison. Rather, CADDIS recognizes that all ecosystems are in stages of dynamic equilibrium, and that management of those ecosystems requires understanding thresholds of impact from multiple stressors rather than incremental impact from individual stressors. Future catchment management will require ongoing assessment of stressorresponse relationships in order to identify the impact of current as well as future stressors to water quality. Our increasing understanding of ecosystem process complexity and connectedness suggests that management endpoints are not static, nor

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are they absolute. Rather, catchment management is anthropocentric, a human endeavour. The use of reference conditions for setting environmental endpoints does not stand up well to scientiﬁc or even historic scrutiny. The goal in ecosystem management is improvement, not perfection. Landscapes within catchments are continuously changing, so management strategies must also change.

5.5 New Operational Insights Development of effective extension and education programmes plays an important role in improving farmer awareness of nutrient loss as an environmental issue, in demonstrating and implementing sustainable catchment management, and overcoming barriers to uptake (Table 5.3). Some farmers are reluctant to admit that nutrient loss originates from their farm, even

Farm management planning can reduce nutrient applications

A survey of 127 farms (90% of all farms) in two north-eastern Wisconsin catchments offers some insight into how successful nutrient management has been in reducing nutrient applications and thereby catchment losses (Shepard 2005). Farmers with a nutrient management plan (53% of farms) applied less N (139 kg ha−1) and P (31 kg ha−1) than farms without a plan (188 kg N and 44 kg P ha−1), but only half the farmers credited on-farm manure N, and only 75% fully implemented their plans on a majority of their acres. Targeted BMPs can decrease nutrient export The importance of targeting conservation management within a catchment is shown by several studies in the Little Washita River catchment (54,000 ha) in central Oklahoma (Sharpley and Smith 1994). Nutrient export from two sub-catchments (2 and 5 ha) were measured from 1980 to 1994, while conservation practices were installed on about 50% of the main catchment. Practices included construction of ﬂood control impoundments, eroding gully treatment and conservation tillage. Following conversion of conventional-till (moldboard and chisel plow) to no-till wheat in 1983, N export was reduced 14.5 kg ha−1 yr−1 (3-fold) and P loss reduced by 2.9 kg ha−1 yr−1 (10-fold; Sharpley and Smith 1994). A year later, shaping eroding gullies decreased P loss ﬁvefold and construction of an impoundment decreased P loss from the sub-catchments by 13-fold (Sharpley et al. 1996). There was no effect of BMP implementation, however, on P concentration in ﬂow at the outlet of the main Little Washita River catchment. Thus, a lack of effective targeting of nutrient management and control of major sources of nutrient P export likely contributed to ﬁeld or sub-catchment scale responses not being translated to reductions in nutrient export from the main Little Washita River catchment.

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Table 5.3 Potential barriers to uptake of catchment management options (adapted from Dwyer et al. 2002; Withers et al., 2003b; Sharpley et al., 2005) Potential barriers Awareness

Skepticism Willingness

Ability

Practicality Cost

Effectiveness Complexity Mechanism

Description Lack of awareness by farmers of ground and surface problems, and potential N and P loss from agriculture, in terms of their existence, nature, causes, solutions, ﬁnancial impact, legislation and impending regulatory controls Farmer skepticism as to effectiveness and/or legitimacy of policy mechanisms, distrust of certain information sources or to the need to manage N and P inputs. For example, over-regulation and high administration costs Lack of willingness to act by farmers because they either do not take responsibility for contributing to the water quality problem, or do not consider it important enough or directly relevant to the farm business. Clearly relates to lack of awareness and perception above, but also includes willingness to overcome the ‘hassle’ factor, and adopt a greater level of farm management Limited ability of farmers to plan, manage, and implement certain BMPs, without specialist training, advice or equipment. Higher level of skill required in using farm decision support tools, understanding some recommendation systems or using specialist equipment Practicality of various BMPs to suit different sites. This will differ among catchments and requires integration of farmers’ ideas. Options need to be realistic, practicality needs to be demonstrated and linked to effectiveness Cost of implementing different BMPs across a range of farm types, particularly during periods of low proﬁtability. Some options may require access to capital grants to change farm infrastructure when farmer has ﬁnancial outlay Effectiveness of BMPs in reducing N and P loss. Linked to lack of research data and demonstration of effectiveness Often complexity of schemes is too great and not co-ordinated. For example, written instructions that are too long, too complex or generally inaccessible. Overlap and incompatibility between schemes on the same farm Mechanism or strategy for implementing different BMPs needs to be considered carefully to ensure adoption (devil is in the detail)

though they accept there are water quality problems in their nearest watercourse. Many farmers are still not fully aware of the nutrient value of applied manure, or indeed the role of soil testing in deciding on fertilizer use, but both are management practices that may increase farm proﬁtability by reducing dependence on fertilizers. Most evaluations of BMP effectiveness at reducing nutrient export from catchments conclude, however, that nutrient management is the single most effective measure for controlling nutrient losses (Sharpley et al. 2006). 5.5.1

Potential BMP tradeoffs within agricultural catchments

In the past, separate strategies for either N or P have been developed and implemented at farm or catchment scales, such that advice given to control P loss may also conﬂict with advice given to control N loss. This needs to change. Because

of different critical sources, pathways and sinks controlling N and P export from catchments, remedial strategies directed at either N or P control can negatively impact the other nutrient. For example, basing manure application on crop N requirements to minimize nitrate leaching to groundwater can increase soil P and enhance potential P losses (Sharpley et al. 1998; Sims 1997). By contrast, reducing surface runoff losses of P via conservation tillage can enhance nitrate leaching. Similarly in the UK, a BMP for reducing P loss is to rapidly incorporate land-applied manure in the autumn when soil conditions are dry, but this practice is prohibited in Nitrate Vulnerable Zones due to the greatly increased risk of N leaching loss. Increased pesticide usage associated with establishing coarse seedbeds to reduce runoff risk, and increased sediment or N loss where cultivation is needed to incorporate manures to reduce runoff P losses are other exam-

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Box 5.4 Shifts in agricultural production can affect water quality Shifts in agricultural production often occur due to external pressures. For instance, the drive for biofuel production to be a greater share of consumed energy led to a 6.5 million ha (16 million acre) increase in corn acreage in the USA from 2006 to 2007. Most of this increase came from land currently in soybean, Conservation Reserve, and pastures (Simpson et al. 2008). Assuming fertilizer application rates will be maximized to obtain optimum yields as a consequence of high corn prices, it is expected that the potential for N and P loss will increase compared with losses from pre-corn land use. Further, dry distiller grain (DDG; 0.8–0.9% P), a byproduct of ethanol production, can be used in animal feed. Even with <20% DDG supplementation of dairy cow diets, this elevates ration P to 0.5% P (0.33–0.36% P recommended), offsetting reductions gained through feed management. This will increase the P content of manure and potential P loss in runoff if land applied. Clearly, increased ethanol production has become a necessity and cellulosic biofuel production will eventually increase. However, catchment management strategies must plan to minimize the potential for further water quality degradation associated with such production changes.

Box 5.5

Level of BMP implementation affects nutrient loss reduction

Land application of dairy manure in the LaPlatte River Basin, Vermont (8832 ha) was identiﬁed as an important source of P to Lake Champlain (2.2 kg P ha−1 yr−1; Meals 1990). To address this, remedial measures, such as control of barnyard runoff, milk-house waste treatment and construction of waste storage facilities, were implemented in the basin. There was little decrease in P runoff with increasing percent of animals in a catchment under a conservation practice (Meals 1990). If the runoff P for catchments where less than 50% of the animals were under conservation practices are not considered, then both dissolved and total P in runoff were decreased signiﬁcantly (r2 of 0.68 and 0.75, respectively; P < 0.05). The low values of implementation (<42%) represent the initial years of land treatment when nutrient management implementation was incomplete. Apparently, there is a minimum threshold level of implementation, which must be achieved before a signiﬁcant response to conservation practices occurs. ples (Dampney et al. 2002). Conﬂicting advice can lead farmers to question the reliability and philosophy of such programmes, as well as a reluctance to use recommended management practices. 5.5.2 Nutrient trading and agricultural catchments A system of buying and selling pollution credits within a given catchment, similar to that adopted for air quality control in the USA in the early 1990s, has been suggested in order to develop catchment-wide management programmes aimed at achieving water quality goals (US Environmental Protection Agency 2003). The appeal of trading is its potential to achieve water quality goals at a

lower cost to society than a regulatory approach (Shortle and Horan 2001). In a well-functioning trading market, operations with lower control costs will emit less nutrients, while those with higher control costs will emit more. The potential trading between point and nonpoint sources of nutrients has received more attention. While successes have been achieved in trading air quality permits, signiﬁcant challenges have been encountered in applying this mechanism to water quality. For instance, water quality emissions from non-point sources are difﬁcult or impossible to observe and are related to uncertain events such as weather. It is not clear, therefore, what basis should be used to measure performance or outcomes. In addition, potential buyers and sellers to non-point source trades can face

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Box 5.6 Nutrient trading and water quality The World Resources Institute, in collaboration with the United States Department of Agriculture, developed an internet-based demonstration market for nutrient trading (Woodward et al. 2006). The internet site is http://www.nutrientnet.org/index.cfm. The presumption is that trading of nutrients will be motivated by point sources seeking to reduce costs of compliance with nutrient criteria in their National Pollution Discharge Elimination System (NPDES) permits. This market-based approach integrates a broker into the market to reduce transaction costs in increased transaction rigor. The governing criteria for nutrient trading under this construct is that trading must reduce total load of the nutrients of concern to the target water body. Guidelines for trading include restrictions of all trades for a given market to a speciﬁc basin (no trans-basin trading of loads) and incorporation of a non-point–point source safety factor of 2:1 for increased conﬁdence of impact. The trades are constructed to be market contract transactions, enforceable through contract law, and not regulated per se. The presumption is that the non-point source generators will be responsible to the point source generators through the trading contract, and that regulatory management will incorporate a suitable mechanism for arbitration and restoration should a violation occur.

signiﬁcant costs of organizing and negotiating trades. Although few have studied the costs as a barrier to nutrient trading, it is possible that brokers may act as a ‘go-between’ in this process. To demonstrate the potential of nutrient trading programmes, the USEPA is piloting programmes in a number of catchments that involve farming (US Environmental Protection Agency 2003). Other programme options involve possible trading among farmers able to limit nutrient loss below recommended levels, who could sell credits to a farmer unable to meet these levels. The number of credits a farmer has could be linked to farm area, crop production, and where appropriate number of animals. As a result, nutrient export from a catchment may be kept within predetermined limits by sharing management responsibilities among farmers. It should be noted, however, that ‘pollution trading’ has been criticized by some environmental groups because it is perceived as allowing wealthier operations to buy the ‘right to pollute’. Heated debate will doubtless precede widespread adoption of pollution credits for agriculture, hence comprehensive analysis and planning to justify their value and need is required. Despite these challenges, the lack of progress with actual trades of water pollution permits, as

well as environmental and others concerns noted above, there continues to be much effort to study and promote trading in the water quality arena in the USA. Two recent federal level developments related to trading are the water quality trading guidance (US Environmental Protection Agency 2003) and trading credit provisions in the 2002 Farm Bill (US Department of Agriculture 2003). 5.5.3 Implementing remedial strategies within catchments Because of the time and expense involved in assessing nutrient export from agricultural catchments, models are often a more efﬁcient and feasible means of evaluating management alternatives. In their most comprehensive form, models can integrate information over a catchment scale to identify BMPs and critical source areas where BMPs are most likely to affect catchment-scale nutrient losses. Many complex models are available and are gaining greater acceptance with managers and planners, as computers become more powerful and we are generally more comfortable using them. Because models yield clear numerical results with which to gauge progress, however, they have a strong appeal to policymakers and managers, an appeal

Managing Agricultural Catchments that can sometimes bring false conﬁdence and misconceptions (Boesch et al. 2001). It has been said that while all models are wrong, some are useful (Silberstein 2006). It is of critical importance that modellers clearly deﬁne what his or her model is useful for and what it is not designed to do. Likewise, users must decide what they want to accomplish with a model. For example, one must consider the scale (ﬁeld, catchment or basin), time (ﬂow event, annual or multi-year) and level of accuracy (0.1 or 10 kg ha−1 yr−1) that needs to be simulated, as well as the amount of parameterization data available. Thus, a key to useful simulation of nutrient loss is selection of the appropriate model and data to run it. If, for instance, one only needs to identify areas in a catchment at greater risk for N or P loss to target remedial BMPs, then more simplistic site vulnerability tools are appropriate (Heathwaite et al. 2000). More reliable and process driven simulation of ﬂuvial processes to estimate nutrient transfer in stream channels is critical to the accurate prediction of land management or BMP impacts on the loss of N and P to receiving waters and thus, biological response. Further, the accurate simulation of the forms of N and P transported in

Box 5.7

123

streams is even vital to selection of appropriate BMPs or remedial measures that would most effectively bring about an improvement in surface water quality. For example, if a stream system is acting as an active source of either dissolved or particulate P, such that it contributes a relatively large proportion of P input to surface waters compared to edge-of-ﬁeld losses, then remedial measures should be targeted to the stream channel. Even so, it is clear that there can be a great deal of uncertainty in model computations. Uncertainty arises in connection with an imperfect representation of the physics, chemistry and biology of the real world, caused by numerical approximations, inaccurate parameter estimates and data input, and errors in measurements of the state variables being computed. Whenever possible, this uncertainty should be represented in the model output (e.g. as a mean plus standard deviation) or as conﬁdence limits on the output of a time series of concentrations or ﬂows. The tendency described earlier for decision makers to ‘believe’ models, because of their presumed deterministic nature and ‘exact’ form of output, must be tempered by responsible use of the models by engineers and scientists, such that model computations or predictions are not over-

Non-point source catchment models

There are three types of models that simulate the runoff and water quality from catchments: • Process-based models – Models that explicitly simulate catchment processes, albeit usually conceptually. These models typically involve the numerical solution of equations that are a mathematical representation of processes such as rainfall runoff; inﬁltration leaching; N and P application method, rate, and timing; land management; fate and chemical transformation of added N or P in soil, etc. • Export coefﬁcient models – Models that rely upon land use categorization (sometimes through a linkage to a GIS evaluation) coupled with export coefﬁcients or event mean concentrations. These models calculate nutrient export from catchments as the sum of individual loads from each source in the catchment. As export coefﬁcients are empirical, these types of models are as accurate as input data (as are also process-based models). • Statistical or empirical models – Models that involve regression or other techniques, which relate water quality measures to various characteristics of the catchment. These models range from purely heuristic regression equations (Driver and Tasker 1990) to relatively sophisticated derived-distribution approaches for prediction of the frequency distribution of loadings and concentrations (DiToro and Small 1984; Driscoll et al. 1989).

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sold or given more weight than they deserve. Above all, model users should determine that the model computations are reasonable in the sense of providing output that is physically realistic and based on input parameters that are within accepted ranges. Clearly, the role of modelling will be more and more important over the next decade in making catchment management and policy decisions related to conservation programmes and water quality. As required nutrient loss reductions from agriculture increase, so does the cost and complexity of remediation. If one is striving for a 40% reduction for example, the ﬁrst 30% may be relatively inexpensive to achieve compared to the remaining 10%. Thus, it will be Box 5.8

this remaining 10% which will present one the greatest challenges. So who pays? Catchments are naturally leaky and thus part of the responsibility should be borne by the public who require clean water along with cheap food. To a large extent this is being accomplished at a ‘grass roots’ level via voluntary alliances and partnerships among all vested stakeholders within a catchment. Voluntary and non-proﬁt catchment management groups started to develop in the 1990s to act as mediators between ‘action agencies’ and land users in a catchment and to stimulate change among diverse land users and their needs within a catchment. In doing this, stakeholder alliances are better able to build trust and support

Stakeholder alliances

Discovery Farms At a state level, the Discovery Farms Program is conducting research on privately owned Wisconsin farms in different geographical areas, facing different environmental challenges (see http://www. uwdiscoveryfarms.org/new/index.htm). The Discovery Farms Program has been very successful at gaining farmer support in at-risk catchments in efforts to ﬁnd the most economical solutions to overcoming the challenges environmental regulations place on farmers. The Chesapeake Bay State, federal and local groups in the Chesapeake Bay work together and with the public to work together to identify critical problems, focus resources, include catchment goals in planning, and implement effective strategies to safeguard soil and water resources (see http://www.chesapeakebay.net/index.cfm; Chesapeake Bay Program 2007). The Chesapeake Bay Program works within a collaborative organizational structure: members from partner organizations participate in a series of committees that drive and implement the Bay Program’s efforts. There are three main types of committees, which govern the Bay Program and guide policy changes, provide external perspectives on current issues and events, and work internally to coordinate restoration activities. This structure is designed to encourage partner organizations to share information and ideas and work cooperatively on Bay restoration. The Illinois River Watershed Partnership At a catchment level, the Illinois River Watershed Partnership (see http://www.irwp.org/index. html) was established in 2005 to improve and protect water quality in the Illinois River in Arkansas and Oklahoma by working at a grassroots level with catchment citizens and other organizations. Similarly, in New York a Watershed Agriculture Council was formed of farmers, civic leaders and representatives from the New York City Department of Environmental Protection to help guide management in the New York City Watershed (Revkin 1995; also see http://www.nycwatershed. org/).

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Fig. 5.9 Conceptual diagram of stakeholder management process for dynamic and complex catchment. (From Sabatier et al. 2005.)

within a catchment, share responsibilities for decisions or actions, create more cost-effective solutions more likely to be adopted, strengthen working relationships, and enhance communication and co-ordination of resources (Tetra Tech, Inc. 1999). Stakeholder alliances encourage collaborative not adversary relationships among the groups involved. Such alliances have been formed in response to recent public health issues related to the nutrient enrichment of waters in the eastern USA. For instance, ﬁshing and tourist industries have been severely affected in the Chesapeake Bay and inland waters of the Delmarva Peninsula by periodic, massive algal blooms as well as the encroaching dominance of undesirable aquatic biota. Sabatier et al. (2005) proposed a dynamic process of catchment management through stakeholder engagement (Fig. 5.9). This process

imposed a structure of technical engagement between policy and management authorities and catchment property owners/stakeholders. The process prescribes the creation of institutions that support and maintain the infrastructure necessary for collaborative engagement. The process also established a clear set of legitimacy criteria that must be met for the process to function. Collaborative stakeholder engagement is increasingly understood as difﬁcult and timeconsuming, but also necessary to avoid costly and destructive litigation (Matlock et al. 2002). Clearly, catchment partnerships and stakeholder alliances will play an increasingly important role in not only developing but also implementing environmental management strategies within agricultural catchments. As environmental pressures within catchments become greater and more politically charged, catchment alliances

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Ecological services by category (United Nations 2005)

Provisioning services Food Fibre Fuel Genetic resources Biochemicals, natural medicines, and pharmaceuticals Ornamental resources Fresh water

Regulating services

Cultural services

Air quality regulation Climate regulation Water regulation Erosion regulation Water puriﬁcation and waste treatment Disease regulation Pest regulation Pollination Natural hazard regulation

will likely become critical in mediating a consensus for future development in agricultural catchments.

5.6 Future Research and Operational Needs Future research must address both the common and unique aspects of ecosystem management at catchment scale across broad ranges of conditions. From an heuristic perspective, the development of the concept of ecological services has provided a unifying language for ecosystem managers and researchers (Daily 1997). Ecological services are those goods and services we (humans) derive from ecosystem structure and function, i.e. they are the things we get from the landscape. In 2005 the United Nations issued the Millennium Ecosystem Assessment (MEA), an international assessment of Earth’s life support system for humans (http://www.millenniumassessment. org/en/index.aspx). This report was prepared by 1360 experts from 95 countries, reviewed by an independent board of 80 editors, and represents the most extensive and current understanding of the condition of the Earth from a human survival perspective (United Nations 2005). The ecological services were divided into four major categories: 1 provisioning services; 2 regulating services;

Recreation and ecotourism Aesthetic values Educational values Knowledge systems Social relations

Supporting services Soil formation Photosynthesis Primary production Nutrient cycling Water cycling

Cultural diversity Spiritual and religious values Cultural heritage values Sense of place

3 cultural services; 4 supporting services. Each of these categories includes a series of explicit goods and services that humans derive from ecosystems (Table 5.4). The assessment of ecological services found that 60% of the ecosystem services evaluated were degraded or were being used unsustainably. Three major problems associated with management of Earth’s ecosystems are familiar to catchment managers: 1 Degradation and loss of ecological services is increasing; 2 basic ecosystem shifts are not well understood or predictable; and 3 the costs of these declining ecological services are disproportionately borne by the poor and disenfranchised, resulting in a disconnect between policy and implementation. These ﬁndings will deﬁne the human condition for the next 100 years; they should drive our catchment-scale research and management strategies. There exists a direct relationship between the constituents of human well-being (security, basic materials, health, relations and freedom) and ecosystem services; as ecosystem services decline, so does human well-being (Fig. 5.10). For real and lasting changes to occur in agricultural production, emphasis is needed on consumer-driven programmes and education, rather than assuming that farmers will absorb the burden of costs associated with implementing

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Fig. 5.10 The interconnectedness of ecosystem services with constituents of well-being. (Adapted from Millenium Ecosystems Assessment, United Nations 2005.)

remedial practices. Remembering that, except for farm-gate measures, BMPs are only a temporary solution to minimizing the environmental impacts of land management, farm-level nutrient management research should verify the impact of farm-gate imbalances of N and P on water quality. This is a difﬁcult proposition because farm boundaries are irregular, not usually contiguous, and often cross catchment boundaries. Despite these challenges, education programmes will depend on sound evidence for the need to change problematical management practices. Planning is needed for regional as well as national agricultural infrastructures to control

nutrient inputs to farming systems and assess large-scale nutrient balances. Infrastructure components might include education and extension programmes. For example, cost-share monies for conﬁned animal feeding operations in northeastern USA catchments are now linked to farmers demonstrating that nutrient inputs to the farm are reduced by feeding animals at a level consistent with National Research Council requirements. This exempliﬁes how public policy can address the source or cause of excess P concentrations and how public investments can provide a long-term mechanism for overcoming infrastructural barriers.

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Despite advantages to each of the nutrient management options, none should be seen or used individually as the primary mechanism by which a farmer reduces N and P losses. Moreover, implementation of BMPs over broad areas of a catchment does not always reduce nutrient exports from the catchment as a whole (Meals 1990; Sharpley et al. 1996; Sharpley and Rekolainen 1997). More information is needed on the performance of speciﬁc BMPs applied to critical source areas within a catchment. This information should lead to development of more effective remedial strategies. These assessments must go hand in hand with basic monitoring of nutrient export from pristine and nutrient-rich or nutrient-surplus catchments. With budgetary constraints in the EU and USA, however, this has been one of the main areas to receive funding cuts. Without baseline monitoring, it will be difﬁcult to document a change in water quality and, more importantly, the effectiveness or lack of in response to catchment remediation efforts. The lag time between BMP implementation and water quality improvements can often exceed the monitoring period due to limited long-term funding opportunities for this type of research. Despite our knowledge of controlling processes, it is difﬁcult for the public to understand or accept this lack of response. When public funds are invested in remediation programmes, rapid improvements in water quality are usually expected and often required. Thus, implementation of effective nutrient management BMPs should consider the re-equilibration of catchment and lake behaviour, where nutrient sinks may become sources of N and P with only slight changes in catchment management and hydrological response. The short-term goals of most research funding agencies preclude the opportunity for the research required to better understand nutrient sinks and sources (i.e. dynamics during transport) on a catchment scale. This may be the least understood component of implementing BMPs at the catchment scale, and deserves the attention of funding agencies and research scientists.

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1975–1995. Journal of Environmental Quality, 31, 17–24. Sabatier, P., Focht, W., Lubell, M., Trachtenburg, Z., Vedlitz, A. and Matlock, M. (2005) Swimming Upstream: collaborative approaches to watershed management. MIT Press, Cambridge, MA. Sharpley, A.N. (ed.) (2000) Agriculture and Phosphorus Management: the Chesapeake Bay. CRC Press, Boca Raton, FL. Sharpley, A.N. and Rekolainen, S. (1997) Phosphorus in agriculture and its environmental implications. In: Tunney, H., Carton, O.T., Brookes, P.C. and Johnston, A.E. (eds.), Phosphorus Loss from Soil to Water. CAB International Press, Cambridge, pp. 1–54. Sharpley, A.N. and Smith, S.J. (1994) Wheat tillage and water quality in the Southern Plains. Soil Tillage Research, 30, 33–38. Sharpley, A.N., Smith, S.J., Zollweg, J.A. and Coleman, G.A. (1996) Gully treatment and water quality in the Southern Plains. Journal of Soil and Water Conservation, 51, 512–517. Sharpley, A.N., Meisinger, J.J., Breeuwsma, A., Sims, J.T., Daniel, T.C. and Schepers, J.S. (1998) Impacts of animal manure management on ground and surface water quality. In: Hatﬁeld, J.L. and Stewart, B.A. (eds), Animal Waste Utilization: effective use of manure as a soil resource. Ann Arbor Press, Boca Raton, FL, pp. 173–242. Sharpley, A.N., Withers, P.J.A., Abdalla, C. and Dodd, A. (2005) Strategies for the sustainable management of phosphorus. In: Sims, J.T. and Sharpley, A.N. (eds), Phosphorus; agriculture and the environment. American Society of Agronomy Monograph. American Society of Agronomy, Madison, WI, pp. 1069–1101. Sharpley, A.N., Schmidt, J.P. and Hergert, L. (2006) Nutrient management practices. In: Schnepf, M. and Cox, C. (eds.), Environmental Beneﬁts of Conservation on Cropland: the status of our knowledge. Society of Soil and Water Conservation, Ankeny, IA, pp.149–193. Shepard, R. (2005) Nutrient management planning: is it the answer to better management? Journal of Soil and Water Conservation, 60, 171–176. Shortle, J.S. and Horan, R.D. (2001) The economics of nonpoint pollution control. Journal of Economic Surveys, 15, 256–289. Silberstein, R.P. (2006) Hydrological models are so good, do we still need data? Environmental Modelling Software, 21, 1340–1352.

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Image facing chapter title page: Courtesy of Lynn Betts, USDA Natural Resources Conservation Service.

6

Efﬂuent Management

ALAN JENKINS1 AND ROBERT C. FERRIER2 1

Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire, UK 2 The Macaulay Institute, Craigiebuckler, Aberdeen, UK

6.1

Introduction

Historically, rivers have provided a dual function of water supply and efﬂuent disposal. The recognition that these two functions are somewhat incompatible has led to the development of legislation to enable both functions to continue whilst retaining the integrity and quality of the water at a standard which is acceptable for human use. More recently it has been recognized that rivers provide ecosystem services which also demand a certain water quality. This development of understanding, coming in parallel with urbanization and industrial development resulting from worldwide economic and population growth, is perhaps further advanced in the developed regions of the world. Interestingly, as population and industrial activity increase so the demand for water increases and as a result the efﬂuent ﬂux returning to the river also increases. This tends to lead to increased pollution and decreased water quality, particularly, and increasingly, moving further downstream. The natural tendency, therefore, is to invoke increasingly tighter controls on the sources of pollutants, usually by demanding reduced pollutant ﬂuxes and often implemented via technological solutions or so called ‘end-of-pipe’ control measures. This has

been possible to a large extent since clean-up technology has improved at the same time as economic development has increased. It is the water engineers, therefore, who have been effectively charged with the duty of ensuring that water for supply attains an appropriately high standard and that efﬂuent is treated to remove pollutants to bring the water to an acceptable quality for return to the river. It is at the catchment scale, however, that water resources, including water quality, are managed and this process has to take into account all sources of pollutants and all abstractions and uses as well as the preservation of the natural ecosystem function. This chapter does not address the advances in technology, including wastewater engineering processes and so called ‘end-of-pipe’ treatments that have historically provided a mechanism to improve the quality of efﬂuent discharges to date. The development of such technological advances will undoubtedly continue in the future. Neither does this chapter consider the management of efﬂuent per se. It is the management of point source efﬂuent discharges in catchment systems where their impact needs to be considered alongside all other catchment scale inﬂuences, including diffuse pollution sources that are addressed here.

6.2 Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

Historical Perspective

The organic matter content of natural waters was for a long time considered the principal criterion

136 Table 6.1

alan jenkins and robert c. ferrier Potential damage to receiving waters by CSOs

Period of time

Kind of stress

Indicator/parameter

Short (few hours)

Hydraulic Chemical Physical Biochemical

Flow, shear stress, erosion, drift Ammonia, toxic substances Suspended matter Oxygen starvation in water and sediment (esp. easily degradable organic solids)

Delayed (days, weeks)

Hygienic Hydraulic Chemical Biochemical

Bacteria, viruses Erosion, morphology Ammonia, toxic substances Oxygen starvation in water and sediment (esp. organic solids)

Long term (weeks, years)

Hygienic Aesthetic Hydrological Chemical Biochemical

Bacteria, viruses Floatables, solids, waste, oil Hydrological regime, morphology Heavy metals, persistent organic substances Inorganic/organic solids/sediments; oxygen starvation by eutrophication

of acceptable quality. The principal sources of organic matter in rivers are domestic wastes with a typical Biochemical Oxygen Demand (BOD) in raw efﬂuent of 400–600 mg L−1 and industrial wastes with a wide range of BOD in raw efﬂuent of 800–50,000 mg L−1 depending on the industrial activity. Runoff from land surfaces can make a signiﬁcant contribution in agricultural catchments during rainfall events but this is generally a smaller and relatively intermittent source. In most urban areas, surface drainage systems are combined with sewerage systems to perform the dual role of moving sewage to waste water treatment plants (WWTP) to maintain public health and to divert waste water into a nearby water course during extreme rainfall to prevent urban ﬂooding. These are termed combined sewer overﬂows (CSOs) and can provide signiﬁcant inputs of organic pollutants and chemicals during periods of heavy rainfall when the transport capacity of the sewer system or the process capacity of the WWTP is insufﬁcient. Only much later, following the construction of most CSO networks in major urban centres, did the aspect of pollution control in the receiving waters become important leading to the introduction of treatment facilities. For example, in Paris, the ﬁrst evidence of CSO impacts on receiving waters was found in the 1960s but it was not until the 1990s that reducing the ﬂuxes from CSOs became a concern for the engineers (Even et al. 2007).

This was because most visible dry-weather pollution had been removed following the systematic construction of WWTPs and thereby turning attention to the CSO contribution to the pollutant load. The major problems of CSOs are the acute short-term impacts of dissolved pollutants, bacteria and viruses which lead to ﬁsh death and public health concerns. Additionally, CSOs can have a delayed and longer-term ecological effect linked to certain contaminants (Table 6.1). The concern about the effect of organic matter on river waters was centred on deoxygenation. Early work established a connection between the bacterial decomposition of organic matter and the deoxygenation of the receiving water. The concentration of dissolved oxygen downstream of an organic discharge results from two competing effects, deoxygenation and reaeration, which combine to produce the classic oxygen sag curve (Fig. 6.1). Together with some knowledge of the ﬂow regime of the receiving river (to estimate dilution) this has historically formed the basis for calculating discharge consents. These consents represent the agreed levels of discharge between the polluter (or polluting activity) and the catchment water managers with statutory responsibility. Similar principles have been adopted, although using different rate coefﬁcients, to consider other inorganic pollutants. In the UK, it was not until the second half of the nineteenth century that legislation was

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137

Fig. 6.1 The oxygen sag curve downstream of an organic matter efﬂuent. Initially the deoxygenation is high causing DO concentration to fall. As the organic matter is biologically degraded, the rate decreases rapidly. At the same time, as the DO concentration falls, the saturation deﬁcit increases and so the rate of re-aeration rises and exceeds the rate of deoxygenation. As this causes the DO level to return to saturation the re-aeration gradually declines.

introduced to treat sewage and other efﬂuent prior to discharge as a necessary measure for eliminating the threat of waterborne diseases. In 1861 an amendment to the 1848 Public Health Act in England and Wales was the ﬁrst legal requirement on polluters to treat sewage and other efﬂuent such that its discharge would not lead to deterioration of the receiving water. Since this time, there has been a steady increase in efforts to protect water quality by introducing legislation focused on the various uses to which the receiving water would be used, for example, drinking, recreation and ﬁsheries. This use-related legislation in the UK has been mirrored in the EU which from 1975 to 1991 issued Directives concerned with setting common standards for environmental quality, emissions, waste treatment and disposal, etc. The Directives that have had an impact upon the control of discharges are largely focused on protecting water

for prescribed uses and against certain key pollutants. All of this legislation has impacted the way in which efﬂuent discharges are considered and some have inﬂuenced the manner in which they are managed. It is worth noting that UK legislation since 1975 has reﬂected the EU Directives since the interpretation and implementation of a Directive is the responsibility of each individual Member State. The general approach to the control of efﬂuent discharges to both surface and ground waters in the UK has been to set discharge limits which are individually based on local circumstances. The alternative is to set uniform standards to all discharges of a particular size or type. This, however, runs the risk of not protecting the environment or placing too big a cost burden on the polluter to provide the speciﬁed treatment in circumstances where it may be unnecessary. Discharge requirements, therefore, are speciﬁed to the needs of the individual receiving water.

138

Box 6.1

alan jenkins and robert c. ferrier

Summary of early UK pollution control legislation.

Public Health Acts 1848–1872: placed responsibility for provision and maintenance of sewers and treatment works on local boards of health and made provision for protection of water supply. Subsequent amendments demanded treatment of sewage efﬂuent such that it would not affect the receiving water. Salmon Fisheries Act 1861: prohibited pollution of salmon waters with substances harmful to ﬁsh. Rivers Pollution Prevention Act 1876: prohibited discharge of solid matter into surface waters, established quality criteria to which the discharge should comply and allowed industrial pollutant discharges to sewerage systems. Public Health Act 1936: consolidated public health legislation concerning discharges that were dangerous. River Boards Act 1948: established the responsibility for pollution control at whole catchment level. Rivers (Prevention of Pollution) Act 1951: introduced the concept of discharge consents. Clean Rivers (Estuaries and Tidal Waters) Act 1960: empowered River Boards to treat discharges to estuaries and tidal waters in the same way as inland waters. Rivers (Prevention of Pollution) Act 1961: brought further discharges into compliance conditions via consents. The Water Resources Act 1963: introduced a consent system for underground discharges and provided for river authorities to sample any efﬂuent. The Water Act 1973: established regional water authorities and circumvented the problem of them issuing consents for their own sewage discharges. The Control of Pollution Act 1974: created a public register of discharge consents and water quality monitoring data. The Water Act 1989: enabled privatization of water industry for water supply and sewerage and established the National Rivers Authority as the regulatory agency.

The other cornerstone of discharge control policy in the UK has been the deﬁnition of water quality objectives which relate to the environmental need or end use of the receiving water. These water quality objectives were either statutory (as in the case of the EC Directive on the Discharge of Dangerous Substances to the Aquatic Environment) or non-statutory (e.g. the General Quality Assessment for comparative assessment of water quality in England and Wales). Once the water quality objective was deﬁned, however, it was possible to determine the efﬂuent quality necessary to achieve it, given knowledge of the dilution capacity (ﬂow or volume of water) of the receiving river, lake or groundwater.

The fact that point source pollution has been controlled through legislation for many decades in Europe, and indeed in developed countries around the world, as described above means that a signiﬁcant source of information exists to inform assessment of the industry sectors providing the major source of that pollution. As part of the implementation of the EU Water Framework Directive (see next section), an analysis of pressures and impacts has been conducted in all EU Member States. In Scotland, for example, this indicated that 17% of rivers from a total length of c. 25,000 km were at risk from point source pollution. Sewage and refuse disposal industrial sectors were identiﬁed as being responsible for this risk in c. 65% of these rivers. Manufacturing

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Box 6.2 Summary of EC Directives impacting on the control of discharges 1976–1991 1975 (75/440/EEC) Directive on quality requirements for surface waters intended for the abstraction of drinking water. – Deﬁned the water quality requirements for water to be used for abstraction to potable supply. The requirements impact on discharges upstream of the abstraction point to limit the concentration of key determinands discharged in the efﬂuent taking into account the assimilative capacity of the water body. 1976 (76/160/EEC) Directive on the quality of bathing water. – Required the improvement of bathing water standards primarily by the control of sewage discharges. 1976 (76/464/EEC) Directive on pollution caused by certain dangerous substances discharged into the aquatic environment of the community. – Under this and a series of further (daughter) Directives several hazardous and dangerous substances were controlled in efﬂuent discharges to meet pre-deﬁned environmental quality standards. 1978 (78/659/EEC) Directive on the quality of fresh waters needing protection or improvement to support ﬁsh life. – This Directive speciﬁed the designation of areas which support ﬁsh life, quality parameters and monitoring requirements. Baseline quality objectives for pH, ammonia, dissolved oxygen and zinc were established. This Directive was a particularly important consideration in determining discharge consents for sewage works. 1979 (79/923/EEC) Directive on the quality required of shellﬁsh waters. – Established quality standards for the protection of shellﬁsh populations rather than the health of consumers. 1980 (80/68/EEC) Directive on the protection of groundwater against pollution caused by certain dangerous substances. – This Directive aimed to eliminate or reduce hazardous substances in groundwater with the emphasis on imposing control measures on the discharge rather than on setting a standard that the receiving water needed to achieve. 1991 (91/676/EEC) Directive on the protection of waters against pollution caused by nitrates from agricultural sources. – Aimed at reducing the contribution of agricultural activities, especially non-point or diffuse sources, to nitrate pollution. 1991 (91/492/EEC) Directive laying down the health conditions for the production and placing on the market of live bivalve molluscs. – Speciﬁed bacterial quality standards for classiﬁcation of shellﬁsh beds. The bacterial quality of shellﬁsh may be affected by sewage discharges and the locations of outfalls and CSO discharges were needed to be taken into consideration with respect to classiﬁed shellﬁsh beds. 1991 (91/271/EEC) Directive on urban waste water treatment. – Addresses the collection, treatment and discharge of urban waste water (sewage) and the treatment and discharge of wastes from certain industrial activities. Established different levels of sewage treatment necessary (primary, secondary or tertiary) based upon the sensitivity of the receiving waters to pollution.

industry (10%), mining and quarrying (12%), ﬁsh farming (6%) and agriculture (6%) were the other major industrial sectors contributing to the identiﬁed risk (SEPA 2004). For standing water bodies, or lakes, 18% of a total of 334 were assessed as being at risk from point source pollution and again a dominant contributing industrial sector was sewage and refuse disposal (48%). Only one

other sector contributed to the risk, however; point sources associated with the operation of ﬁsh hatcheries and ﬁsh farms (51%). Of the 26% of groundwater bodies deemed to be at risk from point source pollution, sewage and refuse disposal was again the dominant contributor (45%), with mining and quarrying (30%) and manufacturing (15%) the other signiﬁcant contributors.

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alan jenkins and robert c. ferrier 6.3

Current Solutions

In essence, the historical framework for controlling point source efﬂuent discharges to rivers and groundwater still underpins existing legislation. The EU Water Framework Directive (WFD) sets out a classiﬁcation based on ecological status with the aim for all water bodies to achieve ‘good’ ecological status by 2015. Ecological status reﬂects both biological and chemical status and the hydromorphological characteristics of the water body itself. Clearly, the water quality required to achieve the classiﬁcation of ‘good’ chemical status provides a river water quality objective which must be met. Good chemical status demands that certain general water quality parameters such as nutrient concentration, pH, etc. are within deﬁned limits but also that a number of speciﬁed Dangerous Substances are absent. In the case of the general water quality characteristics, limit levels for good status are set in relation to ecological targets rather than on ecotoxicology. For example, the boundary between moderate and good status for phosphorus concentration is set at 100 µg L−1 to protect against eutrophication. Within this new legislation the mechanism of management of efﬂuents

Box 6.3

remains as discharge consents and compliance is assessed through downstream monitoring. Historically, then, there has been a dichotomy in approach to pollution control in Europe. There have been controls based on what can be achieved at source taking account of best available technological solutions and others dealing with the speciﬁc requirements of the receiving waters. Each approach has potential ﬂaws. Source controls alone can allow a cumulative pollution load which is severely toxic to the environment to accumulate over time or in locations where there is a concentration of pollutant sources. Quality standards to protect speciﬁc uses, on the other hand, tend to underestimate the effect of a substance on the ecosystem due to limitations in scientiﬁc knowledge regarding dose–response relationships and the mechanisms of transport, fate and behaviour within the environment. The WFD provides a combined approach and under Article 16 lays down the new procedures for identiﬁcation of polluting substances and development of control measures. On the source side, it requires that as part of a programme of basic measures at catchment scale, all existing technology-driven source-based controls must be implemented as a ﬁrst step. In

Calculation of a discharge consent

The most common approach to calculating numerical consent conditions is the method of combining distributions (CD) based on the concept of mass balance: T = FC + fc F+f where: T = concentration of substance downstream of the discharge; F = river ﬂow upstream of the discharge; C = concentration of the substance in the river upstream of the discharge; f = ﬂow of the discharge; c = concentration of the substance in the discharge. A single application of this equation cannot be used to calculate the consent needed to meet river quality targets since the variables in the equation represent values at a particular instant in time. The 95 percentile of T, therefore, is derived from the distributions for F, f, C and c either mathematically or by simulation. Computer models are also used in the determination of consents, particularly for discharges to estuaries for which CD methods are not appropriate. Computer modelling at river catchment scale is also used to look at the cumulative impact of efﬂuent discharge downstream (Boorman 2007).

Efﬂuent Management addition to this, a framework for further developing such controls is established. This framework comprises the development of a list of priority substances for action at EU level, prioritized on the basis of risk and then the design of the most cost-effective set of measures to achieve load reduction of those substances, taking into account both product and process sources. In current urban wastewater systems it is usual to consider the operation of the sewer system, the WWTP and the receiving water as separate systems. The transfer across the interfaces between the systems is generally governed by static rules. For example, the ﬂow to the WWTP under wet weather conditions is limited to a value on the order of twice the peak dryweather ﬂow and combined water is discharged to the receiving water or to a retention tank when the upstream ﬂow rate crosses some threshold level (Rauch et al. 2005). The interface to the receiving water is hardly considered at all and the design of the system depends on the characteristics of the urban catchment rather than on the capacity of the receiving water. Operation of the system is, therefore, along a rigorous emissionbased approach. The goal for managing urban drainage systems and, therefore, CSO discharges, is system integration. In an integrated system, information from the entire system based either on measurements or models is used to take control actions. For example, ﬂow in the sewer system and processes in the WWTP would be controlled according to model-based predictions of the receiving water quality and how it changes through time. Around the world, there are examples of integrative approaches in some aspects but few including the entire system. In many countries it is common to consider storm water inﬁltration as the main option for dealing with rainwater. This serves to recharge the groundwater and helps to reduce CSO spill frequencies and volumes and to increase the fraction of sewage that is treated at the WWTP rather than being discharged to the river via the CSO. In the UK, a water quality based procedure is used for stormwater management whereby the

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pollution impact to the receiving water is considered using models and duration/frequency curves to help deﬁne the design of the CSO and retention tanks. Any interaction with the processes in the WWTP is not included. In the Netherlands and Flanders, combined systems are common whereby the ﬁrst option during rainfall events is to increase the loading of the WWTP, but this again is driven by static rules. Switzerland plans its wastewater systems with explicit inclusion of the morphology of the receiving water and the ecosystem it supports giving the option of improving the drainage system through improvements in the river itself. In the USA, regulation of wastewater discharges from all point sources is water quality based to comply with Total Maximum Daily Load (TMDL) discharge limits. The exception to this is stormwater from which pollutants must be removed ‘to the maximum extent practicable’. For those substances, either as individual or groups of pollutants, that are estimated to present a signiﬁcant risk to the aquatic environment, the WFD requires a progressive reduction in concentrations. For priority hazardous substances, the phasing out and ultimately cessation of discharges, emissions and ﬂuxes is required. In doing this, the appropriate cost-effective and proportionate level and combination of product and process controls for both point and diffuse sources need to be identiﬁed. In addition, for point sources, environmental quality standards (EQSs) for all waters affected by discharges are required to be met given due consideration of the technical options available for reducing the substances in discharges. A list of 33 priority substances and their environmental quality standards are speciﬁed in the directive ‘on environmental quality standards in the ﬁeld of water policy and amending Directive 2000/60/EC’ (Table 6.2). The setting of EQSs is an attempt to ensure a high level of protection against risks to, or via, the aquatic environment. The EQSs stated in the new Directive were based on a maximum allowable concentration to avoid serious irreversible consequences for ecosystems due to acute exposure in the short term and the annual average

142 Table 6.2

alan jenkins and robert c. ferrier Environmental quality standards for priority substances in inland surface waters Name

AA-EQSa

MAC-EQSb

Alachlor Anthracene Atrazine Benzene Cadmium and its compounds C10-13 Chloroalkanes Chlorfenvinphos Chlorpyrifos 1,2-Dichloroethane Dichloromethane Di(2-ethylhexyl)phthalate (DEHP) Diuron Endosulfan Fluoranthene Hexachlorobenzene Hexachlorobutadiene Hexachlorocyclohexane Isoproturon Lead and its compounds Mercury and its compounds Naphthalene Nickel and its compounds Nonylphenol Octylphenol Pentabromodiphenylether Pentachlorobenzene Pentachlorophenol Polyaromatic hydrocarbons (PAH) Simazine Tributyltin compounds Trichlorobenzenes Trichloromethane Triﬂuralin

0.3 0.1 0.6 10.0 0.08 0.4 0.1 0.03 10.0 20.0 1.3 0.2 0.005 0.1 0.01 0.1 0.02 0.3 7.2 0.05 2.4 20 0.3 0.1 0.0005 0.007 0.4 0.05 1.0 0.0002 0.4 2.5 0.03

0.7 0.4 2.0 50.0 0.45 1.4 0.3 0.1 not applicable not applicable not applicable 1.8 0.01 1.0 0.05 0.6 0.04 1.0 not applicable 0.07 not applicable not applicable 2.0 not applicable not applicable not applicable 1.0 0.1 4.0 0.0015 not applicable not applicable not applicable

‘Hazardous’c X

X X

X X X X

X

X X X X X

EQS expressed as an annual average in µg/L. EQS expressed as a maximum allowable concentration in µg/L. Where this is ‘not applicable’ the annual average values are also protective against short-term pollution peaks since they are signiﬁcantly lower than the values derived on the basis of acute toxicity. c Indicates those substances identiﬁed as priority hazardous substances. a

b

EQS is designed to avoid irreversible consequences in the long term. It is to be noted, however, that acute exposure can also have longterm consequences. Priority substances have been identiﬁed through a process of prioritization of risk involving aquatic and human toxicity via aquatic exposure routes. This process is guided by scientiﬁc

principles and taking particular account of evidence regarding the intrinsic hazard of the substance concerned, evidence from monitoring of widespread contamination and other factors which may indicate the possibility of widespread environmental contamination, such as production or use volume of the substance concerned and use patterns.

Efﬂuent Management Developments in chemical and biological engineering have paved the way for improved ‘end of pipe’ reduction technologies. This has enabled the implementation of ‘Maximum Feasible Reductions’ (MFR) or Maximum Technically Feasible Reductions (MTFR) whereby pollutant sources are required to utilize the best available technology to reduce pollutant discharges to the lowest achievable level. This concept has signiﬁcant cost implications and the economics of the polluting activity may not support such a pollution control methodology. In this case the concept of Best Available Technology Not Incurring Excessive Cost (BATNIEC) has been adopted as the economic compromise in some countries. In the UK, the majority of sewage is treated either by biological ﬁlter or by activated sludge. In fact, activated sludge is the major sewage treatment process used for large population centres in much of the developed world. This is an intensive biological treatment by which bacteria are suspended in a tank and vigorously aerated. Hydraulic retention times of the process are conventionally of the order 5–20 hours (Cooper and Downing 1998). As a result of the process, liquid raw sewage is converted into a solid which is separated from the efﬂuent through sedimentation. This is achieved using a large tank, a seed of bacteria and an air supply to promote the biodegradation process. This process removes 90–95% of all the organic material in the liquid sewage. The widespread use of activated sludge treatment has improved many grossly polluted rivers. Biological (also termed trickling or percolating) ﬁlter plants comprise a tank with a bioﬁlm supported on coarse media upon which the sewage liquor is sprayed following primary sedimentation (Boon et al. 1997). Water contact time with the bioﬁlm is relatively short, around 1 hour, and to improve efﬂuent quality some form of tertiary treatment is used. These usually involves additional ﬁltration or an additional biological treatment step. Sewage treatment, however, was never designed to remove all of the organic compounds in the sewage and a proportion of these enter

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rivers in the ﬁnal efﬂuent. This presents a problem for the river environment in that several of these micro-organic compounds are increasingly recognized as having a signiﬁcant impact on the physiology of ﬁsh, even at very low concentrations. The presence of oestrogenic chemicals in surface waters has been increasingly linked to endocrine disruption in ﬁsh. The main oestrogenic chemicals present in sewage treatment efﬂuent have been identiﬁed as the natural oestrogens oestrone (E1) and oestradiol (E2) and the synthetic steroid oestrogen ethinyl oestradiol (EE2). In addition there are signiﬁcant contributions to the total oestrogenicity of rivers from particular industries, notably the textile industry, from alkylphenolic chemicals such as nonylphenol (NP) and its ethoxylates. Based on current knowledge it is most likely that steroid oestrogens and alkylphenolic chemicals are the main cause of the endocrine disruption caused in ﬁsh in most situations. It is increasingly clear that endocrine disruption in ﬁsh is occurring throughout the world (Jobling and Tyler 2003) and that regulatory authorities would like to reduce or eliminate this problem. A review study of the steroid oestrogen removal efﬁciency from different sewage treatment processes has indicated the beneﬁts of employing an advanced or extended biological tertiary treatment process to either activated sludge or biological ﬁlter processes (Johnson et al. 2007). In a study based on UK sewage treatment plants, the activated sludge process was found to remove around 65% (standard deviation = 50) of E2 and 72% (standard deviation = 48) of E1. The biological ﬁlter plants removed 70% (standard deviation = 36) of E2 but only 30% (standard deviation = 31) of E1. Incorporation of a biological tertiary treatment process, however, showed removal rates of 89% (standard deviation = 6) for E2 and 74% (standard deviation = 29) for E1. The data available for this study were rather limited, however, and robust conclusions regarding the comparative efﬁcacy of different sewage treatment processes could not be drawn.

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Box 6.4

alan jenkins and robert c. ferrier

Efﬂuents and endocrine disruption in ﬁsh

Contamination of freshwaters by chemicals is an ever present issue for managers concerned with protecting the environment and for those concerned with the supply of potable water to the public. Over time the chemicals of concern have changed as water treatment and efﬂuent disposal techniques have improved to solve various issues associated with those chemicals. In the UK there is currently an increasingly persuasive body of evidence that sexual disruption of wild ﬁsh is resulting from sewage efﬂuent and the biologically active, endocrine-disrupting, substances contained within them. The most potent of these substances are believed to be the natural (oestrone and oestradiol) and artiﬁcial (ethinyl oestrodiol) steroid oestrogens of anthropogenic origin, although other nonsteroidal substances have also been implicated such as alkylphenols (e.g. nonylphenol and its ethoxylates). Steroid oestrogens are excreted in micrograms per day amounts by humans and sewage treatment plants are not able to remove all of the load from large population centres. Non-steroidal substances are associated with particular industrial sectors and efﬂuents from these contribute directly to the total oestrogenicity of a river. Many studies have shown that exposure to oestrogens and their mimics has caused the synthesis and secretion of vitellogenin, a female speciﬁc protein, in male ﬁsh. Vitellogenin is now a widely accepted biomarker of exposure to oestrogenic substances, and yet the signiﬁcance of elevated vitellogenin in the blood of male ﬁsh is unclear. In vertebrates, oestrogens play an important role in many reproductive and developmental processes including sexual maturation and differentiation. Exposure to oestrogens, or oestrogen mimics, during sexual differentiation has been shown to induce sexual reversal and/or intersexuality, while exposure during sexual maturation can inhibit gonadal growth and development. The possibility exists therefore that these effects may occur in wild populations of ﬁsh that are exposed to industrial and municipal efﬂuents entering rivers (Jobling et al. 1998).

6.4 New Scientiﬁc Insights At the scale of a large catchment the water quality is determined by the mix of pollutants from point source discharges and diffuse or nonpoint, largely agricultural sources. Determining the relative contribution of these sources provides unique information for catchment management not simply through identifying individual polluters with a view to prosecution under water legislation but also to enable consideration of the effort and resource required locally to achieve the best improvement per unit of resources invested. There exists a clear need, therefore, to develop techniques to identify and quantify pollutant sources at the catchment scale. In European rivers, phosphorus is generally considered to be the limiting nutrient that controls eutrophication. The EU Water Framework Directive requires widespread control of phos-

phorus inputs to rivers to sustain or improve the ecological status. The major sources of phosphorus to rivers are point sources of sewage and industrial efﬂuent and diffuse agricultural sources. The relative contribution from each of these sources is traditionally determined from annual ﬂux budgets and these ﬂuxes are in turn dominated by diffuse agricultural sources during storm events. As a consequence, around 50% of the annual phosphorus load to UK rivers is estimated to be from agricultural runoff. In rivers, however, the annual load of phosphorus has less ecological relevance since the time of greatest ecological sensitivity to high phosphorus availability is during spring and summer low ﬂow conditions when water residence times are high, light levels are high and water temperatures are high all acting together to promote rapid algal growth. The annual ﬂux approach provides only a partial management solution in this case. Some

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Fig. 6.2 Source apportionment (point and diffuse). Soluble reactive phosphorus (SRP) plotted against boron (B) concentrations for water samples collected from 54 lowland river monitoring sites within the Tweed, Wear, Humber, Great Ouse, Thames, Wye and Hampshire Avon catchments in the UK. The strong positive relationship and the fact that most points fall below the line of the SRP-B gradient for sewage efﬂuent with no tertiary treatment, indicates the importance of point source contributions. (After Jarvie et al. 2006.)

tracer of sewage efﬂuent is necessary to provide catchment scale information on the relative ‘instantaneous’ contributions of sewage efﬂuent and agricultural sources. In this respect, sewage inputs of phosphorus have been characterized by using a relatively unreactive chemical tracer, boron, which is derived predominantly from detergents discharged in sewage efﬂuent (Jarvie et al. 2006). Results from water samples collected weekly over several years at 54 monitoring stations around the UK demonstrated a strong positive correlation between boron and phosphorus (total and soluble reactive forms) concentrations (Fig. 6.2). The conclusion drawn, therefore, is that the inputs controlling phosphorus concentrations in these UK rivers are closely linked to the sources of boron, which are predominantly sewage efﬂuent derived. Measures to reduce phosphorus inputs to UK rivers currently focus on reducing efﬂuent concentrations from large sewage treatment works (serving up to 10,000 population equivalents) by tertiary treatment (phosphorus stripping) and by minimizing diffuse phosphorus losses in agricultural runoff. New measures are proposed to manage phosphorus inputs more efﬁciently and to control erosion and sediment losses

from agricultural land under a government sponsored initiative on catchment sensitive farming (Defra 2004). Some of these control measures, however, have the potential to cause signiﬁcant disruption to agricultural practices and rural economies. Since the research detailed above implicates sewage efﬂuent as the primary ecological risk factor for phosphorus, targeting point sources would seem to offer the most signiﬁcant ecological beneﬁts. Installing phosphorus stripping at sewage treatment works serving smaller settlements, which discharge into ecologically sensitive rural tributary streams, needs consideration. In summary, research in the UK to date has demonstrated that the important starting point for reducing phosphorus concentrations in rivers is to ensure robust assessment of point and diffuse contributions within individual catchments. The evidence indicates that better control over point source inputs, particularly smaller efﬂuents discharging into ecologically sensitive streams is required. 6.4.1 Total Maximum Daily Loads (TMDLs) Under the Water Framework Directive (WFD), EU countries must aim to achieve and maintain

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good ecological water quality in rivers, lakes and coastal waters by the year 2015. This requires sources of pollution to be addressed at the catchment (watershed or river basin) scale. In recent years, considerable progress has been made in reducing pollution from point sources. The greatest risk of failing to meet WFD requirements is now perceived to come from diffuse sources. Ecological water quality targets are now being set under the WFD, but there is no currently accepted methodology within the UK that enables maximum pollutant loads to be derived from these targets. In the USA, this requirement has been met by the ‘Total Maximum Daily Loads’ (TMDL) approach, which has been developed for catchment management purposes under the US Clean Water Act (CWA). A single TMDL can encompass several pollutants and multiple water bodies. Once established, responsibility for reducing pollution from point and diffuse sources is assigned. Overall, that responsibility lies with everyone who lives and works within the catchment. The TMDL approach is very similar to other catchment management planning process that has already been implemented and comprises the following steps: 1 List waters that are impaired or threatened by pollutants. 2 Identify key factors and background information that describe the nature of the ‘impairment’, for each listed water body. 3 Identify numeric or measurable indicators and target values that can be used to evaluate attainment of water quality standards. 4 Identify sources of pollutant loading to the water body and characterize by type, magnitude and location. 5 Deﬁne the cause-and-effect linkages between selected water quality indicator(s) or target(s) and the identiﬁed pollutant sources; use these linkages to estimate total loading capacity. 6 Determine pollutant loadings that will not exceed the loading capacity of the waterbody and will lead to attainment of water quality standards; allocate these loadings among the signiﬁcant sources of the pollutant concerned.

7 Develop a monitoring and evaluation plan to determine whether the TMDL has resulted in attainment of water quality standards. 8 Develop an implementation plan. The approach is based on the amount of a pollutant, or pollutants, that a particular water body can receive and still meet water quality standards. It comprises the sum of the allowable loads of each pollutant from all contributing point and non-point sources, and must include ‘a margin of safety’. It must also take seasonal variation in water quality into account, if appropriate. Originally, TMDLs were developed for the management of point sources of pollution. Since then, they have been adapted to include pollution from diffuse sources, which is more difﬁcult to monitor and control on a daily basis. So, although the word ‘daily’ remains in the name for historical reasons, this is misleading. In reality, TMDLs usually address much longer time steps such as monthly, seasonal or annual. Establishing a TMDL for each pollutant is only part of the TMDL process. It sits within a much wider, catchment-based management framework that encompasses and formally links many of the stages that already form part of the UK’s strategy for implementing the WFD. These include risk assessment and target setting. However, it also provides a means of complying with some requirements of the WFD that have not been addressed so far, such as source apportionment, source targeting and the assessment of mitigation strategies at the catchment scale. 6.4.2 Reduction of gross pollution from urban areas and CSOs The development of catchment scale models which explicitly incorporate CSO outputs represents a key tool for the managers and decision makers. On the River Seine in Paris, the PROSE model indicated that 32% of the benthic oxygen demand for a typical summer month was due to the contribution of CSOs (Even et al. 2007). The use of models to obtain and conﬁrm a good understanding of the response of the system to urban discharges is clearly a critical step in the

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design of improved sewer systems (e.g. Irvine et al. 2005). Another modelling system developed on a modular basis within the EU is CITY DRAIN (Achleitner et al. 2007). This freeware is developed using a blockwise approach to modelling different parts of the urban drainage system (catchment, sewer system, storages, receiving water, etc.). Each block represents a system element with different underlying modelling approaches for hydraulics and mass transport. The different elements can be arranged and connected to describe an integrated urban drainage system. One of the key problems in the construction, calibration and testing of appropriate spatial and temporal scale models is the availability of data from sewers and CSOs. This situation is changing with the development of instruments and techniques capable of real-time monitoring of both concentrations and loads (e.g. Gruber et al. 2005). The results of over 3000 studies completed between 1968 and 2001, including dry weather ﬂow/wastewater, storm water runoff, combined sewer ﬂow and CSO, have now been compiled into a single database by Brombach and Fuchs (2003). Another novel management option for reducing the impact of CSO is constructed artiﬁcial wetlands (Uhl and Dittmer 2005). These have an advantage over more conventional storage tanks in that they enable the removal (at least partial) of soluble substances as well as ﬁne particles. The need for a fully integrated approach to wastewater system management, design and planning has been clearly summarized by Rauch et al. (2005).

given the low concentrations. An alternative approach has been developed whereby the concentrations of chemicals in rivers can be modelled to provide a prediction of whole lifetime exposure of aquatic organisms. The most promising modelling approach with respect to catchment management is a development of the GREAT-ER model (Geo-referenced Regional Exposure Assessment Tool for European Rivers) to predict concentrations of the target chemicals E1, E2, EE2 and NP. This model can calculate the distribution of chemicals in surface waters for both individual river reaches and entire catchments (Schowanck et al. 2001). The model requires as inputs an estimate of the discharge concentrations of the chemicals from sewage treatment plants, ﬂow data and degradation coefﬁcients (based on an assumption of ﬁrst-order decay kinetics). The model has been tested for boron and a surfactant, LAS (linear alkylbenzene sulfonate), and has demonstrated accurate simulation of concentrations in the river system provided accurate hydrological and chemical fate models are used. The output from the model identiﬁes locations (‘hot-spots’) where predicted environmental concentration (PEC) values are of concern and where adverse biological effects might be expected. The methodology has been developed to not only produce catchment scale ‘effect maps’ for each chemical but has combined the effect of each individual chemical into a single ‘mixture’ effect. This is based on the assumption that the chemicals act dose additively and that the response can be accurately predicted by the model of concentration addition (Brian et al. 2005).

6.4.3 Assessment of chemical risk

6.4.4 Determination of ecological risk

There is an increasing literature documenting a variety of analytical approaches to directly measure the concentrations of steroid oestrogens and alkylphenolic chemicals, especially nonylphenol, in both sewage treatment efﬂuent and river water (Sumpter et al. 2006). Measuring actual chemical concentrations in surface waters, however, is both time consuming and expensive

Chemicals that are released into the environment can have direct consequences in relation to exposure (acute) and bio-accumulation (chronic) effects for both humans and other biota. Dose–response relationships are fundamental and essential concepts in toxicology, correlating exposures and the spectrum of induced effects. Much early toxicity testing on aquatic organisms

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focused on the development of exposure concentration (EC50) thresholds for speciﬁc pollutants for a number of indicator aquatic species, but the ability to determine more chronic exposure and ecological effects was more elusive. The drive to include a more ecological perspective within toxicology involved greater assessment of impacts on species and their associated communities, and multispecies tests become more commonplace. Community level responses representing ecosystem ‘health’ identiﬁed the importance of ecological indicators of system function and consolidated ecotoxicological development. Recent emergent technologies however, such as genomics, proteomics and metabolomics, have provided additional tools in determining sub-lethal effects on aquatic biota, at molecular, cell and organism levels. This development has opened up new opportunities in evaluating the interaction of potentially hazardous substances on systems level functioning in aquatic ecosystems. Such ecosystem level impacts analyses are promoted in legislation such as the WFD, and represent an evolution of traditional toxicological approaches into the science of stress ecology (Van Stratten 2003) and the development of novel biomarkers.

6.5

Implications for Catchment Management

In relation to catchment management, recent research has clear implications and helps provide the tools with which managers will be able to assess the options available for maintaining water resource quantity and quality and for remediation of polluted waters or degraded water sources. The ability to identify upstream pollutants and attribute them to particular land uses or point source efﬂuent discharges provides key management information for the assessment of the most cost-effective control strategy to be implemented. In some catchments this will lead to a reassessment of the existing management of efﬂuent sources in light of improved knowledge of sources and impacts. One clear beneﬁt of the increased ability to attribute pollution to sectors

of industry or sources is the improvement offered to the polluter pays principle. Having demonstrated the scientiﬁc methodology for producing effects maps for both single chemicals and mixtures, these tools now provide unique information for catchment managers charged with reducing or eliminating the problem of endocrine disruption in ﬁsh. The catchment scale maps can be used to target and prioritize the installation of enhanced treatment technologies into those sewage treatment plants predicted to cause the most severe effects. This would enable the most rapid improvement in the quality of the aquatic environment. The methodology also enables the ranking of chemicals responsible for a particular effect. Modern approaches to regulation of water quality are now being proposed by the Environment Agency in England and Wales which are based on preventing and/or minimizing environmental impacts and achieving high standards of environmental management. Direct regulation in the form that is currently used to control abstractions from the aquatic environment and efﬂuent inputs to the aquatic environment will continue to play an important role and underpin catchment management. But, most importantly, risk-based approaches will be further developed. Overall, the future regulatory regime will aspire to be transparent (clear rules and processes), accountable (with regard to catchment management), consistent (same approaches used between sectors and over time), proportionate (risk-based with respect to the scale of outcomes to be achieved and the risks involved) and targeted (focused on optimizing environmental outcomes). In the USA, the Environmental Protection Agency supports the establishment of water quality trading following a catchment scale evaluation of TMDLs, and source apportionment. This trading is a voluntary exchange of pollutant reduction credits, and a facility with a higher pollutant control cost can buy a pollutant reduction credit from a facility with a lower control cost thus reducing their cost of compliance. The EPA’s 2003 National Water Quality Trading

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Efﬂuent Management Policy as part of the Clean Water Act proposes basic rules for trading (what pollutants and how to set baselines, etc.) and under what conditions trading can occur. A key component is that the Trading Policy states trading must occur within the same watershed. The EPA supports trading of nutrients and sediment load as well as cross-pollutant trading of oxygen-demanding pollutants. It may consider supporting trades of other pollutants but believes that these trades require a higher level of scrutiny, but it does not support trading of persistent bioaccumulative toxics (PBTs) except on a pilot basis. A permitted facility or a point source can trade with another point source or with a nonpoint source. The partners can trade directly, or through a third party. All stakeholders interested in their local watershed, including conservation organizations and watershed groups, should be involved in the development of the trading programme. Credits can be generated by a point source over-controlling its discharge or by a nonpoint source installing best management practices (BMPs) beyond its baseline When possible, edge-of-ﬁeld or ambient monitoring should be conducted to gauge the water quality impacts of BMPs. Modelling is another option; however, it should be combined with some monitoring to verify the model. Field testing of BMPs, the use of conservative assumptions for BMP efﬁcacy, and uncertainty ratios can also reduce uncertainty. Recent scientiﬁc developments have paved the way for the further development of the riskbased approach to efﬂuent discharge control proposed above. To this end, the EA in England and Wales is investigating the possibility of introducing a national risk-based Operator and Pollution Risk Appraisal (OPRA) system to the water quality regulatory regime. The proposal is for a method of screening risk which ﬁts within the overall framework for assessing and managing environmental risk and provides an objective and consistent risk assessment of the discharges that require regulation. An important aspect of the OPRA scheme is that it provides a single measure of overall risk posed both by the activities on the

Box 6.5 The model for ‘modern regulation’

Define outcomes

Choice of instruments

Evaluate and inform

Compliance and enforcement

The EA modern regulatory model comprises four segments: 1 Deﬁne outcomes – based on legislative, policy and environmental requirements and coupled with principles of best practice, speciﬁc sector and geographical characteristics, and scientiﬁc understanding. 2 Choice of instruments – deﬁned by the nature of the environmental impacts and the risks and sometimes using instruments in combination. 3 Compliance and enforcement – concentrate resources where risks to the environment are highest and also take account of operators’ performance, ensuring that operators are using the most appropriate monitoring systems and techniques and using enforcement and prosecution where necessary. 4 Evaluate and inform – assessing how well the required environmental outcomes are being delivered and how well operators are managing.

site and the management by the site operator. Crucially, it must be stressed that OPRA will not change the need for operators to comply with the requirements of any existing or new consents which are put in place to protect the environment. Furthermore, it does not imply modiﬁcation of existing consents. The OPRA scheme has been used in England and Wales since 2003 under the Pollution Prevention and Control Regulations. Extending

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the use of the scheme to discharge consents under the Water Resources Act (1991) will help in: targeting the resources of the regulators (in this case the EA) at activities and discharges that pose the greatest risk to the environment; provide a clear and fair regulatory process for all involved; help determine the appropriate type of compliance monitoring and assessment; encourage operators to improve performance beyond marginal compliance, and maintain the ‘polluter pays’ principle. Perhaps most importantly, with respect to catchment management, the scheme allows full use of the increased information available at catchment scale regarding the current status of water quality and the options for improvement and remediation. 6.6

Future Research Requirement

6.6.1 Climate change impacts Future climate change presents some signiﬁcant challenges for the management of efﬂuent discharges within the context of Integrated Water Resources Management. This is likely to be especially signiﬁcant in those regions where river ﬂows are predicted to be reduced either seasonally or annually. A decrease in river ﬂow will lead to a reduction in efﬂuent dilution and, potentially, the need for tighter discharge consent and, therefore, a requirement for improved treatment technology to reduce the pollutant load of the discharge. It is also possible that increased ﬂow seasonality (drier in summer and wetter in winter) will lead to a requirement for seasonal discharge consents; perhaps being tighter during the summer and less stringent during the winter. In addition, future climate change may well lead to changes in land use within the catchment with consequent impacts on diffuse pollutant loadings. This potential change in the ‘background’ water quality will have a further impact on calculated discharge consents. These aspects of managing efﬂuent discharges at the catchment scale will require consideration through the research use of catchment scale water quality models and a range of future rainfall-runoff

scenarios. Such models will need to be processbased, dynamic and spatially distributed to take account of diffuse and point source impacts, groundwater contributions, abstractions for water supply and any other signiﬁcant inﬂuences on catchment water quality and quantity. 6.6.2 Nanoparticles Engineered oxide nanoparticles (SiO2, TiO2 and ZnO) are widely used in cosmetics and personal care products (e.g. sunscreens, toothpastes, skin creams) and in topically applied pharmaceutical formulations (e.g. creams for local and transdermal drug delivery). Engineered oxide nanoparticles in these formulations provide active ingredients (e.g. pigments for UV absorption), improve rheology and transparency of creams and promote dermal absorption. Whereas many other engineered nanoparticles, such as fullerenes, are yet to ﬁnd any widespread commercial applications, oxide nanoparticles in cosmetic and topically applied pharmaceutical formulations are currently routinely being released into wastewaters. There is a much wider potential, therefore, for exposure of aquatic ecosystems to oxide nanoparticles than many other engineered nanoparticles. Although many oxide compounds are widely regarded as possessing a low toxicity, when formulated as nanoparticles they are known to generate free radicals and reactive oxygen species which may disrupt proteins, lipids and DNA. The UK Government’s ﬁrst research report on nanoparticles in the environment and other recent reviews have highlighted the pressing need for research on the environmental fate and behaviour of nanoparticles in topically applied pharmaceutical and consumer products, because of their widespread use and disposal and their physico-chemical properties which are typically tuned to maximize bioavailability. Despite this, the only published studies on the environmental fate of oxide nanoparticles are restricted to movement of TiO2 and SiO2 nanoparticles through porous media (simulating transport through groundwater/soils). However, the major emission route for pharmaceutical and oxide nanoparticles

Efﬂuent Management to aquatic systems is via sewage efﬂuent discharges to surface waters. There is, therefore, a key strategic requirement to assess loadings of oxide nanoparticles to wastewaters, the efﬁciency of removal of oxide nanoparticles in sewage treatment processes and the structure, transport, behaviour and fate of pharmaceutical and cosmetic oxide nanoparticles in surface waters. Previous research suggests that discrete nanoparticles and small aggregates are likely to be more bioavailable/toxic to aquatic organisms than larger aggregates and that factors promoting ﬂocculation and sedimentation of nanoparticles are likely to mitigate ecotoxicity and reduce dispersion of nanoparticles within aquatic systems. 6.6.3

Stress ecology

The strengthening discipline of ecotoxicogenomics is moving current thinking on deﬁning the multivariate normal operating range (NOR) of individual indicator organisms and how they respond to environmental pressures. Deviation from a typical gene expression proﬁle is used to deﬁne this stress. This relies heavily on recognizing that the complexity in an inherent internal system property, and that the state of the environment must be analysed in terms of multiple measurable variables. Bioinformatics supporting the expanding genomics revolution is providing the necessary tools to link ecological response to changing environmental conditions. Linkages between the historical and newer toxicological tools are being developed to assess and predict risk. Being able to classify chemical and other stressors based on the effects that they have on sub-organism level promotes a systems level approach to understanding ecosystem functioning and resilience (Neumann and Galvez 2002). This rapidly developing ﬁeld aims to produce costeffective and comprehensive toxicological testing approaches for evaluating environmental hazards. 6.6.4 Urban storm drainage The dynamics of CSOs remains at a level of understanding which is below requirement for

151

the establishment of integrated management and control systems. There exists a clear need for increased monitoring and evaluation of ﬂuxes from these sources. A further key research requirement lies in the evaluation of highly polluted sediments which are held in the sewer system for long periods and released during extreme events. In addition, their impact on the natural environment is far from clear and both acute and chronic effects require closer investigation, again with a view to optimizing the management options. There is a need to examine the potential for developing more integrated management and control systems which are capable of, for example, increasing the inﬂows to WWTP during wet weather conditions or real-time controls based on pollutant loads to enable discharge of the least polluted combined sewer-storm water rather than sending all to storage and potentially to overﬂow. 6.6.5 Advanced ‘end of pipe’ technologies Many of the micro-organic compounds with the capability for endocrine disruption in ﬁsh, emanating from sewage works, are biodegradable and so it is possible to conceive of a novel ‘end-ofpipe’ treatment process to remove them. There are, however, practical difﬁculties in adapting activated sludge plant to remove them with high efﬁciency. The removal technologies currently under consideration include: treatment with ultraviolet light, contacting with ozone, microﬁltration and reverse osmosis and activated carbon adsorption. All of these techniques are well tested and used in the water supply industry. It is widely held, however, that whilst these technologies offer promise they may in fact be too costly to apply. New research is required to develop novel approaches. The further development of catchment chemical risk maps requires consideration of seasonality and the consequences for biological effects. River ﬂow clearly varies with season and weather conditions. High ﬂows are most likely to be associated with dilution of chemical concentrations and so reduced effects. By contrast, low ﬂows in

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summer will lead to potentially higher concentrations of oestrogenic chemicals and so potentially more pronounced effects, or at least higher risk of effects. Improved assessment of chemical degradation rates would also greatly enhance value of the model applications.

6.6.6 Impacts on historically immobilized material Whilst agriculture provides a major contribution to annual phosphorus ﬂuxes it is clear that point source contributions dominate UK rivers at times of eutrophication risk. To target investment in technology for improved and extended phosphorus stripping, a better understanding of the impacts and time-scales of chemical and ecological response to remediation is needed. Nevertheless, although point sources currently dominate phosphorus concentrations in the UK, it is essential for whole catchment management that mechanisms to control diffuse phosphorus from agriculture are developed and assessed. In this respect, a key future research need is for the establishment of ﬁeld-scale manipulation experiments to determine the nature, scale and timescales of ecological responses to different phosphorus remediation strategies and to provide better estimates of ecologically relevant water quality standards for phosphorus.

References Achleitner, S., Möderl, M. and Rauch, W. (2007) CITY DRAIN © – An open source approach for simulation of integrated urban drainage systems. Environmental Modelling and Software, 22, 1184–1195. Brian, J., Harris, C.A., Scholze, M. et al. (2005) Accurate prediction of the response of freshwater ﬁsh to a mixture of estrogenic chemicals. Environmental Health Perspectives, 113, 721–728. Boon, A.G., Hemfry, J., Boon, K. and Brown, M. (1997) Recent developments in the biological ﬁltration of sewage to produce high-quality nitriﬁed efﬂuents. Journal of the Chartered Institution of Water and Environmental Management, 11, 393–412.

Boorman, D.B. (2007) Towards benchmarking an instream water quality model. Hydrology and Earth System Sciences, 11, 623–633. Brombach, H. and Fuchs, S. (2003) Datenpool gemessener Verschmutzungskonzentrationen in Misch- und Trennkanalizationen. KA – Abwasser Abfall, 4, 441–450. Cooper, P.F. and Downing, A.L. (1998) Milestones in the development of the activated sludge process over the past eighty years. Journal of the Chartered Institution of Water and Environmental Management, 12, 303–313. Defra (2004) Developing measures to promote catchment-sensitive farming. A joint Defra-HM Treasury Consultation. http://www.defra.gov.uk/ environment/water/dwpa/index.htm Even, S., Mouchel, J-M., Servais, P. et al. (2007) Modelling the impacts of Combined sewer overﬂows on the river Seine water quality. Science of the Total Environment, 375, 140–151. Gruber, G., Winkler, S. and Pressl, A. (2005) Continuous monitoring in sewer networks an approach for quantiﬁcation of pollution loads from CSOs into surface water bodies. Water Science and Technology, 52(12), 215–223. Irvine, K.N., Perrelli, M.F., McCorkhill, G. and Caruso, J. (2005) Sampling and modelling approaches to assess water quality impacts of combined sewer overﬂows – the importance of a watershed perspective. Journal of Great Lakes Research, 31(1), 105–115. Jarvie, H.P., Neal, C. and Withers, P.J.A. (2006) Sewageefﬂuent phosphorus: a greater risk to river eutrophication than agricultural phosphorus? Science of the Total Environment, 360, 246–253. Jobling, S. and Tyler, C.R. (2003) Endocrine disruption in wild freshwater ﬁsh. Pure and Applied Chemistry, 75, 2219–2234. Jobling, S., Nolan, M., Tyler, C.R., Brighty, G. and Sumpter, J.P. (1998) Widespread sexual disruption in wild ﬁsh. Environmental Science and Technology, 31, 2498–2506. Johnson, A.C., Williams, R.J., Simpson, P. and Kanda, R. (2007) What difference might sewage treatment performance make to endocrine disruption in rivers? Environmental Pollution, 147, 194–202. Neumann, N.F. and Galvez, F. (2002) DNA microarrays and toxicogenomics: applications for ecotoxicology. Biotechnology Advances, 20, 391–419. Rauch, W., Seggelke, K., Brown, R. and Krebs, P. (2005) Integrated approaches in urban storm drainage: where

Efﬂuent Management do we stand? Environmental Management, 34(4), 396–409. Schowanck, D., Fox, K., Holt, M. et al. (2001) GREATER: a new tool for management and risk assessment of chemicals in river basins. Contribution to GREATER, Chapter 10. Water Science Technology, 43, 179–185. SEPA (2004) Pressures and Impacts on Scotland’s Water Environment, Report and Consultation. Scottish Environment Protection Agency, Stirling. Sumpter, J.P., Johnson, A.C., Williams, R.J., Kortenkamp, A. and Scholze, M. (2006) Modeling effects of mix-

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tures of endocrine disrupting chemicals at the river catchment scale. Environmental Science and Technology, 40, 5478–5489. Uhl, M. and Dittmer, U. (2005) Constructed wetlands for CSO treatment: an overview of practice and research in Germany. Water Science and Technology, 51(9), 23–30. Van Stratten (2003) Ecotoxicology becomes stress ecology. Environmental Science and Technology, 37, 324–330.

Image facing chapter title page: Courtesy of Centre for Ecology and Hydrology.

7

Managing Urban Runoff J. BRYAN ELLIS1

1

Urban Pollution Research Centre, Middlesex University, The Burroughs, Hendon, London, UK

7.1

Introduction

Most urban rivers throughout the world have suffered heavy modiﬁcation and regulation which has eliminated their ‘natural’ ﬂow and morphological status, imposing a characteristic ﬂashy and unsteady ﬂow regime. A typical urban watercourse has a ﬂuctuating ﬂow and quality regime, poor sediment quality and a suppressed biological diversity exacerbated by short nutrient and food retention times. This perturbed ‘equilibrium’ describes what may be called the normal (or natural) ﬂow, quality and ecological balance of a heavily modiﬁed urban receiving water. The physical impacts of increased impermeability, the introduction of ﬂow obstacles (railway/ road piers, culverts, etc.) and channelization generate increased runoff rates as well as increased ﬂow and peak volumes (Fig. 7.1) which pose increased ﬂood risks and downstream erosion potential. The prime objectives of urban management are, therefore, targeted at peak ﬂow volume and water quality control as well as ecological and morphological improvements. It is the optimization and integration of these core regime components which provides the key to sustainable best practice management of urban watercourses.

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

7.1.1 Urban ﬂow regime The transformation of a watershed from a rural to an urban condition produces four major changes in the hydrological characteristics of receiving waterbodies: 1 An increase in ﬂow volume; resulting from increased impermeable surfacing, with increased runoff volume being greatest for frequently occurring small, intense storms. 2 A decrease in lag time; due to a combination of impervious surface runoff and the expansion of the urban drainage net which reduces hydraulic roughness and increases the velocity of overland ﬂow. 3 An increase in peak discharge; this is the combined result of increased runoff volume and decreased lag time. A full urbanized watershed with some 50% impervious cover will increase the peak discharge of a 2-year storm by approximately four times. 4 A reduction in both shallow and deep inﬁltration with losses of up to 20% in terms of percolation to deep groundwater levels. The increased ﬂow volumes caused by the introduction of impermeable urban surfacing increases the probability that runoff of a given magnitude will occur more frequently following urbanization. The recurrence interval (RI) is a statistical expression of the probability of a runoff of a given magnitude occurring with the RI, or return period being the average time between storms of a given

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Box 7.1

j. bryan ellis

Flow and quality sources and pathways

Urban runoff includes separately sewered discharges from impervious surfaces, combined sewer overﬂows (CSOs) as well as ﬂows from illicit connections. The sources contributing to surface ‘dirt and dust’ accumulation come from vehicular wear and trafﬁc emissions, rooﬁng, highway and construction materials, litter and plant/leaf debris, spillages, animal and bird droppings in addition to atmospheric deposition. It is only in the last twenty years that national efforts have been made to individually characterize and quantify these individual sources and to assess their relative contribution to urban runoff ﬂows and water quality (as shown in the ﬂow diagram below). The sinks or outputs receive contributions from unintentional pathways (shaded boxes) whereby ﬂows leave the sewer network via exﬁltration or where elevated groundwater levels act as sources and add ‘clean’ water into the sewer system via pipe inﬁltration. Exﬁltration losses are generally less than 2% of total ﬂow volume due to joint sealing by sediment and bioﬁlm growth in the pipe (Ellis et al. 2004a).

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Fig. 7.1 The hydrological effect of urbanization.

magnitude. Thus a 10-year return period is often described as having a frequency of 1 : 10 years. Following urbanization, the probability of a 1 : 100 ﬂow event, with an exceedance probability

Box 7.2

of 0.01, increases by 1.8 times and a 1 : 5 event, with an exceedance probability of 0.2, increases by 3.0 times the runoff from undeveloped land.

Recurrence interval and exceedance probability

The recurrence interval (RI) is the average period of time within which an event will be equalled or exceeded. The probability, p, (or risk) that the event will equal or exceed a design storm (or other lifetime period) at least once over a period of n years is given by: p = 1 − (1 − 1/T)n, where T is the return period (or RI). The probability of occurrence in any one year is the reciprocal of the recurrence interval. Thus the 10-year storm has a 10% chance (0.1 probability) of occurring in any one year; the 100 year storm has a 1% chance (0.01 probability). The table below shows the risk of at least one exceedance occurring during a design life for given RIs (or exceedance probabilities).

RI (years) 2 5 10 25 50 100

Design life (years)

Exceedance probability

5

10

50

100

0.5 0.2 0.1 0.04 0.02 0.01

0.97 0.67 0.41 0.18 0.1 0.5

1.0 0.89 0.65 0.34 0.18 0.1

1.0 1.0 0.99 0.87 0.64 0.39

1.0 1.0 1.0 0.98 0.87 0.63

For a 1 : 100 year storm event, over a period of 100 years the probability or risk of exceedance is: p = (1 − 1/100)100 = (1 − 0.366) = 0.634. This means that the chance of the 100-year ﬂood being equalled or exceeded at least once in 100 years is 63.4%, i.e there is roughly a 2 in 3 chance that the 100-year event will be equalled or exceeded (or a 1 in 3 chance that it will not).

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As the recurrence probability increases to 0.5, i.e. a 2-year storm, the runoff after urbanization is about 40–60 times more than the undeveloped rate. It is the annual (or 1.5-year RI, ﬂow) and more frequently occurring smaller storms, however, that are the dominant channel-forming events and which generally shape the watercourse as well as being responsible for delivering the majority of the pollutant load to the receiving water. Given the need to capture and convey large ﬂoods up to the 100 RI storm event that results from increased surface runoff, traditional engineering of urban waterways has been to employ straightened concrete-lined channels with safety walls or gabion buttresses and ﬂood storage ponds to attenuate the peak ﬂow volumes. Flows along urban watercourses are ﬂashy and potentially hazardous to local residents, thus longitudinal slopes need to be reduced using grade control drop structures, low (trickle) ﬂow channels, on-line rifﬂe/ pool sequences, sinuous low-ﬂow paths, etc. 7.1.2

Urban runoff pollutants

The range of pollutant concentrations and loadings associated with impermeable stormwater runoff indicates that discharges from surface water outfalls (SWOs) can be highly variable in quality (Table 7.1), with standard deviations frequently being 75% (equivalent to a coefﬁcient of variation, Cv, of 0.75) of the average event mean concentration (EMC) value. EMCs are frequently close to, if not exceeding, the minimum no observable effects limit (NOEL) value (Table 7.1) and thus potentially present a problem to receiving water ecology. The relationship between urban land use activities and pollutant loadings, however, tends to be site speciﬁc and regional extrapolations on a continental scale cannot be readily applied. Further, the impact is varied with organism and some are victim to chronic low level exposure, others to acute higher concentration ﬂushes. Nevertheless, a land use EMC value times runoff volume approach provides a convenient and appropriate screening-level methodology for annual load estimations (Marsalek 1991a; Ellis and Mitchell 2006).

Mean EMCs can be calculated from observations or transferred from existing databases such as that of the US EPA National Best Management Practice (BMP) Database (www.bmpdatabase. org); runoff volume is produced from hydrological modelling. The volume is then multiplied by the mean EMC to obtain the loading and estimate bounds derived from multiplication by the upper and lower conﬁdence limits. Properties of EMCs are of further interest in load and impact estimations. The US Nationwide Urban Runoff Program (NURP) data (United States Environmental Protection Agency 1983) indicates that geographical location is of little use in explaining site-to-site variability, or in predicting data for unmonitored sites. Alternative approaches were, therefore, necessary to allow comparison of data between sites. Under such circumstances, bestﬁt EMCs were found to be statistically independent of runoff event volumes, which implies that loads can be derived by multiplication of the mean EMC by runoff volume, and furthermore, when sampling runoff events, randomly selected events of any magnitude are acceptable. Finally, analysis of EMC data in the NURP program and other European studies, have indicated that stormwater pollutant concentrations as well as their EMCs are typically log-normally distributed. This distribution enables the prediction and comparison of mean, standard deviation and quantile values for site pollutant EMCs and is an appropriate basis for the probabilistic modelling of urban runoff quality (Adams and Papa 2000). Many urban water quality planning and design tasks require spatially and temporally distributed data on stormwater and combined sewer ﬂows. Such tasks require the use of urban simulation models which have greatly evolved during the past 20 years. There are many such models currently in use, but a number of software packages have become well known and stand out in urban drainage modelling practice. These include InfoWorks CS of UK Wallingford (www.wallingfordsoftware.com), the Danish Hydraulic Institute software MOUSE (www.dhi.com), the US EPA Storm Water Management Model SWMM (www. epa.gov/ceampubl) and the Australian eWater Ltd

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Table 7.1 Pollutant concentrations and loadings for urban stormwater runoff (compiled from: United States Environmental Protection Agency 1983, 2004; Deutsch and Hemain 1984; Marsalek 1991b; House et al. 1993; D’Arcy et al. 2000; Mitchell 2001; Maestre and Pitt 2007) Event mean concentration and range (mg L−1) Pollutant parameter Total suspended solids BOD COD NH4-N Total nitrogen Total phosphorus Total lead Total zinc Total hydrocarbons Polyaromatic hydrocarbons Faecal coliforms (E. coli)

Load per unit area (kg imp.ha−1 yr−1)

Residential and and commercial

Motorways and trunk roads

Residential and and commercial

Motorways and trunk roads

Minimum concentration causing observable biological effects

190 (1–4,582) 11 (0.7–220) 85 (20–365) 1.45 (0.2–4.6) 3.2 (0.4–20.0) 0.34 (0.02–14.3) 0.21 (0.01–3.1) 0.30 (0.01–3.68) 1.9 (0.04–25.9) 0.01

261 (110–5700) 24 (12.2–32.0) (128–171)

487 (347–2340) 59 (35–172) 358 (22–703) 1.76 (1.2–25.1) 9.9 (0.9–24.2) 1.8 (0.5–4.9) 0.83 (0.01–1.91) 1.15 (0.21–2.67) 1.8 (0.01–43.3) 0.002

(815–6289)

25 mg/l

(90–172)

N/A

(181–3865)

N/A

(0.8–6.1)

1.7 µg L−1

6,430 (40–500,000) MPN 100 ml−1

10–103 MPN 100 ml−1

(0.02–2.1)

0.96 (2.41–34.0) 0.41 (0.17–3.55) 28 (2.5–400) (0.03–6.0)

N/A N/A (1.1–13.0)

12.26 µg L−1 30 µg L−1

140

2.1 (0.9–3.8) × 109 counts ha−1

N/A

BOD, biochemical oxygen demand; COD, chemical oxygen demand; MPN, mean probable number; N/A, not applicable.

MUSIC model (www.toolkit.net.au) as well as STORM and MIKE URBAN CS. These models are modularly structured and, in general, calculate surface runoff and wastewater ﬂows and their quality, route the ﬂows and water quality constituents through transport, storage, management and treatment facilities and simulate efﬂuent fate in the receiving waterbody. In general, these readily available models provide for the basic needs of urban drainage modellers and facilitate easy input data import from GIS and other databases (Zoppou 2001). The range of EMC values associated with global combined sewer overﬂow (CSO) discharges varies in geographical and climatic situation (Table 7.2). The average EMC values shown (in italic) for the UK are the formal values used in

the national sewer design and rehabilitation guidance manual (Water Research Centre 2001). This SRM Manual also provides factors to convert dry weather ﬂow (DWF) concentrations to give estimated average concentrations for wet weather storm ﬂow. The pollutants in CSOs come from domestic sources, especially biochemical oxygen demand (BOD), total suspended solids (TSS) and nutrients, as well as trade efﬂuents, especially fats, grease, metals and synthetic organic compounds and additionally from surface water runoff during wet weather events. Concentrations can vary substantially on a diurnal basis, both within and between stormﬂow events as well as from community to community. In addition to faecal and pathogenic bacteria, sewage also contains enteric

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Table 7.2 Combined sewer overﬂow pollutant concentration (compiled from: Ellis 1986; Water Research Council 1991; Arnberg-Neilsen et al. 2000; EPA 2004) TSS (mg L−1) US Canada UK Europe

237–635 190 425 176–647 105–721

BOD (mg L−1) 43–95 90 43–225 39.9–200

COD (mg L−1)

Cd (µg L−1)

Cu (µg L−1)

120–560

Zn (µg L−1)

Ptotal (mg L−1)

Ntotal (mg L−1)

E. coli (100 ml)

870 100–1070 357–1070

1.4 10 6.5–14.0 2.4–4.0

2.9–4.8 8.3 8.3 2.1–28.5 2.1–14.4

106–108

150–290

260–507 148–530

Pb (µg L−1)

1.1–9.6

37–170

250 80–450 42–450

107–108

BOD, biochemical oxygen demand; COD, chemical oxygen demand; TSS, total suspended solids.

viruses and parasites such as Gardia and Cryptosporidium. Median concentrations of BOD in CSOs are some ﬁve times higher than in urban stormwater and TSS concentrations are also much higher. Overﬂow concentrations of copper, lead and zinc can also frequently exceed target values for protection of aquatic ecosystems. The prime factors inﬂuencing pollutant concentrations are the length of the preceding dry weather period and the intensity and duration of the wet weather event. Aesthetic pollutants such as sanitary products, toilet tissue and faeces also characterize CSO discharges with loads depending on the magnitude and frequency of overﬂow events, catchment characteristics as well as population size and character. Floatables, including sanitary products, litter and detritus, are also characteristic of CSOs and can have an adverse impact on wildlife, primarily through entanglement or ingestion as well as having adverse aesthetic impacts. A recent concern has arisen over the incidence of sewage contaminants associated with pharmaceuticals and personal care products (PPCPs) such as chelating agents (e.g EDTA), antibiotics, anti-inﬂammatories, steroids and endocrine disrupters which are being found at levels well above the widely accepted 1 µg l−1 limit (Marsalek et al. 2002; Ellis 2006). Combined sewer and surface runoff networks should not be regarded simply as conveyance systems as they also serve as physical and chemical reactors having the potential to alter and modify the quality of received urban surface

runoff. The sudden ﬂow inﬂux into a CSO or storm water outfalls (SWO) brought on by a rainfall (or snowmelt) event can create a ﬁrst-ﬂush effect which occurs when pollutants washed from impermeable urban surfaces combine with septic pollutants re-suspended from in-pipe sediment. This combination can produce peak pollutant concentrations at the beginning of the overﬂow or outfall event, particularly if the rainfall is intense in nature. First-ﬂush effects are typically observed during the ﬁrst 30–60 minutes of a CSO/SWO discharge and are generally more pronounced after an extended dry period, in sewer systems having low gradients and systems having a small contributing catchment area. In-pipe sediments are a major source of septicity usually accompanied by gas and corrosive acidity production. Studies of combined sewer entry–exit mass loads have shown that exchanges with in-pipe pollutant stocks make up a principal source of wet weather ﬂow pollutants for solids, BOD/ chemical oxygen demand (COD) and soluble metals such that they present a prime source of acute oxygen depletion in the receiving water (Ashley et al. 2004). 7.1.3 Receiving water impacts Urban receiving water impacts are caused by discharges from both separately sewered SWOs and CSOs with the nature and magnitude of the impact being dependent on the characteristics of the generating catchment and the interactions with the receiving waterbody (Ellis and

Managing Urban Runoff Hvitved-Jacobsen 1996). Such impacts need to be evaluated in terms of speciﬁc characteristics at each site, including physical habitat changes, hydraulic and water quality changes, sediment and toxic pollutant impacts, impacts on biological communities and groundwater impacts (Ellis

Box 7.3

161

et al. 2005). Discharges of faecal bacteria also pose health risks, particularly during and immediately after wet weather events from both combined and surface water sewers (Ellis 2004). Such physical, chemical and biological effects will operate at varying temporal and spatial

Calculating CSO design ﬂows

The traditional CSO setting has been 6DWF with ‘Formula A’ used to determine combined sewer ﬂows and weir settings for carry-on ﬂow to treatment; ﬂows in excess of the Formula A discharge being discharged untreated to the receiving water. Formula A dates from the work of a UK 1970 technical committee on storm overﬂows and sewage disposal. The traditional overﬂow setting of 6DWF allows for diurnal variations in wastewater ﬂows plus an additional allowance for stormwater runoff. Formula ‘A’ = DWF + 1360P + 2E where DWF = PG + I + E and P is population served, G is water consumption/head/day (typically 120–150 L day−1) and D is inﬁltration rate (typically 50–75 L day−1, although this can be much higher in ageing sewers or where elevated groundwater inﬁltration occurs) with E being trade efﬂuent. Formula A approximately equates to 7.8DWF with instantaneous dry weather ﬂows equating to 3DWF during peak diurnal periods. A combined sewer catchment serves a population of 60,000 and has an impervious area (Aimp) of 20 hectares and conveys an efﬂuent ﬂow of 300 L head day−1. Determine the CSO setting for both 6DWF and Formula A, given a rainfall intensity (Rfint) of 30 mm h−1 for the 1-year, 30-minute duration event. Given that the upstream river ﬂow averages 1.5 m3 s−1 and the average BOD concentration is 2.5 mg L−1 with the overﬂow BOD being 250 mg L−1, also determine the overﬂow setting from mass balance of the receiving water quality immediately downstream of the overﬂow. The receiving river quality standard is limited to a BOD value of 5 mg L−1 for this storm return period and duration. Inﬁltration and trade efﬂuent are negligble. DWF = 60,000 × 300 = 18 × 106 L day−1 or 208 L s−1 6DWF setting = 6 × 208 = 1250 L s−1 Formula A setting = 18 × 106 + 1360 × 60,000 = 1153 L s−1 Runoff ﬂow-rate = (RfintAimp) = 30/3600 × 20 × 104 = 1667 L s−1 By mass balance: BOD load: river upsteam + overflow = river downstream

(1500 × 2.5) + (QCSO × 250) = (1500 + QCSO ) × 10 QCSO = 17 L s−1 DWF + runoff = setting + QCSO 208 + 1667 = setting + 17 CSO setting = 1858 L s −1

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scales. Temporal scales correspond to the nature of short-term, acute and longer-term chronic (accumulative) impacts, with fundamental intermittent standards (FIS) for discharges normally related to exposure duration and return period of the impact-causing event as well as the pollutant concentration (Foundation for Water Research 1998). The typical recovery time after a CSO event is in the order of 5–7 days. With a spillage frequency of 5–10 events per year this could imply a local violation of traditional standards during some 25–70 days per year. For separately sewered surface runoff, the recovery period is much shorter being typically 1–2 days. It cannot be concluded, however, that compliance with such criteria will provide guaranteed long-term protection as continued episodic exposure and perturbation can lead to a permanent weakening of the aquatic ecosystem and prevent ultimate recovery. For receiving waters, having a dilution greater than 9 : 1 and with no or limited interaction with other discharges, a simple mass balance assessment can be applied to set and manage impact consent for CSOs and SWOs, with minimum pass-forward ﬂows from CSOs based on Formula A. For lower dilutions and sensitive waters, stochastic impact modelling is normally recommended to ensure compliance with 90 or 95 percentile river class standards. The various impacts of urbanization on the water cycle are independent and they have a synergy which reinforces each other and lead to a general deterioration and loss in water use. This yields the paradox that it is urban areas which have the greatest requirements in terms of water and aquatic amenity use, water quality and ﬂood protection but are characterized by the highest ﬂood and pollution risks and have the most highly degraded aquatic environments.

7.2

Historical Background

Flood protection, drainage and sanitation have always ranked highly in the needs of urban society. Even in early civilizations, cities such as

Ur and Babylon of the second century millennium BC Mesopotamian Empire possessed sophisticated wastewater collection and stormwater drainage systems, the remains of which can still be found. Signiﬁcant advances in urban drainage technology were introduced during the period of the Roman Empire with roadway drainage, underground conduits and sewer networks, primarily intended for ﬂood mitigation and the drainage of lowlands. The Roman linkage of urban water supply and drainage systems marks one of the earliest examples of establishing an urban water cycle which became common in the late nineteenth century in Europe and the USA. Sanitation practices deteriorated after the decline of the Roman Empire with surface drains and streets used in the Middle Ages as the only means of conveyance and disposal of all kinds of waterborne wastes. Stormwater and foul sewage streams were thus indiscriminately mixed and became so noxious that they had to be covered and turned into sewers, giving rise to the birth of the ‘combined’ sewer principle. The beginnings of modern urban drainage practice, however, can be traced to the need for control of contaminated runoff from impermeable urban surfaces in European cities following the numerous typhoid and cholera epidemics of the 1830–1870 period. Inlets, gutters and sewers replaced open street channels and planned sewerage networks were introduced in cities such as Paris (1810–1839), Hamburg (1843), Chicago (1850), London (1859– 1865) and New York (1880). The perspective of urban drainage also changed from a design standpoint during the late nineteenth century. Most sewers constructed before the nineteenth century were not planned or designed by an engineer using numerical calculations but by a trial-and-error process. The development of empirical design methods for sizing drainage pipes was ﬁrst introduced in the design of the Hamburg sewer system in 1843. Roe’s Tables, indicating catchment area (in acres) that could be drained by circular sewers of speciﬁed sizes laid at various slopes, were ﬁrst used in the construction of the London sewer system. A

Managing Urban Runoff number of simple equations such as those of Bazalgette, Adams and McMath were also developed at this time to calculate the volume of rainfall-runoff produced by differing rainfall amounts. Such intuitive reasoning about conversion of rainfall into runoff led to the emergence of the Rational Method through the work of Mulvaney, Kuichling and Lloyd-Davies. Essentially the method contends that the rate of ﬂow (runoff) from a surface exposed to rainfall may be calculated from the product of the rainfall intensity and the surface area, with a reduction for the loss of water due to surface losses and storage, which allows for both evaporation and seepage below the surface. This reduction factor is known as the runoff coefﬁcient (Crf). By the end of the late nineteenth century, engineers, therefore, possessed various design concepts and methods for wastewater disposal systems and for the next 100 years these would become the standard tools used in urban drainage design throughout the world. The key paradigm was that stormwater and other wastewater should be collected in urban areas but with disposal and treatment undertaken outside of the urban environment and as completely as possible. Nevertheless, treatment of urban drainage was still limited within Europe and the USA with, for example, only 27 cities in the USA having wastewater treatment works by 1892. 7.2.1 Evolution of modern drainage practice The Rational Method approach dominated engineering drainage practice until the late 1960s although it is still widely used for certain applications such as small drainage areas with simple tree-branch type sewer systems having no controls, no storage or backing-up of ﬂows (Butler and Davies 2000). Since the 1960s, rapid developments have occurred in urban drainage practice and this can be directly linked to the introduction of the computer and associated electronics. Firstly, a number of runoff computation methods were developed and reﬁned, advancing the basic Rational Method to provide estimates of

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maximum and peak ﬂow rates. In addition, various hydrograph methods were developed to account for the actual variation in runoff ﬂows as rainfall intensities change during a storm. Currently it is possible to calculate ﬂows in sewer and drain networks with high precision and resolution to support cost-effective design, analysis and operation. Nevertheless, issues still exist in relation to the modelling of initial losses, accurate determination of impervious areas and the deﬁnition of urban land use types (Wastewater Planners User Group 2004). In addition, urban drainage and BMP design is still largely based on the use of 1D design proﬁle storms. Whilst these advances have helped to cope with urban ﬂooding and have substantially improved the health of urban citizens, progress in water quality considerations and particularly those addressing the impact of increasing urban populations and their activities upon both surface and groundwaters, have been much slower. Initially, research interests were focused on pollutant characterization and transport by stormwater and combined overﬂows which serve as ‘relief valves’ during wet weather conditions. Considerable advances have now been made in understanding the changes that occur in the quality of urban drainage waters during transport, storage and treatment and their impacts on receiving waterbodies (Ellis and Hvitved-Jacobsen 1996). Unfortunately the processes that control water quality in drainage systems are much more complex and less deterministic than those which control ﬂow rates (Ashley et al. 2004). Hence, many challenges remain in developing a full understanding of drainage quality processes and their impacts on receiving waters in terms of both acute (short-term) and chronic (long-term) exposure. A number of modelling approaches have been developed to identify and quantify ‘down-thedrain’ pollutant distributions in order to derive templates for risk assessment on a catchment scale (Ellis and Mitchell 2006; Keller 2006). Such approaches represent basic screening tools to support the objectives of receiving water ‘Good

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Ecological Status’ (GES) and Total Maximum Daily Loads (TMDLs) required respectively under the EU Water Framework Directive (WFD) and US Clean Water Act (CWA) regulations. The current state of knowledge has thus enabled the development of a variety of computer-based modelling approaches, which after calibration against ﬁeld data have become generally acceptable for most engineering and management tasks. A common thread in dealing with wastewater efﬂuent in the industrialized world is the separation and removal of the ‘problem’ from the community. Stormwater, domestic and industrial efﬂuents are removed both effectively and efﬁciently by conventional sewered ﬂows and once they disappear ‘down-the-hole’ the general public has little idea, or interest in, where they are conveyed and disposed; the ‘out-of-sight-out-ofmind’ philosophy. It is fortunate for much of the population in the industrial nations that their forebears invested so heavily in the construction of robust and long-lasting drainage systems. These have outlived the needs of the generations that paid the taxes to build and operate them and continue to provide services requiring minimal, though ever-increasing, re-investment. There is a growing acknowledgement, however, that a new urban drainage paradigm is required which recognizes that ‘waste’ water is in fact ‘resource’ water. 7.2.2

Towards sustainable integrated urban drainage management (IUDM)

More recently, major changes in drainage design and operation philosophy have been introduced as a result of: • the introduction and adoption of the concept of sustainable development; • improved understanding of drainage effects on receiving waterbodies; • acceptance of the need to consider urban drainage, wastewater systems and receiving waters in a holistic, integrated manner; • the continuing development of computing power and an associated range of new analytical and real-time control techniques; • the need to involve a wider group of stakehold-

ers, including local communities, in the decisionmaking process on urban drainage infrastructure and urban water resource management; • the concept of ecological integrity contained within a watershed-wide framework to control sources of both diffuse and point pollution. Whilst the latter two points are of particular relevance to European countries under the EU Water Framework Directive (WFD), similar legislation is in place in other countries such as Australia and New Zealand and has also emerged in the USA under the CWA and the National Pollutant Discharge Elimination System (NPDES) regulations. IUDM is an integration across different functional forms and types of water management (including water quality, ﬂooding and water resource management); between land and water management; and, across catchments as coherent hydrological units (Fig. 7.2). There has also been a general shift from the traditional management autonomy to the holistic IUDM approach with the key to success being the bringing together and interaction of relevant stakeholders to overcome institutional, administrative and legislative barriers. Only such collaborative, catchment-scale initiatives will support meaningful partnership engagement to achieve appropriate spatial planning strategies and sustainable master planning of urban drainage infrastructure. The existence of a lead partner to champion and facilitate sustainable drainage solutions is also a pre-requisite for successful stakeholder negotiations. Methods for the design, construction and maintenance of sustainable urban drainage systems are being developed and tested in many parts of the world. Alternative drainage concepts such as source control, water-sensitive design and low-impact development (LID) are inﬂuencing urban development practices to minimize the impacts of such development on receiving water drainage. The LID strategy is based on creating hydrologically functional equivalent design features that replicate pre-development conditions through the use of pollution prevention techniques, precision engineering and integrated micro-scale BMPs distributed across a site.

Managing Urban Runoff

165

Fig. 7.2 Shift from conventional to integrated urban drainage management.

One key hydrological feature of the LID design approach is to re-establish the pre-development water balance (National Science Center for Environmental Protection 2000). LID solutions have been principally applied to new highdensity (clustered) urban developments (Ellis et al. 2004b) but the general principles are also applicable as retroﬁt techniques for existing urban areas. In addition, alternative on-site wastewater treatment strategies are being implemented as sustainable ‘blackwater’ and ‘greywater’ options to centralized wastewater management facilities. Urban communities are searching for innovative harvesting techniques to capture, detain and re-use rainwater on-site instead of seeking to construct hard engineering structures. Many urban communities throughout the world are now

developing watershed-wide stormwater management masterplans to meet the triad objectives of ﬂood prevention, water quality control and ecological/amenity improvement. Urban drainage has indeed expanded signiﬁcantly during the past decade, well beyond the technical challenges involved in expeditiously draining urban areas, to include considerations of environmental, regulatory, political and socio-economic factors. In addition, considerable effort is being made to encourage public consultation and participation in the decision-making process leading to the choice and implementation of urban drainage infrastructure. The inclusion of a wider stakeholder participation in urban drainage issues is seen as an essential element to achieve workable long-term sustainable integrated urban water management (IUWM).

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Urban drainage has thus evolved at the beginning of the twenty ﬁrst century to become much more than the simple transport and treatment of urban runoff, and the time is ripe for the introduction of new paradigms based on more longterm sustainable strategies which can address future trends in urban and demographic growth, consumerism and privatization as well as climate change (Butler 2004). A major future issue for IUDM will be the control of ﬂow and water quality within existing high-density urban areas for which there are more limited solutions than for new greenﬁeld or brownﬁeld sites. Solutions must necessarily focus on retroﬁtting and municipal land use controls, e.g. trafﬁc and parking management, street and gully cleaning, controls on conversion of gardens to hard standings etc., as well as national campaigns on product substitution to reduce toxic pollutants and on the problem of vehicle contributions to surface water pollution.

7.3

Current Solutions

7.3.1 The development of urban drainage standards Impermeable surface runoff presents a complex drainage problem in comparison to foul drainage systems which are more stable in terms of the magnitude, duration and frequency of their ﬂow regime. The design ﬂow rates (Q) for stormwater runoff systems are essentially a function of rainfall intensity (Irf) and storm return interval (RI) and follow a basic general equation: Q = (RI × Crf × Cr × A × I rf ) where Crf is the runoff coefﬁcient, Cr is the storage routing coefﬁcient and A is catchment area. Flow rate increases with return period and with reduced storm duration. The universal standard design criteria for ﬂood return period has been the 1 : 100 year event, although varying up to 1 : 1000 in Australian practice. However, 1 : 25 or 1 : 30 RI are widely

adopted performance standards for surface water drainage whilst a design level of 1 : 1 or less is required for the capture and treatment of frequent events associated with the majority of urban runoff pollutants. The generally preferred return period design criteria for differing types of ﬂooding with a 10–20 year service level is a minimum level of protection against area ﬂooding and a 20–30 year service level range for property ﬂooding (Table 7.3). Such levels of service protection are intended to protect the development site from ﬂooding of the drainage system. In designing drainage systems there is an expectation that the return period of the design storm equates to the ﬂooding performance of the sewer system, but there is little evidence to support this assumption. The event design approach is based on the concept of ‘acceptable’ levels of risk, but there are uncertainties associated with rainfall durations, antecedent condition as well as climate change uplift levels, with wet weather parameters being particularly uncertain. From a modelling perspective it is easier to take average (or veriﬁed) conditions and then improve the factor of safety in design by just increasing the rainfall return period, but is the safety factor the same as acceptable level of risk? In any case, it may be more appropriate to base urban drainage and BMP design on 2D time series events rather than on 1D design storm proﬁles. For receiving water quality protection, a much lower trigger level of service protection is required. This would require a return period of less than 1 : 1, typically being in the range of 0.25–0.5 RI with interception storage also required for the initial 10–15 mm of effective rainfall runoff. Interception of the initial 5 mm, however, can be a more cost-effective target level. There should also be provision on- or offsite for surface storage (and/or surface routing) of infrequent, high intensity (>1 : 100 year) extreme events and this is frequently overlooked in the design of some source control devices. Most countries have introduced regulations covering controlled or permitted activities as applied to urban developments having impermeable surfaces greater than about 1000 m2 which

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Managing Urban Runoff Table 7.3

Urban design return periods Flooding type Roof drainage

Internal ﬂooding External ﬂooding Road ﬂooding

Trigger level

Target level

Flat roof; 1 : 1 Pitched roof; >1.5 × design life of building up to max. probable rainfall (or 1 : 100) 1 : 10 (10%) 1 : 30 (3%) (also 1 : 20; 5%) (1 : 20 residential areas; 1 : 30 city centre areas) 1 : 5 (20%) 1 : 20 (5%) (1 : 2 gardens, parks etc.) 1 : 1 (100%) 1 : 10 (10%)

require best practicable technology or BMP to minimize and mitigate any deleterious discharge effects. Such practice frequently refers to the nature of the discharge and receiving water, technical feasibility as well as cost implications. Set against the objectives of ﬂood and pollution potential and ecological status, the regulatory criteria usually relate to pre-development peak ﬂow rates for various rainfall probabilities (1% to 50%, i.e. 2 through 10 to 1000 RI), and very often have a 70–75% TSS removal baseline target as a long-term average. The capture of rainfall runoff for 80–90% of all storm events will normally meet the 70% TSS removal criteria. For ponds and wetlands, treatment performance should also consider the pollutant decay rate within the control structure, where the removal rate becomes a function of the plug ﬂow kinetics as well as the hydraulic retention time (Ellis et al. 2003). Storage and/or attenuation of excess ﬂow volumes, however, tend to focus on the initial 10–25 mm of total rainfall (or 5–10 mm of effective rainfall-runoff) over a minimum 24hour period for pollution control. The adoption of such criteria has required onsite interception, attenuation and/or inﬁltration where the ≤1 : 1 greenﬁeld peak needs to be controlled to a discharge rate perhaps as low as 2 L s−1 ha−2. A two-tier design discharge standard has, therefore, become more acceptable, with total runoff split between very low rates during wet weather peak ﬂow receiving water periods and a more extended higher ﬂow rate (up to 5/7 L s−1 ha−2). In addition, consideration of design

exceedance during short-term, high intensity (<150 mm h−1) rainfall events has also become a recommended criterion. Under such conditions, overland ﬂow contribution from normally permeable surface areas will occur, as well as additional exceedance ﬂows from surcharged gullies. A 10% uplift is also generally applied to drainage conveyance and routing design to allow for climate change up to 2025, increasing to 20% for designs to the end of the century. Such allowances mean that future storage volumes must be up to 50% greater than allowed under present day rainfall conditions. There will still be a need for hydraulic design to incorporate 24–48 hours drain-down times which will require consideration of time series rainfall rather than event-based design storms. For development sites over 2 ha, stormwater storage, therefore, must normally be provided as a range of distributed treatment train units rather than as single ‘stand-alone’ devices at the downstream end of the urban catchment. There are also a variety of national (or state/provincial) guidelines for the design, layout and gradients of hard surfaces such as car parks and pavements as well as for roadside gutter channels and spacing of road gulleys (inlets). 7.3.2 Best management practices Two categories of urban BMPs for impermeable surface runoff can be generally identiﬁed as: (i) non-structural or source control BMPs which seek to prevent and/or minimize runoff

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Fig. 7.3 BMP in-series treatment (surface water management) train.

generation and the introduction of pollutants; and (ii) structural or treatment BMPs aimed at runoff storage and/or inﬁltration with or without pollutant removal. This BMP concept and term was originally developed for the mitigation of diffuse pollution impacts and is rooted in publications emanating from the US Environmental Protection Agency (United States Environmental Protection Agency 1993). The original conceptualization has evolved to the recognition of a categorization of a BMP hierarchy for the control and treatment of urban stormwater runoff which is normally integrated in terms of device design into series to maximize the quality beneﬁts (Fig. 7.3). It is the use of BMPs in series that has

become commonly termed the surface water ‘treatment or management train’ although the performance of in-series BMPs in comparison to individual, stand-alone BMP devices, has rarely been demonstrated. The concept of sustainable urban drainage systems (SUDS in the UK or Water Sensitive Urban Design, WSUD, in Australia) has become subsequently adopted to advocate the design integration of ﬂow quantity, quality and amenity aspects as an essential and holistic element of stormwater management (Campbell et al. 2005). This SUDS triangle concept recognizes the signiﬁcance of all three interests in integrated stormwater management, and although one driver (e.g peak ﬂow volume)

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Managing Urban Runoff Table 7.4

Low impact development hydrological design components

LID practice Flatten slopes Increase ﬂow paths Increase roughness Minimize disturbance Flatten swale slopes Inﬁltration swales Filter strips Disconnect impervious area Reduce kerb/gutter Rainwater barrels/tanks Green roof storage Bioretention (rain gardens) Re-vegetation

Reduction in postdevelopment runoff

Increase in Tc value

Provision of retention

Provision of detention

X X X X X X X X X

X X

X X X X X X X

X X X

X X X X

X

Tc, time of runoff concentration.

might be of primary importance, the ﬁeld design should be capable of ‘scoring’ on all three components. Contemporary views, however, would consider urban surface water drainage as essentially comprising an integrated management train having a ‘toolbox’ of drainage techniques that include ‘semi-natural’ and proprietary BMPs as well as conventional pipe drainage, but set within the context of LID. There is growing conviction that non-structural approaches, which engage and involve stakeholders, can be more effective than increased infrastructure. The use of site/lot clustering can free up land for the introduction of structural drainage controls and community open space on greenﬁeld developments to provide enhanced beneﬁts (Ellis et al. 2004b). The effectiveness of LID practices can be measured through the hydrological effectiveness provided by runoff potential, as measured by the reduction in runoff volumes (and curve number, CN), the time of runoff concentration (Tc) and by the amount of retention and detention provided (Table 7.4). As an example of the effectiveness that a LID approach can achieve, a 32-ha (80-acre), 199-dwelling development site at Somerset in Maryland saved over $30,000 and gained six additional plots of 1000 m3 for dwelling units through its

application (Richards 2001). Such development gains provide a ﬁnancial advantage (and incentive) to the developer and construction companies as well as providing enhanced quality-of-life environments for residents. There is also interest in LID approaches for new developments given national and state planning guidance relating development and ﬂood risk as well as the need for increasing house densities of up to 30–40 dwellings per hectare. Such high densities will require LID clustering approaches to reduce the building ‘footprint’ and leave intervening and adjoining land available for LID-type drainage components and open space (Ellis et al. 2004b). In the USA, BMPs have been advocated since the inception of the 1972 Federal Water Pollution Act, although they have been rather slower to take off in Europe, apart from in Scandinavia. Such surface water BMPs (SUDS or WSUDs) are now either required or recommended drainage systems for new urban development in Europe and North America although there are still barriers to their introduction primarily on grounds of heath and safety, operation and maintenance issues as well as long-term performance and adoption practice. The principles of BMP and associated LID construction are now well recognized and there are

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numerous guidance manuals available for their design and operation produced at local, state/provincial and national level. The majority still follow the basic generic design parameters and criteria developed for the Maryland state stormwater management manual by Schueler (1987). This baseline guidance manual has strongly inﬂuenced most of the follow-on US state/ municipal manuals, e.g Portland Bureau of Environmental Services (2004), or the series of stormwater BMP handbooks produced by the Californian Stormwater Quality Association (CASQA) over the period 1993 to 2003 (www. casqa.org) as well as the 2004 three volume national US EPA Stormwater BMP Design Guides (EPA/600/R-04/121). The inﬂuence of US design criteria and parameters can also be seen in those published elsewhere such as in Canada (Ontario Ministry of Environment 2003), the UK (Woods-Ballard et al. 2007), Australia (Wong 2005), New Zealand (Auckland Regional Council 2003) or Malaysia (River Engineering Division, Dept. of Irrigation & Drainage – Malaysia 2000). A detailed critique of various national BMP design manuals can be found in Ellis et al. (2008). It is important to bear in mind the fundamental difference between the BMP and SUDS/LID approaches. The former were essentially developed in the USA to address problems arising from small-scale, frequent ﬂow and/or quality events with any ecological or amenity beneﬁts being considered to be a subsidiary ‘added’ value. The SUDS and LID approaches are intended to address all three regime components and to provide for the full range of ﬂows and quality conditions that might be encountered on an urban site. 7.3.3 Regulatory approaches for urban runoff The most comprehensive regulations covering urban stormwater discharges are those included in the US EPA NPDES permit program under Section 402 and 101 requirements (Ryan 2003). Water quality standards are contained in Section 303(c)

of the Federal Water Pollution Control Act as applied to designated receiving water uses. The 1987 Phase I and follow-on 1999 Phase II conditions of the NPDES permitting procedure now apply to all urban areas having total populations exceeding 10,000 and with densities greater than 1,000 persons m−2. To date, 275 general permits have been issued for municipal separate storm sewer systems (MS4s) which embraces some 70% of the US urban population. The emphasis of such MS4 permits is on the application of BMP technology to limit receiving water pollutant exposure to the maximum extent practicable through the development and implementation of stormwater pollutant prevention plans (SWPPPs). These plans must identify minimum control measures for: construction site runoff and post-construction stormwater management; illicit discharge detection and elimination; municipal ‘house-keeping’ (or source control) procedures; and, public participation, education and outreach. The US approach has, therefore, focussed on minimum technology based standards, relying on best professional judgement, site-speciﬁc factors and associated monitoring, acknowledging the inherent variability in urban surface runoff. As currently designed, total maximum daily load (TMDL) limits for non-point discharges are based on the assumption that there is a direct correspondence between the total mass of waste loads and ambient water quality and that an annual cap on total wasteload allocation (WLAs) can sufﬁciently maintain acceptable urban water quality and ecologic status. Information and guidance on the NPDES regulations and procedures together with BMP detail and sample outreach materials can be viewed on the US EPA website (www.epa.gov/npdes/stormwater). There is considerable promotion of watershed-based approaches within SWPPPs and which involve coordinated public and private efforts to address highest priority water quality problems. This is intended to encourage ﬂexible thinking and innovative methods such as LID to be explored (Dodson 1999). The LID approach utilizes local vegetation, landscaping and

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Managing Urban Runoff small-scale treatment systems to treat and inﬁltrate at source, and which yield ﬂow control ‘credits’ to municipal administrations. In addition, as previously argued, the LID approach is applicable to both new and existing development; retroﬁtting opportunities providing the key to the latter situations. In terms of control and management of combined sewer discharges, the modiﬁed Rational Method is still widely in use as a minimum sewerage performance criteria for wet weather discharges deemed of low impact signiﬁcance. New designs of sewer overﬂows were adopted in the late 1960s, however, to provide for retention of pollution in the sewer pipe, such as the use of storage, stillage ponds and high-side weir overﬂows. This, together with increased sewer overloading resulting from continued urban development, led to deteriorating CSO performance (Ellis and Crabtree 1999; United States Environmental Protection Agency 2004). US and European wet weather CSO discharge control policy is now essentially based on receiving water quality standards linked to use-related objectives (Arsov et al. 2003). Such standards are normally based on consideration of the duration– frequency relationships for key pollutant concentrations such as TSS, dissolved oxygen (DO), ammonia (NH3-N) and bacteria (Table 7.5). As an alternative to the fundamental intermittent standards (FIS) approach, the UK Urban Pollution Manual (UPM2), identiﬁes the 99 percentile criteria as related to the appropriate river ecosystem (RE) classiﬁcation. Whilst comparable levels of receiving water protection provided by the FIS and 99 percentile criteria cannot be directly quantiﬁed, they do offer alternative descriptive (or narrative) standards for the acceptability of extreme wet weather events, although the FIS approach only focuses on acute biochemical impacts and ignores chronic, long-term toxic effects. In practice, however, the FIS standards are usually modiﬁed by the use of correction factors to take into consideration speciﬁc site conditions and likely pollutant interactions. Very similar

episodic wet weather standards have been applied in Scandinavia and will certainly be more widely considered within Europe as the WFD catchmentbased management process is progressively implemented. 7.3.4

Urban runoff risk assessment

The outcomes of the Article 5 EU assessment procedure required under the WFD regulations for the preliminary characterization of receiving waterbodies has clearly identiﬁed diffuse (nonpoint) urban pollution as a risk to both surface and ground waters (Table 7.6.). About 25% of river lengths and 14% of groundwaters are considered to be ‘at risk’ or ‘probably at risk’ in England and Wales from such drainage sources with some 18,175 km2 of aquifers being potentially impacted. The assessment procedure was based on consideration of urban land use activity, source pressure, exposure pressure and impact data with outcomes categorized as high, moderate, low or no exposure pressure. This pressure category was then converted to a risk factor using impact data, where available, together with a measure of conﬁdence. This preliminary risk assessment will inform input to the development of appropriate Programme of Measures (PoMs) for the next River Basin Management Plan (RBMP) stage. In order to support sustainable control and treatment planning strategies, procedures for identifying and quantifying the spatial distribution of source pollutant loadings within urban catchments are, therefore, needed to further evaluate the potential risks to receiving water objectives under both ‘good ecological’ and ‘healthy’ status required respectively by the EU WFD and US CWA regulations. Thus, a similar policy driver underpins the need for information on pollutant ﬂuxes on a catchment scale as a basis for mitigation and/or remediation measures. The methodology for risk assessment for unit area loadings allows ‘hotspot’ source loadings to be located more accurately and can be used as a basis for the strategic planning of new or retroﬁt surface runoff controls and/or enhanced road

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Table 7.5 UK fundamental intermittent standards applied to urban runoff in respect of dissolved oxygen (DO) and ammonia (NH3-N) Salmonid ﬁshery

Cyprinid ﬁshery

1h

6h

1 yr

1h

1 month: DO NH3-N

5.0 0.065

5.5 0.025

6.0 0.018

4.0 0.15

5.0 0.075

5.5 0.03

3 months: DO NH3-N

4.5 0.095

5.0 0.035

5.5 0.025

3.5 0.225

4.5 0.125

1 year: DO NH3-N

4.0 0.105

4.5 0.040

5.0 0.030

3.0 0.25

4.0 0.95

Table 7.6

6h

Marginal cyprinid ﬁshery 1 yr

1h

6h

1 yr

3.0 0.175

3.5 0.1

4.0 0.05

5.0 0.05

2.5 0.25

3.0 0.15

3.5 0.08

4.5 0.065

2.0 0.3

2.5 0.2

3.0 0.14

Waterbody pressures in England and Wales resulting from diffuse urban sources Rivers

Groundwaters

No. of WBs

River length (km)

% River length

No of WBs

At risk Probably at risk Probably not at risk Not at risk Not assessed

176 895 2222 4523 0

3,441 14,103 20,087 32,587 0

5 20 29 46 0

12 65 190 89 0

1,513 16,662 89,803 22,732 0

1 13 69 17 0

Total

7816

70,217

100

356

130,710

100

Risk of failing WFD EQOs

Area of WBs (km2)

% Area of WBs (km2)

EQO, environment quality objective; WB, waterbody; WFD, Water Framework Directive.

surface and gullypot cleaning programmes. The averaged UALs express the exposure pressure exerted by source land use activities within the urban catchment and are consistent with the risk assessment framework required by both the CWA and WFD regulations. A catchment hazard (or risk assessment) map can be produced by comparing the ratio of the estimated micro-catchment UALs (UALMC) with permitted (or targeted) EQS maximum acceptable unit area loads (MUALMC) for speciﬁc water uses impacted by the micro-catchment discharges. The micro-

catchment load kg ha−2 yr−1 is calculated as: UALMC = (∑loadcells(s))/(Area) and the MUALMC as: MUALMC = (EQS × QSC ) ( AreaSC ) where EQS is the maximum allowable pollutant concentration by use class and Q is the observed annual mean 30 year sub-catchment (SC) discharge. The pollutant hazard is then calculated as: UALMC MUALMC

Box 7.4

Risk assessment for unit area loadings

A number of risk assessment approaches have been suggested including that proposed in Ellis and Mitchell (2006) based on a volume–concentration procedure. The method uses the product of site runoff volume and the pollutant EMC for the relevant urban land use activities (as identiﬁed in Table 7.1) in a semi-distributed stochastic modelling procedure. The runoff volume (Runvol) is deﬁned by the Wallingford modiﬁed rational method: Runvol = [{0.829PIMP + 25.0SOIL + 0.078UCWI − 20.7}100] × Rfa × Ac where PIMP is the percentage impervious area, SOIL is the rainfall acceptance potential, UCWI is an urban catchment wetness index, Rfa is the annual average rainfall and Ac is the contributing catchment area. The ﬁgure below illustrates the unit area load (UAL; g ha−1 year−1) of zinc for a 6310-ha sub-catchment of the River Lee, a north-bank tributary of the River Thames in Greater London, based on 50 m × 50 m grid cells derived from this methodology and developed within a GIS database.

Zinc UALs for an urban sub-catchment of the lower River Lee in north-east London, UK It can be seen that the highest unit zinc loadings are associated with linear land uses such as heavily trafﬁcked highways (in this case the east–west urban M25 orbital motorway and the north-south A10 trunk road) with other elevated UAL ‘hotspots’ associated with intensive commercial/industrial and high housing density locations. Thus many of the zinc sources shown in the ﬁgure are related to linear street sources and associated gullypot accumulations which are ﬂushed into the surface water pipes during wet weather conditions. The capture, treatment and inﬁltration of these key sources would undoubtedly provide a major beneﬁcial component to any PoMs intended to achieve receiving water EQS.

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The percentile classes produced for hazard maps (Ellis and Mitchell 2006) would facilitate the identiﬁcation of those diffuse source areas likely to pose the greatest threat to relevant beneﬁcial water uses and at which mitigating PoMs may be cost-effectively targeted at water which may also be a prime cause of EQS non-compliance. There is increasing national effort throughout the world in the development and implementation of catchment-based ﬂood risk management plans and associated improved administrative coordination and public communication. This reﬂects, for example, that in the UK some 1.8 million residences and 140,000 commercial properties (equating to 4–5 million people) are at risk from ﬂooding and it is estimated that by the 2080s, urban ﬂooding costs could be anything between £1 and 10 billion pounds per year (www. foresight.gov.uk). In the last 10 years, Europe has suffered over 100 major damaging ﬂoods causing some 700 fatalities and 25 billion Euros in insured economic losses. Central and state governmental activity has been stimulated by federal regulations such as the EU Floods Directive (2007/60/ EC) as well as public concern regarding ﬂood insurance coverage. There is a need, however, to ensure that ﬂood and water quality protection are appropriately balanced to best meet the interests of sustainable development and that links, rather than overlaps, are made between various legislative directives and regulations. Joined-up thinking is needed to ensure that programmes separately implemented for ﬂood and water quality/ecology mitigation do not compound the overall risks, and that intraurban sewer ﬂooding and pollution is included in catchment analysis and management plans. It is important that the spatial-temporal timeframes as well as supporting legislation, technical and administrative solutions are compatible. At the present time, the techniques and technologies to enable a fully integrated risk-based assessment of urban ﬂooding and pollution with concurrent appraisal of strategic catchment-level portfolios of control options, are not generally available. Uncertainty is increasingly being used in place of ‘risk’ and instead of ‘acceptable risk’ there

is a tendency for strategic policy to seek to manage all risks. The likely result will be a layered system of ﬂood and water quality risk management rather than reliance upon any single strategy. 7.4 Insights and Implications 7.4.1 Socio-economic drivers In any analysis of trends and drivers affecting urban water systems, sustainable solutions provide only one, albeit important, inﬂuencing factor. Population and demographic trends are of equal signiﬁcance to scientiﬁc and technical innovation. Despite major demographic shifts associated with reducing household sizes and ageing populations, higher per capita water and enhanced wastewater infrastructure demands are quite likely. The rise of consumerism, individualism and increased disposable income in western societies will also generate new drivers for future water and wastewater supply and management as exempliﬁed in the substantial rise of bottled water supplies, water-using domestic appliances, separated waste streams and waste re-cycling schemes. These will be coupled with increased expectations of improvements in environmental quality-of-life within riparian urban corridors and adjacent open space. Chocat et al. (2004) have proposed the evolution of four possible future scenarios for urban drainage. Their ‘green’ scenario is dominated by decentralized source-control BMP approaches with minimum sewer connections and extensive wastewater re-cycling and water conservation. This scenario is equivalent to the community-based, local stewardship scenario envisioned in the UK Foresight ﬂood report (Ofﬁce of Science and Technology 2004). The conservative ‘technocratic’ scenario, adopting centralized advanced technology, monopolistically retained and managed within the public sector, reﬂects the Foresight regionalized, national enterprise socio-economy. The ‘privatization’ scenario is clearly consumer and world market oriented and represents the predominant strategy operating within many developing countries at present. The ﬁnal scenario suggested by

Managing Urban Runoff Chocat et al. (2004) is that of ‘business-as-usual’ which stumbles between the technocratic tradition and green ideas as well as picking up degrees of privatization. The scenario analysis approach provides no ﬁrm indication of which socio-economic future might be more probable than another. It is important to note, however, that technical solutions do not map clearly onto any particular socioeconomic future and that the high existing ‘sunk’ asset of urban drainage means that there is always likely to be considerable inertia and conservatism in the water industry. Thus it is feasible to visualize a ‘no-change’ (or little change) future scenario with urban water resource managers having a lack of control with respect to land use planning and chemical usage as well as being underfunded in respect of rehabilitation and maintenance. In decision-making, the process of choice is becoming more central with stakeholder engagement seen as a core issue, rather than consultation which implies a more passive role. Concomitantly, there is a strong move to decentralization of government and devolution of planning and regulatory guidance with a move towards multi-criteria analysis (MCA) as a means of enabling stakeholder debate. In the use of such analysis, there are questions of whether local or national ‘weights’ should be applied to the beneﬁt considerations. Alternative perspectives have been widely canvassed, with some advocating a complete revolution in future urban water systems in terms of new paradigms, new contexts and new methodologies (Maksimovic and Tejada-Guibert 2001). This thinking is based on the introduction of multi-disciplinary, integrated approaches to urban drainage incorporating sustainable principles, the adoption of network, risk and vulnerability analysis, complex modelling, incorporation of educational and social values as well as anticipatory and contingency scenarios for addressing new impacts such as climate change. 7.4.2

Climate change

Projected changes in rainfall and temperature over the lifetime of BMP/SUDS structures may

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well lead to problems of overloading and poor performance, increased erosion and transport of sediment and pollutants by urban runoff and adverse receiving waterbody impacts including pond and in-stream siltation as well as groundwater contamination. The emergence of chronic delayed ecosystem impacts presents another climate change danger as well as the evolution of infectious diseases by vectors inhabiting urban wetlands, ponds and vegetated BMPs. New emerging chemicals and potential GMO use in urban landscaping might also pose additional problems as well as the impact of increased public amenity water uses. The UK Foresight review of ﬂood defence (Ofﬁce of Science and Technology 2004) has developed various scenarios for future ﬂood risk and CSO spillage, although the report says very little on surface water ﬂows and associated pollution. Volume uplifts of 33% to 40% are predicted for the 10- and 30-year RI storms, and given that a large percentage of storm drains are less than 100–300 mm and only designed to a 50 mm h−1 storm, they have very little spare hydraulic capacity. The inadequate capacity of surface water drains was highlighted by the Pitt Review (Pitt 2007) as being the major cause of urban pluvial ﬂooding during the intense storm events of summer 2007 in the UK. Standard surface water drainage design and sewer models are not well suited to 1 : 100 extreme return period analysis, yet it is precisely these extreme events that stakeholders are interested in. Substantial increases in future ‘headwater’ ﬂooding within urban catchments can be expected both in winter (following longer wet periods) and in summer from increased short duration, high intensity events. The problem will be exacerbated as there will be less (∼30%) inﬁltration capacity in the winter period, and a 7–10% urban creep factor (from inﬁll development and paving over drives and gardens etc.) which also makes the problem worse. The loss of inﬁltration capacity, especially in winter, will render inﬁltration control devices less efﬁcient (up to 50% are thought to fail at present anyway), and thus a move towards on-site surface control approaches such as car park ponding are likely to be preferred.

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Damage costs are likely to rise by a factor of between 2 and 14 over the next decade with stormwater outfall discharges increasing by up to 35–40%. These increased, climate-driven ﬂows have knock-on water quality effects with enhanced and more frequent ‘ﬁrst-ﬂushes’ and mixing of stormwater with sewage in manholes leading to problems of surcharging and consequent health hazards as well as adversely affecting in-stream ecology. There are, of course, considerable uncertainties involved in these predictions, e.g receiving waters are quite likely to have higher dilution capacities and so may be better capable of absorbing enhanced pollution impacts, although the hydraulic impacts will remain. It is clear that under these climate scenarios, operation and maintenance activities will become much more important as will performance measures to monitor sewer and BMP condition and capacity. 7.4.3

Advanced solutions

There are a number of challenges for future urban drainage management arising from the consideration of further technical and strategic reﬁnements to existing BMP controls and regulatory approaches. One important issue is that of managing urban ‘retreat’ as a means of reducing ﬂood and pollution risks whereby roads, parks and open spaces are allowed to ﬂood during storm events (say up to 100 mm for storms exceeding the 1 : 1 event), with ﬂood prone property being bought up for ﬂood storage areas. In delivering IUDM it is likely that integration will need to be administered through a geographically overlapping, functional mosaic of institutions which might be difﬁcult to achieve. This organizational issue has been recognized, for example, in a recent UK review of surface water drainage (Department for Environment, Food and Rural Affairs 2008) which recommends a focussed line management for stormwater control and planning with clear powers and boundaries of responsibility. Equally, IUDM will have to be delivered through functional budgets with different institutions being responsible for the differ-

ent ﬂood pollution, groundwater, ecological and recreation/amenity beneﬁts. Different institutions, however, will be responsible for these different beneﬁts with varying rules covering their budgetary outgoings. The problem is thus not evaluating the beneﬁts but reaching agreement over shared payment and administration. If legal and ﬁnancial powers are more weakly deﬁned then accountability can become an issue, and this can also present a problem in the stakeholder engagement process when the most articulate group (or those who have most to gain or lose), are most likely to engage in the process but will be spending other people’s money. There has been a reliance in most BMP/SUDS design handbooks and manuals on stand-alone devices and/or hybrid combinations for stormwater management with little if any consideration for more advanced solutions. This is of particular relevance given the ‘cocktail’ nature of urban runoff, the ﬁrst-ﬂush phenomena and primacy of ﬁne particulate-associated toxic pollutants. More advanced designs based on consideration of unit operating processes offer possibilities for control devices which are able to specify the efﬂuent in order to meet required water quality standards and permitted target levels. Within such design constraints it should also be possible to estimate lifecycle costs as well as long-term BMP/SUDS cost-performance. One such approach is the integration of stormwater measures into the built form extending BMP application into the urban architectural design. Such approaches can help to ‘close-the-loop’ between surface (stormwater) and indoor (greywater) water budgets through reconﬁguration of the building drainage. Such integration as well as including green roofs, rainwater harvesting and greywater treatment, should also address nonstructural measures such as indoor water saving measures which might require changes to prevailing policy, ﬁnancial and administrative frameworks. The application of land use control measures in conjunction with such built-form measures provides a further advanced stormwater management approach. Such land use controls include

Managing Urban Runoff restricting road widths in low-density residential areas, footpaths restricted to one side of the street, use of shared accessways, minimization of on-street parking as well as ‘daylighting’ piped gutter systems at gully inlets to provide pocket wetlands and rain gardens which can also serve as chicane trafﬁc calming measures. The use of customized, lightweight honeycombed modular cells incorporating mixed-media ﬁltration units and constructed from recycled PVC, LDPE or HDPE materials can also be applied to inﬁltration and porous surfacing BMPs. Such structures have a high void ratio (>97%) with high compressive strength (1000 KNm−2) and low water ﬂow resistance. The integration of such proprietary devices with conventional BMP controls (and with hydrodynamic devices, sorbent chamber ﬁlter inserts, slot drains, etc.) can provide a more advanced treatment of the urban runoff efﬂuent ‘cocktail’ and can be particularly effective as retroﬁt solutions. Multi-chambered treatment trains (MCTT) suitable for small (0.1–1.0 ha) urban sites, including critical pollution hotspots such as service stations, car washes, vehicle maintenance areas, commercial centres, etc.) can also be valuable retroﬁt options. The MCTT is a three-chambered tank with an initial grit chamber (with volatile release provided by column aerators), a main settling chamber providing bubble aeration and sorbent pillows, followed by sand/pear mixed-media ﬁltration for sorption and ion exchange of ﬁlterable toxicants with ﬁnal gravel underdrainage. Such tiered or multiple treatment series follows the basic unit operating process (UOP) hierarchy in terms of particle size and contaminant association and removal mechanisms. High removal rates of both solid and dissolved pollutants have been achieved in ﬁeld trials with tests in Wisconsin on garage and public works maintenance yards yielding outﬂow discharges of <10 mg L−1 TSS, <0.1 mg L−1 Ptot 3–10 µg L−1 HMs and <0.1 µg L−1 total PAHs. The further development of the MCTT to incorporate inorganic primary coagulants and/or synthetic polymers to target speciﬁc toxic species is under ﬁeld testing

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(Field et al. 2006). Such systems might be used as mobile ‘packaged’ treatment plants for the combined use of high-rate sedimentation, disinfection and hydrodynamic technologies to treat ﬁrst-ﬂush hotspots and for the removal of metals, hydrocarbons or other priority pollutants. Upﬂow ﬁltration units are being developed which can reduce clogging potential and allow high ﬂow rates to be maintained. Such units can be installed into conventional catchpits (gully chambers) and use sedimentation, screening and ﬁne solids sorption and ion exchange mechanisms to remove solids and dissolved pollutant fractions. A variety of mixed-media including zeolites, activated carbon as well as compost and sand/peat mixes have been tested and results suggest removal down to 5–210 mg L−1 TSS, 0.02– 0.1 mg L−1 Ptot and 0.1–3 µg L−1 metals (<20 µg L−1 zinc) can be regularly achieved. Upﬂow ﬁltration units, with sump and inter-event drainage capabilities, can also be used as follow-on polishing systems to conventional BMP storage and bioﬁltration devices such as swales or inﬁltration trenches. The retroﬁt application of such small-scale advanced BMPs offers many opportunity costs for market-based trading runoff allowances (or credits) supporting the strategic assignment of dispersed runoff controls throughout an urban catchment.

7.5 Future Research Requirements Future knowledge and programme needs in the urban runoff ﬁeld span across a variety of disciplines including science, engineering, planning, social science and management and may conveniently be grouped under four major categories. 7.5.1 Urban runoff processes There remains a need to develop a better understanding of urban runoff processes to support future regulatory requirements and planning strategies as well as to reﬁne and optimize exist-

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ing controls and modelling procedures in order to achieve a more sustainable cost-performance for urban drainage infrastructure. Principal areas of knowledge which would beneﬁt from further research include: • Sources, source inventories, pathways and fate of toxic contaminants and microbial pollutants in the urban environment and in particular a knowledge of the occurrence and impact of priority substances on which only limited research has been undertaken to date. Much further knowledge is required of the processes, routing and impacts associated with extreme wet weather events for which only very limited data exists. This issue of exceedance ﬂow is of particular relevance given the increased likelihood of extreme event occurrence with future climate change. The impact of extreme events in general, such as large oil and detergent spillages and their effect on porous surface performance, also need addressing. • The effectiveness and long-term performance of control measures in protecting receiving waterbodies and both aquatic and terrestrial ecosystems as well as human health. This requires investigation of both technical source control solutions as well as strategic policy approaches and their cost-effectiveness at plot level (BMPs/ SUDS), community level (site end-of-pipe level) and at sub-regional catchment level. Given the recognition of the toxic ‘cocktail’ nature of urban runoff, more knowledge is needed on the advantages of applying more ‘advanced’ source control technologies for the treatment and control of urban runoff, e.g the use of multi-media ﬁltration systems for porous paving and kerb drainage. What are the ‘better’ BMP/SUDS control solutions for the detention and treatment of speciﬁc pollutant groups and species? Is it feasible to identify appropriate design criteria for the achievement of speciﬁed outﬂow standards in terms of concentration limits and occurrence frequency? The long-term hydrological and water quality impacts resulting from the accretion and superimposition of dispersed site level controls and attenuated storage on the downstream catchment regime also warrants further examination.

Such knowledge will contribute to the eventual substitution of efﬂuent criteria with ecological risk assessment of receiving waters as required under the EU WFD and US CWA regulations. • The increased vulnerability of urban aquatic ecosystems resulting from the secondary effects of stormwater management such as risks of groundwater contamination, heating of wetlands and storage ponds, transformation and release of contaminants from sediments and risks to wildlife through uptake of contaminants and exposure to pathogens. • The cumulative and combined effects of urban efﬂuents on receiving waterbodies and their ecosystems with respect to intermittent exposures of varying concentration. At present there is only limited knowledge on the nature of those processes which control the inherent variability and uncertainties associated with urban runoff discharges. A better knowledge of the interactive process effects during wet weather ﬂow conditions within and through BMP/SUDS devices would undoubtedly help to improve design parameters and the robustness of their long-term performance as well as optimizing their control and treatment potential. 7.5.2 Data and monitoring needs • To overcome the deﬁciencies in available data and the limited data on priority pollutants such as pathogens, pesticides, herbicides, dissolved metals and hydrocarbons as well as other persistent trace organics such as POPs, fuller and more up-to-date biochemical and microbiological descriptors of both surface water and combined sewers are required. National harmonized data inventories with standardized sampling and analytical protocols would provide more accurate, robust and comparable databases to help reduce uncertainty. • An increased data effort to quantify new priority chemicals including natural and synthetic hormones, industrial chemicals having oestrogenic potential, pharmaceuticals and personal care products (PPCPs), genetically modiﬁed organisms (GMOs), road salts, used crankcase

Managing Urban Runoff oils and their various breakdown and daughter products. • Appropriate databases identifying the longterm cumulative impacts of urbanization including geomorphic channel and associated regime changes as well as extreme overland ﬂows, habitat degradation and ecological niche improvements. • Long-term performance capabilities and design criteria for both stand-alone and hybrid source control and BMP/SUDS measures and their impact on the status and performance of existing drainage systems, extending from catchment headwaters to the point of receiving discharges. 7.5.3 Integrated management of urban water • The development, testing and implementation of appropriate sustainability criteria for the understanding of total urban water cycle management. • The testing and implementation of appropriate tools to model complex urban drainage systems and exceedance ﬂows and routes and to facilitate the delivery of procedures for integrated ﬂood and water quality risk management approaches. The key innovation here would be risk-based methods capable of exploring the performance of multiple urban drainage strategies within a single coherent and dynamic analytical framework. • The development of mutual supporting processes and administrative structures for urban master planning and drainage infrastructure based on a better scientiﬁc understanding of operational hydraulic and water quality processes and which provides maximum environmental protection and community beneﬁts. • The development of ﬂexible, adaptive and costeffective approaches to urban water management in the light of increased economic and environmental risks and future climate change. • Enhanced frameworks and structures for collaborative stakeholder partnerships in urban drainage master planning with associated networks for increased public education and awareness of both surface and wastewater issues and their receiving water impacts.

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7.5.4 Maintenance and renewal of infrastructure • Protocols for the protection of surface and ground water drinking supplies from sewer exﬁltration, illicit connections and cost-effective control approaches for inﬁltration into sewers. Tighter regulatory programmes of measures (PoMs) for speciﬁc urban land use activities also need to be identiﬁed and implemented such as public vehicle washing, garden pesticide/herbicide and fertilizer applications, car-jet washing, used car oil disposal, etc. • Development of national standards for the efﬁcient design and operation of urban drainage systems in both environmental and economic terms which takes into full consideration structural life expectancies and discount rates as well as community beneﬁts. • The assessment of alternative modes of infrastructure ownership, asset and operational management as well as service provision, user service fees (including stormwater taxes/tariffs) and establishment of cross-stakeholder drainage utilities and agencies. Nevertheless it must be stressed that whilst drainage professionals and water resource managers are still held to be primarily responsible for identifying and implementing sustainable drainage solutions, they have only limited control over urban land use activities or chemical and other material inputs into the drainage system. It may thus be inappropriate or even naive to expect the emergence of ‘sustainable drainage’ without the concept of sustainable cities being ﬁrst developed. In addition, the rehabilitation and maintenance of urban drainage infrastructure is chronically underfunded and is unlikely in the future to be highly ranked in the competition for national funding resources. The major objectives for future urban drainage will remain public hygiene, ﬂood protection and environmental enhancement although the emphasis in developed countries is still ﬁrmly placed to date on primary ﬂood protection with secondary pollution control.

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Adams, B.J. and Papa, F. (2000) Urban Stormwater Management Planning with Analytical Probabilistic Models. John Wiley & Sons, Chichester, UK. Arnberg-Neilsen, K., Hvitved-Jacobsen, T., Johansen, B.N. et al. (2000) Stormwater Concentrations in Foul Sewers. Milloproject 532, Danish EPA. Arsov, R., Marsalek, J., Watt, E. and Zemen, E. (eds) (2003) Urban Water Management. NATO Science Series. Springer, Paris. Ashley, R.M., Bertrand-Krajewski, J.L., HvitvedJacobsen, T. and Verbanck, M. (2004) Solids in Sewers. Scientiﬁc & Technical Report No. 14. IWA Publishing, London. Auckland Regional Council (2003) Stormwater Management Devices: design guidelines manual. TP10. Auckland Regional Council, Auckland. Butler, D. (2004) Urban water: future trends and issues. In: Maksimovic, C. (ed.), Hydrology: science and practice for the 21st century, vol. II. Proceedings BHS National Conference, July 2004. Imperial College, London. IAHS Press, Wallingford. Oxon. UK, pp. 233–247. Butler, D. and Davies, J.W. (2000) Urban Drainage. E. & F.N. Spon, London. Campbell, N., D’Arcy, B.J., Frost, A., Novotny, V. and Sanson, A. (eds) (2005) Diffuse Pollution: an introduction to the problems and solutions. IWA Publishing, London. Chocat, B., Ashley, R., Marsalek, J. et al. (2004) Urban drainage: out-of-sight-out-of-mind? In: Proceedings of the 5th International Conference on Sustainable Techniques and Strategies in Urban Water Management (NOVATECH June 6-10 2004), Lyon, France, pp. 1659–1690. D’Arcy, J.B., Ellis, J.B., Ferrier, R.C., Jenkins, A. and Dils, R. (eds) (2000) Diffuse Pollution Impacts: the environmental and economic impacts of diffuse pollution in the UK. Terence Dalton Publishing (CIWEM), Lavenham. Department for Environment, Food and Rural Affairs (2008) Improving Surface Water Drainage: consultation document on the Government’s future water strategy. Defra, London. Deutsch, J.C. and Hemain, J.C. (1984) Main results of the French National Programme of Urban Runoff Quality Measurement. In: Proceedings of the 3rd International Conference on Urban Storm Drainage, Chalmers University, Gothenburg, Sweden, pp. 939–946. Dodson, R.D. (1999) Stormwater Pollution Control, 2nd edn. McGraw Hill, New York.

Ellis, J.B. (1986) Pollutional aspects of urban runoff. In: Torno, H., Marsalek, J. and Desbordes, M. (eds), Urban Runoff Pollution, NATO Technical Series. Springer-Verlag, Berlin, pp. 1–34. Ellis, J.B. (2004) Bacterial sources, pathways and management strategies for urban runoff. Journal of Environmental Planning and Management, 47, 941–956. Ellis, J.B. (2006) Pharmaceutical and Personal Care Products (PPCPs) in urban receiving waters. Environmental Pollution, 144, 184–189. Ellis, J.B. and Crabtree, R.W. (1999) Organizational issues and policy directions for urban pollution management. In: Trudgill, S.T., Walling, D.E. and Webb (eds), Water Quality; Processes and Policy. John Wiley & Sons, Chichester, pp. 357–364. Ellis, J.B. and Hvitved-Jacobsen, T. (1996) Urban drainage impacts on receiving waters. Journal of Hydraulic Research, 34, 771–783. Ellis, J.B. and Mitchell, G. (2006) Urban diffuse pollution: key management issues for the Water Framework Directive. Water and Environmental Journal, 20, 19–26. Ellis, J.B., Shutes, R.B.E. and Revitt, D.M. (2003) Guidance Manual for Constructed Wetlands. Environment Agency R&D Technical Report P2-159/TR2. Urban Pollution Research Centre, Middlesex University, London/Environment Agency, Bristol. Ellis, J.B., Revitt, D.M., Blackwood, D.J. and Gilmour, D.J. (2004a) Leaky sewers: assessing the hydrology and impact of exﬁltration in sewers. In: Maksimovic, C. (ed.), Hydrology: science and practice for the 21st century, vol. II. Proceedings BHS National Conference, July 2004. Imperial College London. IAHS Press, Wallingford, Oxon, pp. 266–271. Ellis, J.B., Revitt, D.M. and Scholes, L. (2004b) Sustainable approaches for urban development and drainage in the 21st century. Journal of the Institute of Civil Engineers (Municipal Engineer), 157 (ME4), 245–250. Ellis, J.B., Marsalek, J. and Chocat, B. (2005) Urban water quality In: Anderson, M.G. (ed.), Encyclopaedia of Hydrological Sciences. John Wiley & Sons, London, pp. 1479–1491. Ellis, J.B., Revitt, D.M., Scholes, L. and Shutes, R.E.B. (2008) Review of Best Practice Guidelines for Stormwater Management. Urban Pollution Research Centre, Middlesex University, London. Deliverable WP2.1, EU 5th Framework SWITCH Programme (www.switchurbanwater.eu)

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Pitt, M. (2007) Learning Lessons from the 2007 Floods: interim report. The Cabinet Ofﬁce, London. Portland Bureau of Environmental Services (2004) Stormwater Management Manual. Environmental Services City of Portland, Oregon. Richards, L. (2001) Protecting Water Resources with Higher-Density Development. Report 231R06001. US Environment Protection Agency, Ofﬁce of Water, Washington DC. Ryan, M.A. (ed.) (2003) The Clean Water Act Handbook, 2nd edition. Section of Environment, Energy and Resources, American Bar Association, Chicago. Schueler, T.R. (1987) Controlling Urban Runoff: a practical manual for planning and designing urban BMPs. Metropolitan Washington Council of Governments, Washington DC. United States Environmental Protection Agency (1983) Results of the Nationwide Urban Runoff Program, vol. 1. Final report PB84-185552. Water Planning Division, USEPA, Washington DC. United States Environmental Protection Agency (1993) Guidance Specifying Management Measures for Sources of Non-point Pollution in Coastal Waters. Ofﬁce of Water, USEPA, Washington DC. United States Environmental Protection Agency (2004) Impacts and Control of CSO and SSOs. Report to Congress EPA 833-R-04-001. Ofﬁce of Water, USEPA, Washington DC. Wastewater Planners Users Group (2004) Urban Rainfall and Runoff. Report of Wastewater Planning Users Group (WaPUG). Workshop held on 30 April 2004, Coleshill, Birmingham. Foundation for Water Research, Marlow Bucks., UK. Water Research Council (1991) Sewer Quality Archive Data. Report FR0203. Foundation for Water Research, Medmenham, UK. Water Research Centre (2001) Sewerage Rehabilitation Manual (SRM), 4th edition (version 2). Water Research Centre, Swindon, Wilts. Wong, T. (ed.) (2005) Australian Runoff Quality: a guide to water sensitive urban design. Institution of Engineers, Barton, Australia. Woods-Ballard, B., Kellagher, R., Martin, P., Jefferies, C., Bray, R. and Shaffer, P. (2007) The SUDS Manual. RP697. Construction Industry Research & Information Association (CIRIA), London. Zoppou, C. (2001) Review of urban stormwater models. Environmental Modelling and Software, 16, 195–231.

Image facing chapter title page: Courtesy of the Centre for Ecology and Hydrology.

Catchment to Coast Systems – Managing Microbial Pollutants for Bathing and Shellﬁsh Harvesting Waters

8

D A V I D K A Y 1, A D R I A N M C D O N A L D 2, C A R L S T A P L E T O N 1, M A R K W Y E R 1 A N D JOHN CROWTHER3 1 2

Centre for Research into Environment and Health, Aberystwyth, Wales, UK Faculty of Environment, School of Geography, University of Leeds, Leeds, UK 3 Centre for Research into Environment and Health, Lampeter, Wales, UK

8.1

Introduction

Catchment management impacts on the quality of coastal and estuarine waters through pollutant loadings carried from catchment systems to nearshore waters. The principal pollutants of interest are (i) the nutrient parameters and (ii) the microbial pathogens impacting on shellﬁsh harvesting and bathing waters. The nutrients (principally nitrogen in the context of marine ecosystems) are fully utilized in ‘natural’ pristine marine aquatic environments over annual cycles and their availability normally ‘limits’ algal growth after early season growth (Bowen et al. 2007; Conley et al. 2007; Philippart et al. 2007; Saraiva et al. 2007; Savchuk and Wulff 2007). Where land runoff is contaminated with nutrients from, for example, farming systems or human sewage efﬂuents, this can ‘over-fertilize’ nearshore marine waters producing high primary productivity and potentially unbalancing the natural ecological system to a ‘eutrophic’ condition. This is characterized by an Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

excess of primary producers (algae) which cause problems when they begin to breakdown in the water, a process which can lead to removal of dissolved oxygen in localized areas and the generation of algal toxins. The latter are associated with paralytic and diarrhoeic shellﬁsh poisoning which can present a signiﬁcant health risk to the human population (Pavela-Vrancic et al. 2002; Vale and de Sampayo 2002; Economou et al. 2007). Microbiological ﬂuxes to nearshore waters principally derive from agriculture activities such as livestock farming, and from human sewage disposal either directly to the coast or into rivers discharging to sea. There are other, probably more minor, sources of microbial contamination from wildlife excretion, faeces of domestic animals and pets, boat and ship discharges and industrial efﬂuents from pharmaceutical, food processing and paper plant discharges (World Health Organization 2003). The principal danger from microbes in coastal waters occurs where the microbial ﬂora contains species which are pathogenic (i.e. cause illness) in humans (Prüss 1998). The pathogens are passed out of the human body in faecal wastes discharged by individuals with symptomatic or asymptomatic

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infection. Contact with nearshore waters by bathing and other types of high contact recreation or consumption of shellﬁsh can provide a means of ingesting pathogens, thus facilitating the ‘chain of infection’ for faecal-oral illnesses such as viral gastroenteritis. This then initiates further pathogen production inside the human alimentary canal (Kay et al. 2004; Wade et al. 2006; Kay and Fawell 2007). 8.2 Historical Background Both nutrient and microbial pollution have caused signiﬁcant problems and major international initiatives to control their impacts can be seen in the OSPAR Convention covering the north-east Atlantic area (OSPAR 2007) and international directives such as the EU Urban Waste Water Treatment Directive (Stapleton et al. 2000) targeted at nutrient control in coastal waters. Regulation of bacterial pollution is a more recent development which has been largely driven by health concerns associated with bathing and shellﬁsh harvesting waters (Council of the European Communities 1976; United States Environmental Protection Agency 1986; World Health Organization 2003). The key difference between nutrient and microbiological parameters is that regulation of the former has always adopted a catchment or drainage basin based approach involving integrated control of point and diffuse sources (Fig. 8.1). By contrast, regulation of microbiological contamination of coastal waters has, until very recently, tended to focus on anthropogenic point source discharges mainly derived from human sewage systems. Control strategies to achieve water quality targets in bathing and shellﬁsh waters have depended on extremely costly engineering interventions to enhance treatment by, for example, storing storm ﬂows to reduce intermittent storm discharges and disinfection of treated efﬂuents with UV light, chemical dosing and/or microﬁltration ( Tessele et al. 2005; Neto et al. 2006; Zanetti et al. 2006). Deployed in isolation, these ‘technical ﬁxes’ have proven ineffective in achieving microbiological water quality

standards in many coastal waters because the nonhuman contribution remains unaltered and this is generally sufﬁcient to cause ‘non-compliance’ (EU) or ‘impairment’ (US) of water quality (Kay et al. 2006b; Kay et al. in press). In response, both the EU and US administrations have enacted catchment-based legislation which is designed to address water quality deterioration through integrated management of diffuse and point sources of pollution. In the USA, this is enshrined in the Federal Water Pollution Control Act (United States Environmental Protection Agency 2005) and the total maximum daily loads (TMDLs) concept (United States Environmental Protection Agency 2005; Kay et al. 2006b). The latter is a formal plan agreed after consultation with stakeholders to improve ‘impaired’ waters, i.e. those which fail standards in force. Section 303(d) of the Federal Water Pollution Control Water Act requires that states identify ‘impaired’ water bodies that do not meet deﬁned water quality standards. The TMDL process investigates these water quality problems and designs actions in consultation with stakeholders to effect remediation. The EU Water Framework Directive (WFD) (Council of the European Communities 2000) has a direct parallel of this process (Kay et al. 2006a) whereby member states are required to manage both point and diffuse sources of pollution to achieve ‘good’ ecological status and ‘good’ water quality by 2015. The chemical and microbiological status required of ‘good’ water quality is speciﬁed in daughter Directives. This is illustrated in the text of the 2006 Bathing Water Directive (European Community 2006) where a beach to be classiﬁed as ‘good’ must have a 95th percentile intestinal enterococci concentration of less than 200 per 100 ml. Member states are required to achieve this standard through implementation of the provision’s of the Water Framework Directive (Council of the European Communities 2000). The approach represents a radical change from traditional point source efﬂuent quality regulation towards ambient water quality control at the point where the water is used for ecosystem maintenance, water supply, recreation or ﬁsheries. Whilst correction of poor efﬂuent

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Catchment to Coast Systems Nitrogen

Atmospheric deposition and fixation Atmospheric deposition and fixation

Combustion

Mariculture Non agriAmmonia cultural land volatilization

Fodder

Interflow

Leaching Storage in soil and aquifer Removal in soil and aquifer Removal in groundwater

Combustion

Sludge Industry Freshwater Commercial and fish farm manure fertillizer

Sewage treatment plant

Soil erosion Surface runoff Artificial drains

Upper and deeper groundwater

Stormwater runoff Retention in freshwater (streams, rivers, lakes, reservoirs) Denitrification

Urban areas

Combustion

Households not connected to public sewage

Bank erosion

Leaky sewers

Storage in sediment

Fig. 8.1 Nitrogen sources and pathways, diffuse sources to the left and point sources to the right. (Source: OSPAR 2007.)

quality at point source discharges can be addressed through traditional ‘engineered’ solutions at treatment plants, maintenance of ambient water quality criteria, or good ecological status, requires integrated river basin management (IRBM) involving a much wider community of stakeholders covering industrial dischargers, the farming community and land managers. Consultation with stakeholders and the wider public is required by the WFD which presents further challenges for those involved in implementation. Engineers, social scientists and environmental professionals throughout Europe, therefore, face new challenges and opportunities

as the implications of the new approach become apparent. The WFD requires designated agencies in member states to identify pressures and impacts on water bodies (Article 5) and then to design a ‘programme of measures’ to achieve ‘good’ water quality and ‘ecological status’ within all EU drainage basins (Article 11). This directly parallels the Clean Water Act TMDL process and applies principles ﬁrst explored by the National Rivers Authority in the UK (National Rivers Authority 1991). Catchment nutrient ﬂuxes are addressed elsewhere in this book (Chapter 5). This chapter focuses on the microbial component. It seeks to

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explain how the emerging ‘catchment’ paradigm is being used to understand, quantify, mange and regulate coastal water quality to achieve perhaps the most ‘health-related’ environmental standards used today covering human exposures to bathing and shellﬁsh harvesting waters. The emerging catchment paradigm has generated a demand for the development of new tools and approaches and has brought into question the adequacy of historical data resources to underpin the management questions raised by the demands of integrated catchment management. Stated simply, these new requirements of the science and policy communities are as follows: 1 Quantiﬁcation of the ﬂux of microbial parameters from catchment systems to coastal waters. 2 Microbial source apportionment within complex catchment systems. 3 Assessment of the compliance implications for bathing and shellﬁsh harvesting waters. 4 Cost–beneﬁt assessment of the remediation options available, both farm-based best management practices (BMPs) and larger point source control strategies such as raw sewage storage schemes and treated efﬂuent disinfection. 5 Design of amended farm support instruments to encourage implementation of appropriate BMPs at the farm scale. 6 Modelling tools able to bridge the scale-divide (chasm) between individual ﬁeld-scale BMPs and catchment ﬂux assessments, thus to inform the policy community on the ‘compliance’ implications of amended farm support mechanisms. 7 New research and policy agendas driven in part: (a) by novel management tools of real-time risk assessment and modelling; (b) rapid methods for microbial enumeration; and (c) environmental forensic tools for source tracking of faecal indicator organisms and pathogens.

8.3 Current Perspectives Internationally, the microbial compliance parameters are generally indicators of faecal contamina-

tion rather than pathogenic bacteria and viruses. These include organisms such as coliform bacteria, Escherichia coli (E. coli), and intestinal enterococci. These are currently measured by ‘culture’ methods, i.e. they are grown on selective media. This takes time and, generally, microbial enumerations become available perhaps 24–48 hours after the sample is delivered to the laboratory. Additionally, the organisms are not ‘conservative’ parameters, their concentrations tend to reduce in the environment after they are ‘voided’ in faeces from contributing animals. The reason for this is that they have evolved to exist in the alimentary canal which: (i) is dark; (ii) has low oxygen tension; (iii) is approximately 37 °C in the human; and (iv) has high nutrient status. Thus, in environmental waters, or soil, they are effectively starving but have a slower metabolic rate than in the gut. Hence, they will die more slowly if it is cold, wet and dark and reduce more quickly in warm water with high light penetration. These considerations are also major factors in ensuring scientiﬁcally credible sample acquisition, storage and the time between sample acquisition and laboratory analysis. In general, this means aseptic hand sampling (i.e. not autosamplers), refrigerated storage and a maximum period of 24 hours between sampling and analysis is required by regulatory agencies (Standing Committee of Analysts 2002). In some tropical soil environments re-growth has been observed (Fujioka et al. 1999; Byappanahalli and Fujioka 2004). Additionally, the faecal indicator organisms (FIOs) used for recreational and shellﬁsh compliance assessment are know to re-grow to high concentrations in industrial waste waters which are warm with high organic carbon content (Byappanahalli and Fujioka 2004; Beauchamp et al. 2006; Barros et al. 2007). These factors are important considerations in the interpretation of FIO data. Their utility derives from the rather simplistic assumptions that: 1 They are always excreted from the gut of humans and other warm blooded animals: this is certainly true. 2 All animal sources present a similar risk of pathogen presence:

Catchment to Coast Systems this is patently not true in the case of human speciﬁc viruses and/or zoonotic (i.e. animal derived) pathogens. 3 Where there is disease in the contributing population (i.e. symptomatic and asymptomatic), they will index the risk of pathogen presence and concentration which, probably, decay in the environment at similar rates to the associated enteric faecal indicators: again this is patently untrue in the case of viral parameters and pathogens which produce resistant spores or oocysts as is the case for Cryptosporidium spp. 4 The faecal indicators will not grow outside the alimentary canal: which is again questioned particularly for semi-tropical environments where growth of faecal indicators in soil has been demonstrated (Fujioka et al. 1999). In fact, farm, wild and human contributors of faecal indicator bacteria provide very different pathogen loadings, generally the viruses of interest, e.g. norovirus, will derive principally from human faeces but many other human pathogens are zoonotic (i.e. mammalian and/or avian) in origin, e.g. Cryptosporidium spp., Giardia, spp. and Campylobacter spp. In addition, farm animals contribute, perhaps, the dominant FIO loading in livestock farming areas and non-faecal catchment sources are, indeed, proven. Thus, any direct health-risk-quantiﬁcation derived FIO quantiﬁcation should always be interpreted with some caution., FIOs, however, do represent the most useful and accepted regulatory tool for ensuring health protection of bathers and shellﬁsh consumers coming into contact with or consuming products grown in environmental waters world-wide. FIO data available for policy interpretation, therefore, reﬂect the environment, at best, 48 hours prior to the data being available and its interpretation is not simple. A further complication derives from the highly episodic nature of FIO pollution in catchment systems (DaviesColley et al. 2004; Jamieson et al. 2004; Wilkinson et al. 2006; Kay et al. 2007b; Oliver et al. 2007), (Fig. 8.2). This is commonly characterized by

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Fig. 8.2 Changes in faecal indicator organism (in this case Escherichia coli) concentration and turbidity in a small New Zealand stream during two high ﬂows draining an area of pastoral agriculture. (Source: Davies-Colley et al. 2004.)

∼3 log10 orders change in concentration during rainfall-induced storm events when ﬂow may increase in small catchment streams by >1 log10 order. The net effect of this pattern is to produce, perhaps, >97% of the FIO ﬂux from rural livestock farming areas in the very short episodes (generally <10% of the ﬂow sequence) when streams exhibit elevated ﬂows caused by rainfall events. It should be noted, however, that a different pattern is evident where human point sources of pollution, from sewage treatment plants in urban areas, dominate the total catchment loading and this is addressed below (Wyer et al. 2003; Wither et al. 2005; Stapleton et al. 2008).

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The combined effect of these characteristics of FIO data is that the policy community rarely has real-time information relevant to ‘current’ levels of compliance or associated risk in the environment. Furthermore, routine monitoring programmes which have generated the historical data archives available for analysis rarely or never target episodic high ﬂow events and, thus they will under-represent such conditions in the data archive which will, as a result, systematically underestimate both total catchment ﬂux of FIOs and, more importantly, the resultant risk at relevant receiving waters. For these reasons, much of the available archive data, worldwide, are of limited utility in addressing the new agendas and requirements. Policy agendas in the UK, principally compliance with EU bathing water standards, have driven a number of microbial source apportionment studies (Fig. 8.3) and these have generated an approach to quantitative microbial source apportionment (QMSA) which can be characterized as follows (Wyer et al. 1996, 1998b, 2001; Stapleton et al. 2007; Kay et al. 2008a): 1 Base-ﬂow and high-ﬂow concentrations are characterized at the catchment outlet and in feeder streams draining key sub-catchments, this is done over a ﬁxed period (e.g. the bathing season) by opportunistic sampling of speciﬁed locations. 2 The arithmetic mean log10 statistic is used to characterize the FIO concentration. 3 Flow is estimated by measurement or modelling at each of the ﬂux assessment locations. 4 The stream ﬂow pattern for the period is split into low and high ﬂow conditions by detailed analysis of hydrograph records (e.g. hourly time series). 5 The appropriate arithmetic mean log10 concentration is applied to each site and ﬂow condition. 6 This is used to generate a ﬂux microbial ﬂux time series from each location in the catchment discharging to the receiving water. 7 This is represented graphically as pie charts which clearly separate high and low ﬂow conditions (Fig. 8.4) and/or proportionate temporal ﬂux

diagrams (Fig. 8.5), in this case representing the ﬂux of faecal indicators to the Ayrshire coast of Scotland, UK (Kay et al. 2008a). This description relates speciﬁcally to surface water stream and river ﬂux assessment. It is clearly essential, however, to examine the corresponding ﬂux of FIOs from the sewage system which comprises both treated efﬂuents and untreated (often termed ‘raw’) efﬂuent. Untreated efﬂuent can ‘spill’ to the environment (i.e. receiving water) at times of plant breakdown (e.g. a pumping station shutdown) or when ‘combined’ sewerage systems are receiving surface water drainage as well as foul sewage above the ﬂow capacity of the system, creating the need to discharge the excess in a pre-determined and relatively safe manner (i.e. to prevent the sewerage system discharging through domestic sanitary systems into properties) through combined sewer overﬂows (CSOs: see Chapter 6 on efﬂuent management). In addition to these spills of diluted but raw sewage efﬂuent QMSA studies should also quantify the ﬂux from treated efﬂuents. Here, FIO concentrations can vary greatly between both ﬂow conditions and treatment types (Kay et al. 2008c). The ﬂow varies in response to rainfall even within the treatment plant. For ‘combined’ sewerage systems most plants will be designed to treat efﬂuent through the plant until the discharge reaches perhaps three to ﬁve times the ‘dry weather ﬂow’ (DWF). Beyond this ﬂow the efﬂuent will be either stored (e.g. in storm tanks) for later treatment where such infrastructure exists, or it will be allowed to overﬂow to the nearest river or other environmental receiving water. In a modern well designed combined sewerage system, most rainfall events will cause a variation in the ﬂow through the plant between the DWF and the plant’s treatment capacity, thus limiting the spill frequency of the sewerage system. The reduced retention time caused by higher ﬂows between DWF and three to ﬁve times DWF, however, can cause differences in the FIO concentration in the ﬁnal efﬂuent discharged from the plant (Kay et al. 2008c). Because of this, it is necessary to characterize dry and wet weather FIO concentrations in all point

Catchment to Coast Systems sources of ﬁnal (i.e. treated) efﬂuent as well as the quality of the spilled efﬂuents. The ﬁnal efﬂuents present sampling challenges similar to those experienced at riverine sites but the spills from CSOs and storm tank overﬂows (STOs) are notoriously difﬁcult to sample because of the highly intermittent nature, often difﬁcult access, and very short periods of operation which presents particular challenges for aseptic hand sampling. A surprising information gap in the peer reviewed literature is information on the FIO concentrations produced by different levels of sewage treatment. The historical reason for this is that modern sewage treatment systems in the developed nations were designed to achieve standards ﬁrst established in the UK by the 1912 Royal Commission on Sewage Disposal. This was established principally to address low oxygen tension, ﬁshery and aesthetic problems in UK rivers. The principal efﬂuent standards established therefore related to biochemical oxygen demand (BOD mg L−1), suspended solids (SS mg L−1) and ammoniacal nitrogen (NH4-N mg L−1). Because these related to treated sewage efﬂuents they have traditionally been termed the ‘sanitary parameters’. This is, in semantic terms, misleading because they have no direct sanitary, or public health, signiﬁcance. Microbial standards for FIOs in sewage efﬂuent are generally rare meaning that regulatory monitoring programmes of sewage efﬂuents rarely test for these parameters leading to a poor data resource. This becomes an increasingly signiﬁcant gap as we seek to address the WFD and US Clean Water Act requirements for QMSA. Kay et al. (2008c), however, have reported UK FIO concentrations in high-ﬂow and low-ﬂow efﬂuents from a range of treatment types and in storm water discharges (CSOs and STOs) using data acquired from 12 locations (Fig. 8.3). The sample numbers for each treatment type and the resultant descriptive statistics (Tables 8.1 and 8.2, respectively) were made available in the peer reviewed literature to allow FIO concentration estimates for domestic efﬂuent point source discharges to be made. Empirical data were unavailable although, clearly, such data would be

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preferable particularly where the climate or sewerage infrastructure differed from that experienced in the British Isles which are characterized by a maritime west-European temperate climate and, generally, ‘combined’ sewerage systems serving both rural and urban areas. Clearly, integration of the infrastructure inputs into the pie chart (Fig. 8.4) or temporal ﬂux diagram (Fig. 8.5) requires discharge information, ideally as hourly time series, from the sewage treatment plant and for CSOs and STOs in the sewerage system. This is often available for sewage treatment plants where modern process control systems require data on the current plant loading (often termed ﬂow to full treatment or FFT) which is gathered in real time by rated ﬂume structures at the plant inlet and/ or outlet. Even where such instrumentation does not exist, knowledge of the population served by the plant and a conversion factor for that nation’s sewage generation per head (e.g. 185 L person−1 day−1 has been used in the UK) can be used to give a realistic estimate of FFT, along with an estimate of groundwater inﬁltration to the system. Very few storm spill discharges are measured, however, and this element is perhaps the most difﬁcult to quantify. For speciﬁc targeted investigations, ﬂow monitors can be installed at key sites and this may be less logistically difﬁcult than ﬁrst imagined. For example, Stapleton et al. (2007) installed CSO and STO monitors at only 15 sewerage systems in the 1500 km2 Ribble catchment in the UK, which has a population of over 5 million, yet they were able to characterize discharges making up over 95% of those from all the contributing works by selecting the most signiﬁcant locations with the assistance of stakeholders and a pre-investigation desk study (Wither et al. 2005).

8.4 New Scientiﬁc Insights The approach outlined has been used successfully to quantify microbial ﬂux from several sources impacting on recreational or shellﬁsh harvesting waters. This approach can address

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Fig. 8.3 Location of quantitative microbial source apportionment studies in the UK 1996–2007. (Source: Kay et al. 2007a.)

Catchment to Coast Systems Table 8.1

191

Types of sewage treatment/efﬂuent (source Kay et al. 2008a) Level of treatmenta Untreated sewage (69) Primary treatment (12)

Secondary treatment (67)

Tertiary treatment (14)

Speciﬁc efﬂuent typesa Crude sewage discharges (16) Storm sewage overﬂowsb (53) Primary settled sewage efﬂuent (7) Stored settled sewage efﬂuent (2) Settled septic tank efﬂuent (3) Trickling ﬁlter efﬂuent (38) Activated sludge efﬂuentc (17) Oxidation ditch efﬂuent (3) Trickling/sand ﬁlter efﬂuent (1) Rotating biological contactor efﬂuent (8) Reedbed/grass plot efﬂuent (6) Ultraviolet-disinfected efﬂuent (8)

a

Figures in brackets indicate number of different treatment plants sampled (numbers of valid enumerations (n) are shown in Table 8.2). Comprise treatment plant inlet overﬂows, stormwater retention tank overﬂows and combined sewer overﬂows (CSOs); high-ﬂow data only. c Includes deep-shaft activated sludge efﬂuent at one site. b

Fig. 8.4 Pie chart representation of base ﬂow and high ﬂow ﬂux of faecal coliform bacteria to bathing waters at Staithes, Yorkshire, UK.

questions concerning the relative magnitude of FIO ﬂuxes at speciﬁc times. If, for example, a bathing water or shellﬁsh harvesting area is experiencing elevated FIO concentrations after rainfall events, then the temporal ﬂux diagram (Fig. 8.5) can indicate the likely sources, i.e. it pro-

vides temporally referenced microbial source apportionment when constructed for a speciﬁc catchment. This provides, however, a complex tool requiring extensive empirical data acquisition and it does not, in isolation, offer scope for the use of remotely sensed data for QMSA inves-

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Table 8.2 Summary of faecal coliform organism concentrations (cfu 100 ml−1) for different treatment levels and individual types of sewage-related efﬂuents under different ﬂow conditions: geometric means (GMs), 95% conﬁdence intervals (CIs)a; and results of t-tests comparing base- and high-ﬂow GMs for each group and typeb; and (in footnote) results of t-tests comparing GMs for the two untreated discharge types and the two tertiarytreated efﬂuent types (source Kay et al. 2008a) Indicator organism Treatment levels and speciﬁc types Untreated Crude sewage discharges Storm sewage overﬂows Primary Primary settled sewage Stored settled sewage Settled septic tank Secondary Trickling ﬁlter Activated sludge Oxidation ditch Trickling/sand ﬁlter Rotating biological contactor Tertiary Reedbed/grass plotd Ultraviolet disinfectiond

Base ﬂow conditions: c

Geometric mean

252 252

1.7 × 107*(+) 1.7 × 107*(+)

n

High ﬂow conditions:

Lower 95% CI

Upper 95% CI

1.4 × 107 1.4 × 107

2.0 × 107 2.0 × 107

n

c

Geometric mean

Lower 95% CI

Upper 95% CI

282 79

2.8 × 106*(−) 3.5 × 106*(−)

2.3 × 106 2.6 × 106

3.2 × 106 4.7 × 106

203

2.5 × 106

2.0 × 106

2.9 × 106

4.6 × 106*(−) 5.7 × 106

2.1 × 106 –

1.0 × 107 –

1.0 × 107*(+) 1.8 × 107

8.4 × 106 1.4 × 107

1.3 × 107 2.1 × 107

14 8

25

5.6 × 106

3.2 × 106

9.7 × 106

1

8.0 × 105

–

–

42

7.2 × 106

4.4 × 106

1.1 × 107

5

4.8 × 106

–

–

3.3 × 105*(−) 4.3 × 105 2.8 × 105*(−) 2.0 × 105 2.1 × 105

2.9 × 105 3.6 × 105 2.2 × 105 1.1 × 105 9.0 × 104

3.7 × 105 5.0 × 105 3.5 × 105 3.7 × 105 6.0 × 105

5.0 × 105*(+) 5.5 × 105 5.1 × 105*(+) 5.6 × 105 1.3 × 105

3.7 × 105 3.8 × 105 3.1 × 105 – –

6.8 × 105 8.0 × 105 8.5 × 105 – –

80

1.6 × 105

1.1 × 105

2.3 × 105

2

6.7 × 105

–

–

179 71

1.3 × 103 1.3 × 104

7.5 × 102 5.4 × 103

2.2 × 103 3.4 × 104

8 2

9.1 × 102 1.5 × 104

– –

– –

108

2.8 × 102

1.7 × 102

4.4 × 102

6

3.6 × 102

–

–

127 60

864 477 261 35 11

184 76 93 5 8

CIs only reported where n ≥ 10. t-tests comparing low- and high-ﬂow GM concentrations only undertaken where n ≥ 10 for both sets of samples; only statistically signiﬁcant (p < 0.05) differences between base- and high-ﬂow GM concentrations are reported: indicated by *, with the higher GM being identiﬁed as *(+) and the lower value by *(−). c n indicates number of valid enumerations, which in some cases may be less than the actual number of samples. d t-tests comparing the GM concentrations between the two tertiary-treatment efﬂuent types show GM FC concentrations to be signiﬁcantly higher (p < 0.001) in reedbed/grass plot efﬂuents than efﬂuents from UV disinfection for base-ﬂow conditions (there are too few high-ﬂow samples for these tertiary efﬂuents for meaningful comparisons to be made for high-ﬂow GM concentrations). a

b

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Fig. 8.5 Temporal ﬂux plot representation of high and low ﬂux volumes and faecal coliform ﬂux for the Irvine Bay bathing waters in Ayrshire, Scotland, UK.

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tigations. In other areas of catchment science, considerable effort has been expended on examination and development of general principles and data from which export models could be developed. Practitioners have, for example examined the export of nutrients from different land use systems and quantiﬁed the coefﬁcients of nutrient export in terns of kg ha−1 yr−1, leading to the availability of such coefﬁcients in the peer reviewed literature. Given the much later development of catchment microbial dynamics as a central discipline, this resource is, again, unavailable for large regions with only UK data published to date in the peer reviewed literature (Anon 2006; Kay et al. 2008b). One exception is seen in Scotland where a consortium commissioned by the regulators and government to produce a screening tool as part of the preparation for WFD implementation reviewed a series of FIO modelling approaches. The screening tool is based on research carried out at the Scottish Agricultural College (SAC) (Ogden et al. 2001; Vinten et al. 2002; McGechan and Vinten 2003; Vinten et al. 2004). It provides what is termed a ‘smart dynamic export coefﬁcient’ approach. It is driven by data at a 1 km2 resolution and provides export estimates for each 1 km2 grid based on: 1 areas of farm impervious surfaces (i.e. yards and roof surface); 2 soil applied organic wastes; 3 livestock numbers or soil burden; 4 channel contribution related to ditch and small stream density in each cell; 5 a ﬂow-dependent mobility factor incorporating sub-surface and overland components; and 6 the probability density function of overlandand through-ﬂow in each 1 km2 cell. The Screening Tool authors note the lack of empirical ground truth data on riverine faecal indicator concentrations which would be needed to assess the predictive accuracy of this approach. They produce,however, annual FIO (i.e. E. coli) runoff loadings for each 1 km2 grid cell covering the whole of Scotland and Northern Ireland suggesting annual ranges in FIO export of between

<1 × 1013 cfu km−2 yr−1 and >1 × 1014 cfu km−2 yr1 (see European Community 2006). The only environmental FIO ground truth data available to the Screening Tool authors were the bathing beach compliance data collected as required by Directive 76/160/EEC (Council of the European Communities 1976). They sought to test the modelling approach by parametric correlation analysis between the number of samples (i.e. of the 20 samples collected each bathing season at each compliance point in the UK) achieving the Directive 76/160/EEC Guide value for faecal coliform (i.e. 100 cfu 100 ml−1) and the annual mean (i.e. arithmetic mean) value of the E. coli export from all 1 km2 grids which fell within the hydrological contributing catchment which was thought to affect speciﬁc coastal bathing water compliance locations. The FIO export models for 60 correlation pairs gave poor explained variance of the number of compliant samples with explained variance (r2) values ranging between 0.01 and 0.33 (Anon 2006). This is, perhaps, unsurprising given: (i) the unpredictable, and inherently dissimilar, nearshore dilution and transport effects which would be found at different coastal locations linking riverine inputs to the bathing water compliance point at which water quality is measured; (ii) the seasonal mismatch between the ‘summer’ compliance data (dependent) and the ‘annual’ loading (predictor) variables which were used in the correlation analysis; (iii) the probable right skew in the predictor variable; and (iv) the effect of using the mean loading for each 1 km2 grid to characterize catchment derived ﬂux to the bathing water (i.e. at the catchment outlet which may have a greater proportion of high FIO export land use than, for example, afforested headwater areas). UK empirical studies have suggested signiﬁcant statistical associations (p < 0.05) between land use and faecal indicator concentrations in catchment streams with credible explained variances (r2 = 0.5–0.7), particularly where land use is used to predict the key ‘high ﬂow’ concentration (Crowther et al. 2002, 2003). Kay et al. (2008b) have calculated catchment export coefﬁcients

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Catchment to Coast Systems Table 8.3

Catchments and data sources reported in present study (source Kay et al. 2008a) Number of subcatchmentsa

Land use datab

Catchment

Year

England River Leven/Crake Holland Brook River Ribble Staithes Beck Windermere (lake) inputs

2005 1998 2002 1995 1999

30 14 40 4 25

CEH2000/OS Field mapping/OS ITE1990/OS Field mapping/OS CEH 2000/OS

Scotland Brighouse Bay inputs Ettrick Bay inputs River Irvine/Garnock Killoch Burnc River Nairn Sandyhills Troon coastal inputs

2004 2004 1998 2004 2004 2004 2000

2 (2) 3 (2) 30 4 (3) 1 4 (4) 6

Estimated SE Field mapping + MLCMS SE SE SE Estimated

Wales Afond Nyfer Afond Ogwr Afond Rheidol/Ystwyth

1996 1997 1999

2 18 22

Field mapping/OS Field mapping/OS Field mapping/OS

a

Numbers of subcatchments used in summer/winter comparisons are shown in parentheses. Land use data sources: Estimated = estimates for the two key land use types: built-up land (from OS 1 : 50,000 maps) and improved pasture (from ﬁeld reconnaissance); Field mapping + MLCMS = land use mapping during study period of part of the catchment, supplemented by the 1988 Macaulay Land Cover Map of Scotland, calibrated through ﬁeld mapping; Field mapping/OS = land use mapping during study period, supplemented by Ordnance Survey 1 : 50,000 digital map information for built-up land and woodland; ITE1990/OS = Institute of Terrestrial Ecology Land Cover for 1990, calibrated using ground truth data from the ﬁve study areas in England Wales where ﬁeld mapping was undertaken, and supplemented by Ordnance Survey 1 : 50,000 digital map information for built-up land and woodland; CEH2000/OS = Centre for Ecology and Hydrology Land Cover Map for 2000, supplemented by Ordnance Survey 1 : 50,000 digital map information for built-up land and woodland; SE = land use data generated by Scottish Executive. c Killoch Burn is located within the headwaters of the River Irvine/Garnock catchment. d ‘Afon’ (Welsh) = ‘River’. b

and related these to land use for >200 subcatchments studied in the empirical investigations (Tables 8.3, 8.4). Given the extreme seasonality of FIO concentration in UK catchment streams (Rodgers et al. 2003; Kay et al. 2005a, 2007a), separate analyses are required for winter and summer predictions and for high-ﬂow and lowﬂow conditions. These export coefﬁcients include an ‘urban’ predictor but they could be adapted to provide just the ‘agricultural’ component of diffuse pollution ﬂux. These empirically based relationships between land use and FIO export/concentration have been

used to calibrate large catchment models of FIO ﬂux from the Ribble system in the UK which is the UK’s single sentinel catchment for WFD research investigations (Kay et al. 2005a). This study illustrates the practicalities of using satellite and digital map derived land use data for large catchment modelling (Fig. 8.6). In the Ribble context, the sub-catchment regression modelling was used to deﬁne subcatchments with residual outputs lower and higher than expected to facilitate objective choice of candidate areas for remedial attention by the UK environmental regulator (Fig. 8.7).

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Table 8.4 Summary of geometric mean faecal coliform organism export coefﬁcients (log10 cfu km−2 h−1) under base- and high-ﬂow conditions at the 205 sampling points and for various subsets, and results of paired, 1-tailed t-tests to establish whether there are signiﬁcant elevations at high ﬂow compared with base ﬂow (source Kay et al. 2008a) FIO Base ﬂow

High ﬂow

n

Geometric mean

Lower 95% CI

Upper 95% CI

Geometric meana

Lower 95% CI

Upper 95% CI

All subcatchments

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5.5 × 108

4.1 × 108

7.2 × 108

3.6 × 1010**

2.7 × 1010

4.8 × 1010

Degree of urbanizationb Urban Semi-urban Rural (<2.5% built-up land)

20 60 125

2.8 × 109 1.2 × 109 2.9 × 108

1.1 × 109 7.4 × 108 2.1 × 108

7.2 × 109 1.9 × 109 4.0 × 108

1.3 × 1011** 4.6 × 1010** 2.6 × 1010**

4.8 × 1010 2.5 × 1010 1.9 × 1010

3.6 × 1011 8.6 × 1010 3.5 × 1010

4.3 × 108 1.1 × 108 4.7 × 106 1.2 × 107 3.8 × 106

1.6 × 109 5.7 × 108 8.2 × 107 9.0 × 107 1.9 × 107

1.2 × 1011** 2.5 × 1010** 3.3 × 109** 3.6 × 109** 3.8 × 108**

6.5 × 1010 1.1 × 1010 1.3 × 109 1.3 × 109 1.3 × 108

2.2 × 1011 5.5 × 1010 8.8 × 109 9.7 × 109 1.1 × 109

Subcatchment land use

Rural subcatchments with different dominant land uses 15 ≥75% Improved pasture 8.3 × 108 13 ≥75% Rough grazing 2.5 × 108 6 ≥75% Woodland 2.0 × 107 13 ≥75% Rough grazing 3.3 × 107 6 ≥75% Woodland 8.5 × 106

Signiﬁcant elevations in export coefﬁcients at high ﬂow are indicated: ** p < 0.001. Degree of urbanization, categorized according to percentage built-up land: ‘Urban’ (≥10.0%), ‘Semi-urban’ (2.5–9.9%) and ‘Rural’ (<2.5%).

a

b

The signiﬁcant drawback of the catchment modelling approach, however, is its lack of deterministic process (Novotny 1999). This is essential if the modelling community is to provide the policy maker with the ability to address key ‘what if’ questions related to the implementation of speciﬁc BMPs designed to achieve compliance at waters ‘impaired’ by elevated FIO concentrations. In a recent review for the UK Government of the science available to underpin UK agricultural diffuse pollution control, Haygarth et al. (2005) noted the emerging policy imperative to understand and predict catchment microbial ﬂuxes and they described this area as ‘the challenge of the 21st century’. They also noted the lack of research and regulatory attention to the microbial pollution area when compared to, for example, the nutrient parameters (Heathwaite 2003; Heathwaite et al. 2005a, 2005b).

8.4.1 Early process-based deterministic modelling developments to underpin BMP design Process-based, deterministic, catchment models of faecal indicator ﬂux which link the hillslope phase to in-channel processes have not, to date, been produced. Individual processes have been studied which offer preliminary ﬁndings for model calibration. For example, Oliver et al. (2005, 2006) examined environmental die-off rates in agricultural and catchment matrices and studied the transfer of FIOs to surface waters from grassland. Bales et al. (1995) examined phage (a viral tracer) and FIO transport in an aquifer at Cape Cod, USA and reported higher phage attenuation than for the bacteria. The phage, however, was easily remobilized by enhanced ﬂow, which suggests continued viabil-

Catchment to Coast Systems

Fig. 8.6 Satellite derived land use information and sampling sites in the Ribble catchment, UK. (Source: Kay et al. 2005a)

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Fig. 8.7 Residuals analysis of higher and lower FIO export than predicted by the modelling outputs.

Catchment to Coast Systems ity whilst adsorbed to the sandy aquifer matrix. Sinton et al (2005) reported vertical microbial tracer transport from areas of slurry spreading to a 16.8-m-deep aquifer in New Zealand. The breakthrough curves suggested travel speeds of between 15.7 and 39.2 m h−1. The tracers were attenuated rapidly which was attributed to early exclusion from macropore ﬂow but signiﬁcant groundwater contamination was considered possible from animal waste applied to land. Byappanahalli et al. (2003) suggested that signiﬁcant environmental reservoirs of faecal coliform organisms in mid-west stream bed sediments existed indicating protracted survival and potential re-growth outside the alimentary canal within freshwater stream environments. Jamieson et al. (2003) noted the probable existence of a stream bed store of FIOs in a 1000-ha watershed in Ontario and noted a general pattern of exceedence of Canadian recreational water quality criteria in catchment streams draining the livestock farming and residential areas studied. McDonald et al. (1982), Wilkinson et al. (1995, 2006) and Muirhead et al. (2004) have used artiﬁcial releases of water to assess the in-channel sediment contribution to stream water concentrations. Bai and Lung (2005) adopted a similar approach but then applied a model based on Environmental Fluid Dynamics Code (EFDC) to model sedimentary and FIO transport. They suggest that the approach facilitates quantiﬁcation of the channel sediment store contribution and the catchment derived inputs at speciﬁc points, although the published model validation data appear much better for the sediment than the FIO parameters measured. Collins and Rutherford (2004) developed a catchment simulation model to predict coliform concentration in New Zealand streams draining livestock grazing areas. The model used daily livestock data and simulated surface and sub-surface FIO ﬂuxes as well as direct deposition at locations where livestock could access the stream channel. They note the uncertainty regarding a number of these processes and the sensitivity of the model to the location of faecal inputs to the catchment. A

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scenario analysis suggested that riparian buffer areas could be effective both through livestock exclusion and microbial attenuation. Land use impacts on catchment-scale FIO budgets have been reported by Fraser et al. (1998) who used a GIS-based sediment delivery model (SEDMOD) calibrated for 12 sub-watersheds of the Hudson River, New York. The sediment delivery model, together with a GIS layer describing livestock density, explained 50% of the variance in ‘average’ faecal coliform output. Storm ﬂow water quality, discharged to coastal waters, was also examined by Lewis et al. (2005) in a study of on-farm remedial measures (BMPs) implemented in California. Tong and Chen (2002) report an application of the USEPA BASINS model calibrated against 11 water quality parameters including faecal coliforms in Ohio. FIO monitoring data will commonly have runs of low ﬂow and low concentration interspersed with occasional high values caused by agricultural diffuse pollution and intermittent discharges from sewers. The authors report that: ‘In the data set for … and fecal coliform, there are a few outliers (in the form of extremely high values). These outliers as well as missing data and “zero” values were deleted from the data set.’ Information on the proportion of data deleted is not given, but modelling the episodic nature of FIO ﬂux would need to address the high values which are the periods of peak contamination and likely impact with respect to health risk (see section 8.1 above). In one of the rare ‘before and after’ catchmentscale studies designed to quantify the effect of a BMP (in this case cattle exclusion from catchment streams) on FIO ﬂux from a 56.7-ha drainage basin, Line (2003) reported data derived from a 7.5-year sampling period which suggested 65.9% and 57.0% reductions in faecal coliform and enterococci export, respectively. They also reported that the provision of an alternate water supply without fencing was not effective in producing FIO reduction, i.e. removing the need for stock access to streams for drinking purposes was not effective in actually preventing stock access and consequent voiding of faeces in the riparian

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zone (see also Shreeram and Mostaghimi 2002). In a 2-year UK investigation of FIO export through a period of de-stocking due to an outbreak of foot and mouth disease, Chalmers et al. (2005) and Sanders et al. (2005) report a surprisingly slow improvement in water quality following the most drastic BMP of >95% stock removal from the 254.6-ha Caldew catchment in Cumbria, UK. A study at Brighouse Bay in Scotland, UK, examined the effects of BMPs on water quality in catchment streams and at an adjacent bathing water beach. The principal BMP was stream bank fencing to create a riparian buffer strip (RBS), with associated provision of drinking troughs. Farm dirty water containment was also implemented. The stream water quality data suggested extreme seasonality, with the summer period having markedly higher FIO concentrations in catchment streams. Comparison with an unmodiﬁed adjacent control catchment, however, suggested a 66% reduction in E. coli summer high-ﬂow export coefﬁcient (in cfu m−2 h−1) with a parallel 81% reduction in intestinal enterococci export. Detailed monitoring through a rainfall event in the post-remediation period suggested that even this improvement would be insufﬁcient to guarantee compliance with the European Bathing Water Directive Directive 160/76/EEC (Dickson et al. 2005; Kay et al. 2006a). The separate effects of RBS and farm dirty water control have been addressed in a study of 60 monitored catchments in Scotland by Kay et al. (2005a). Here, signiﬁcant improvements were recorded in FIO ﬂux when compared to ‘control’ catchments but a relatively high intensity of ‘measures’ was required (i.e. >30% of stream bank length protected by RBSs). The regional (i.e. multi-catchment) scale sources of agricultural diffuse pollution impacting on bathing waters along the Ayrshire coast in Scotland has been assessed by Aitken et al. (2001). Following from this work, the impacts of faecal indicator ﬂuxes from the catchment hinterland was examined using three modelling strategies; the ﬁrst a soil transport model, the second a regression model and the third a more distributed catchment model (PAMIMO). The

regression model gave the best prediction of bathing water quality and the authors concluded that preventing surface runoff would prove most protective of bathing water quality (Vinten et al. 2004). 8.4.2 Current developments in microbial source tracking to inform catchment studies Microbial source tracking (MST) has been employed by Hyer and Moyer (2004) to inform TMDL studies in the USA and Pond et al. (2004) provide an overview of the potential for the source tracking methods currently available to contribute to FIO ﬂux source apportionment. These methods use either: (i) species and or subspecies of organisms thought to be associated with faecal matter from humans or deﬁned animal groups; or (ii) chemical markers indicative of human sewage. There is currently no single and deﬁnitive approach with which to identify exact proportions of human and animal derived FIOs, but this area is developing rapidly and may provide operationally useful data in the medium term. Where the latest molecular source tracking tools have been applied to a catchment where existing QMSA data were available (Stapleton et al. 2008), however, they have not provided any additional explanation. At this stage, the available tools are largely ‘qualitative’ and they can suggest animal or human origin for a surface water with relatively high contamination. As most practical questions relate to ‘mixed’ catchment systems, the information that the sample contains indications of both human and animal molecular markers, but in unknown or imprecise proportions, is the likely outcome, which is generally unsurprising and of limited operational utility. Until more ‘ground truth’ calibration studies have been conducted to assess this approach it cannot be considered a ‘replacement’ for QMSA, although the considerable current research efforts may yield more quantitative approaches in the medium term (Field and Samadpour 2007; Gawler et al. 2007).

Catchment to Coast Systems 8.4.3 Innovative regulatory approaches driving the management of microbial exposures in catchment systems There is a strong evidence-base linking numerical values for the microbiological compliance parameters with speciﬁc health risk outcomes (Cabelli et al. 1982; Kay et al. 1994; Fleisher et al. 1996; Wade et al. 2006; Wiedenmann et al. 2006). This evidence-base has been peer reviewed (Prüss 1998) and used by WHO, USEPA and the EU to establish microbiological standards for recreational waters world-wide (United States Environmental Protection Agency 1986; World Health Organization 2003; European Community 2006). In the WHO and EU standards the probabilistic exposure approach developed by Wyer et al. (1999b) and Kay et al. (2004) was utilized to derive the risk-based numerical FIO standards for regulation of bathing water quality. During the development of the WHO Guidelines (World Health Organization 1998) early ﬁndings of the UK QMSA studies were becoming available (Wyer et al. 1994, 1996, 1997, 1998a, 1999a). These data suggested that diffuse catchment sources of faecal indicator organisms from agricultural sources could easily produce non-compliance of bathing waters measured against the ‘relatively lax’ EU standards then in force, which could not claim a clear healthevidence basis (Council of the European Communities 1976). The early UK QMSA studies demonstrated that new tighter, and clearly health-evidence-based standards might not be achievable where the source of pollution was from the livestock industry. A further implication of this ﬁnding was that if the new water quality criteria, being developed by WHO, were adopted as regulatory limit values with legal force by the competent authorities worldwide, then traditional remediation options such as sewage treatment might have little or no impact and the only path to compliance might be drastic land use change in large rural areas where livestock farming was practised and where the catchment outlet streams impacted on bathing waters. This problem was addressed at a meeting of the

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WHO technical advisory group on bathing waters in Annapolis in 1999, hosted by the USEPA (World Health Organization 1999). This led to the Annapolis Protocol which forms one of the most innovative developments in environmental regulation seen since the EU WFD, the principles of which have been incorporated into the new EU Bathing Water Directive (European Community 2006). The Annapolis Protocol is based on the following observations and principles: 1 Gross pollution of bathing waters by human sewage from casual discharges, CSOs and STOs or treated efﬂuents is always unacceptable and should be controlled and regulated. 2 There will always be a ﬂux of FIOs from catchment systems to coastal waters even in the absence of any human sources (e.g. a sheep excretes nearly 10 times more FIOs per day than a person), assuming secondary biological processes with no disinfection (Table 8.2) treatment of human sewage and the normal attenuation produced by, for example, activated sludge treatment, then the catchment loading in one day of 100 sheep is approximately equivalent to 1000,000 people, other species common in catchment systems such as cows, ducks and gulls all excrete more FIOs per day than humans). 3 This catchment-derived ﬂux is highly episodic and driven by rainfall events reﬂected in stream high ﬂows impacting on bathing waters (Fig. 8.2). Thus, the risk period is often short, clearly deﬁned and predictable. 4 If this episodic pollution is animal derived, it is likely to present a lower health risk to the human population than would be evident if the FIOs came from sewage where more human pathogens would be expected. 5 There would, however, be the expectation that even animal-derived FIOs were associated with zoonotic pathogens excreted by livestock or wildlife which could cause infection in the human population. It would be inappropriate, therefore, simply to ignore this exposure or imply zero risk. 6 The appropriate regulatory response to this episodic pollution exposure was to:

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(a) predict the risk in real time; (b) develop a management system to inform the public and thus ensure ‘informed choice’ by potential users of recreational waters; (c) conduct QMSA investigations to ensure that the microbial sources were indeed nonhuman; and (d) develop policies over time to bring down the background pollution from all sources within feasible management and regulatory scenarios. In the subsequent published guidelines, the World Health Organization (2003) suggested that, given the implementation of the Annapolis principles of real-time prediction and a management system in place, regulatory samples taken at the times when the public were advised against recreational water contact would not be used in the numerical compliance assessment of the bathing water (in the WHO and EU standards this is a calculation of a 95th percentile of a data set for a speciﬁc beach or bathing area). Thus, given real-time prediction, public information and appropriate management, occasional high FIO concentrations can be ‘discounted’ for regulatory purposes. The effect of this approach is that tight health-evidence-based numerical limits can be applied to the bathing water which cover the period of exposure. When ‘natural’ episodic variability causes deviation from the desired water quality the public is advised of this risk and exposure is reduced or prevented during these short periods, thus public health is maintained, i.e. the principal function of the regulator is to ‘predict and protect’. This radical principle was carried forward into the 2006 European Bathing Water Directive (European Community 2006) in which up to 15% of regulatory samples could be discounted. This may seem a relatively minor concession, but the UK Government’s Regulatory Impact Assessment for the new Directive suggested that this element alone could reduce the UK compliance cost of the new Directive from >4000 × 106 UK£ (8000 × 106 US$) to almost zero (Department for Environment Food and Rural Affairs 2002) whilst maintaining the public health beneﬁts.

There is a developing literature on the tools available for real-time prediction of FIO concentrations in environmental waters (Crowther et al. 2001; McPhail and Stidson 2004; USGS 2006) and an operational advisory system, using real-time electronic signs, has been deployed in Scotland (UK) in preparation for the implementation of the 2006 Bathing Water Directive (EC 2006). Real-time regulation and management of any environmental quality parameter is rare but, if proven robust, it offers signiﬁcant advantages in protecting public health and avoiding the credibility gap which can open when the public ﬁrst comprehend that they or their children were exposed to polluted water in, perhaps, June when a beach is conﬁrmed to have been non-compliant with health-based standards in, perhaps, November.

8.5 Future Research Catchment outputs become coastal inputs in most nearshore environments which explains why estuaries, in their ‘natural’ state, have very high rates of net primary productivity caused by the nutrient and energetics ‘subsidies’ produced by ‘pollutants’ in the land runoff. These pollutants can become a problem where excess nutrients cause eutrophication and microbiological concentrations present a threat to human health and/or restrict human use of the environment by creating impairments or non-compliance with environmental regulation in force. Microbiological pollution is an emerging area of investigation with only a sparse ‘catchment’ literature compared to the nutrient parameters. However, the practitioners have, perhaps, been less constrained by: (i) historical approaches; (ii) data availability from harmonized monitoring programmes; and (iii) well worn regulatory paradigms. Many of the questions are similar for both communities and could be phrased: (a) what advice do we give to the farmer wishing to implement ﬁeld-scale BMPs to reduce pollutant ﬂux? (b) how do we model the impact of such ﬁeld scale advice at the large catchment scale? and (c) what

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Catchment to Coast Systems will be the effect on receiving environments, principally the nearshore zone of different scales of action withing contributing catchments? It is not evident that the more mature nutrient community is any nearer addressing these questions than the emerging new community of workers in the area of catchment microbial dynamics even after very considerable effort and expenditure over many years worldwide. The achievements of this new community include: 1 Novel real-time regulatory approaches, based on relatively simple prediction of ‘episodic’ pollutant elevations, have been developed in the last decade and are already enshrined in legislation. 2 Health-evidence-based regulatory standards have been agreed at the international level and now underpin legally binding EU standards. 3 Simple black box models predicting catchment ﬂux from satellite land use data have been demonstrated with credible explained variances. 4 Export coefﬁcient data are becoming available to underpin diffuse source modelling approaches which build on the black box land use modelling approach. 5 Process studies are underway worldwide which will assist in the parameterization of more deterministic modelling approaches. The microbial community, however, faces the same scale problems as other catchment researchers for the nutrient parameters (Heathwaite 2003; Heathwaite et al. 2005a, 2005b). The relative importance of catchment microbial dynamics is clearly demonstrated by the league table of US Clean Water Act TMDL investigations and a similar pattern is likely as the implications of the WFD dawn on the EU regulators and Governments. Thus, rapid progress is required and, where possible, harmonization and effort-sharing between the catchment science communities across artiﬁcial lines which have been deﬁned by parameters and disciplines. Therefore, what Haygarth et al. (2005) termed ‘the challenge of the twenty ﬁrst century’ will hopefully produce operationally useful scientiﬁc advances in a much shorted timescale than has been the case for other areas of catchment science.

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Edinburgh. http://www.scotland.gov.uk/Resource/ Doc/1057/0012161.pdf (last veriﬁed 3 October 2008). Economou, V., Papadopoulou, C., Brett, M. et al. (2007) Diarrheic shellﬁsh poisoning due to toxic mussel consumption: the ﬁrst recorded outbreak in Greece. Food Additives and Contaminants, 24, 297–305. European Community (2006) Directive 2006/7/EC of The European Parliament and of The Council of 15th February 2006 concerning the management of bathing water quality and repealing Directive 76/160/ EEC. Ofﬁcial Journal of the European Union, L64, 37–51. Field, K.G. and Samadpour, M. (2007) Fecal source tracking, the indicator paradigm, and managing water quality. Water Research, 41, 3517–3538. Fleisher, J.M., Kay, D., Salmon, R.L., Jones, F., Wyer, M.D. and Godfree, A.F. (1996) Marine waters contaminated with domestic sewage: nonenteric illnesses associated with bather exposure in the United Kingdom. American Journal of Public Health, 86, 1228–1234. Fraser, R.H., Barten, P.K. and Pinney, D.A.K. (1998) Predicting stream pathogen loading from livestock using a geographical information system-based delivery model. Journal of Environmental Quality, 27, 935–945. Fujioka, R., Sian-Denton, C., Borja, M., Castro, J. and Morphew, K. (1999) Soil: the environmental source of Escherichia coli and enterococci in Guam’s streams. Journal of Applied Microbiology, 85, 83S–89S. Gawler, A.H., Beecher, J.E., Brandao, J. et al. (2007) Validation of host-speciﬁc Bacteriodales 16S rRNA genes as markers to determine the origin of faecal pollution in Atlantic Rim countries of the European Union. Water Research, 41, 3780–3784. Haygarth, P., Granger, S., Chadwick, D., Sheperd, M. and Fogg, P. (2005) A Provisional Inventory of Diffuse Pollution Losses. ADAS Report for Defra, London. Heathwaite, A.L. (2003) Making process-based knowledge useable at the operational level: a framework for modelling diffuse pollution from agricultural land. Environmental Modelling and Software, 18, 753–760. Heathwaite, A.L., Dils, R.M., Liu, S. et al. (2005a) A tiered risk-based approach for predicting diffuse and point source phosphorus losses in agricultural areas. Science of the Total Environment, 344, 225–239. Heathwaite, A.L., Quinn, P.F. and Hewett, C.J.M. (2005b) Modelling and managing critical source areas of diffuse pollution from agricultural land using ﬂow

Catchment to Coast Systems connectivity simulation. Journal of Hydrology, 304, 446–461. Hyer, K.E. and Moyer, D.L. (2004) Enhancing fecal coliform total maximum daily load models through bacterial source tracking. Journal of the American Water Resources Association, 40, 1511–1526. Jamieson, R.C., Gordon, R.J., Tattrie, S.C. and Stratton, G.W. (2003) Sources and persistence of fecal coliform bacteria in a rural watershed. Water Quality Research Journal of Canada, 38, 33–47. Jamieson, R., Gordon, R., Joy, D. and Lee, H. (2004) Assessing microbial pollution of rural surface waters – a review of current watershed scale modeling approaches. Agricultural Water Management, 70, 1–17. Kay, D. and Fawell, J. (2007) Standards for Recreational Waters: ROC brieﬁng paper. Foundation for Water Research, Medmenham. Kay, D., Fleisher, J.M., Salmon, R.L. et al. (1994) Predicting the likelihood of gastroenteritis from sea bathing – results from randomized exposure. Lancet, 344, 905–909. Kay, D., Bartram, J., Pruss, A. et al. (2004) Derivation of numerical values for the World Health Organization guidelines for recreational waters. Water Research, 38, 1296–1304. Kay, D., Wilkinson, J., Crowther, J. et al. (2005a) Monitoring the Effectiveness of Field and Steading Measures to Reduce Diffuse Pollution from Agriculture to Bathing Waters in the Ettrick, Cessnock, Nairn and Sandyhills Catchments. Scottish Executive Rural Affairs Department Report ENV/7/4/04, March 2005, Edinburgh. Kay, D., Wyer, M., Crowther, J. et al. (2005b) Predicting faecal indicator ﬂuxes using digital land use data in the UK’s sentinel Water Framework Directive catchment: the Ribble study. Water Research, 39, 3967–3981. Kay, D., McDonald, A.T., Stapleton, C.M., Wyer, M.D. and Fewtrell, L. (2006a) The challenges of the water framework directive. Proceedings of the Institution of Civil Engineers. Water Management, 159, 58–64. Kay, D., Stapleton, C.M., Wyer, M.D., McDonald, A.T. and Crowther, J. (2006b) Total Maximum Daily Loads (TMDL). The USEPA approach to managing faecal indicator ﬂuxes to receiving waters: lessons for UK environmental regulation? In: Gairns, L., Crighton, C. and Jeffrey, B.E. (eds.), Agriculture and the Environment VI; Managing Rural Diffuse Pollution. Proceedings of the SAC/SEPA Biennial

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Prüss, A. (1998) Review of epidemiological studies on health effects from exposure to recreational water. International Journal of Epidemiology, 27, 1–9. Rodgers, P., Soulsby, C., Hunter, C. and Petry, J. (2003) Spatial and temporal bacterial quality of a lowland agricultural stream in north east Scotland. Science of the Total Environment, 314–316, 289–302. Royal Commission on Sewage Disposal (1912) Eighth Report. Standards and Tests for Sewage and Sewage Efﬂuent Discharging to Rivers and Streams. Eighth Report, Cd 6464. HMSO, London. Sanders, B.S., Anthony, S.G., Stapleton, C.M., Kay, D., Crowther, J. and Wilson, J.G.M. (2005) The Impact of De-Stocking on the Microbiological Quality of Rivers in the Caldew Catchment, vols 1, 2. Science Report SC020045/SR. Environment Agency, Bristol. Saraiva, S., Pina, P., Martins, F., Santos, M., Braunschweig, F. and Neves, R. (2007) Modelling the inﬂuence of nutrient loads on Portuguese estuaries. Hydrobiologia, 587, 5–18. Savchuk, O.P. and Wulff, F. (2007) Modeling the Baltic Sea eutrophication in a decision support system. Ambio, 36, 141–148. Shreeram, P.I. and Mostaghimi, S. (2002) A long-term watershed-scale evaluation of the impacts of animal waste BMPs on indicator bacteria concentrations. Journal of the American Water Resources Association, 38, 819–833. Sinton, L.W., Braithwaite, R.R., Hall, C.H., Pang, L., Close, M.E. and Noonan, M.J. (2005) Tracing the movement of irrigated efﬂuent into an alluvial gravel aquifer. Water Air and Soil Pollution, 166, 287–301. Standing Committee of Analysts (2002) The Microbiology of Recreational and Environmental Waters 2002. Methods for the Examination of Waters and Associated Materials. The Environment Agency, Bristol. Stapleton, C.M., Kay, D., Jackson, G.F. and Wyer, M.D. (2000) Estimated inorganic nutrient inputs to the coastal waters of Jersey from catchment and waste water sources. Water Research, 34, 787–796. Stapleton, C.M., Wyer, M.D., Kay, D. et al. (2007) Microbial source tracking: a forensic technique for microbial source identiﬁcation. Journal of Environmental Monitoring, 9, 427–439. Stapleton, C. M., Wyer, M. D., Crowther, J. et al. (2008) Quantitative catchment proﬁling to apportion faecal indicator organism budgets for the Ribble system, the UK’s sentinel drainage basin for Water Framework Directive research. Journal of Environmental Management, 87, 535–550.

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Image facing chapter title page: Courtesy of the Centre for Research into Environment and Health.

Irrigation Management in a Catchment Context

9

SHAHBAZ KHAN1 1

9.1

UNESCO Division of Water Sciences, UNESCO, Paris, France

Introduction

The global water cycle, land management and food security are intimately linked. The global food system has responded to the doubling of world population by more than doubling food production during the past 50 years. Many of the gains came from productivity growth (70%) and area expansion (10%), and irrigated agriculture has played a major role. About 40% of the global harvest comes from just 20% of croplands that are irrigated (Food and Agriculture Organization 2003; Khan et al. 2006). The expansion in irrigated agriculture and productivity growth during the Green Revolution period virtually shielded communities against episodes of hunger and famine. Alongside irrigation, development provided other beneﬁts such as: gains in employment and income; better nutrition, health and education; human capital formation; improvements in equity in favour of the poor, particularly those having good access to land and water resources; better access to related rural infrastructure and services (Narayanamoorthy and Hanjra 2006). The world’s population continues to grow rapidly but much of the potential land suited to agriculture has already been developed. Frontiers are being reached in crop productivity growth and Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

area expansion. Feeding a growing and wealthier population with more diversiﬁed diets poses signiﬁcant challenges for food security and environmental sustainability in the coming decades (Abdullah 2006). Producing more food requires more water; richer and more nutritious diets require even more water. The scope for expansion in cropped or irrigated area, as in the past decades, has become limited. Much of the additional food production must, therefore, come from the intensiﬁcation of land and water systems (Food and Agriculture Organization 2003; Khan et al. 2006). This would, however, exert unprecedented pressure on ecosystems that currently provide a range of beneﬁts to mankind including: food, ﬁbre, timber, fuels, climate regulation, biodiversity conservation and regulation of water ﬂows and quality. Even greater pressures would emanate from globalization of the world food supply system. In the half century following the Second World War, the irrigated area of the world tripled from 90 million ha to 270 million ha, an annual compound growth rate of over 2.5%. There is, however, wide variation among continents in the proportion of cropland irrigated, ranging from 30% in Asia to between 6% and 12% in other continents. About 60% of irrigated cropland is in Asia and nearly half that area is devoted to rice production. Sustainable irrigation management remains a daunting task (Khan et al. 2006). Irrigation induced water-logging and salinity reduce

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agricultural production and impose other economic, social and environmental costs (see Ragan et al. 1999; Wichelns 1999, 2002; Hussain et al. 2004; Houk et al. 2005). The costs are often not conﬁned to a paddock or private property but spread over a wider scale. There can be a range of off-farm impacts including damages to roads and infrastructure, loss of aesthetic values and reduced biodiversity (see Characklis et al. 2005). Other impacts include displacement of labour, out-migration, income disparity and overall reduction in food output. These off-site costs or externalities have important policy implications regarding land and water resource management and how taxes are deployed to mitigate salinity and create social beneﬁts for all. A range of complex natural resource management (NRM) issues are challenging traditional irrigation management approaches across Australia (Clarke et al. 2002; Proust 2003) and elsewhere. In many cases, the loss of agricultural productivity impacts resource-using industries, while the ﬁnancial costs of changing management practices and ensuring the future sustainability of land and water resources appear substantial. Restoring ecosystem health and maintaining biodiversity conservation would entail further costs. Appropriately targeted investments, such as those under the Australian National Water Security Plan, can eliminate many of the inefﬁciencies in water resource management in a cost-effective way. One requirement to eliminate inefﬁciencies in regional and on-farm water delivery systems would be to map the areas where inefﬁciencies exist (irrigation hotspots) and then target investments to those areas. This entails a whole-of-system diagnostic analysis of the water use to prepare a water balance sheet of net inﬂows and outﬂows across the system linked with a spatial account of water losses. Often, the information on spatial water use efﬁciencies does not exist or does not feed into the water policy process to guide investments. Such information, along with measures of physical and economic productivity of water on-farm and across uses, may help to identify real opportuni-

ties to ‘save’ water, and to tailor and prioritize investments to realize those savings. Turbulence in global ﬁnancial markets can easily cause impacts on the water economy, dampening investments. In water-scarce settings like Australia, where water resources are fully developed and ﬁnancial and storage capacities are not constrained, enhancing the efﬁciency of available water through more efﬁcient delivery systems and onfarm production technologies holds the key to resolving water scarcity issues. Signiﬁcant gains in on-farm water use efﬁciency (WUE) and water productivity are possible through appropriate interventions. These gains are often assumed, rather than identiﬁed, at various spatial scales and within irrigation systems. Without proper water accounting for the whole irrigation system, misguided investments to ‘save’ water can reduce return ﬂows and can be detrimental to the environment and other users (e.g. see Meijer et al. 2006). This chapter provides a generic framework for sustainable management and development of water resources in Australia, both to boost the productivity of land and water resources and to conserve the quality of the natural environment. Future investment scenarios, both to meet food demand and achieve strategic regional and global food security goals, are presented. This chapter concludes with an overview of emerging challenges and the approaches and investment needs to address those challenges head on.

9.2 Current Solutions 9.2.1 A conceptual framework for managing water and salt balance in an irrigation area Irrigation generates on-farm beneﬁts for irrigators but can also impose costs on third parties, including downstream communities and the environment. The issue can be modelled as a negative externality problem. The externalities exist in situations where the activity of one person affects

Irrigation Management in a Catchment Context Table 9.1

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Salt and water inﬂows and outﬂows in irrigation areas Salt balance

Salt inﬂows Dissolution of native deposits Surface runoff Ocean and wind spray Fertilizers Irrigation water Capillary rise from groundwater

Salt outﬂows MINUS

Leaching to low lying or deep drainage areas Uptake by vegetation Rainfall as insoluble compounds

EQUALS

Net change in salt content of the system

Water balance Water inﬂows Irrigation water applications Rainfall Capillary upﬂows

Water outﬂows

MINUS

Evapotranspiration Surface runoff Leakages/deep drainage Change in soil water content

or spills over onto another, without compensating the latter (Baumol and Oates 1993). In many irrigation areas, the return ﬂows containing dissolved salt and other pollutants are either directly shunted to streams or enter the unconﬁned underground and adjacent aquifers – a common property resource (Table 9.1). Farmers have an incentive to use more water to enhance production and hence often over-irrigate which exacerbates these issues. Society, on the other hand, has the opposite incentive to reduce water use to mitigate potential negative impacts on the environment (Luquet et al. 2005). Where the aquifer recharge exceeds the natural assimilative capacity over long periods, damage to the supportive capacity of the resource base becomes inevitable, compromising productivity. In terms of irrigation management for waterlogging and salinity mitigation, the sustainability goal would mean a net zero change in the groundwater table and salt balance. 9.2.2 Farm-scale management tool In any irrigated system, a paddock or farm remains the basic unit to make land and water management decisions – the decisions that determine net recharge at the farm level and add up

Net groundwater recharge EQUALS

to determine the net recharge at the zonal, vis-àvis the system, level. The biophysical, hydrological and economic information at paddock scale, however, is generally not available. Furthermore, the factors that underpin land and water management decisions at the farm level may not be uniform across farms, due to heterogeneity in resource endowments and personal preferences for certain farming types that may in turn impact net recharge on an individual farm. This calls for farm-speciﬁc information to address the net recharge and soil salinity issue head on. A uniﬁed approach across a farm, guided by a common vision for sustainable land and water management, would ensure a win–win outcome for all. This, nevertheless, requires farm-scale net recharge analysis. The SWAGMAN® Farm model affords such insights. This state-of-the-art hydroeconomic model clearly captures economic and environmental trade-offs in adopting different land and water management options at farm scale, to help farmers decide sustainable irrigation intensities, while optimizing proﬁts. The development of a paddock-scale net recharge metric is perhaps the novel contribution of this model. The management of paddock-scale net recharge targets and optimization of proﬁts achieves a win–win outcome for all farmers. The

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Box 9.1

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SWAGMAN models – Australia and Pakistan

Carrying capacity concepts were used as policy goals to achieve regional sustainability to develop the Salt Water And Groundwater MANagement (SWAGMAN®) series of models. SWAGMAN® Farm is a lumped water and salt balance model that integrates climatic, agronomic, hydrogeological and economic aspects of irrigated agriculture under shallow watertable conditions at a farm scale (Khan 2000). The model balances the economic and environmental trade-offs from alternative land and water management options, and enables irrigators to decide sustainable irrigation levels and optimal crop choices. Regional groundwater investigations, surface–groundwater interaction models and the SWAGMAN® Farm models are strategic tools that can be used to develop strategies such as improving WUE, reducing net recharge and monitoring changes in environmental conditions on a spatial basis. Farmers can determine paddock-scale recharge targets using the SWAGMAN® Farm model, available online as well as in a PC version. The results from the SWAGMAN® Farm model runs show that under conditions of average water availability, increasing salinity would reduce the total gross margin by $3 million annually from nearly $35 million to nearly $32 million over the next 20 years in the Coleambally Irrigation Area in the southern Murray-Darling Basin of Australia. Carrying capacity based land and water management options, however, generate a positive net present value (NPV) of $3.4 million (over 20 years at 5%) – that is, the external management costs are fully recovered. The analysis, thus, suggests that the tradable groundwater/salinity recharge credits policy would generate net farm-level beneﬁts to the community. Restrictions on rice-cropped area generate smaller net beneﬁts but these are not sustainable over the 20-year period. Sensitivity analysis, using a discount rate of 2.5–10%, does not alter these conclusions. The results are also robust to other sensitivity tests including: 10% lower than assumed salinity damages; inclusion of waterlogging damages alongside salinity damages; and above average variations in rainfall. The model has a dedicated stand-alone and web-based user-interface to help input data and seamlessly visualize results. This model has been used to distill management options such as net recharge management for the control of shallow watertables, focusing speciﬁcally on managing the net recharge beneath the root zone in relation to the vertical and lateral regional groundwater ﬂows. The SWAGMAN® Farm model computes the lumped estimates of the water and salt balance components for the cropping and fallow periods for a range of irrigated crops (such as rice, soybean, maize, sunﬂower, fababean, canola, wheat, barley, lucerne hay, grazed lucerne, annual and perennial pasture) as well as dryland wheat and uncropped areas, for different irrigation, soil, climatic and hydrogeological conditions. Farmers have structured their environmental management plan around on-farm net recharge management using the SWAGMAN® Farm model. The net recharge management implementation process builds on strong research partnerships among stakeholders over a few decades. This innovative decision-making tool has helped promote sustainable land and water management options; it also provides an institutional instrument to monitor changes in WUE and environmental conditions. The model could serve as a milestone for deﬁning net recharge target/credits to individual farmers that could in turn help achieve overall groundwater and salinity management targets and promote a rational environmental management dialogue between the farmers and other stakeholders. Apart from Australia, the SWAGMAN modelling tools have been calibrated in the Indus Basin of Pakistan.

Irrigation Management in a Catchment Context recharge management metric enables the conversion of diffuse source recharge to point source recharge at paddock scale, enabling deﬁnition of private property rights within a common pool issue and assigning individual responsibilities for its management. Put alternatively, net recharge targets implemented at paddock scale would achieve overall sustainable management of groundwater recharge and salinization of the underlying aquifers – a common pool resource.

9.3 Case Study Insights 9.3.1 Groundwater management zones Shallow groundwater salinity dynamics have a bearing on crop productivity but are sensitive to irrigation management practices. A whole-ofcatchment water and salt balance model can be a useful framework for evaluating the carrying capacity based on sustainable irrigation management options. The farm- or paddock-scale net groundwater recharge can be linked to a systemwide carrying capacity constraint, accounting for spatial variations in various components of the salt-water system. The modelling results from an irrigated area in the southern Murray-Darling Basin show that groundwater management zones hold the key to deciding on-farm and regional actions to achieve the most effective management options. The whole irrigation system can be divided into various groundwater zones, each with distinct salinity and groundwater dynamics (Fig. 9.1). The regions where groundwater ﬂows downwards with respect to the watertable are called ‘recharge zones’ and regions where groundwater ﬂows upwards are called ‘discharge zones’. An extreme environmental problem in the recharge zones may appear as a result of saline groundwater ﬂows to the soil surface causing waterlogging in adjoining landscapes. Likewise, the discharge zones may be impacted due to poor drainage and insufﬁcient removal of salts from the root zone. The analysis of groundwater dynamics shows that:

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• The groundwater ﬂows from the northern area are directed outside the area towards the northern zone. • The ﬂows from the central area are directed towards the western area. • In general the ﬂow patterns in the southern area are not directed towards the western area, although there are smaller ﬂows from the central to the southern area. • The overall drainage ﬂows from the southern area are directed towards the edges. • The ﬂows towards the western area are mainly contributed by the central area. • The groundwater ﬂows from the northern area are not contributing to the accessions in the southern, central and western area. The typology of groundwater dynamics in Figure 9.1 suggests that where recharge to groundwater is avoidable, the pricing of recharge above leaching requirements can be an option to provide mechanisms to make inefﬁcient users pay for reduced system capacity and loss of production from a community’s perspective. This may apply to areas where the watertables are very shallow (below 2 metres) and effective porosity of soils is low (below 5%), such that small amounts of excess irrigation and rainfall can raise the watertable and compromise the carrying capacity of the system by causing waterlogging and salinity problems. The maximum allowable groundwater recharge quota should be based on the regional salt and groundwater dynamics and, hence, should be location-speciﬁc, not uniform. Systemwide uniform regulations may reduce some better quality recharges, which may be beneﬁcial to deeper aquifers and, hence, reduce potential future groundwater reserves, impacting longterm sustainability. This suggests that blanket or one-size-ﬁts-all approaches are unlikely to address real issues in sustainable environmental management. A carefully crafted science-based policy can provide a sound basis for sustainable resource management. Policies aimed at improved management of water resources in irrigated areas or seeking the development of new irrigation

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Zone 1

Zone 3

Zone 2

Zone 4

Zone 5

Irrigation Area Boundary GW Zone Boundary GW Contours

Groundwater Flow Direction

Fig. 9.1 Groundwater management zones.

schemes must be guided by the shallow groundwater salinity dynamics and changes in salt and water balances of the hydrological cycle. The modelling evidence from the Philippines supports this mindset.

9.3.2 Crop choices and soil salinity Crop choices and on-farm land and water management actions inﬂuence net changes in salinity (Table 9.2). For instance, crops such as rice often lead to excess groundwater recharge due to

Irrigation Management in a Catchment Context

Box 9.2

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The science of water savings – Philippines

Improper water accounting can lead to erroneous decisions (Hafeez et al. 2007). Scale effects of increase in water use and water productivity are often unclear and poorly documented. A large fraction of water inputs to rice ﬁelds are believed to be lost by seepage and percolation. These losses can be reused downstream and do not necessarily lead to true water depletion at the irrigation system level. Because of this high potential for reuse, the general efﬁciency of water use can increase with increasing spatial scale, as shown by the results of a multi-scale water accounting study undertaken in the rice-based irrigation system in the Philippines. Daily measurements of all surface water inﬂows and outﬂows, rainfall, evapotranspiration, and amounts of water internally reused through check dams and shallow pumping were summed into seasonal totals for 10 spatial scale units ranging from 1500 ha to 18,000 ha. The options for improvement in water productivity were evaluated. The results show that water reuse by pumping and check dams was 7% and 22% of the applied surface water at irrigation district level. The reuse of surface water through check dams increased linearly with 4.6 Mm3 per added 1000 ha. Reused water from pumping was equivalent to 30% of the water lost through rice evapotranspiration during the dry season. Water reuse plays a dominant role in growing a rice crop during the dry season; more than 1500 farmers, from a total of 10,000, used 1154 pumps to draw water from shallow tube wells (or from drains and creeks) for supplementary irrigation at the district level, thus supporting a proﬁtable rice crop during the dry season.

ponded conditions, whereas other crops such as lucerne may extract water from a shallow watertable. It is possible to devise a spatial mix of recharging and discharging crops and tree plantations to achieve an area-wide water balance and to keep the watertable at a desired level. For instance, restrictions on cropped area for recharging crops may be desirable in some settings while not in others; some crops such as lucerne and permanent plantations may be more desirable in areas affected by shallow ground watertable and high salinity. Water policy must be tuned to the local carrying capacity needs and farmers’ cropping preferences, conditioned by the hydrological and economic environment. 9.3.3

Sustainable irrigation intensities under a water trading regime

Irrigation requirement includes the evapotranspiration needs of crops and the water necessary to leach out excess salts. Restricting irrigation to meet those needs can help reduce groundwater accessions. The records of irrigation volumes used by best performing farmers can serve as benchmarks to guide this policy. The concept

must consider variability in soil, depth to the watertable and regional groundwater discharge, all of which may explain the differences in water use across farms. In shallow groundwater areas, minimum irrigation intensities may be highly desirable to keep the salts pushed down below the root zone, while in deep groundwater areas such minimum irrigation intensities may be helpful, but not essential, for improved salinity management (Fig. 9.2). Watertable rise and salinity dynamics are sensitive to farmers’ crop choices and irrigation practices, and farmers react well to the signs of salinity and economic incentives (see Huffaker and Whittlesey 2003; Kijne 2006). Together, this fuels the hope that inﬂuencing farmers’ on-farm land and water management practices via economic incentives holds the greatest promise for salinity management. Market-based approaches can also be used to inﬂuence farmers’ crop choices and land and water management practices. Under this model, a limit can be placed on the overall level of an activity or pollution associated with the environmental damage, and ‘rights’ to the agreed individual level of activity can be allocated among

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Table 9.2 Watertable and salinity impacts of different crop growing options. The upper and lower values for respective column are shown in bold. The leakage rates are 0.2, 0.5 and 1.0 ML ha−1 for the South, Central and Northern zones respectively

Scenario 1 2 3 4 5 6 7 8 9 10 a

Net recharge/discharge under land use (ML a)

Depth of watertable (m)

South Zone-3

1 3 1 3 1 3 1 3 1 3 1 3 1 3 1 3 1 3 1 3

177 260 355 262 121 270 224 262 322 324 94 170 −225 42 587 466 288 210 −14 −14

Central Northern Zone-2 Zone-1 108 191 286 193 52 201 155 193 253 255 25 101 −294 −27 518 397 219 141 −83 −83

−7 76 171 78 −63 86 40 78 138 140 −90 −14 −409 −142 403 282 104 26 −198 −198

Net recharge (ML ha−1) South Zone-3 0.8 1.1 1.5 1.1 0.5 1.2 1 1.1 1.4 1.4 0.4 0.7 −1.0 0.2 2.6 2.0 1.3 0.9 −0.1 −0.1

Central Northern Zone-2 Zone-1 0.5 0.8 1.2 0.8 0.2 0.9 0.7 0.8 1.1 1.1 0.1 0.4 −1.3 −0.1 2.3 1.7 1.0 0.6 −0.4 −0.4

0.0 0.3 0.7 0.3 −0.3 0.4 0.2 0.3 0.6 0.6 −0.4 −0.1 −1.8 −0.6 1.8 1.2 0.5 0.1 −0.9 −0.9

Salt from irrigation (tonnes) 106 106 164 164 132 132 161 161 149 149 168 168 132 132 192 192 154 154 125 125

Net salt into root zone (tonnes) 486.2 4.8 184.2 44.8 661.9 30.4 510.4 43.2 335.2 8.5 335.2 3.3 737.4 16.8 90.6 0.9 90.6 58.2 391.7 3.84

ML = million litres.

Fig. 9.2 Model of estimated minimum irrigation intensity at different watertable depths.

Average watertable change (m)

Average salt conc. change (dS m)−1

−1.0 0.6 0.2 0.9 −1.2 0.6 −0.6 0.9 −0.4 1.1 −0.4 1.4 −1.5 0.6 0.6 2.6 0.6 1.1 −0.6 0.7

0.8 0.0 0.3 0.0 1.1 0.0 0.8 0.0 0.5 0.0 0.5 0.0 1.2 0.0 0.1 0.0 0.1 0.0 0.6 0.0

Irrigation Management in a Catchment Context users. Through the trading of these rights, greater efﬁciency, effectiveness and ﬂexibility can often be achieved relative to other policy instruments. Market-based approaches, however, cannot automatically balance efﬁciency and environmental sustainability goals. For instance, interstate water trading, which transfers large volumes of water out of the mature irrigation areas, can impact soil salinity and environmental sustainability locally and regionally, as discussed later. 9.4

New Scientiﬁc Insights

9.4.1 Sustainable environmental management technologies Waterlogging and salinity are twin evils of irrigation that can be managed, not controlled. Key on-farm salinity management strategies include: • improved WUE (Wichelns 1999, 2002); • better irrigation practices (Tarboton et al. 2004); • better crop choices (Ashraf and Saeed 2006); • integrated agroforestry and crop farming systems (Dunin 2002);

Box 9.3

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• re-vegetation and tree plantations on barren land (Hatton et al. 2002; Hajkowicz and Young 2002); • conjunctive use of surface and ground water (Ahmad et al. 2002; Kijne 2006); • mechanical solutions, such as improved drainage (Datta et al. 2000) and integrated approaches including evapotranspiraion basins and biological concentration of salts (Rundle and Rundle 2002; Abdel-Dayam et al. 2004). Serial biological concentration (SBC) is an innovative technology to manage salts in agricultural drainage (Fig. 9.3). This approach utilizes saline drainage water as a resource to produce marketable crops and may provide a method to manage salts in an economically viable manner. SBC technology is currently being adopted in Australia and Pakistan. The economic evaluation of SBC trials in Australia indicates that with salinity levels below 2.5 dSm−1, the SBC operation can be an economically viable option for managing saline drainage. Economic evaluation of SBC trials in Pakistan, however, did not show promising results. This was related to the low economic value of additional production, which translated

China and Pakistan – urban efﬂuent reuse for peri-urban agriculture

Untreated urban and industrial efﬂuents, which result in water quality deterioration of receiving water bodies, can be treated before disposal without much additional investment on treatment plants. A groundwater modelling framework was applied to assess the hydraulic loading and salinity impacts on underlying the groundwater system of a large-scale land-based efﬂuent treatment system for treating urban sewage of the Yanggao County located in the north of Shanxi Province, China, as a case study. This land-based treatment system is called ‘Filtration and Irrigated cropping for Land Treatment and Efﬂuent Reuse’ abbreviated as FILTER. The FILTER operations include intensive irrigation of crops with the urban efﬂuent, ﬁltration through the soil matrix, and pumping and reuse of improved quality drainage water. Three sizes of FILTER plot were simulated using groundwater modelling techniques to: (i) up-scale the operations and beneﬁts of an experimental FILTER site to a larger FILTER system to handle the larger volume of efﬂuent; and (ii) assess the potential effects on the regional groundwater. Vertical drainage wells were used as this was a preferred option to suit the conditions of deep groundwater levels and presence of suitable aquifers. The boundary ﬂux analysis carried out to quantify the net impact of FILTER on the surrounding areas indicated that a well-designed and aptly managed vertical drainage system can be used in a catchment-scale FILTER system for managing domestic wastewater in peri-urban settings. Results of this innovative and cost-effective technology from the ﬁeld trials in Australia, Pakistan and China were promising (Khan et al. 2008).

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into incomes lower than costs, making the SBC system uneconomical. Accounting for the third party impacts of treating salt could result in better economic results for both Australia and Pakistan. This could be achieved by developing sustainable public–private partnership for ﬁnancing the technology. The ﬁnancing options could be based on the polluter-pays principle in terms of ‘salinity credits’. In terms of salinity management, it is the amount of money that the polluter is willing to pay for treating salt to minimize third party impacts, or the amount of money they would like to accept for treating salt. The application of the polluter-pays principle in terms of salinity credits for treating a tonne of salt offers considerable incentive for public–private investment for managing salinity. 9.4.2 Sustainable environmental management policies In many countries, the current thrust of water policy is to use water markets as a viable mechanism for reallocating water resources most efﬁciently to their highest valued uses. For instance,

in Australia the water markets were introduced in many states as a tool for reallocating water to its most efﬁcient use (Topp and McClintock 1998; Brennan and Scoccimarro 1999; Crase et al. 2004). Water markets can boost productivity by reallocating water from low- to high-value uses, and by enhancing the availability and reliability of water supplies. The economic gains from water trading within an individual water district may be relatively modest; interdistrict water trading may, however, produce larger economic gains (Bjornlund 2004; Peterson et al. 2005). However, water trading policy must take into account: • the third party effects of trading that result from changes in return ﬂows; • declining economic activity in the origin/net selling districts, and its ﬂow on impacts on equity and welfare of the local population; • lower returns to irrigation infrastructure and other ﬁxed investments; • in particular, the impacts on salinity and overall health of the river system (Easter et al. 1999; Etchells et al. 2006). Water markets can lead to substantial environmental externalities due to hydrological and

Fig. 9.3 Conceptual framework for the serial biological concentration (SBC) of drainage. NB: The stages and volumes are indicative only; actual SBC stages, drainage volumes and salinity would depend on the ﬁeld situation.

Irrigation Management in a Catchment Context environmental constraints and, thus, have the potential to alter ﬂow regimes and impact surface and ground water salinity levels and riverine environments – particularly where permanent and large transfers of water upstream are involved. For instance, an approximately thirtyfold increase in the magnitude of salinity is likely as water trade moves water from a low impact zone, where groundwater moves slowly towards the river, to a high impact zone, where salt additions to the river system per unit of water use are the greatest (Duke and Gangadharan 2005). In particular, interstate water trading should not result in reductions in environmental ﬂows, degradation of the natural environment or increased levels of salinity. Put simply, water markets should ensure that each transfer is at least ‘salinity-neutral’, that is, has a zero salinity impact. The transfer of water to different locations or to new irrigation developments has the potential to affect salinity levels in a river system in three ways. First, the transfer of water entitlement changes the volume of water in a river system at different locations. This would affect a river’s dilution capacity and, hence, changes the salinity level of the water. Second, expanded or new irrigation developments have the potential to increase or decrease the amount of saline drainage that enters a river system. Third, individual transfers of water across farms within an irrigation scheme can impact salinity and groundwater accession in the destination zone due to individual farm characteristics and management practices. This may impair drainage and salt removal in the origin districts by curtailing irrigation to below optimal levels required for ﬂushing salts. Currently, permanent trade in water entitlements is limited: less than 1% of diversions were traded in 2001–2002. The majority of trade in the Murray-Darling Basin has been temporary trade in seasonal water allocations, with around 10–20% of the seasonal allocation being traded within irrigation districts per year (Beare and Heaney 2001). The cumulative net effects of water trading on salinity are likely to be substantial in future as interstate water trade expands

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beyond the current limit of 10 gigalitres (GL) for permanent interstate trades, or moves water over greater distances or involves farmers in many irrigation districts currently not participating and extends to different levels of water security/entitlements, both low and high security entitlements. How water markets in aggregate will impact on salinity and the overall health of the Australian river system is yet to be fully understood, due to both limited experience with water markets (speciﬁcally those involving large upstream transfers), as well as the complexity in modelling the dynamics of irrigation and salinity development in a catchment context. Approaches to model the nexus are fewer, but evolving (Tisdell 2001; Etchells et al. 2006; Heaney et al. 2006). In some areas, the charges for preventing water exit from irrigated areas were introduced (Australian Competition and Consumer Commission 2006) but these were without any sound scientiﬁc basis and were subsequently abolished and replaced with a termination fee. The nexus between water trading and onfarm shallow groundwater salinity was explored taking selected irrigated settings in the southern Murray-Darling Basin as an example. A management model, cast in a dynamic programming format to integrate a detailed biophysical and hydrological model with an economic model, was applied to estimate the productivity, proﬁtability and sustainability of irrigated agriculture under salinity-neutral resource management scenarios with water trading. The results show that salinity-neutral water trading offers signiﬁcant gains in economic efﬁciency over a business-as-usual (non-salinity-neutral) scenario (Table 9.3). The modelling results can be used to articulate the idea of quantifying and applying the minimum/optimal irrigation intensities required to maintain groundwater salinity levels and watertable at suitable limits in a given irrigation area with known soil types under given climatic conditions. Water trading rules must comply with these minimum irrigation intensities to ensure salinity-neutral water trading.

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shahbaz khan The economics of alternative water trading and salinity management scenarios

Item Total gross margin per year Total net beneﬁt (20 years)a NPVb a b

Scenario 2 Area cap, with water traded out

Scenario 3 Water cap, no water trade

Scenario 4 Water cap, water traded out only

Scenario 5 Water cap, with full water trade

$33,484,357 $844,622 −$2,845,326

$24,795,853 −$181,613,973 −$114,241,969

$33,427,573 −$347,853 −$3,573,369

$33,973,879 $11,124,586 $3,430,911

It generates a total gross margin of $34,919,667 in year 1 and $31,933,285 in year 20. The values are reported as difference from the business-as-usual scenario.

Agricultural water markets can facilitate adjustments to water scarcity and competition, and enhance economic efﬁciency but markets cannot automatically balance efﬁciency and environmental sustainability goals. Results from an irrigated area in Australia show that minimum irrigation intensities must be retained to push the salts out of the root zone, especially in shallow watertable/high salinity impact areas. Such minimal irrigation intensities are helpful, but not essential, in deep watertable/low salinity impact areas. Should water markets lead permanent water transfers out of the irrigated areas to violate the minimum irrigation intensity needs (especially in the high salinity impact areas), substantial negative impacts on resource quality and agricultural productivity are likely. Water trading that adds to salinity cannot be economically viable in the long run. The trade in water entitlements would likely conﬂict with environmental sustainability and rural renewal goals, especially where large and permanent water transfers are involved. Trade-offs between economic efﬁciency promised by the water markets and environmental goals will need to be evaluated before embarking on any water trading programme. These conclusions have important policy implications for salinity management and water trading in other areas and water policy reﬁnements. Upstream water transfers can impose signiﬁcant externality costs, requiring differential adjustments to water abstraction caps for the origin versus destination zones to account for the potential impacts on salinity. At the same time,

it also implies that irrigators who resort to water sales out of distress must be protected by governments against losing their water entitlements and leading to deterioration of the quality of their land. 9.4.3 Optimal irrigation investments Investments in water savings boost agricultural productivity, create employment, propel regional development and, therefore, contribute to overall economic development. The beneﬁts of investments in water resources accrue both within and outside the catchment and irrigation districts. The long-term beneﬁts are more widely shared and are often several times greater than the shortterm beneﬁts (World Bank 2005). The realization of such beneﬁts on a sustained basis requires a minimum platform of infrastructure (Grey and Sadoff 2006): • well-functioning hydrological and physical infrastructure to convey a required ﬂux of water in the time and space needed; • institutional infrastructure to operate the water system and to protect and safeguard the rights of all parties, including the environment; • policy infrastructure to guide national water vision and underwrite incentives for sustainable water resources management. Where water infrastructure is dysfunctional (Perry 1995) or the institutional and policy framework fails to provide the right incentive (Saleth and Dinar 2005), lower WUE and wasteful use results in lost economic opportunity. Water resources development for collective economic

Irrigation Management in a Catchment Context good and social beneﬁt requires investment to maintain the health of river systems and to rehabilitate and modernize ageing infrastructure, as well as building new infrastructure to enhance efﬁciency in water delivery and use across the system. Equally important are on-farm investments in water-efﬁcient irrigation and cropping technologies, as well as ﬁnancial and pricing incentives to promote reallocation of water from low- to high-value-added uses (Berbel and Gómez-Limón 2000). The two major socioeconomic functions of Australian irrigation systems are: • to provide food and ﬁbre for Australia’s present and future population; • to boost agricultural exports to serve Australia’s strategic regional/global interests. For a current population of around 20 million, and considering an adequate agricultural water requirement of 1000 m3 capita−1 yr−1 (Falkenmark 2007), then there is a need to provide some 20,000 GL yr−1. Using a food security approach to provide food for around 20 million Australians at 3000 kcal capita−1 day−1, and assuming a calorie to water ratio of 1 : 1, i.e. one calorie litre−1 of water (evapotranspired) on average (Molden et al. 2007), Australia needs around the same amount of water. Considering a future population level of around 30 million by 2050, Australia will need 50% more water. One option would be intensive agriculture, including expansion of irrigated agriculture from about 2.5 million ha to over 5 million ha, relinquishing some of the current area under rain-fed agriculture. This will result in a lower environmental footprint for agriculture, while meeting future food demand. In terms of Australia’s global and regional strategic role, there would be greater demand for Australian agricultural commodities due to physical scarcity of water in populous countries such as India and China. This may require additional land and water resources to the current area of Australian agriculture. The real potential to meet the global/regional strategic interests is through intensiﬁed water-efﬁcient irrigated agriculture in suitable locations through public-private investments.

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Irrigated agriculture makes up 70% of Australia’s consumptive water use. Given that Australia’s water resources are fully allocated, and in some places over-allocated, the only way to ensure enough water for future irrigation development is to use available water more efﬁciently at both farm and catchment scales. Water can be saved through better management of its delivery and application. In an average year, irrigated agriculture uses 14,000 GL, which is about 70% of all water use in Australia. This water is not used as efﬁciently, however, since: • 10–30% of the water diverted from rivers into irrigation systems is lost before it reaches the farm gate. • Up to 20% of water delivered to the farm gate may be lost in distribution channels on farms and around 60% of water used for irrigation on farms is applied using high volume, ineffective gravity irrigation methods. • More than 10–15% of water applied to crops is lost through overwatering, whereas scheduling tools and observational data could more precisely match water application to crop water requirements. In 2007, the Australian Prime Minister announced a $10 billion National Plan for Water Security (‘the National Plan’) to boost investment in water resources in order to achieve a 10-point agenda over 10 years (Government of Australia 2007). The National Plan included an 80 : 20 public– private investment proposal for a 50 : 50 sharing of water savings. Under the National Plan, the Commonwealth will become solely responsible for sustainable management of Australian water resources on behalf of all states and territories, for the common good of all citizens and the environment. The public–private investment proposal is expected to culminate in a framework to select the most deserving irrigation infrastructure projects. Appropriately targeted investments can eliminate many of the inefﬁciencies cost effectively (Wichelns 2000). One requirement to eliminate inefﬁciencies in regional and on-farm water delivery systems would be to map the areas where inefﬁciencies exist and then target invest-

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ments to those areas. This entails a whole-ofsystem diagnostic analysis of the water use to prepare a water balance sheet of net inﬂows and outﬂows across various reaches within the system. Often the information on spatial water use efﬁciencies does not exist or does not feed into the water policy process to guide investments. Such information, along with measures of the physical and economic productivity of water on-farm and across uses, may help to identify real opportunities to ‘save’ water and to tailor and prioritize investments to realize those savings. This study used results from a whole-ofsystem water accounting framework for the Murrumbidgee Catchment, the Coleambally Irrigation Area and the Murrumbidgee Irrigation Area, all located in the Murray-Darling Basin in Australia. The aim was to prepare a water balance sheet to identify unaccounted ﬂows in the river system zone, leakage/inefﬁciency in the nearfarm water supply zone and on-farm zone that could be assessed economically to bolster overall WUE. The water balance sheet so developed was linked to investment needs and to anticipated beneﬁts to make a business case for investments in Australian water resources (Fig. 9.4). The ﬁndings show that annually some 1334 GL of water remain unaccounted for in the Murrumbidgee Valley, which represents potential water savings. The on-farm zone offers the greatest potential to save some 607 GL of water annually and to enhance WUE, with win–win outcomes for the farmers, investors and the environment. Some 168 GL of water lost annually in the near-farm water supply zone could partly be brought back into productive or environmental use through investments in infrastructure rehabilitation and modernization. Enhancing overall WUE offers investment opportunity of $293–824 million and in annual farm-gate income depending on the intended end use of the saved water. The hydro-economic rankings of on-farm and offfarm investment options (Figs 9.5, 9.6) demand a systems analysis of real water losses and savings and pathways to achieve cost-effective water savings. There is a need to prioritize investments

which represent economically justiﬁable and proﬁtable opportunities for public and private investors. Enabling conditions for building community–public–private partnerships to foster equity investments in the Australian water sector are identiﬁed below. 9.4.4 Investing in tomorrow’s irrigation Irrigation and environmental sustainability in irrigated catchments have, to date, been managed as two competing enterprises under separate and divergent control. In Australia, there is an increasing quest, and support, for a ‘harmonized’ business approach to sustainable use of land and water resources to achieve enduring business partnerships in water management. The ‘system harmonization’ approach seeks to identify business opportunities for irrigators to become an integral part of the expanding environmental services industry and, in so doing, support a truly sustainable and diversiﬁed irrigation business environment. A good understanding of system-wide harmonization can be gained from how irrigation systems are linked with the catchment water cycle and how life support systems and regional economies depend on them (Fig. 9.7). The irrigation system involves many subsystems, all of which are intrinsically linked to one another, not only physically but also with the environment and to the society within which they exist. In addition to establishing the base physical, economic and social position of the region, some of the key pressure points in the system and the constraints they impose can be identiﬁed. In particular, these relate to the capacity to optimize on-farm and near-farm irrigation system performance and water demand patterns to boost productive and environmental dividends and to strengthen the backward and forward linkages with interlinked economic systems to boost the indirect value chain impacts they generate. The key pressure points in a water system are not necessarily biophysical; they can also be economic, social, environmental or institutional. These are the changes in these key pressure

Irrigation Management in a Catchment Context

Fig. 9.4 Whole system water balance accounting sheet for the Murrumbidgee irrigation area.

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Fig.9.4 Continued

Irrigation Management in a Catchment Context

227

Fig. 9.5 On-farm water savings investment options.

points that need to be assessed in a comprehensive and systematic way to enhance the multifunctional productivity of the irrigation system. The phases involved in developing and implementing system harmonization processes are shown using a conceptual-operational framework (Fig. 9.8). New science and knowledge needed for harmonizing irrigation systems within their operating environments are developed in the ﬁrst three phases of the process. This knowledge feeds into subsequent phases, which are designed to operationalize the process. The term ‘feasibility stage’ is used to refer to the steps involved in designing and delivering the system harmoniza-

tion framework. This approach is used to clarify the conceptual understanding of this process. However, these boxes constitute permeable boundaries that allow information, concepts and data to ﬂow across them in a seamless fashion to ensure shared learning. Another important element of this framework is the iterative rather than linear nature of the process. The ﬁrst three steps of the process are designed to inform the business development phase which in turn relies on the analytical and modelling framework to evaluate and explore alternative business options in an iterative fashion. In the ﬁrst phase the hydrological system is analysed and characterized. This entails

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Fig. 9.6 Off-farm water savings investment options.

Fig. 9.7 Identiﬁcation of key pressure points (hexagon shape) in the irrigated catchment water cycle.

Irrigation Management in a Catchment Context

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Fig. 9.8 The phases involved in developing and implementing system harmonization processes.

the quantiﬁcation of stocks and ﬂows of water, pollutants and other constituents within the entire catchment water cycle. The understanding of the system dynamics achieved in this process is crucial in analysing the existing hydrological situation of the system and in formulating, calibrating and modelling predictive tools to assess system responses to alternative operational policies designed to achieve productive and environmental improvements. In the second phase the economic, social and environmental outcomes from the hydrological systems are assessed. At this feasibility stage the markets for ecosystem services and economic trade-offs and synergies between productive and environmental outcomes are evaluated. It is taken as a given in this phase that the purpose of an irrigation scheme is to satisfy society’s demands. The desire for environmental outcomes is no different to economic outcomes. These outcomes need to be measured in physical terms prior to being valued. In the third phase the institutional aspects and mechanisms for change are investigated. This analysis is intended to understand the existing institutional framework and identify potential barriers to the implementation of change and

capture of productive and environmental opportunities. In the two boxes in Figure 9.8 business plans are developed to capitalize on the ﬁndings from the research. A key element of the system harmonization approach that can bring stakeholders together is the fact that irrigation generates myriad beneﬁts that are not conﬁned to the irrigators or the sector itself; rather these accrue to others and/or other sectors within and beyond the catchment through multiplier effects. Studies show that the multiplier values are generally larger for developed economies, as high as 5–6 for Canada and Australia (Hill and Tollefeson 1996) which means that a $1 increase in agricultural income will generate another $4–5 dollars in non-farm goods and services, beneﬁting all. This generates widely shared beneﬁts for the community as a whole and can serve as a magnet to attract shared investments to harness these beneﬁts. 9.4.5 Future research requirements In the future, irrigation faces many traditional ongoing and emerging challenges. The traditional challenges give plenty of reasons to undertake smarter investments. These include:

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• managing salinity hazard by matching landscape capability while making decisions on locations for irrigation development; • managing extraction impacts by quantifying both positive and negative externalities from different irrigation areas by evaluation, auditing and benchmarking of irrigation industry by sectors; • balancing irrigation and environmental ﬂow demands through real savings due to improved distribution and on-farm WUE and alternative cropping options; • managing drainage water quality and the minimization of impacts on rivers and ecosystems; • managing negative environmental impacts, such as methane and nitrous oxide emission, salinity, water pollution (abuse of pesticides), stagnation – especially in intensive crop production systems. More efﬁcient use of inputs (water, fertilizer, pesticides and labour) to reduce negative impacts on the environment, to reduce production costs and to bolster productivity growth. Alongside these, the emerging challenges that provide additional rationale to invest include: • growing and wealthier population demand more food and water, as well as better quality rivers for recreational purposes; • climate and land use change and their impacts on water vulnerability; • global warming impacts on transboundary water co-operation; • ecosystem services and markets; • biofuel competition for land and water; • water, energy and greenhouse linkages to rationalize the use of pressurized irrigation systems. Transparent debate should surround future irrigation investments and development, emphasizing that irrigation must be treated in infrastructural terms, rather than as a commodity. The debate underpins the belief that water security and services must improve while reducing costs and vulnerability, regionally and rapidly. In the water-scarce catchment settings of Australia, where water resources are fully developed and ﬁnancial and storage capacities are not constrained, enhancing the efﬁciency of available water through more efﬁcient delivery systems and on-farm production technologies holds the

key to water scarcity issues. Signiﬁcant gains in on-farm WUE and water productivity are possible through appropriate interventions. These gains are often assumed, rather than identiﬁed, at various spatial scales and across irrigation systems. Without proper water accounting for the whole irrigation system, misguided investments to ‘save’ water can reduce return ﬂows and can be detrimental to the environment and to other users. Investments to boost the multifunctional productivity of water resources must build on the concept of irrigation system harmonization: an approach that seeks to align various components of the irrigation system, namely, the river system, the near-farm delivery network and the on-farm water management system. This will boost water productivity in a catchment and river basin context, based on business principles. It entails reinventing irrigation and related enterprises on a business model designed to seize competitive commercial opportunities through technological innovation and improved eco-efﬁciency, while achieving better environmental stewardship in time–space dimensions. The concept is not new but it has never been applied to irrigation management. It acknowledges the existence of trade-offs between irrigation and environmental sustainability but asserts that the two are not purely antagonistic and positive synergies can be realized by adopting a harmonized and uniﬁed business approach to water management. New science for synchronizing irrigation systems with their operating environments and for creating ecosystem markets are the core agenda items that deﬁne various phases of the system harmonization process. The proposed interventions and actions are fully solvent, in that they are ﬁnanced on a business model, and offer rewards to shareholders through new value added uses of available water resources. A shared vision for economic gains from system-wide improvements in multifactor productivity brings diverse stakeholders together. Regional ‘irrigation business partnerships’ are proposed as an institutional media to implement the system harmonization model.

Irrigation Management in a Catchment Context Acknowledgements The author acknowledges discussion of ideas with the System Harmonization Researchers and Dr Thierry Facon of the Food and Agriculture Organization. Funding support from the Pratt Water Group, CSIRO Water for Healthy Country Flagship, the Cooperative Research Centre for Irrigation Futures and the Coleambally Irrigation Cooperative Limited for different aspects of this work is acknowledged. Aspects of this work are concurrently being published in the Australian Farm Policy Journal and ICID Journal of Irrigation and Drainage. References Abdel-Dayam, S., Hoevenaars, J., Mollinga, P.P., Scheumann, W., Slootweg, R. and Van Steenbergen, F. (2004) Reclaiming Drainage: toward an integrated approach. Agriculture and Rural Development Department, The World Bank, Washington DC. Abdullah, K. (2006) Use of water and land for food security and environmental sustainability. Irrigation and Drainage, 55, 219–222. Ahmad, M.-u-D., Bastiaanssen, W.G.M. and Feddes, R.A. (2002) Sustainable use of groundwater for irrigation: a numerical analysis of the subsoil water ﬂuxes. Irrigation and Drainage, 51, 227–241. Ashraf, M. and Saeed, M. (2006) Effect of improved cultural practices on crop yield and soil salinity under relatively saline groundwater applications. Irrigation and Drainage Systems, 20, 111–124. Australian Competition and Consumer Commission (ACCC) (2006) A Regime for the Calculation and Implementation of Exit, Access and Termination Fees Charged by Irrigation Water Delivery Businesses in the Southern Murray–Darling Basin. Advice to the Australian, New South Wales, South Australian and Victorian Governments, Commonwealth of Australia. http://www.accc.gov.au/content/index. phtml/itemId/771300 (last veriﬁed 10 October 2008). Baumol, W.J. and Oates, W.E. (1993) The Theory of Environmental Policy. Cambridge University Press, New York. Beare, S. and Heaney, A. (2001) Irrigation, Water Quality and Water Rights in the Murray Darling Basin, Australia. ABARE Conference Paper no. 15. Paper presented to the International Water and

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Resource Economics Consortium and Seminar on Environmental and Resource Economics, Biannual Conference, Girona, Spain, 3–5 June. ABAREEconomics, Canberra, Australia. Berbel, J. and Gómez-Limón, J.A. (2000) The impact of water-pricing policy in Spain: an analysis of three irrigated areas. Agricultural Water Management, 43, 219–238. Bjornlund, H. (2004) Formal and informal water markets: drivers of sustainable rural communities? Water Resources Research, 40, W09S07. Brennan, D. and Scoccimarro, M. (1999) Issues in deﬁning property rights to improve Australian water markets. Australian Journal of Agricultural and Resource Economics, 43, 69–89. Characklis, G.W., Grifﬁn, R.C. and Bedient, P.B. (2005) Measuring the long-term regional beneﬁts of salinity reduction. Journal of Agricultural and Resource Economics, 30, 69–93. Clarke, C.J., George, R.J., Bell, R.W. and Hatton, T.J. (2002) Dryland salinity in south-western Australia: its origins, remedies, and future research directions. Australian Journal of Soil Research, 40, 93–113. Crase, L., Pagan, P. and Dollery, B. (2004) Water markets as a vehicle for reforming water resource allocation in the Murray-Darling Basin of Australia. Water Resources Research, 40, W08S05. Datta, K.K., de Jong, C. and Singh, O.P. (2000) Reclaiming salt-affected land through drainage in Haryana, India: a ﬁnancial analysis. Agricultural Water Management, 46, 55–71. Duke, C. and Gangadharan, L. (2005) Regulation in environmental markets: what can we learn from experiments to reduce salinity? The Australian Economic Review, 38, 459–469. Dunin, F.X. (2002) Integrating agroforestry and perennial pastures to mitigate water logging and secondary salinity. Agricultural Water Management, 53, 259– 270. Easter, K., Rosegrant, W. and Dinar, A. (1999) Formal and informal markets for water: institutions, performance, and constraints. World Bank Research Observer, 14, 99–116. Etchells, T., Malano, H.M. and McMahon, T.A. (2006) Overcoming third party effects from water trading in the Murray-Darling Basin. Water Policy, 8, 69– 80. Falkenmark, M. (2007) Shift in thinking to address the 21st century hunger gap: moving focus from blue to green water management. Water Resources Management, 21, 3–18.

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Food and Agriculture Organization (FAO) (2003) World Agriculture: towards 2015/2030: an FAO perspective. FAO/Earthscan, Rome. Grey, D. and Sadoff, C. (2006) Water for growth and development: a theme document of the 4th World Water Forum. In: Thematic Documents of the IV World Water Forum. Comision Nacional del Agua, Mexico City. http://siteresources.worldbank.org/ INTWRD/Resources/FINAL_0601_SUBMITTED_ Water_for_Growth_and_Development.pdf (last veriﬁed 8 October 2008). Government of Australia (2007) A National Plan for Water Security. Government of Australia, Canberra, Australia. http://www.pm.gov.au/docs/national_ plan_water_security.pdf (last veriﬁed 8 October 2008). Hafeez, M.M., Bouman, B.A.M., van de Giesen, N. and Vlek, P. (2007) Scale effects on water use and water productivity in a rice-based irrigation system (UPRIIS) in the Philippines. Agricultural Water Management, 92, 81–92. Hajkowicz, S. and Young, M.D. (2002) An economic analysis of revegetation for dryland salinity control on the Lower Eyre Peninsula in South Australia. Land Degradation and Development, 13, 417–428. Hatton, T.J., Bartle, G.A., Silberstein, R.P. et al. (2002) Predicting and controlling water logging and groundwater ﬂow in sloping duplex soils in western Australia. Agricultural Water Management, 53, 57–81. Heaney, A., Dwyer, G., Beare, S., Peterson, D. and Pechey, L. (2006) Third-party effects of water trading and potential policy responses. Australian Journal of Agricultural and Resource Economics, 50, 277–293. Hill, H. and Tollefeson, L. (1996) Institutional questions and social challenges. In: Pereira, L.S., Feddes, R.A., Gilley, J.R. and Lesaffre, B. (eds), Sustainability of Irrigated Agriculture. Proceedings of the NATO Advanced Research Workshop, Vimeiro, Portugal, 21–26 March 1994, NATO ASI Series, E, Applied Sciences, vol. 312. Kluwer Academic Publishers, Dordrecht, Netherlands, pp. 47–59. Houk, E., Frasier, M. and Schuck, E. (2005) Irrigation technology decisions in the presence of waterlogging and soil salinity. Global Business and Economics Review, 7, 343–352. Huffaker, R. and Whittlesey, N. (2003) A theoretical analysis of economic incentive policies encouraging agricultural water conservation. International Journal of Water Resources Development, 19, 37–53.

Hussain, I., Mudasser, M., Hanjra, M.A., Amrasinghe, U. and Molden, D. (2004) Improving wheat productivity in Pakistan: econometric analysis using panel data from Chaj in the Upper Indus Basin. Water International, 29, 189–200. Khan, S. (2000) Research Project Information from CSIRO Land and Water, Sheet No. 22, SWAGMAN® Series. CSIRO, Australia. Khan, S., Tariq, R., Yuanlai, C. and Blackwell, J. (2006) Can irrigation be sustainable? Agricultural Water Management, 80, 87–99. Khan, S., Rana, T. and Hanjra, M.A. (2008) A cross disciplinary framework for linking farms with regional groundwater and salinity management targets. Agricultural Water Management, 95, 35–47. Kijne, J.W. (2006) Salinization in irrigated agriculture in Pakistan: mistaken predictions. Water Policy, 8, 325–338. Luquet, D., Vidal, A., Smith, M. and Dauzatd, J. (2005) More crop per drop: how to make it acceptable for farmers? Agricultural Water Management, 76, 108–119. Meijer, K., Boelee, E., Augustijn, D. and Molen, I. (2006) Impacts of concrete lining of irrigation canals on availability of water for domestic use in southern Sri Lanka. Agricultural Water Management, 83, 243–251. Molden, D., Oweis, T.Y., Steduto, P. et al. (2007) Pathways for increasing agricultural water productivity. In: Molden, D. (ed.), Comprehensive Assessment of Water Management in Agriculture. Water for Food, Water for Life: a comprehensive assessment of water management in agriculture. Earthscan, London, UK/International Water Management Institute (IWMI), Colombo, Sri Lanka, pp. 279–310. Narayanamoorthy, A. and Hanjra, M.A. (2006) Rural infrastructure and agricultural output linkages: a study of 256 Indian districts. Indian Journal of Agricultural Economics, 61, 444–459. Perry, C.J. (1995) Determinants of function and dysfunction in irrigation performance, and implications for performance improvement. International Journal of Water Resources Development, 11, 25–38. Peterson, D., Dwyer, G., Appels, D.C. and Fry, J. (2005) Water trade in the Southern Murray-Darling Basin. The Economic Record, 81, S115–127. Proust, K. (2003) Ignoring the signals: irrigation salinity in New South Wales, Australia. Irrigation and Drainage, 52, 39–49. Ragan, G.E., Young, R.A. and Makela, C.J. (1999) New evidence on the economic beneﬁts of controlling

Irrigation Management in a Catchment Context salinity in domestic water supplies. Water Resources Research, 36, 1087–1095. Rundle, P.J. and Rundle, B.F. (2002) A case study of farm-based solutions to water logging and secondary salinity in southwestern Australia. Agricultural Water Management, 53, 31–38. Saleth, R.M. and Dinar, A. (2005) Water institutional reforms: theory and practice. Water Policy, 7, 1–19. Tarboton, K.C., Wallender, W.W. and Raghuwanshi, N.S. (2004) Farm salinity appraisal with water reuse. Irrigation and Drainage Systems, 18, 255–273. Tisdell, J.G. (2001) The environmental impact of water markets: an Australian case-study. Journal of Environmental Management, 62, 113–120. Topp, V. and McClintock, A. (1998) Water allocations: efﬁciency and equity issues. Australian Commodities, 5, 504–510.

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Wichelns, D. (1999) An economic model of waterlogging and salinization in arid regions. Ecological Economics, 30, 475–491. Wichelns, D. (2000) A cost recovery model for tertiary canal improvement projects, with an example from Egypt. Agricultural Water Management, 43, 29–50. Wichelns, D. (2002) An economic perspective on the potential gains from improvements in irrigation water management. Agricultural Water Management, 52, 233–248. World Bank (2005) Shaping the Future of Water for Agriculture: a sourcebook for investment in agricultural water management. Agriculture and Rural Development Department, The International Bank for Reconstruction and Development/The World Bank, Washington DC.

Image facing chapter title page: Courtesy of http://en.wikipedia.org/wiki/File.Irrigation1.jpg

Managing Potable Water Supplies

10

BERNARD BARRAQUE1 1

10.1

CIRED-Agroparistech, Paris, France

Introduction

The fundamental issue of this chapter is the dialectics between drinking water and public health on the one hand and the good quality of the aquatic environment on the other. One of the reasons for the ongoing argument is that humans have a much longer life expectancy than nonhumans. Not only do we all want to take minimal risks with diseases carried by water in the short term (like diarrhoea epidemics) but we also want to protect people from diseases caused by an accumulation of toxic substances in the long run (heavy metals, pesticides, etc.). In this respect, it is better to maintain the aquatic environment in a good ecological status and this is the basis of the EU Water Framework Directive (European Union 2000). While sanitary engineering has developed sophisticated treatment technologies, the present ‘complexiﬁcation’ of potable water criteria has led an increasing fraction of the water supply community to advocate better land use control so as to beneﬁt from less polluted water. Indeed the ﬁrst European Directive concerning water (EC/75/440) dealt with this issue, mandating that surface water should be clean enough to be used for drinking purposes despite treatment capacities.

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

10.2 Historical Perspective Two hundred years ago, the European urban population was served by water from public fountains, public or private wells and/or from water vendors who took water from rivers or from fountains. There were no public water supply systems as we know them today; in Paris, the ﬁrst concession contract for a piped water system connecting private houses was granted in 1781 to the Perrier brothers. Their project was to pump water from the Seine River into a reservoir using a steam engine and to distribute it under pressure to private subscribers. The scheme went bankrupt within just a few years. Yet, soon the skill of British ‘engineering’ made it possible to progressively generalize the pressured water service during the nineteenth century and later to extend it to rural areas. Other European countries soon followed the example and engineers and hygienists also invented sewer systems to handle both waste water and rainwater and transport it away from the city. During the ﬁrst half of the twentieth century, chemical engineering supplemented ﬁltration with technologies to treat drinking water which increased the autonomy of cities vis-à-vis their environment. Today, Europe is the only continent where the great majority of (quasi all) the population is connected to a public water supply (PWS). Most of the population is connected to a centralized public sewage collection and treatment (PSCT); most of those who are not connected to a PSCT live in low density areas

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and rely on increasingly efﬁcient decentralized on-site sanitation systems (e.g. a septic tank with a ﬁltration bed). This success story is the product of several interlinked developments: public health and medical knowledge; involvement of successive layers of engineering technologies and science; innovative solutions to the ﬁnancing issue (which is the issue behind the well known public vs. private debate); and, last but not least, development of multi-level governance, beyond some initial conﬂicts between central and local governments. The development of public water services can be broadly presented in three stages. In the nineteenth century, or rather until the discoveries by Florence Nightingale (1820–1910), Edwin Chadwick (1800–1890), Robert Koch (1843–1910) and Louis Pasteur (1822–1895) were popularized, the PWS developed on the assumption that water should be drawn from natural environments far from the cities. Large metropolizes, in particular, would have to get water from further and further aﬁeld. At the end of the nineteenth century, however, this initial strategy started meeting resistance in some places because of competition with local uses or competing demand from other cities. Eventually, however, the discovery of bacteriology resulted in the invention of water treatment which then allowed local authorities to provide good sanitary quality water from much closer sources, in particular from the rivers on which they were located. Municipalities then took the lead and in many cases terminated contracts with private companies which were often ﬁnancially unable to connect the whole population. Once the water services became a mature industry, municipalities also had to face the issue of long-term infrastructure maintenance and capital reproduction. Some were then led to reintroduce or to develop an industrial and commercial status, or to delegate services to private companies, including to some that they would create and own themselves. At the end of the twentieth century, the issue of cost recovery led to the questioning of some of the technological choices of the ‘chemical engineering’ period and attempts to solve the sustainability issue through

new strategies inspired by environmental and ecological engineering. 10.2.1 Bringing water to cities and connecting people When industrialization started in Europe, the prevailing legal system was still largely inherited from the feudal period: according to feudal law, water rights differed between landlords and peasants. Communities had a right to use water for domestic and husbandry purposes and landlords had a right to use and abuse (alienate, export, destroy) water provided they would respect the user rights of communities. In expanding cities, newcomers from the countryside imported their customs and requested free water of good quality from the public fountains. The growing needs could only be met, however, through some sort of water transfer from a distant and ‘pure’ natural environment and this implied the need for the landlords’ consent. Later, with the formation of

Box 10.1 Finland In Finland, the General Fire Assistance Company of the Grand Duchy was established in 1832 and it funded the cities’ water systems. Cheap loans from this insurance company (average about 6% in the second half of the nineteenth century) played a large role in the expansion of city water works. But there were other important forms of funding, especially taxes from spirit distilleries. In each locality a company was given the exclusive right to distillate spirits against the payment of a liquor tax. From this tax, capital was raised over time for the establishment of a water works. This amounted to about 10% of the total required and most came through taxes and quite substantial donations and willed sums. Loans were also taken from local banks where necessary. A loan from the ﬁre insurance company was nevertheless generally the largest single source of funding, and the interest charged was clearly lower than with other creditors (Juuti and Katko 2005).

Managing Potable Water Supplies ‘Nation-states’, national governments usually took over the full sovereignty of water and their intervention was often necessary for transferring water to cities. In several areas, like New England (Anderson 1988), insurance companies also contributed to the development of water services since the generalization of standpipes allowed for an increased capacity to ﬁght ﬁres. The largest water use however often remained connected to ‘washing the city clean’ from all sorts of wastes. This could be done from close-by rivers, but needed some sort of pumping. Long-distance aqueducts were eventually employed to bring water to public fountains and to a few public buildings or aristocratic homes. Domestic water supply was initiated by private companies who tried to recover their initial investment from customer subscriptions. Most of these were industrial and not ﬁnancial companies: their primary aim was to install infrastructure and to get repaid for it, not to operate services in the long term. They tended to prefer relatively inexpensive solutions such as pumping untreated water from nearby rivers. Overall, however, in the same period in the nineteenth century, few people were ready or able to pay a water bill. There were growing conﬂicts between companies, customers and municipalities: alleging an insufﬁcient payment of their services, companies often postponed investments to extend the service or to maintain it which in turn attracted criticism, and even legal prosecution, from the municipalities. There were also debates about tapwater quality which was not yet monitored on a scientiﬁc basis. Drinking water criteria were then very simple and organoleptic. For example, until the end of the nineteenth century, the French Superior Public Health Council considered water as ‘good and potable, if it is fresh, clear, odourless, quasi-tasteless, neither unpleasant, nor saltered, insipid, or sicky sweet; when it contains few alien parts and enough dissolved air, when it dissolves soap without forming lumps and it cooks vegetables well’. Increasingly, once they were convinced of water being an essential asset for public health, municipalities terminated concession contracts and chose to run the services under direct labour.

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In some cases, national governments sided with the private companies and maintained them in operation, as was the case in Lisbon where central government distrusted the municipality’s capacity to operate the service. But in most cases, governments either let municipalities take over or even provided ﬁnancial support. There were other possibilities for local government to obtain ‘cheap money’, for example, from the early popular savings banks that they controlled, whose bonds were attractive to the public. Bills or subscriptions by private companies were partly, or totally, replaced by the payment of rates on property values or of local taxes by all citizens which provided revenue that was independent of actual water consumption and eventually had a cross-subsidizing effect in favour of the poor. The Montreal case illustrates this (Fougères 2004) where a PWS was started by a private company as early as 1798 but it did not work well. It was transformed into a tax-paid direct labour service. As soon as it was incorporated, the municipality purchased the private company (which was losing money and was willing to sell). Soon the city realized that it had difﬁculties in raising funding to connect the periphery, and in 1853 it obtained a new authorization from the crown to compel all the Montréalais to connect to the PWS and to charge the service through local taxes. Today, the payment of water supply through rates is still largely dominant in Britain and in many countries; sewerage has long been paid through local taxes. Public procurement was bound to succeed better than private. The technical solution of delivering untreated water from distant sources remained dominant in the New World due to government investment in long distance infrastructure. Through the successive construction of local, state and federal projects, California has become the largest artiﬁcial river basin in the world, part of the water coming from the Colorado more than 500 miles away. The Colorado itself is a very wild river in a huge river basin (628,000 km2) which is bigger than the combined area of France and Belgium but provides one-tenth of the yearly ﬂows in these two. Yet, the construction of some

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Box 10.2

Glasgow

Glasgow offers a good example: direct municipal provision seemed to offer several advantages to the city. The existing private company had (…) outdated infrastructure [and] consequently was unable to cope with the demands of the rapidly growing population (…). Moreover, the company was not in a position to raise the necessary capital for improvements, unlike the town council, whose extensive community assets made it eminently credit worthy. Public accountability meant that unpredictable market forces could be over-ridden and a stable service provided (…). Loch Katrine was located in the Perthshire highlands, some 55 km from Glasgow and thus well away from the polluted city (…). The ofﬁcial opening by Queen Victoria on an appropriately wet autumn day in 1859 was an event of enormous signiﬁcance for Glasgow (…). Loch Katrine was unquestionably the prime municipal showpiece for the city, combining the wonders of Victorian technology with the nurturing quality of pure Highland water (Maver 2000).

Joel Tarr (1996) has illustrated this broad approach in the USA: getting cleaner water from further on the one hand and using the rivers as sewers on the other, this latter decision relying on the assumption of natural dilution and self puriﬁcation of rivers.

of the biggest reservoirs in the world (Boulder Dam, and later Glen Dam, which stores around 30 km3) allows for the storage of approximately four times the mean yearly ﬂow. In turn this allows an export of 90% of the water, leaving a quite insufﬁcient portion for the Mexicans in the now shrinking estuary delta. California gets the lion’s share, with up to 5.4 million acre-feet (one acre-foot is around 1235 m3) and within it, irrigation gets around 75% of the total imported. This model was extended to the rest of the world after the Second World War, due to the co-occurrence of international ﬁnancing institutions offering cheap money and of various (Keynesian or socialist) forms of support for gov-

ernment intervention in infrastructure provision. In the 1950s and 1960s large hydraulic projects were increasingly devoted not only to cities but also to irrigated agriculture for the export market. The symbol of these ‘multipurpose projects’ is the Tennessee Valley Authority. Today, many developing countries still base their water policy on large water transfers so as to indirectly subsidize the production of irrigated cash crops. In some cases the maximum extractable water resources have been reached and sometimes irrigation is given priority at the expense of public water supply. To summarize, government intervention allowed the transfer of enough water from distant

Managing Potable Water Supplies sources through largely subsidized aqueducts. Local authorities started taking over the infrastructure which had been initially created by private companies and the frequent ﬁnancing of connections by local taxes or rates allowed a reduction in the tension between urban population and water suppliers and made public services acceptable. But, municipalities were then dependent upon higher levels of government and upon distant water sources which eventually would become unsustainable. The innovation of water treatment helped alleviate this situation. 10.2.2

Treating water

In the early twentieth century, in the heart of industrial Europe, growing population densities, the increase of per capita consumption and limited resources increased competition for pure water, while the development of biochemical analyses showed growing contamination. The issue slowly shifted from quantity to quality. Eventually it became clear that whatever the water source it should be ﬁltered (end of nine-

Box 10.3

239

teenth century) and later chlorinated, ozonized or disinfected through granulated activated carbon (GAC) beds or UV (these technologies were developed around the ﬁrst world war). Groundwater was still not ﬁltered or treated at all. The question then arose, if water was to be treated whatever its source, then why fetch it from far away? Taking it from the river just upstream of the cities would not make a difference in terms of public health and would save a lot of investment. This had been the choice of many British cities which, due to unfavourable hydrogeology, had to rely on rivers for water supply. So in that early period, large European cities changed their strategy from investment aimed at increasing available water quantity to investment aimed at improving water quality which resulted in growing operation costs. But, at the same time, with the rise of the middle classes the status of domestic delivery of water pressure changed from a luxury good to a normal commodity and made it possible to have customers pay water bills to cover those costs.

Paris 1

This is exactly what happened in the city of Paris a century ago. From the third Empire (1850–1870), the idea had prevailed that Paris should get water from distant sources and indeed, engineer Belgrand worked towards securing longer distance sources of water (around 100 km away from Paris). In 1890, engineer Duvillard came up with a project to draw water from Lake Geneva, i.e. 440 km away from Paris! Proponents of this project soon came up with all sorts of arguments to convince the Paris City Council and the State: a ‘capital of the world’ would need at least 1000 litres per capita per day (lcd), i.e. ﬁve times more than the most optimistic standards of the time. This would allow Paris to have more luxurious fountains, more street cleaning, better domestic comfort and hygiene. Besides, such a quantity of water would extend navigation possibilities in drought periods, help ﬂush waste water from the new sewer system away to the river Seine and then to the sea, etc. In the end, they argued, a huge transfer would make Paris’ PWS reliable forever and the bigger it would be, the cheaper each cubic metre would become! However in 1899 a disease epidemic broke out and bacterial analyses put the blame on one of the distant natural intake points: so even distant ‘pure’ water could be contaminated and should be ﬁltered and treated! In 1902, proponents of treatment won, with the new water ﬁltration plant in Ivry pumping water from the Seine just upstream of Paris. A long lasting choice was being made. Chlorination was added after the First World War. Water demands were growing incrementally at the time and the big jump then appeared to be too risky. Lastly, the Lake Geneva aqueduct was discarded by Paris City Council in 1919 for fear of a water war: destruction by the Germans!

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The invention of water treatment by chemical and sanitary engineers allowed many cities to turn to nearby surface water and thus to complete the networks and serve the population through a mostly local solution. Eventually this model would evolve into inter-municipal joint boards. Treating water also meant growing operational costs and supported the idea of water as a commodity to be paid by water bills, increasing the ﬁnancial autonomy of the utilities. This commodiﬁcation was better accepted in the twentieth century when the urban population got accustomed to the supply of tapwater under pressure. After the Second World War, rising pollution in rivers was felt to be increasingly unacceptable and cities were constrained to build sewage treatment plants so as to reduce polluting discharges. Two types of plants, water and sewage works, materialized the frontiers of the ‘networked city’ (Tarr and Dupuy 1988) and allowed for the development of an institutional, legal and economic system quite separate from the issue of water resources allocation which would eventually be extended to rural areas and would help to invent the concept of local public services. Conversely, in the Mediterranean part of Europe, central governments often remained active and sponsored the development of regional bulk water transfer institutions. Local authorities would then get very cheap water to distribute to their population which in turn helped maintain the tradition of cheap, but relatively unreliable and incomplete, services. Today in these countries there is, paradoxically, less open criticism over the service quality than elsewhere in Europe (e.g. in UK and France) but water users express a sort of ‘silent distrust’. This creates a problem for raising water prices or rates since authorities must build the trust of their citizens with an improved service. Despite these differences, water supply services eventually became a mature industry in Europe: the initial infrastructure, which had frequently been subsidised, now needs replacement, while environmental performance has to improve and this creates the need for a further change. But possibly the most severe constraint is the obliga-

tion to respect the same drinking water criteria all over Europe: the Drinking Water Directive was adopted in 1980 (EC/80/778) and was reinforced in 1998 (EC/98/83) (European Council 1980, 1998). This body of legislation not only provides a list of contaminants, it also deﬁnes the monitoring process that should be observed and makes it mandatory to inform water users of the quality of their water. 10.2.3 Mature services, new issues In continental Europe, the decision to change the status of PSCT from an imposed administrative service to a commercial one and then to have it paid for as part of water bills caused the bills to subsequently increase dramatically (depending on the degree of subsidy PSCT still receive from governments). For example, in Germany and Finland, water bills more or less doubled. In France, sewage works were developed through earmarked levies on drinking water which were contained a quasi mutual banking system, the well known Agences de l’eau. Funds were then used to subsidise new sewage works, upstream reservoirs, etc. In the same period, PWS itself became a mature business and had to face the issue of renewing an ageing infrastructure without further subsidy. This is the fundamental reason why public procurement had to evolve in various ways towards legal private status: under traditional public accounting it was neither encouraged to depreciate the assets nor to make renewal provisions while private accounting could. In Germany, the Netherlands, Switzerland and the Nordic countries, a long and quiet process of separation of water services from the municipal corporation and its main budget took place. This is largely related to the tradition of subsidiarity which encourages lower levels of government to get involved in the local economy. Mixed economy companies developed and eventually evolved into public limited companies, the capital of which was wholly owned by public authorities. In Germany, and to a lesser degree in other countries, a single multipurpose local

Managing Potable Water Supplies urban services provision company, called the Stadtwerk or Querverbund, is common. This has many advantages: cross-subsidising services that are losing money (e.g. public transport) by proﬁtable ones, paying smaller taxes on proﬁts to the land government while paying important professional taxes to the local level. Critics, in turn, tend to blame the high overall cost of this type of structure, the overmanning linked to the monopoly situation and the lower innovation capacity, etc. There are, however, few benchmarking studies available because it is indeed very difﬁcult to compare situations stemming from different institutional contexts. It is not obvious, however, that the German water services model is as inefﬁcient as a World Bank report suggested (Barraqué 1998). Conversely, centralized and liberal countries, like France and the UK, were more reluctant to allow local authorities to develop mixed economy systems and then, not so surprisingly, the private sector could develop and make a come back. In France, because more than 36,000 local authorities had a high degree of sovereignty, it was difﬁcult to merge them and this explains why there were still more than 12,000 PWS utilities in 2007. Many of these are ﬁnancially weak, they lack technical capacity and they often have to contract out services. The most important of these are joint boards which can be very large (Paris suburbs) but most importantly, this voluntary concentration process often followed the development of a new form of private company involvement whereby infrastructure was not privatized but operation and maintenance were. In this case, the contract deﬁnes which renewal investments are made. Private companies are usually responsible for changing the parts that wear out quickly (less than 15 years) while authorities have to invest in the renewal of pipe systems. After some corrupt affairs were disclosed, and also in the context of the global debate on privatization in the Third World, this private-public partnership (PPP) formula has come under criticism and there is an open debate in France on the ‘return to public procurement’. This remains limited and indeed the most impor-

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tant outcome is the growing involvement of the Conseils généraux (County Councils) in a concentration process taking place silently. Evolution in Britain was different and was embedded in a historical process of reducing municipalism. Like other public services sectors, there was a move towards regionalization from the 1929 crisis onwards, ultimately resulting in a merging of water planning, water policing and water industry into ten regional authorities in 1974. Since these were public institutions, however, they were subjected to rate and borrowing capping in the subsequent period of economic crisis. Finally, the water industry itself made a plea for privatization. This was also to reduce the criticism that they were ‘poachers and gamekeepers’ at the same time. In 1989, full privatization of the industry was decided in which planning and policing were handed to a National Rivers Authority, soon to become the Environment Agency. This model of private services under the supervision of national level independent regulators (EA, OFWAT and DWI) is a model of stateliberal policy and it has inﬂuence at the European level. It is in Britain and in France where private companies play the largest role, that consumers’ criticism is indeed the highest. What this short analysis reveals is that the supra-local concentration process is probably more important than the public vs. private issue. In the Netherlands, for instance, water supplies were voluntarily concentrated to a point where only ten remain, all being private companies owned by joint boards of local authorities, provinces or a mixture thereof. A further issue is then whether these utilities should be merged with the pluri-century old waste water and drainage management institutions, the famous water boards or Waterschappen. In Italy, the Galli law of 1994 is now in the process of being implemented which will reduce the number of utilities from above 6000 to less than 100. Most of the new entities will be at the scale of a Provincia (a county) and they must be private limited companies, eventually owned by the local authorities. Direct labour should be phased out and in order to foster increased

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investment, a price setting formula has been adopted following the British OFWAT example. In Portugal, the evolution is less systematic: the country was eligible for EU subsidies until 2007 and it set up a national water company to channel the funds. Local authorities were then encouraged to form large joint boards at the scale of regions or metropolitan areas and to join the national water company into a mixed regional company (50% national capital, the rest split between member municipalities). This regional venture is usually in charge of what is locally called Agua Alta, i.e. bulk water production and treatment and sewage works plus large interceptors; while Agua Baixa, i.e. local water distribution and sewerage, remains in the hands of the local authorities. Many of these now adopt the municipal company formula and some end up charging waste water, garbage collection, stormwater and beach cleaning as part of the water bills. Together with concentration, privatization provided for increased investment but also increased ‘commodiﬁcation’ of water services and, of course, meant further increases in water bills because of depreciation and renewal provision. Additionally, governments are now under the inﬂuence of economists who support full, or at least fair, cost pricing and subsidies are phased out. In turn, water bills have risen dramatically; for example, in Paris and the rest of France, bills doubled on average between 1990 and 2004, largely because of PSCT improvements. The unexpected outcome is that an increasing number of large users (industry, services) either change their processes or invest in leakage control. This largely explains the recent stagnation of volumes sold. In some countries, even domestic consumers have reduced their demand for PWS through changes in ﬁxtures and domestic appliances, different garden design and rainfall storage or utilization of other alternative sources of water for non-drinking uses. Such cases have been identiﬁed in several places in France, in Belgium and in Germany (Cornut 2000; Montginoul et al. 2005) and more recently in Central and Eastern Europe (Juuti and Katko 2005). Eventually, these

new attitudes will threaten the ﬁnancial balance of public services.

Box 10.4 Paris 2 After the Second World War, the prefect of the Seine département (county) took advantage of a severe ﬂood to obtain the construction of three large upstream reservoirs on the Seine, the Marne and the Aube, in fact to increase summer ﬂows and meet Paris water demands even in very serious droughts (as the 1976 one). Interestingly enough, a fourth upstream reservoir was planned by mayor Chirac’s councillors in the 1990s but it was abandoned for the same reasons (we must and can purify the water anyway, said the giant water supply companies) and because water demand in Paris went down by 16% between 1990 and 1998 (Cambon-Grau 2000). Another demand reduction since 2004 led the Paris mayor to increase water bills by 7% in 2007, despite a former electoral promise.

At the same time, water suppliers have discovered that it is going to be ever harder to permanently comply with drinking water standards at reasonable costs. The control of eco-toxicologists over the setting of standards tends to support a traditional ‘no-risk’ strategy (Lave 1981) without taking the implied costs into account. For example, in Europe, the lowering of the lead content from 50 to 10 µg/l implies that all lead pipes need to be removed; this will cost up to $35 billion while there is no evidence that under the former standard, lead poisoning from water was observed. Bacteriological criteria are now just a small component of drinking water standards and heavy metals and micro-organic compounds bring the total to 63 criteria, including 20 different pesticides, in Europe; and 84 in the USA. The multiplication of criteria is progressively making the situation over-complex. Year after year, the media can report a growing proportion of people

Managing Potable Water Supplies receiving water that does not comply with the standards even though treatment is improving in the long term (for a history of drinking water criteria and the present result, see, e.g., Okun 1996). This is creating a growing distrust from water consumers and in some cases there is open conﬂict. In a way, municipal water supply is the victim of its own success: people are accustomed to having clean tapwater and they use it without knowing where it comes from, how it is treated, etc. Besides, they use drinking water for all sorts of purposes which leads to the mixing up of sanitary, organoleptic and appliances’ criteria. For instance, people dislike hard tapwater (high calcium content) which damages their washing machines and gas liners, while they buy calciumrich bottled water in the supermarket. New York City, like many US and Canadian cities, was able to follow a different path of water supply development than many European cities because of the abundance of clean water. The tradition grew whereby a lot of water was used and it was taken from further and further aﬁeld, while protecting the water intake points through extensive land use control, for example, in the Catskills. Yet, the US east metropolis might have to catch up with the European story because clean natural resources are not immune from cryptosporidium, a type of protozoa, and other new (lethal) substances. A USEPA Panel of Experts concluded that New York should not be given any further derogation on the need to ﬁlter and treat water extensively, while city engineers were arguing that increased land use control would sufﬁce (Ashendorff et al. 1997; Okun et al. 1997). The controversy is serious since the new treatment required is likely to seriously increase water prices which might lead to social impacts and eventually to a water demand collapse. A further question, in this case then, is why not just pump the water from the Hudson River and forget about the water from Canada! It seems, however, that this issue could be so controversial that the big change is suspended. The city rather turns to increased land use control and compensation to the rural territory for curbing diffuse pollution.

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10.3 Environmental Engineering: New Insights Broadly speaking, water services, as a separate sector from the rest of water policies, developed on the supply side paradigm. The dream was to bring more water of better quality to serve as many uses as possible to take advantage of economies of scale. Now, Europeans realize that if they have to fund long-term capital reproduction of this massive infrastructure solely from their water bills, with no more subsidies, then they may not be able to afford the services, at least politically. It is all the more paradoxical that only 5% of our tapwater really needs to be potable but it is difﬁcult to imagine a dual water supply system because of the high costs involved. In large tropical cities, however, nobody except the poorest drink tapwater and the vast majority rely upon bottled water. Could a similar solution be found in Europe whereby mineral water is used for food and beverage and sub-potable tapwater supplies for all other uses. Today, however, this solution seems unacceptable to most water engineers and they argue that, on top of the old sanitary imperative, even at higher prices, tapwater supply keeps the highest quality–price ratio. To lower the risk of being unable to supply the requirement, water supplies require some demand side management. This is very new and there are few social science studies about what people do with tapwater, how much they really pay, how the situation would evolve if new tariff systems were developed, what types of families would pay more or would pay less, and how would the level of arrears evolve? A whole ﬁeld of research is opening. Water conservation policies tend to develop, however, and they are needed, if only to create an information context where authorities understand why domestic water demand went down for the ﬁrst time in history and whether this will continue or not. In California, for example, conservation is encouraged through the promotion and ﬁnancial support of all sorts of ‘best practices’ in homes while new economic methods allow optimization of the search and the control

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of ‘unaccounted for water’ (not only leaks), and sociologists and anthropologists try to develop global foresight studies rather than simply extending the provision curves as was done before. Despite the labelling, ‘demand management’ is, in fact, not so much derived from a sole pricing system than from the search for a better reciprocal adaptation between supply and demand, including the use of ‘non-conventional resources’: rainwater harvesting and waste water reuse after careful disinfection. In water scarce countries which have a coastline, there is now a boom in desalination to produce additional water in the summer. This technology is cheap in investment, usually quite ﬂexible, but expensive in operation. In Mediterranean countries, where water is overly cheap, it is always easier to pass the operation costs rather than the investment costs on to the water bills. Desalinated water tends to induce new water conservation attitudes and, thus, helps in postponing new investments, like additional upstream reservoirs, which are much less ﬂexible. Environmental engineering encompasses integrated river basin management, including strategies to protect water resources for the sake of drinking water. Even though engineers would be able to turn mud into water, this is very expensive and not safe. On the other hand, the economic value of water in public services is very high, much higher than in agriculture, for instance, and so a deal is possible between the two types of uses. Waste water treatment technologies have developed to reduce the impact of cities and industry on rivers while reducing the time and space needed to stabilize the efﬂuents and reduce their toxicity. Environmental engineering, however, now ﬁnds it may be cheaper and safer in the long run to give more space to rivers so as to naturally reduce ﬂood risks as well as letting natural processes supplement water treatment to achieve a better water resources quality (artiﬁcial wetlands, reed beds, etc.). For several reasons then, the former separation between water as a public service and water as a resource tends to blur. It certainly does in ﬁnan-

cial terms, since both water and waste water costs are increasingly covered by drinking water bills. Indeed, given the very important turnover of the water industry, the water policy community ﬁnds it convenient to use water bills to fund integrated river basin management. Special mention should be made here of land use control policies on areas or catchments where groundwater is abstracted for public water supply. Meeting the drinking water quality standard often requires a change to organic, or at least lownitrates, no-pesticides, agriculture and compensation programmes for farmers. According to Brouwer et al. (2003), the adoption of the Nitrates from Agriculture Directive (91/676 EC), and the issuance of the European regulation 2078/92 introduced agri-environmental programs into the Common Agricultural Policy. But further, some member states authorized, or even supported, the development of cooperative agreements at local level, between authorities, farmers and water utilities, with the aim of going further in the control of diffuse pollution from agriculture. Four German Länder (Lower Saxony, North-RhineWestphalia, Hesse and Bavaria), as well as several Dutch provinces, have developed contracts where utilities, with the eventual support of water abstraction taxes levied by regional governments, fund advice and training with respect to good practices, monitoring of diffuse pollution, as well as the losses due to changes in crop patterns (e.g. turning arable to extensive grassland). In many cases, the context is that farmers could go to court and claim for compensation and the cooperative agreement strategy has the purpose of reducing transaction costs by creating conﬁdence and stimulating an agreement between farmers to instigate ‘best’ practices. This policy turns out to be cheaper than the sophistication of water treatment and allows a positive-sum game with farmers. In Britain, cooperative agreements with farmers were initiated in Nitrate Sensitive Areas in 1994 but this policy was eventually abandoned once Nitrate Vulnerable Zones were created in 1998 where pollution reduction by farmers was made mandatory. Also, the post-privatization control system of investments and prices by

Managing Potable Water Supplies OFWAT does not encourage private water companies to propose measures whose outcome are uncertain and extend over much longer periods than 5 years. In France, the co-operative agreements are not so frequent due to the combination of a vast but underfunded agri-environmental measures programme and of a tradition not to compensate for farmland located in groundwater protection zones (périmètres de protection rapprochés). Yet increasingly, the Agences de l’eau will fund more ambitious contracts where, among several new investments, farmers will be subsidised by other groundwater stakeholders to achieve some conversion. These contract policies, however, raise issues. Paying farmers for not polluting reverses the polluter pays principle which is an EU environmental principle. Also, when applied locally in areas of severe pollution, compensation is obtained by the worst polluters, while those who have already adopted a more environment friendly approach do not get subsidies. Further, cooperative agreements go against the neo-classical economic principle of cost recovery, sector by sector. It may well be that regionally integrated and participative water resource management, another principle put forward in EU water policy, yields better results than simple taxation of externalities and in the end, allows a better level of global cost recovery despite frequent money transfers from drinking water bills to farmers. Besides, German, Dutch or Swiss cooperative agreements correspond to projects limited in time. It is acknowledged that farmers are not solely responsible for the intensiﬁcation of agriculture and, therefore, should be given ﬁnancial support during an adaptation period, even more so since pollution control at source is much cheaper than additional treatment. It is also more sustainable than another policy called ‘sanctuarization’ where water suppliers would purchase the required land and expel the farmers. In the USA and Australia, these territorialbased policies are considered as a mechanism for creating ‘markets for ecosystems services’ (Salzman 2005). In fact, the initial aim was not so much to protect water resources (except may

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be in the case of New York protecting the Catskills water) but to preserve biodiversity. In situations where private property owners and farmers are well respected, it may be more efﬁcient to contract with them for the shelving of part of their land for the sake of nature conservation. In 2004, a federal programme called CRP spent $1.6 billion on diverse actions covering 13.6 million ha. In Australia, the Bush Tender programme achieved a system of reverse auctions to make the best use of a given amount of money to protect ecosystems. As Salzman argues, despite all the difﬁculties and criticisms of these programmes, ‘these experiences have demonstrated that investing in natural capital rather than built capital can make both economic and policy sense’. In the western USA, however, agreements also developed to trade water quantity and not only water quality. When California was forced by the federal government to reduce its abstractions from the Colorado River, to enable a fair share to be provided to other states and to Mexico, an intense bargaining took place to redirect water from irrigated agriculture (which always had the lion’s share) to growing cities. The issue was not only to purchase water, but to take into account the eventual negative impact of the transfers on third parties. This policy is called wheeling. It took several years to ﬁnd an agreement between the farmers, the bulk water supplier of Southern California (MWD), and the city of San Diego, which feared that droughts would leave it without water. The ﬁrst agreement could not be accepted by the environmental movement, however, because it did not include the constraint of maintaining enough fresh water in the river to maintain the level and the salinity of the Salton Sea, where migratory birds and several endangered species live. The ﬁnal contract was agreed under strong pressure from the federal government and now San Diego recovers water at a very high cost, for the sake of ﬁnding a global regional agreement between the parties. Southern Europe illustrates this point differently, since the non-completion of PWS infrastructure prevents raising prices to cost recovery

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levels and this generates irrational allocation. Almeria was supposed to get water from the Pyrenees (Ebro basin) at a cost that is ten times what the city would have to pay to buy this water from farmers. Through a model initially developed under the dictatorship, but now involving European subsidies, Spanish farmers overexploit the aquifers and request increasing volumes of surface water transfers to grow tomatoes and strawberries, part of which will be bought as surplus by the EU. Since the farms are increasingly owned by rich northern agribusinesses, if they were not pushed by this uneconomical set up, farmers would then make more money without working, just reselling ‘their’ water. To summarize, in Europe and other developed countries, because of population density and environmental issues, the water industry has had to progressively supplement the quantity approach (hydraulics and civil engineering), and the quality approach (water treatment and sanitary engineering), with territorial policies for water resources protection and demand management (environmental engineering). This is made possible by the relatively high price of water, which allows utilities to use a small fraction of their turnover to reach co-operative agreements with farmers. This new type of strategy, as well as various forms of cross-subsidisation, creates a new issue in terms of cooperation needs between various levels of government and the public at large. 10.3.1

Cross-subsidization

Water supply and sanitation developed as separate policy sectors at the local level and this model had to be adapted to maintain a high level of sustainability. Authorities, under various modes according to the national politico-historical culture, have organized various forms of averaging out or cross-subsidy to limit the impact on water prices of heavy and lumpy but long lasting investments. Some do it by spatial integration and this is what happened in Britain where, before centralization created the ten regional water authorities in 1974, later privatized in 1989, there had been a sustained concentration

process going on since the Second World War. There is also a concentration process going on all over Europe, which tends to dispossess local authorities of their former responsibilities on water services. Another strategy goes through a temporal averaging process and is exempliﬁed by the establishment of earmarked funds, water banks like the French Agences de l’Eau and the Dutch Waterschaps bank, modernization of public accounting to allow for depreciation, etc. The third type of cross-subsidisation is the German Stadtwerke which incorporates the transverse management of several local utilities at local level. And lastly, social forms of averaging are well known, even if they are not always presented as such. For example, paying for waste water removal via local taxes makes this service more expensive for those who have larger houses but who are usually richer. This creates a redistributive effect (Rajah and Smith 1993). Under municipalism, water suppliers did not pay much attention to the detailed breakdown of potable water uses and to the distributive effects of tariffs since they wanted the best quality water in unlimited quantities to serve all purposes (in a commonwealth vision). Today, water suppliers are still reluctant to really study distributive effects, wanting water services to be customeroriented (through billing instead of taxing) for ﬁnancial reasons and they assert that this is the more equitable approach. Eventually, they develop new forms of tariffs, such as increasing rates by blocks, but the ﬁeld studies showing their outcomes both in terms of social justice and water conservation are still in their infancy. Further the opening of the traditionally closed PWS policy community to a whole range of newcomers, in particular the public in general, makes water engineers feel awkward and insecure. Is there any alternative to this opening, however, if what is at stake is the rebuilding of the general conﬁdence placed by the public in the service and those who provide it? Beyond economic and environmental sustainability dimensions, it is the third dimension, that

Managing Potable Water Supplies of ethics/equity, which is the most crucial today. In other words, are social and political outcomes of the ongoing changes still acceptable in the long term or are we heading for a collapse of the networked model for water? Before water supplies were determined by demand management and pricing became closer to real costs, there was great acceptability of water services. This was based, however, on what Melosi (2000) calls ‘out of sight, out of mind’. Now the limitations of the supply-side model forces managers to study their demand like any business, i.e. through marketing. The breakdown of drinking water demand and the evolution of sector-differentiated uses have to be studied. Local econometric studies show little elasticity of indoor domestic demand to price (Chesnutt and Mitchell 2000), yet there is a slow reduction of per capita demand taking place in the developed world. In the USA, this may be obtained just through information policies and subsidies to individual conservation measures (Dickinson 2000). The point is to maintain a quiet pace of change so as to let people adapt to higher water prices over sufﬁcient time. In northern Europe, historical smaller consumption and higher prices reduce the scope of change compared to North America. Yet there is such a reduction in water uses (mostly from large users and outdoor uses) that the only way to maintain the cost recovery principle is to raise unit prices (e.g. Paris, January 2007). Outcomes may be very serious and in the new eastern Länder of Germany, to compensate for the post-uniﬁcation collapse in water consumption (which poses serious problems for sewer natural ﬂushing), authorities force people in the outer periphery to connect, with connection costs having serious social impacts. Yet at the same time, wealthy and ecologically aware middle class people obtain the right to disconnect from services for the sake of zero environmental footprint. This illustrates the need for more direct public participation, so as to exchange information, in particular on the difference between short-term and long-term sustainability, on how to combine individual and collective demand side management, and on the real causes

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for (formal) privatization. If this does occur the rampant crisis situation generates a growing distrust towards systems that had reached their equilibrium under a ‘municipalist’ type of welfare. In some cities in France, this distrust reaches the level of court cases. In the end, the growing consumerization of water services itself (if not of water resources) is very dangerous if it is not done within a collective re-learning process away from the ‘out of sight out of mind’ approach. For instance, there is a trend in Europe to encourage the generalization of individual water metering and billing, even in ﬂats in small buildings, for the sake of better equity, of making people able to trade off between conserving and paying, etc. In many experiments, however, individual metering does not induce a signiﬁcant change in consumption patterns, in particular after a few months, and in many cases the savings on bills of the most thrifty are offset by the yearly cost of the meter itself (depreciation, reading and billing separately). Consumers can then become furious to see water bills rise when they had been told they would save. According to Rajala and Katko (2004), installing meters in individual apartments might be feasible in new condominiums but it is too expensive in old ones. One of the most interesting recent cases is the Flanders government’s decision to implement the Rio agenda 21 and to give away an initial water volume for free. In fact, the regional authority had decided to charge waste water collection and treatment within the water bill and the free volume was conceived as a sort of compensation. For practical reasons, the free volume was set at 15 cubic meters per capita per year, while extra volumes would be charged in such a way that water supplies would make the same income. The consequences have been studied by the social and economic council of Region Flanders (Van Humbeeck 1998). First, water supplies experience a reduction in total volumes sold because increasing block tariffs have made the last ones very expensive. It is suspected that people reinvest in cisterns and in private wells for watering the gardens, etc. but in the meantime, they are

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obliged to raise the unit prices. Second, it turns out that richer families pay a little less than before, while poorer ones pay a little more. This is due to speciﬁc socio-demographic conditions; richer families are larger than poorer ones and extra children do not take enough water to imply a bigger bill (the free initial volume is per person). This quite complex case illustrates how little we know about domestic water uses. If we are going to tackle the present crisis of water services, it will not be just with the economist’s toolbox and good moral feelings. Anthropologists, sociologists, historians, geographers, etc. and interdisciplinary approaches are needed.

10.4

Future Perspectives

If there are increasing discussions on how to reach a better compromise between the economic, environmental and ethical dimensions of water services sustainability in the EU and in the USA, the development of new tools and indicators can give us some conﬁdence for the future. But what is missing is an indicator translating the degree of overall conﬁdence of the general public into the systems and the water services policy community. The long-term threat to water services is illustrated by the situation in large Third World cities where when the utilities are not fully reliable, in terms of quantity, continuity and quality (in particular sanitary quality), the compensation strategies adopted by various groups of users tend to increase the uncertainty and the unreliability of the services. In this case, trying to apply the new economic approaches, suited to the mature systems of developed countries, can be catastrophic. Indeed privatization and even public– private partnerships at state or regional level, show their limitations in terms of universalizing water services. May be in the end it is impossible to improve deﬁcient utilities in developing countries and may be they will stay with bottled mineral water. Zérah (1997) showed the striking case of New Delhi where, facing unreliability, the wealthy developed redundant home supplies,

thus increasing the reliability of public services. The further down the social ladder, the more expensive the water. But looking at it carefully, there is no signiﬁcant difference in the kind of poor services and private domestic alternatives one can ﬁnd in some areas in eastern Europe and the Mediterranean. In the end, in developed countries, if global conﬁdence in PWS goes on shrinking and if water volumes sold go down signiﬁcantly, we may end up with a growing irregularity of the service, so that it would be the large Third World cities’ systems which would ﬁnally prevail in the long term (Barraqué 2001). Conversely, from this contemporary history of water supply in Europe, one could argue that the end reason why Third World cities have unreliable and uncompleted PWS is that there is no municipality with sufﬁcient legitimacy and capacity to build a reciprocal conﬁdence between the company and the water users. The European example paradoxically illustrates that on top of local authorities’ involvement, co-operation between levels of government and eventually private partners is essential for meeting the complex issues mentioned above. This co-operation is not so frequent in developing countries. So why do not we take a look back to the future of municipalism? Maybe there is something to keep from the past before local welfare and public economy had been thrown into the dustbin of liberal economics. What municipalism achieved was to channel the savings of the upper and middle classes into the ﬁnancing of a longterm solidarity system for all, based on a public economy of urban services, sometimes with the participation of the private sector but without privatization of infrastructure. Slowly, municipally owned water undertakings were made autonomous and covered their costs. In many countries, such undertakings buy goods, services and works from the private sector based on continuous competition without tying their hands with long-term operational contracts (Hukka and Katko 2004). But overall, what current policy analyses reveal is that if water services are locally based, their good governance implies the cooperation between local authorities and with

Managing Potable Water Supplies levels of government. In many Third World countries, what prevails is, in fact, confrontation rather than co-operation. Indeed, the territorial dimension of the issue is more important than the simple public vs. private debate. If we keep hope for Third World cities, we will need to invent similar mechanisms at appropriate territorial levels, depending on the national/local citizenship traditions and community cultures. Such a ‘subsidiary’ system would be the way to offer better guarantees for national and international public investors which in turn would result in having access to cheaper money for water systems. This seems to be the lesson taught by the historical development of European water services.

References Anderson, L. (1988) Fire and disease: the development of water supply systems in New England, 1870–1900. In: Tarr, J. and Dupuy, G. (eds), Technology and the Rise of the Networked City in Europe and in America. Temple University Press, Philadelphia, PA. Ashendorff, A., Principe, M., Seeley, A., Beckhardt, L. and Mantus, J. (1997) Watershed protection for New York City’s supply. Journal of the American Water Works Association, 89, 75–88. Barraqué, B. (1998) Europäisches Antwort auf John Briscoes Bewertung der Deutschen Wasserwirtschaft. GWF Wasser-Abwasser, 139, 360–366. Barraqué, B. (2001) De l’eau dans le gaz à l’usine à gaz. In: Hydrotop 2001, colloque scientiﬁque et technique, Marseille 24–26 Avril, Recueil des Communications, C-068. Brouwer, F., Heinz, I. and Zabel, T. (eds) (2003) Governance of Water-related Conﬂicts in Agriculture, New Directions in Agri-environmental and Water Policies in the EU. Kluwer Academic Publishers, Boston. Cambon-Grau, S. (2000) Baisse des consommations d’eau à Paris: enquête auprès de 51 gros consommateurs. Techniques Sciences Méthodes: Génie Urbain Génie Rural, 2, 37–46. Chesnutt, T.W. and Mitchell, D.L. (2000) California’s Emerging Water Market. Paper presented at the French Académie de l’Eau (available: tom@ antechserv.com).

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Cornut, P. (2000) La Circulation de l’eau Potable en Belgique: Enjeux Sociaux de la Gestion d’une Resource Naturelle. Thèse de Doctorat, Université Libre de Bruxelles, Belgium. Dickinson, M.-A. (2000) Water conservation in the United States: a decade of progress. In: Estevan, A. and Viñuales, V. (eds), La Efﬁciencia del Agua en las Ciudades. Bakeaz y Fundacion Ecologia y Desarrollo. European Council (1980) Council Directive 80/778/ EEC of 15 July 1980 Relating to the Quality of Water intended for Human Consumption as amended by Council Directives 81/858/EEC and 91/692/EEC (further amended by Council Regulation 1882/2003/EC). Council of the European Union, Brussels. European Council (1998) Council Directive 98/83/EC of 3 November 1998 on the Quality of Water Intended for Human Consumption as amended by Regulation 1882/2003/EC. Council of the European Union, Brussels. European Union (2000) Directive 2000/60/EC of the European Parliament and the Council of 23 Oct. 2000 establishing a framework for community action in the ﬁeld of water policy. Ofﬁcial Journal of the European Communities, L327, 1–72. Fougères, D. (2004) L’Approvisionnement en Eau à Montréal: du privé au public, 1796–1865. Sillery, Septentrion, Québec. Hukka, J. and Katko, T. (2004) Liberalization of water sector – a way to market economy or to monopoly market? Water and Wastewater International, 19, 23–25. Juuti, P. and Katko, T. (eds) (2005) Water, Time and European Cities. History Matters for the Futures. Tampere University Press, Tampere. http://tampub. uta.ﬁ/index.php?Aihealue_Id=20 (last veriﬁed 3 October 2008). Lave, L. (1981) The Strategy of Social Regulation: decision frameworks for policy. Brookings Institution, Washington, DC. Maver, I. (2000) Glasgow (Town and City Histories). Edinburgh University Press, Edinburgh. Melosi, M.V. (2000) Sanitary City: urban infrastructure in America from colonial times to the present. Johns Hopkins University Press, Baltimore. Montginoul, M., Rinaudo, J.-D., Lunet de Lajonquière, Y., Garin, P. and Marchal, J.-P. (2005) Simulating the impact of water pricing on households behaviour: the temptation of using untreated water. Water Policy, 7, 523–541.

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Okun, D.A. (1996) From cholera to cancer to cryptosporidiosis. Journal of Environmental Engineering, 122, 453–458. Okun, D.A., Craun, G., Edzwald, J., Gilbert, J. and Rose, J.B. (1997) New-York city: to ﬁlter or not to ﬁlter? Journal of the American Water Works Association, 89, 62–74. Rajah, N. and Smith, S. (1993) Distributional aspects of household charges. Fiscal Studies, 14, 86–108. Rajala, R.P. and Katko, T.S. (2004) Household water consumption and demand management in Finland. Urban Water Journal, 1, 17–26. Salzman, J. (2005) Creating markets for ecosystem services: notes from the ﬁeld. New York University Law Review, 80, 870–961. Tarr, J. (1996) The Search for the Ultimate Sink, Urban Pollution in Historical Perspective, Akron University Press, Ohio.

Tarr, J.A. and Dupuy, G. (eds) (1988) Technology and the Rise of the Networked City in Europe and in America. Temple University Press, Philadelphia PA. Van Humbeeck, P. (1998) An Assessment of the Distributive Effects of the Wastewater Charge and Drinking-water Tariffs Reform on Households in the Flanders Region in Belgium. Report of the SERV (Sociaal-Economische Raad Van Vlandern). Paper presented at the World Bank sponsored Workshop on Political Economy of Water Pricing Implementation, Washington, DC, November 3–5, 1998. Zérah, M.-H. (1997) Inconstances de la distribution d’eau dans les villes du tiers-monde: le cas de Delhi: Unreliability in urban water distribution in developing countries: the case of Delhi Flux, Cahiers Scientiﬁques Internationaux Réseaux et Territoire, 30, 5–15.

Image facing chapter title page: Courtesy of the Macaulay Institute.

11

Managing Catchments for Hydropower Generation

H A A K O N T H A U L O W 1, A R V E T V E D E 2, TOR SIMON PEDERSEN3 AND KARIN SEELOS2 2

1 Norwegian Institute for Water Research, NIVA, Gaustadalléen 21, Oslo, Norway Statkraft Energy Production, Environment and Concessions, Lilleaker, Oslo, Norway 3 Norwegian Water Resources and Energy Directorate; NVE, Licencing and Supervision Department, Majorstua, Oslo, Norway

11.1

Introduction

11.1.1 Water and energy The production of electricity is strongly dependent on water, be it for cooling in thermal power plants, hydropower, the production of biomass or emerging technologies like wave, tidal and osmotic power. Hydropower is, however, the dominant link between water resources and energy. As to water and thermal power; when using surface water for the cooling of towers needed to condense the steam exiting form the steam turbines in the thermal power plants, the return water will be several degrees warmer. The resulting temperature change is capable of affecting aquatic ecosystems. A power plant using a oncethrough cooling approach can kill tons of ﬁsh every year by trapping ﬁsh against intake screens or drawing ﬁsh into the facility. It is important, therefore, to consider the ecological impacts on the aquatic environment. In several major coal-producing countries, including China, parts of India and South Africa and the USA, coal deposits are located in arid

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

areas. In the USA, about 40% of daily freshwater usage is for power generation. Most of this is returned to the source and only about 2% is consumed/evaporated. Water availability is now driving the development of cooling technology for thermal power application. Since emerging water power generation technologies from the ocean, such as waves, currents and tides, are not really in relation with river catchment management, this chapter will mainly focus on osmotic power, which is unleashed when fresh and salt water are mixing close to a river’s estuary. 11.1.2 Hydropower – how does it work? Water falling as rain, hail or snow is transported on the earth’s surface in streams and rivers and temporarily stored in groundwater, snow cover and glaciers before ﬁnally reaching the sea. Hydropower converts the natural ﬂow of water into electricity. Power is generated when the water rushes through turbines on its way to the sea. Thus we beneﬁt from the kinetic energy freed by falling water to produce electricity as a result of transforming the water’s motion into mechanical and electrical energy. A river’s ﬂow volume and the heights from which the water is falling into the turbines, the so-called ‘head of

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Fig. 11.1 Working principles for hydropower. (Renewable Energy 2007. Illustration: Kim Brantenberg.)

water’, determine the potential energy. The head of water is the height difference between water intake and power station outlet. Water is directed into pressure shafts leading down to a power station, where it strikes the turbine runner at high pressure. The kinetic energy of the water is transmitted via the propeller shaft to a generator, which converts it into electrical energy (Fig. 11.1). As hydropower is a very site-speciﬁc technology, hydropower plants can be classiﬁed in a number of ways which are not mutually exclusive. (a) By head (high or low), setting the type of hydraulic turbine to be used Based on the pressure height one can distinguish low- and highhead power plants. Low-head power stations often utilize a large water volume but have a low

head, as in a run-of-river power station. Since regulating the ﬂow of water is difﬁcult, it is used when available. The amount of electricity generated, therefore, increases considerably when the river is carrying more water during the spring thaw or when precipitation is very high. The river is dammed by the power plant to lead the water into one or more turbines. After having ﬂown through the turbines, the water runs out in the river below the power station. High- and low-head hydropower projects Highhead power stations are generally constructed to utilize a high head but smaller volume of water than run-of-river installations. Many types of these power stations store water reservoirs. The water is normally led from the water reservoir in a pressure shaft. At the bottom of the pressure shaft, the water is distributed through pipes to

Managing Catchments for Hydropower Generation (a)

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(b)

Fig. 11.2 High and low head hydropower projects. (a) High head project with penstocks, Tyssedal Power Plant, UNESCO Worlds Heritage site, Western Norway (courtesy of Statkraft). (b) Low head project, Gaur scheme (courtesy of IHA).

the turbines at high pressure, as a result of the large head.. The water pressure drives the turbine and the momentum from the turbine is transferred through a shaft to the generator. Modern high-pressure power plants are normally built into the rock. The power station and regulation reservoir are connected by tunnels through the rock or pipelines down the mountainside. Reservoirs allow a larger proportion of runoff to be used in power production. They usually have a larger installed capacity than run-of-river stations, but a shorter utilization period (Fig. 11.2). (b) By storage capacity (run-of-river or reservoir projects) Run-of-river type plants are using the water ﬂowing in a river and are generating power according to the availability of water in the river. Reservoir projects can beneﬁt from a water storage system which can be natural (i.e. lake) or artiﬁcial (i.e. reservoir). These storage systems allow a more even distribution of water in time and space. They act like a pool to accumulate water in periods with a high inﬂow and low consumption, whereas in periods with low inﬂow and high consumption, this water is then available for use. In other words, the reservoir is a way to store water and thus potential energy until it

is needed. Storage reservoirs have an important function for water and energy security. Not only are reservoirs storing rain in wet seasons to ensure adequate water supply in dry seasons, they also enable the regulatation of power production. Thanks to this regulating ﬂexibility, hydropower projects with storage capacity can stabilize the electric system and provide load control, as they are able to follow the electricity demand rapidly. For example, the power plant with reservoir can produce more electricity during the day, when consumption is at its highest, and less during the night. In these schemes water can also be held back during periods of ﬂood and later be released during periods of drought. Reservoirs can, therefore, have a ﬂood reducing effect. They can be dimensioned to store water for several seasons (Fig. 11.3). A special type of reservoir hydropower project is pumped storage plants. They involve two reservoirs with a signiﬁcant height difference. Water is pumped from the lower reservoir to the upper when there is spare capacity in the electricity network. Once the water is pumped up, it is then again available to generate power, especially at times of peak demand. The power producers can

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(a)

(b)

Fig. 11.3 Run-of-river and reservoir type hydropower plants. (a) Run-of-river plant (45MW), Rivières-des-Prairies, Montreal, Canada (courtesy of Hydro-Québec). (b) Hydropower plant with reservoir (1528MW), Québec, Canada (courtesy of Hydro-Québec).

achieve an economic proﬁt by pumping low-lying water up to higher altitude regulation storage reservoirs because the water’s potential energy increases in proportion with the head of water. When power prices are low, it can be proﬁtable for the producers to use energy to move the water to a higher storage reservoir, so that the water can be used for production in periods with high prices. Such a system uses more power than it generates, but is essential as a ﬂexible reserve and can make an electricity network more efﬁcient (Fig. 11.4). (c) By purpose (single or multipurpose) As hydropower does not consume the water that drives the turbines, it is available for other uses (Fig. 11.5). In fact, a signiﬁcant proportion of hydropower projects are designed to fulﬁll more than one function. Besides generating electricity, they prevent or mitigate ﬂoods and droughts, they offer the possibility of enhancing agricultural yields through irrigation, and to supply water for domestic, municipal and industrial use, as well as improving conditions for navigation, aquaculture, ﬁshing, tourism or leisure activities. However, these different water uses come along with conﬂicting demands on water utilization

Fig. 11.4 Pumped storage power plant, Goldisthal, Thüringen, Germany (courtesy of IHA).

leading to trade-offs which highlight the need for an integrated water management plan per catchment. (d) By size (large, medium, small, mini, micro) Hydropower covers a wide variety of project scales, from less than 0.1 MW (micro hydro) to over 10,000 MW. The classiﬁcation by size is done according to the installed electricity generating capacity expressed in Mega Watts. Although there are some regional variations in

Managing Catchments for Hydropower Generation the deﬁnition of small hydro, this term usually refers to a project that is less than 100 MW, while mini hydro usually has an installed capacity of about 1 MW. All in all, there are several different types of hydropower projects. Each has speciﬁc design characteristics, which enable them to respond to different energy and water needs and to supply a wide range of services. Obviously, each type of project also produces speciﬁc types and magnitudes of impacts (Fig. 11.6). 11.1.3 Impacts of hydropower Since every human activity has its impacts, it is important to minimize negative impacts and maximize positive effects of a project, through

Fig. 11.5 Multipurpose hydropower storage scheme. (a)

Fig. 11.6 Small and large scale hydropower developments. (a) Chamuera, Rätia, Switzerland (0.55MW) (courtesy of IHA). (b) Macagua, Venezuela (15,910MW) (courtesy of EDELCA, Venezuela).

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careful planning, design, construction and operation. Appropriate impact assessment should be done in a holistic way, based on life-cycle assessment and due consideration of all three dimensions of sustainable development: environmental protection, social development and economic viability. The advantages and disadvantages of hydropower are listed in Table 11.1 (International Hydropower Association 2003a). Hydropower uses renewable water supplies, not ﬁnite fossil fuels. In contrast to nuclear power, it leaves no toxic waste to threaten future generations, and in contrast to thermal power, it emits very limited amounts of greenhouse gases. Hydropower contributes to reduced use of fossil energy sources such as oil, gas or coal. Average CO2 emissions from the production of the 21 largest power producers in EU-15 were 358 kg CO2 MW−1 h−1 in 2003 (Price Waterhouse Coopers/ Enerpresse 2003). In comparison, emissions from the production at a typical Norwegian hydropower plant are approximately 0.15 kg CO2 MW−1 h−1 and 0.7 kg CO2 MW−1 h−1 in total, including the construction of the power plant (Statkraft 2002). Seen from the viewpoint of the climate change challenge and the need for a shift from fossil-based to renewable energy, hydropower is very environment friendly. Negative effects of a hydropower project are inevitably borne by the local communities, whereas the beneﬁts in the form of reliable electricity supplies or clean air and emissions reductions are shared by everyone in the nation or (b)

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Table 11.1 Advantages and disadvantages of the hydropower option (International Hydropower Association 2003a) Advantages Economic aspects Provides low operating and maintenance costs Provides long life span (50–100 years and more) Meets load ﬂexibility (i.e. hydro with reservoir) Provides reliable service Includes proven technology Instigates and fosters regional development Provides highest energy efﬁciency rate (payback ratio and conversion process) Generates revenues to sustain other water uses Creates employment opportunities Saves fuel Provides energy independence by exploiting national resources Optimizes power supply of other generating options (thermal and intermittent renewables) Social aspects Leaves water available for other uses Often provides ﬂood protection May enhance navigation conditions Often enhances recreational facilities Enhances accessibility of the territory and its resources (access roads and ramps, bridges) Provides opportunities for construction and operation with a high percentage of local manpower Improves living conditions Sustains livelihoods (freshwater, food supply)

Disadvantages High upfront investment Precipitation dependent In some cases, the storage capacity of reservoirs may decrease due to sedimentation Requires long-term planning Requires long-term agreements Requires multidisciplinary involvement Often requires foreign contractors and funding

May involve resettlement May restrict navigation Local land use patterns will be modiﬁed Waterborne disease vectors may need to be checked Requires management of competing water uses Effects on impacted peoples’ livelihoods need to be addressed

Environmental aspects Produces no atmospheric pollutants and only very few GHG emissions Enhances air quality Produces no waste Avoids depleting non-renewable fuel resources (i.e. coal, gas, oil) Often creates new freshwater ecosystems with increased productivity Enhances knowledge and improves management of valued species due to study results Helps to slow down climate change Neither consumes nor pollutes the water it uses for electricity generation purposes

Modiﬁcation of hydrological regimes Modiﬁcation of aquatic habitats Water quality needs to be monitored/managed Temporary introduction of methylmercury into the food chain, needs to be monitored/managed Species activities and populations need to be monitored/ managed Barriers for ﬁsh migration, ﬁsh entrainment Sediment composition and transport may need to be monitored/ managed

region. The more direct impacts of hydropower schemes fall into two main categories: environmental and social impacts. The environmental impacts of hydropower are related to changes in the ecosystems caused by obstructing a river’s ﬂow with a dam, by modify-

ing ﬂow regimes and water levels as well as by building roads and power lines. Building a dam implies to transform a fast-ﬂowing river environment, into a calm waterbody, similar to a lake. In consequence, the most severe negative impacts are borne by ﬁsh who are migrating or living in

Inundation of terrestrial habitat

Managing Catchments for Hydropower Generation fast-ﬂowing rivers. However, a lot of mitigation and compensation measures are available to ensure that the affected species acknowledge no losses. The effect of ﬂow regulation on ﬁsh and ﬁshing is a complicated interaction between a number of physical and biological factors. The natural habitat of ﬁsh is formed by physical circumstances such as water level, water velocity, refuge possibilities, suitable spawning opportunities and food supply. The amount of water will also affect ﬁsh in different ways, depending on the age of the ﬁsh and the ﬁsh species. Yet, a number of regulated river systems are still very good ﬁshing rivers. Among others, regulated rivers ensure a minimum water ﬂow, which avoids natural survival threats to ﬁsh such as the drying out of rivers. The building of reservoirs can also affect biodiversity, wildlife nourishment areas and migration routes. In tropical areas, standing water can lead to thermal stratiﬁcation, where a layer without oxygen forms in the bottom of lakes and reservoirs. These anoxic conditions favour the activity of bacteria who create methane – a greenhouse gas – by decomposing organic matters. Hydropower often entails changes to the natural variations in the water in a watercourse. Run-of river power plants without water storage reservoirs cause relatively small changes to the level and ﬂow of water, and therefore have little effect on biodiversity. In high-pressure power plants with reservoirs, the impact on biodiversity depends on the regulation height. Changes to the water level throughout the year can lead to scouring of ﬁne particles which affect the water quality and erosion in the regulation zone. When the power plants are built, natural settings are modiﬁed by the construction of roads, worker’s camps and quarries or sand pits. However, these impacts can be offset by requiring that the constructor takes care of revegetation and rehabilitation of the landscape, so that the intervention has as little permanent impacts as possible. Powerlines do seriously affect the landscape, but can hardly be avoided. Powerlines can affect bird populations, either through collision or by short circuiting.. It is possible to put power transmission lines underground. However

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digging and blasting of ditches affects hydrology and vegetation and implies about ten times higher construction costs. As with environmental impacts, social impacts vary considerably from scheme to scheme. There are both advantages and disadvantages. In some countries and areas resettlement in the reservoir area can be the most important issue. For example, the Three Gorges Project in China required the relocation of 13 smaller cities to give room to the reservoirs (see Box 11.5). However, in Canada the total installed hydro capacities is over 70,000 MW and has not required any involuntary resettlement. Other disadvantages are possible restrictions on navigation, modiﬁcation of land use patterns and effects on peoples` livelihoods. The issue of waterborne disease vectors also needs to be carefully managed in tropical regions. Among the positive social impacts ﬁgure improved ﬂood and drought protection. Moreover, hydropower projects with fresh water storage schemes make water available for other uses such as irrigation, aquaculture, water supply and recreation, particularly if the reservoir is planned for multiple purposes right form the onset. Sometimes reservoirs also improve the opportunities of cost-effective water-based transport. In addition, the accessibility of new territory and its resources can be improved through new roads, boat ramps and bridges. A hydropower project provides also working opportunities during construction and operation with normally a high percentage of local manpower. Sustainable hydropower development in developing countries is considered an important tool in poverty alleviation strategies contributing to improve economic viability and enhanced social justice. Hydropower can also be one important solution to meet speciﬁc needs for rural electriﬁcation. 11.2 Historical Perspective 11.2.1 History of hydropower The mechanical use of falling water is an ageold tool. It was used by the Greeks to turn

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waterwheels for grinding wheat into ﬂour, more than 2000 years ago. The availability of cheap slave and animal labour, however, restricted its widespread application until about the twelfth century. During the Middle Ages, large wooden waterwheels were developed with a maximum power output of about 50 hp. Modern large-scale water-power owes its development to the British civil engineer John Smeaton, who ﬁrst built large waterwheels out of cast iron. Water-power played an important part in the Industrial Revolution. It gave impetus to the growth of the textile, leather and machine-shop industries in the early nineteenth century. Although the steam engine had already been developed, coal was scarce and wood unsatisfactory as a fuel. Water-power helped to develop early industrial cities in Europe and the USA until the opening of the canals provided cheap coal by the middle of the nineteenth century. Dams and canals were necessary for the installation of successive waterwheels when the drop was greater than 5 m. Large storage-dam construction, however, was not feasible and low water ﬂows during summer and autumn, coupled with icing during the winter, led to the replacement of nearly all waterwheels by steam when coal became readily available. The earliest hydroelectric plant was constructed in 1880 in Cragside, Northumberland, England. The rebirth of water-power came with the development of the electric generator, further improvement of the hydraulic turbine and the growing demand for electricity by the turn of the twentieth century. By 1920, hydroelectric plants

Box 11.1

already accounted for 40% of the electric power produced in the USA and in the 1940s, hydropower provided about 75% of all the electricity consumed in the West and Paciﬁc Northwest and about one-third of total US electrical energy. In fact, hydropower generation from the Paciﬁc North West in the USA played a pivotal role in the Second World War, by powering the very energy intensive ship and aircraft production required to sustain the allies in Europe. With the increase in development of other forms of electric power generation, hydropower’s percentage has slowly declined. Today hydropower provides about 16% of the world’s electricity supply. Early hydroelectric plants were direct current stations built to power industrial activities and lighting during the period from about 1880 to 1895. The years 1895 through 1915 saw rapid changes in hydroelectric design and a wide variety of plant styles built. Hydroelectric plant design became fairly well standardized after the First World War. By the end of the twentieth century there were over 45,000 large dams in over 140 countries. The period of economic growth following the Second World War saw a phenomenal rise in the global dam construction rate, lasting well into the 1970s and 1980s. At the peak, nearly 5000 large dams were built worldwide in the period from 1970 to 1975. The decline in the pace of dam building over the past two decades has been notable, especially in North America and Europe where most technically attractive sites are already developed. The average large dam today is about 35 years old (World Commission on Dams 2000).

Norway – a hydropower country

The ﬁrst hydropower plant was built in 1877. Up to 1940 a total of more than 2000 hydropower plants were built, producing 9.5 TWh annually The period 1960–1990 was a particularly expansive time. The annual electricity production quadrupled during that period, to 108 TWh annually. Today (2007), average annual hydropower production in Norway is 121 TWh. Hydropower development in Norway was triggered by the need for small- and large-scale industrial development, and the availability of cost-effective/affordable hydropower and timber for forest industry stimulated the development. However, the need to electrify town and villages also promoted hydropower, and especially transmission development. The Rjukan example, with the

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Saaheim power station, commissioned in 1916, is an example of the development of an energy intensive industry close to important hydropower resources. In 1916, the Saaheim power station ﬁgured among the largest in the world with its nine Pelton turbines producing 12 MW each.

Saaheim power station (108 MW in 1916, 189 MW in 2002)

The typical large hydropower plant in Norway today is a high head underground power plant, and a typical small hydropower plant is tailored to the local environment with a powerhouse and a penstock. The the main reservoir of Norway’s largest power plan Kvilldal; a high head underground station, is shown below.

The rockﬁll dam at Blåsjø, the main reservoir for Norway’s largest power station, Kvilldal (1240 MW) in the Ulla Førre hydro scheme, is tailored to the landscape.

Most power plants are publicly owned (87%), and the system today has been developed by county and municipal power companies (large, medium and small hydro) ∼57%, a state-owned power company (large and medium hydro) ∼30%, and private companies, mainly industrial enterprises (large and medium hydro) ∼13% (International Hydropower Association 2003b).

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11.2.2 The role of hydropower in energy generation and economic development Hydropower has played and plays an important role in the economic development of many countries and regions all over the world. The modern history of hydropower is closely linked to the developments of dams which have been promoted as an important means of meeting perceived needs for water and energy services and as long term, strategic investments with the ability to deliver multiple beneﬁts. Some of these additional beneﬁts are typical of all large public infrastructure projects, while others are unique to dams and speciﬁc to particular projects. Today the world’s large dams regulate, store and divert water from rivers for agricultural production, human and industrial use in towns and cities, electricity generation and ﬂood control. Dams have been constructed to a lesser extent to improve river transportation and, once created for other purposes, the reservoirs of many large dams have been used for recreation, tourism and aquaculture. About one-third of large dams serve

two or more purposes (Fig. 11.5). Recent trends have favoured multi-purpose dams and there is considerable regional variation in the functions served by large dams and these functions have also changed over time (Fig. 11.7). Africa is an example where the potential of hydropower development could be key for economic growth and poverty alleviation, since so far only about 7% of its economically feasible hydropower resources have been developed. Africa is an important user of traditional (noncommercial) energy, namely biomass. The per capita electricity consumption is particularly low in Central, East and West Africa.. Hydropower presently provides 22% of electricity generation in Africa; nuclear power provides 2%, while thermal power stations provide 70%. Reliance on hydropower, however, is 80% or greater in Cameroon, the Democratic Republic of the Congo, Ghana, Mozambique, Rwanda, Uganda and Zambia. Because of the enormous potential of Africa’s great rivers, particularly the Zambezi and the Congo, hydropower is seen as the driving force for future development in Africa.

Fig. 11.7 Distribution of existing large dams by region and purpose. (World Commission on Dams 2000.)

Managing Catchments for Hydropower Generation In many countries hydropower has been extremely important for economic development and Norway is a good example in this context. With only 4.7 million inhabitants Norway is the world’s sixth biggest producer of hydropower and the largest in Europe. Hydropower is the main source for electricity production with a share of approximately 99%. Thanks to the development of its ‘white coal’, Norway has evolved from one of Europe’s poorest nations at the turn of the nineteenth century.

Box 11.2

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11.2.3 Environmental and social conﬂicts emerging: international guidelines for sustainable hydropower In the early days of hydropower development, improving the living conditions of people was so imperative that negative impacts on the environment tended often to be overlooked and regarded as less important compared with the beneﬁts offered by development. Large-scale dams, dry riverbeds, roads and transmission

The Alta River conﬂict

The most well-known controversy connected to hydropower development in Europe is the Alta River conﬂict which occurred in Norway during the late 1970s and 1980s. The background for the conﬂict was plans for developing a large hydropower plant on the Alta River, in the county of Finnmark, northern Norway in order to supply this area with sufﬁcient electricity, as Finmark could not import electricity from other Norwegian regions at this time due to a lack of interconnections in the national transmission grid. Finnmark is the traditional area of the Sámi people who were at that time dependent on the migrating reindeer herds. The legal status for the Sámi people and the rights for the land and the water were in addition unclear. Moreover, the Alta River was known worldwide for its stock of large Atlantic salmon. The ﬁrst plans made in the 1960s considered damming the river and ﬂooding a Sámi village. This plan was, never ofﬁcially submitted, but the idea was publicized through the media and created much anger within the Sámi community. In April 1974, an application for concession was forwarded by Statkraftverkene (the State hydropower company) which at that time was part of NVE (Norwegian Water and Electricity Board). NVE at that time was also responsible for the licensing process. The application was only half-ﬁnished when it was forwarded and most of the environmental impact assessments (EIAs) were not ﬁnished or were not even started, including the effects on the salmon. It should be stated that in the 1960s it was not unusual that EIAs in the applications for hydropower licences were carried out while the application was under public processing. Despite the missing EIAs, the Norwegian Parliament agreed to work out a licence in 1978. The resistance against the scheme increased and in 1978 the People’s Action Movement was established in Alta. In 1979, the ﬁrst demonstration camp was established, stopping access to road construction. Hunger strikes were started in front of Parliament and the Government stopped construction work. In 1980, the Government decided that construction work should continue considering that the interests of the majority should not be compromised by the interests of a minority. The demonstrations continued, however, and in February 1981, a Sámi women’s group occupied the Prime Minister’s ofﬁce and a new hunger strike started. In the spring of 1981 the access road construction was once more stopped while archaeological investigations continued. In February 1982, the Supreme Court decided that the licensing procedure had been within the existing law but the guidelines for EIAs should be improved to avoid similar conﬂicts in the future.

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lines, however, brought visual and ecological changes to the environment that became less and less acceptable, once peoples’ basic needs were met. In particular, the creation of large reservoirs without the implementation of appropriate mitigation and measures had a notable negative impact on their immediate surroundings. Damage to the local environment and inadequate provision for those affected in the area has contributed to the opposition by some environmental and human rights organizations towards the dam developers and the hydropower industry. There are several examples of conﬂicts related to hydropower schemes. In Norway, the development of watercourses for hydropower was considered as having only positive impacts until the 1970s. Several schemes caused protests from environmentalists who found increased support from the general public. The protests culminated with the Alta River conﬂict (Fig. 11.8). Lessons learned from the Alta conﬂict were: • The fact that both the environmental and the social conﬂicts emerged from local opposition,

initiated a fervent, local ‘People’s Action Movement’, but with strong national and international support. • Deﬁcient Environmental Impact Assessments (EIA) were part of the reason for the Alta River Conﬂict. • The conﬂict triggered a development towards better EIAs and more openness in the licensing procedure. • Although local resistance was overruled by the government to meet the regional energy needs, the demonstrators and the Sámi people have gained a lot of attention. • Among others, the Sámi Parliament and more focus on Sámi questions is a direct result of the Alta case. The development of the watercourses in Norway for hydropower production took place during a period of 100 years without a coordinated plan for the whole country. However, the increasing conﬂicts and fewer remaining resources initiated the government in the 1970s and 1980s to launch hydropower-related nationwide planning activities. (a) National protection plan for watercourses After the ﬁrst plan was adopted in 1973, the plan was revised and extended protection plans were adopted in 1986, 1993 and 2005. As a result, close to 400 rivers with a hydropower potential of 44 TWh are now protected from development. The purpose of the plans is to maintain the environmental diversity from the mountains to the sea, by protecting whole catchments. Yet, the current plans only protect against hydropower development, whereas a careful policy should also address other kinds of development activities that might have even greater environmental impacts. There is still an opening for the development of mini and micro hydropower (<1 MW) in protected watercourses, but only if the development does not counteract the protection criteria. In practice, the policy is restrictive and permissions are only granted in special cases.

Fig. 11.8 The Alta Dam. (Courtesy of the Norwegian Institute for Water Research).

(b) Master plan for hydropower development The Protection Plan for Watercourses was

Managing Catchments for Hydropower Generation a defensive type of plan ‘taking’ watercourses out of hydropower planning. But conﬂicts due to environmental concerns increased and the Master Plan for Hydropower was initiated in the early 1980s. The ongoing conﬂicts during the Alta concession handling pushed the decision to start the plan. A white paper to the Parliament in 1980, Norway’s Future Energy – use and production, asked for development of a national master plan for hydropower. The Government was in demand for an extended planning and licensing system that took into account not only the particular hydropower scheme, but also hydropower development at a broader scale, including consideration of socioeconomic and environmental issues. The ﬁrst version of the Master Plan passed the Parliament in 1986 and categorizes actual hydropower projects according to economic and techni-

Box 11.3

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cal viability, as well as potential conﬂict levels regarding environmental and cultural values. In its ﬁrst version 340 projects were divided into three categories: • Category I: hydropower projects ready for immediate licensing and consequently ‘go projects’. • Category II: hydropower projects for later consideration. • Category III: ‘no go’ projects due to disproportionately high development costs and/or high degree of conﬂict with other user interests, including environmental interests. Categories II and III were later merged.. Projects with an installed capacity less than 10 MW or an annual production below 50 GWh are, however, exempt from appraisal through the Master Plan system. The Master Plan has been revised twice since 1986.

Increasing runoff and power production at the Alta power station

The Alta power station is mainly a run-of-river type station, but has a medium sized storage reservoir which is used for part of the winter season. The station was put into operation in 1987. In the ﬁnal licence given by the Norwegian Parliament in 1978, the mean annual production was estimated to be 625 GWh. There has been no rebuilding of the Alta power station since 1987. 800 750 700 Planned production

GWh

650 600 550 500 450 400 1988-92

1993-97

1998-02

2003-07

Mean annual power production for the period 1988–2007 is 706 GWh or 81 GWh higher than was estimated in 1978. The increase in production has been largest in the last years, and this trend is continuing in 2007 which gave a record production of 878 GWh. As other factors have been constant during this period, the increase in power production is a result of increasing runoff which is potentially the result of climate change. If the Alta power plant were to be redesigned today, it is highly likely that the turbines in the station would have been planned with increased capacity.

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(c) Regional plans for small hydropower development In Norway, the interest for small hydropower (<10 MW) is growing rapidly. At present the annual production from small hydropower is 5–6 TWh, of a theoretical potential of 25 TWh. More than 200 applications are currently in some stage of the licensing process. The licensing follows the regulations in the Water Resources Act, but is simpliﬁed compared to larger projects. A general description of possible environmental impacts and conﬂicts is required, and a detailed report on biodiversity, with focus on red listed species, is compulsory. The ﬁrst step in the process is to demarcate ‘planning areas’ in each county based on maps for the development potential for small hydropower. The second step is to map interests that are sensitive to small hydropower, such as landscape, biodiversity, recreation and tourism, cultural heritage, ﬁshery, unaffected ‘wilderness’ areas and Sami interests (reindeer husbandry). These sector interests are classiﬁed according to their intrinsic ‘value’, in order to deﬁne possible areas of conﬂict. The ﬁnal step includes development of management policies and regulations based on the systematized information for each of the planning areas. These plans now constitute the framework for individual licensing of hydropower projects in Norway. Internationally, increasing conﬂicts prompted the World Commission on Dams (WCD) initiated by the World Bank and IUCN, the World Conservation Union in 1997. The WCD recommended starting with an options assessment which included the exploration of alternative options to building a dam to secure needs for water, food and energy. The WCD also recommended a comprehensive and participatory assessment of a full range of policy, institutional and technical options. Social and environmental aspects should be given the same signiﬁcance as technical, economic and ﬁnancial factors. Focus was set on increasing the effectiveness and sustainability of existing structures for water, irrigation and energy systems. The WCD suggested a negotiated agreement between the developers

and the stakeholders before decisions to develop could be taken. These principles, as well as the WCD’s core values and strategic priorities, achieved a large consensus among all stakeholders. However, some recommendations published in the ﬁnal WCD report in 2000 without prior consultation were subject to controversies, such as giving a veto-right to minorities, or requiring a legally binding agreement to be signed with each affected individual. Professor José Goldemberg, ex-commissioner of the WCD, stated at the World Summit on Sustainable Development in 2002: ‘Not a single country has adopted the WCD guidelines for the simple reason that literal adoption makes it impossible to build anything.’ As a consequence, even the World Bank as a main sponsor was not favourable to implement the WCD guidelines. However, the multi-stakeholder dialogue instigated by the WCD has been considered to be very valuable and has been carried on by the UNEP hosted Dams and Development Project (2001–2007). Within the hydropower community, more recent guidelines have been worked out in collaboration with other stakeholders by the International Hydropower Association (IHA) in 2004. These guidelines are based on the core values and strategic priorities of the WCD, but are focusing more speciﬁcally on hydropower issues, while the WCD addressed all kind of dams generally. In addition, a pragmatic rating tool has been developed and adopted in 2006, the Hydropower Sustainability Assessment Protocol, to facilitate a comprehensive evaluation of environmental, social and economic key aspects related to hydropower projects in their planning, construction and operation phase. To ensure broad consensus and state-of-the art practice, a multi-stakeholder forum has been established recently in early 2008. Forum members represent developing countries, developed countries, ﬁnancing agencies, the hydropower sector, and well respected NGOs representing environmental, economic and social concerns related to hydropower projects. The Hydropower Sustainability Assessment Forum is ﬁnanced amongst others by the Norwegian,

Managing Catchments for Hydropower Generation Icelandic and German governments, as well as by the World Bank. Rapidly developing countries such as China, India and Turkey frequently argue that their electricity requirements for economic growth and social development outweigh the environmental concerns surrounding hydropower, and that support for large hydropower development is a pro-poor policy. Sustainable hydropower development as a source of renewable, clean energy has received since the beginning of this century increased support in international policy-making, such as the Johannesburg Plan of Implementation (United Nations, 2002), the ministerial declarations issued at the 3rd and 4th World Water Forum (Kyoto in 2003 and Mexico in 2006), the ﬁrst International Conference for Renewable Energy (Bonn in 2004), the UN Symposium on Hydropower and Sustainable Development (Bejing, 2006), the UN Conference of Parties (Montreal, 2005) and lately at the African Ministerial Conference on Hydropower and Sustainable Development (Johannesburg, 2006). Some NGOs, however, are still campaigning to have large hydropower excluded from global efforts to promote renewable energy. Among the arguments advanced for this position are: • including large hydro in renewable initiatives reduces the available funding for new renewable energy technologies; • there is no technology transfer beneﬁt from large hydro, which is a mature technology; • large hydro projects often have major social and ecological impacts; • in tropical areas, reservoirs can emit signiﬁcant amounts of greenhouse gases; • the sustainability of reservoirs can be compromised by sedimentation. Yet, all energy and water supply options do have signiﬁcant impacts and especially large-scale interventions entail major changes. Since in the case of hydropower the management of a vital resource is at stake, it is of prime importance that

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this is done in a open, transparent and sustainable way.

11.3 Current Solutions 11.3.1 The resource base Global energy generation has increased from approximately 5 cm GWh−1 to nearly 16 m GWh−1 since the 1970s (Fig. 11.9). Although the share of hydropower in total word energy supply was only 2.2% in 2002 (Fig. 11.10), hydropower accounted for 19% of all electric power generated. Currently, hydropower makes up approximately 90% of the renewable energy production in the world. Hydropower supplies at least 50% of electricity production in 66 countries and at least 90% in 24 countries. Most of the installed hydropower capacity is found in Europe and North America. The largest potential still unexploited is found in Africa, Asia and South America where respectively only 33%, 7& and 22% of the technically feasible hydropower potential is used. Asia has an economic potential of 2600 TWh−1 yr−1, and an addition of approximately 153,000 MW of new hydropower is planned. China’s production is slightly less than 300 TWh, but the technical/ economical potential equals about 1900 TWh. The total global potential for electricity production from hydropower is about 14,000 TWh (2006). Approximately 60% is considered economically feasible today. There are large differences as to its importance in different countries and regions. For instance hydropower contributes to 15% of the European production. In certain countries like Norway, Austria and Sweden the share of hydropower is crucial (respectively 99%, 60% and 40% of the national electricity production). On the other hand in the UK and Germany its importance compared to the total production is small (respectively 3% and 5%). The importance of hydropower is not only reﬂected in the quantity of produced electricity, but also in terms of unique services that hydro-

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Fig. 11.9 The development of global generation of electricity by source 1971–2001. Global energy generation has increased from approximately 5 m GWh to nearly 16 m GWh since the 1970s (UNESCO 2006).

Fig. 11.10 Total primary energy supply by source, 2002 (International Energy Agency 2005).

power with storage capacity provides to the stability of an electric system. Hydropower not only plays a key role as peak demand regulator, since thermal and nuclear power plants are not able to take care of the rapid switches in demand appearing over a day, but also balances the highly variable electricity supply from other renewable energy sources, such as wind and solar. For the sake of illustration, three hydropower projects are presented in the boxes below: the

Three Gorges in China, the world’s largest hydropower plant to be completed in 2009; Palmiet in South Africa, a pumped storage scheme located in a protected area with particular rich biodiversity; and Nam Ngum Dam in Laos, a hydropower project which also provides several other services to the local communities. The three examples show the issue of hydropower generation in a wider perspective of integrated water resource management.

Managing Catchments for Hydropower Generation

Box 11.4

269

The Three Gorges Dam, China

Under construction since 1994, The Three Gorges Project is one of the world’s largest multi-purpose water infrastructures. It is located in the Yangtze River, which is both China’s and Asia’s longest river, with a length of 3200 km from west to east and for more than 1000 km from north to south and draining an area of 1808,500 km2. The Yangtze river has been dammed with a 2.3-km-long and 185-m-high dam. When all the turbines are in place in 2009, 26 Francis turbines of 18,200 MW in total, will produce as much as 84.7 TWh yr−1. The primary purpose of the Three Gorges Project is to improve ﬂood control, as the very densely polutated Yangtze River valley has been regularly prone to devastating inundations over the last centuries which have caused massive losses of life. Only in 1954, the river ﬂooded 193,000 km2 of land, killing 33,169 people and forcing 19 million people to move.

Km

75

m

Wei YELLOW SEA

CHINA

DABA SHAN

Huai

Han

Nankin

Jialing Shanghai Yichang

Wuhan

Yangzi RUSSIA

Chongqing

Three Gorges Dam Yangzi

Yangzi

MONGOLIA JAPAN CHINA

detail Pacific Ocean

Area of reservoir of proposed Three Gorges Dam

Location: Yangtze River in Sandouping, Yichang, Hubei, PRC. Catchment area: 1.8 million km2 home to more than 358 million of people; nearly 35% of the national population. Population density: 199 per km, 1.86 times of the national average. Multipurpose functions: ﬂood control, navigation, water supply, power generation. Energy production: 84.7 TWh yr−1. Installed capacity: 22,400 MW based on a 2.3-km-long and 185-m-high gravity dam. a 600-km-long reservoir (average width of 1.1 km) covering a surface area of 1,084 km2. Storage capacity: 39.3 billion m3 including 22.15 billion m3 anti-ﬂood storage capacity. (Data from: China Three Gorges Project Corporation (CTGPC).) Flood control and navigation The reservoir’s ﬂood storage capacity will reduce the frequency of major downstream ﬂooding from once every 10 years to once every 100 years. The Yangtze River is China’s watercourse with the most developed inland navigation. The installation of ship locks at the Three Gorges Dam is intended to increase river shipping from 10 million to 100 million tones annually, while cutting transportation costs by 30–37%. Shipping will become safer, since the gorges, 139 strong rapids and the treacherous shoals from Ichang and Chonging were notoriously dangerous to navigate.

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Environmental issues Building such a colossal infrastructure entails major modiﬁcations in the watershed and in the land use of the surroundings. Although the overall rationale of the project is to improve living standards in China, 13 smaller cities have been moved. The signiﬁcant number of more than 2.3 million of people to be resettled has created important social disruption also for the host communities. In addition, cases of corruption are highly likely to occur when signiﬁcant amounts of capital are at stake in rather poor areas, which affects the effectiveness of compensation measures. The creation of the reservoir has on the one hand improved the conditions for navigation and reduced emissions from the transport sector, but involves also a considerable loss of some 1300 archaeological sites and natural gorges. Uncertainties remain about the extent of sedimentation of the reservoir, potentially amplifying the effects of the reservoir’s weight on seismic activities. In order to maximize the utility of the Three Gorges Dam and to diminish sedimentation, China plans to build a series of dams upstream of the Yangtze river, including Wudonge Dam, Baithetan Dam, Xiluodu Dam and Xiangjiaba Dam. The total capacity of those four planned dams is 35,500 MW. Water quality issues are being addressed by increased implementation of treatment plants for industrial and municipal sewage water, since for the time over one billion tons of wastewater are released annually into the river, which presents a major threat to aquatic biodiversity. The Three Gorges Project affects among others the winter habitat of the endangered Siberian crane and contributes to the functional extinction of the Baiji, the Yangtze river dolphin. Special sturgeon reserves and hatcheries combined with a rescue programme for rare species have been established to protect the biodiversity of the Yangtze. The State Council and the Three Gorges Commission have developed a special ecological protection plan named ‘7 + 1’, which comprises seven pilot projects and one modelization initiative. This plan is governing the monitoring of ecological issues in the modiﬁed catchment and reservoir area and provides also a pollution prevention programme in rural and urban areas, in addition to the biodiversity protection project. According to the Yangtze River Water Resources Committee, the Three Gorges Project, despite important negative impacts on the micro-ecology, also provides signiﬁcant positive impacts at a macro-ecological angle. The hydropower produced allows 50 million tons of coal equivalent for China to be saved every year. Thus the Three Gorges Project helps to improve China’s air quality and its related respiratory health problems, by avoiding annual a signiﬁcant amount of atmospheric emissions; approximately 100 million tons of carbon dioxide, 2 million tons of sulphur dioxide, around 10,000 tons of carbonic oxide, about 3700,000 tons of nitrogen oxides, as well as a large quantity of industrial waste discharges, which will greatly alleviate environment pollution and acid rain in China, its neighbouring countries and regions. Although this project has been subject to major controversies, the ﬁrst status report and other sources showcase that the overall beneﬁts for China might well offset the environmental and social burdens to that it will have to bear, provided the sensitive issues continue to be managed in a responsible way. The project has a long history, starting in 1919, and was controversial when the plan was approved by the National Peoples Congress in 1992. In the last year, though, some Chinese ofﬁcials have been expressing concern about the negative impacts on water quality and bank erosion.

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Palmiet, South Africa

Palmiet Pumped Storage Scheme is located close to Grabouw in the Western Cape, near Cape Town between Palmiet and Steenbras Rivers. This is in the middle of a protected area which shows a particular rich biodiversity: the scenic and ecologically sensitive Cape Fynbos Plant Kingdom. The area has recently been given international recognition as the Kogelberg Biosphere Reserve by UNESCO – the ﬁrst such reserve in South Africa.

Generating Mode Upper Reservoir Headrace Tunnel

Surge Shaft

Surface Building Lower Reservoir

Pressure Shaft Pressure Tunnel

Power House Complex

Tailrace Tunnel

Installed capacity: 400 MW including two dams and reservoirs. Rockview Dam (upper reservoir): 20,800 ML of capacity. Kogelberg Dam (lower reservoir): 19,300 ML of capacity. Catchment area Palmiet River: 535 km2. Average production in 3 years: 637 GWh. Multipurpose functions: Electricity production, load balancing and water supply The Palmiet Pumped Storage Scheme was built between 1983 and 1988, and is a joint venture between Eskom, a state-owned power utility, and the Department of Water Affairs and Forestry (DWAF). This partnership enabled the optimization of the use of water. Since South African electricity generation is dominated by coal-ﬁred power stations, the Palmiet pumped storage hydro scheme provides valuable back-up generation in the event of the loss of thermal plant or during periods of peak demand for electricity. Thus, the power station is viewed by the electric grid operator as the key to system stability. The Palmiet River catchment has a much greater run-off than that of the adjoining catchment around Cape Town. Construction of the pumped storage scheme allowed Eskom and the DWAF not only to provide hydro-peaking power but also much needed additional drinking water supply for Cape Town. During winter rainfall months, excess water in the Palmiet River (up to 25 million m3 per year) is pumped to the upper reservoir for transfer to the Steenbras Dam and the Cape Town water consumer. The Palmiet Pumped Storage Scheme is a good example of a hydropower plant built and operated in the context of integrated water management (IWRM). The Palmiet Catchment Management Committee was formed in 1996. The vision of the committee is: ‘To manage the Palmiet River Catchment Area so that optimal use is made of the total resources (land, water and air) to sustain

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the ecological, social and economic requirements and to maintain the unique conservation status and scenic beauty of the area’. As a result of public consultation and participation meetings, representation was drawn from Eskom, local government, recreational organizations, farming organizations, industry/business, environmental bodies, governmental departments, tourism bureau, ratepayers associations, civic and farm workers associations. The committee serves as a forum for local issues ranging from water matters to land use and includes stakeholders from the source to the estuary of the Palmiet River. The committee ensures that ongoing catchment management is implemented and adhered to so that the Palmiet Pumped Storage Scheme remains compatible with its surrounding environment. It is recognized that the Catchment Management Plan for the Palmiet River may serve as a blueprint for other catchments.

Box 11.6

Nam Ngum Dam, Laos

The Nam Ngum Dam is located 60 km north of Vientiane, on the 354-km-long Nam Ngum River, a tributary of the Mekong River. The catchment area is 8460 km2 which is home to about 1 million people.

Population density (people/km2)

China

100 500

Viet Nam

Lao PDR

Flood prone areas

Gulf of Tonkin

Thailand Nam Ngum Dam

Cambodia

Gulf of Thailand

0

South China Sea N 300 kms

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Installed capacity: 150 MW. Yearly production: 960 GWh. Storage area: 400 km2. Laos is one of the world’s poorest and least developed countries, with annual per capita income in 1994 estimated at US$295–335. There is very little industry, with agriculture and ﬁshing the most common commercial activities. Laos’ hydropower potential is one of the key resources of the country. The country has an estimated generation potential of some 18,000 MW from over 60 project sites on tributaries of the Mekong River, of which less than 3% have been developed. The ability to sell electricity to Thailand has signiﬁcant implications for the Laotian economy. During the 1960s, power production in Laos was insufﬁcient to meet the national demand. The Mekong Commission as part of a national development plan constructed Nam Ngum 1, the ﬁrst hydropower scheme in Laos between 1968 and 1971. The plant has multipurpose functions, though the project was originally implemented as a standalone hydropower project. The dam and the reservoir provide several other services to the local communities including: Flood management Current operating rules for the storage are designed to protect downstream populations and agricultural crops. The rules include a requirement to draw down the water level prior to commencement of the wet season, providing a 1 billion cubic meter ﬂood retention capacity. Creation of a commercial ﬁshery The importance of a newly created ﬁshery was an unexpected outcome of the Nam Ngum 1 project. Fishery development projects contributed by the Netherlands and Swizerland included construction of primary schools and water supply networks, the provision of monoﬁlament gill nets and boat engines to selected villages. The Nam Ngum Reservoir Management and Development Organization (NMDO) manage the lake ﬁshery. According to this organization, the annual production of economically important species is about 6000 tons. The ﬁshery has created jobs and signiﬁcant growth in the economy of the region. Of the approximately 16,000 people living around the shoreline of Nam Ngum Lake, it is estimated that over 92% are engaged in commercial ﬁshing activities. Irrigated agriculture The creation of the Nam Ngum Power Station enabled the use of electric pumps, making lift irrigation a viable option. This created more than 13,000 ha of new dry season agriculture, and in 1998 approximately 25% of the country’s rice production came from areas that were cultivated by means of lift irrigation. Tourism More recently, tourist developments have appeared at Nam Ngum, including the Dan Sa Vanh – Nam Ngum resort, boasting a hotel, restaurant, golf course and gaming room. 21,000 tourists visited the resort in 2002, 70% of them foreigners, contributing US$630,000 to the local economy. The resort is expected to expand its operations, and it is projected that this will result in a US$5 million boost to the local economy annually.

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haakon thaulow et al. 11.3.2 The outlook for hydropower: climate change mitigation and adaptation

In Europe, the EU Commission has identiﬁed hydropower as a key renewable energy source. Several EU policy proclamations stress the importance of hydropower; among others the Renewable Energy Sources Directive (2001/77/ EC). Hydropower currently dominates the renewable energy generation in the 25 EU member states. As for future potential of renewable energy sources, recent analyses suggest that wind energy and biomass have an important role. These two technologies can be expected to deliver most of the increase in electricity from renewable sources in the EU for 2010. The production of electricity from these sources, however, will ﬂuctuate over time (wind) and will not have the ﬂexibility (biomass) demanded by the electricity market. Such technologies will be dependent on hydropower as a regulator or stabilizer of the general electricity supply. The choice made by the different European countries to reach their national indicative targets on renewable electricity may vary considerably. Thus, the importance of further development of hydropower is likely to be different in various EU countries. Part of the potential for development may also come from the modernization of existing hydropower facilities.

In this context, hydropower with storage capacity is becoming an increasingly attractive option which is not only providing a special balancing service to the electric grid, but which also offering interesting synergies arising from multi-purpose uses of reservoirs, such as ﬂood and drought protection, water-based transport, aquaculture, irrigation, municipal and industrial water supply and new recreational opportunities. However, policy-making in Europe acknowledges for the time being signiﬁcant contradictions between the aims of the Renewable Energy Sources Electricity Directive and the Water Framework Directive. Therefore, increased efforts at harmonizing these conﬂicting frames will have to be made in the near future. It is obvious that future climate change could bring about a new situation for hydropower production as well as the pattern of demand compared to the present situation. Reduction of CO2 emmissions (20% in Europe within the year 2020, 30% by 2050) will put pressure on development of new hydropower and refurbishing of existing facilities (Fig. 11.11). In some areas, hydropower may beneﬁt from increased precipitation, while in other countries the potential will decrease due to reduced river runoff. Countries in Scandinavia and northern Russia could see an increase of 15–30% in potential, whereas Southern Europe and countries such

Fig. 11.11 The EU targets for energy policy towards the year 2020.

Managing Catchments for Hydropower Generation as Portugal, Spain, Ukraine and Turkey could see a decrease of 20–50%. In areas with increased precipitation and runoff, dam safety may become a problem due to more frequent and intensive ﬂooding events. In this respect, increased storage capacity provided by reservoirs can be a structural measure to better cope with the rising variability of precipitation pattern. The EU Commission, the member states, the hydropower industry and NGOs have (European Union (EU) 2007) concluded that a consequence of increased temperatures in Europe could be increase the need for electricity for cooling and

Box 11.7

275

lower production capacity for thermal and nuclear power plants. In fact, in the summer of 2006 the French electricity producer EDF purchased electricity from the EU wholesale energy market to make up for its lost capacity from thermal plants as a result of increased temperatures. Nuclear power plants could not be operated at full capacity since water law restricted the discharge of cooling water at high temperatures. Hence, hydropower could not only be a tool to improve climate change adaptation, but also a part of the solutions to mitigate climate change thanks to its low carbon footprint.

Hydropower and the environmental goals in the WFD

Hydropower is one of the most important reasons for poor ecological status of freshwater in Europe and is addressed in a thorough way in the WFD and in the consequent River Basin Management Plans (RBMP). Heavily Modiﬁed Water Body (HMWB) The WFD states that speciﬁc hydro-morphological water uses of great importance to society, such as hydropower, should qualify for alternative and less stringent environmental goals. The goals are set based on a scientiﬁc and socioeconomic basis and are set on an individual basis. The consideration on which water bodies qualify as HMWB is subject to an elaborate process. The WFD states that: The beneﬁcial objectives served by the modiﬁed hydromorphological characteristics of the water body cannot, for reasons of technical feasibility or disproportional costs be achieved by other means which are a signiﬁcantly better environmental option.

For the Alta River, the national characterization has classiﬁed a river stretch of 20 km as HMWB. This is based on the modiﬁed hydro-morphological characteristics. Good ecological potential (GEP) The WFD sets an alternative goal for water bodies where the ecological situation is considerably changed by hydro-morphological pressure The acceptable goal is termed Good Ecological Potential (GEP) and is in principle what can be achieved after all possible mitigating measures are considered and those which are technically feasible and socio-economically acceptable. The goals are individual and site speciﬁc and could be governed by national and regional priorities. The environmental objectives should be quantiﬁed with respect to ecological, chemical and hydro-morphological criterion. One of the key issues with respect to GEP is the question of an ‘ecological hydromorphological continuum’. The general rule is that a water body considered HMWB should have a discharge sufﬁcient to provide an environment for water living species to migrate and reproduce.

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11.3.3 Hydropower in the context of water resources management: the EU Water Framework Directive Integrated water resources management is considered a prerequisite for a holistic approach when it comes to ecological and socioeconomic good management practice. Hydropower is one of the most important physical encroachments in watercourses in many countries. In this respect, the EU Water Framework Directive is an important and very ambitious tool within the European Community with the aim to improve the water environment in a socio-economically sound way. The directive is not a protection tool, but rather a tool for the improvement of water conditions. The directive sets speciﬁc ecological environmental goals to a number of indicator species and to chemical and hydromorphological conditions. Hydromorphological pressures are the most important reason for rivers and watercourses not meeting the general goals expected by the communities. In Europe, the characterization done in accordance with the implementation of the Water Framework Directive showed that more than 60% of waterbodies did not meet the demands of the directive. Hydropower together with transportation is the most important reason for that. Characterization of the ecological status in waterbodies form the basis for the River Basin Management Plans (RBMP) in the WFD. The RBMP is a four-step process starting with local analyses which come up with quantiﬁcation of existing ecological conditions, what the desired new condition is and what the ecological goal could be and how it could be achieved. New hydropower will, as any new added pressure in the watercourses, be subject to thorough considerations in the RBMP to make sure they qualify to the exception that designation as a HMWB represents. A workshop in Berlin in June 2007 (Ecologic 2007) addressed issues of hydropower and water resources management in the context of the WFD and recommended the promotion of sustainable hydropower. Some key recommendations from the workshop were:

• Hydropower is a highly reliable, CO2-free and renewable source of electricity production, but there is also a need to maintain the ecological functions of regulated water stretches to strike a good balanced in meeting climate, water and nature protection objectives. It is important to ensure that existing and forthcoming EU policies which encourage the development of hydropower ensure also coherence with the WFD as well as other EU environmental legislation. Hydropower schemes must consider the ecological impacts on the affected waterbodies and the adjacent wetlands. • An environmental assessment based on WFD criteria is required for all waterbodies including those with hydropower plants. This assessment includes other environmental criteria and a socio-economic assessment taking all water uses into account. Hydropower development should take into account future climate change impacts and possible future conﬂicts between new hydropower priorities due to climate change impacts and the aims of the WFD to achieve Good Ecological Status (GES) or GEP. • National and European instruments (such as tradable certiﬁcates, feed-in tariffs, support schemes for renewables or ecolabelling) to support and promote hydropower development should be linked to ecological criteria for the protection of water status. • There should be a clear insight into all costs and beneﬁts of hydropower compared to other available energy options based on a life-cycle assessment. This insight will help sustainable decision-making on hydropower projects and implementing the polluter pays principle. • The EU Commission, the member states and industry recognized the advantages of preplanning mechanisms to facilitate the identiﬁcation of suitable areas for new hydropower projects. These pre-planning mechanisms should take into account WFD and other environmental criteria as well as socio-economic aspects, including other water uses. The use of such preplanning systems could assist the authorization process through faster implementation. • At least three categories of area could be distinguished for pre-planning: suitable, less favour-

Managing Catchments for Hydropower Generation able and non-favourable areas. These categories should be identiﬁed with the involvement of all stakeholders based on transparent criteria and they should be monitored and revised within a period of time. • Small and large hydropower should be treated equally. Assessments should be based on sustainability criteria taking into account basin-speciﬁc as well as site-speciﬁc WFD criteria and global environmental criteria (climate change) and not on the size of the hydropower plant per se. In addition, the need to consider new hydropower in the light of climate challenges, the conference ‘Time to Adapt, Climate Change and the European Water Dimension’ (February 2007) emphasized the importance of seeing different policies in context in the planning process. Beside new hydropower, the possibilities that lie within refurbishment and expansion of existing hydropower facilities must be considered. In Norway alone, the technical potential represents approximately 10 TWh which constitutes about 9% of the nation’s total production.

11.3.4 Key tools in sustainable hydropower management: IHA Sustainability Guidelines and Assessment Protocol Other planning tools and recommendations are provided by the International Hydropower Association (IHA) in the Sustainability Guidelines (2004) and the Sustainability Assessment Protocol (2006). The IHA Sustainability Guidelines state that hydro developers planning a project should try to minimize the following (United Nations (UN) 2006): • health dangers, particularly from waterborne diseases or malaria; • loss of homes, farms and other livelihoods; • disruption of community networks and loss of cultural identity; • changes to biodiversity in the affected area. They should try to maximize the following: • timely consultation at all levels; • the ﬂow of relevant information to all those affected; • negotiated settlement of disputes;

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• timely and adequate payment of any compensation. Where people or communities have to be transferred to new sites, developers should do the following: • investigate possible alternative ways of doing the project; • ensure adequate consultation with the people to be displaced throughout the project; • guarantee equivalent or improved livelihoods at the new location; • provide better living standards and public health at the new location. The IHA (2006) has also agreed upon a Sustainability Assessment Protocol as a support tool to the implementation of the IHA Sustainability Guidelines. It attempts to measure the degree of sustainability during planning, implementation and operation. The tool in the protocol enables the rating of 20 key aspects on a score card at all three project stages. The 20 key aspects are of environmental, social and economic nature and each have one process- and one performance-related criteria wich can be rated according to a score of ﬁve levels. The emphasis in the assessment protocol is on a systematic approach comprising a four-step accomplishment of sustainable projects: (i) Plan it; (ii) Do it; (iii) Check that it works; and (iv) Act to correct any problems. The assessment protocol suggests a thorough auditing of critical issues to make sure that potential and existing hydropower projects are built and operated in a sustainable way. The IHA guidelines coincide in many ways with the ideas of WFD, but go further in addressing other issues than the environment. They especially address general socio-economic considerations and aspects of community acceptance, poverty alleviation and enhancement of living standards. These guidelines also highlight that sustainability is the outcome of an efﬁcient co-operation among all members of society. Whereas the hydropower sector can take responsibility for skilful and conscientious planning, building and operation of a hydropower project, governments also have a signiﬁcant role to play, when it comes

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to governance issues, to arbitrating conﬂicting water interests or to implementing appropriate legal frameworks to ensure a transparent and inclusive decision-making process as well as clear rules within national and regional energy policies.

11.4

New Scientiﬁc Insights and Future Research Requirements

In many ways the existing hydropower technology must be considered to be mature, highly developed and efﬁcient. No big technological breakthroughs are needed in order to develop it further. Any mature technology can be improved, however, but on the technical side research and development (R&D) is mainly a question of improving already known technologies and methods. The book Renewable Energy (Norwegian Water Resources and Energy Directorate, 2007) points out some areas for future developments: • Modernizing by continuously upgrading ageing equipment. • Developing better methods to handle the consequences of disconnection and reducing disconnection periods by shifting ruined equipment effectively. • Developing better methods for risk assessment and prioritization of investments. • Improving technical changes in new markets related to operation. • Integrating hydropower with other renewable and distributed energy technologies such as wind power and hydrogen. Possibly the challenges in R&D lie not so much in the hydropower technology itself, but the relations and mix with other energy sources: renewable or not renewable. Hydropower with a role as peak power supplier and interaction with intermittent producers like wind power and small hydro is challenging. There are further great challenges facing research into hydropower’s role in integrated water resources management and river basin management. New knowledge has been

achieved on environmental and social impacts. IHA has collected much information, for example in the publication ‘The Role of Hydropower in Sustainable Development’ (International Hydropower Association 2003a). The Norwegian Government has launched the programme Energy21 in order to increase the R&D on energy development in Norway (not including the oil and gas industry). Proposals to Energy21 on hydropower include many of the points mentioned before. In addition more knowledge is needed on: • Ecological optimalization of the production facilities. • Increased knowledge about biological, physical or chemical consequences of changes in discharge or water level. • Increased knowledge about mitigating measures and how successful they are. • New production patterns (e.g. more hydro peaking) and the consequences for the water environment caused by climate change. • Shifts in species and a new set of mitigation measures due to climate change 11.5 Implications New insights and future research will have implication on the hydropower option in numerous ways. 11.5.1 Assessment of hydropower design in WFD management plans The WFD puts more focus on ecology, especially on river continuity in the design process. The focus shall be at the catchment level. For new hydropower projects, the tests lined out in Article 4.7 in the WFD must be applied. It may be more or less impossible to get new hydropower projects that will change the rivers/lakes from GES to HMWB. There is, however, still little experience in the EU on the use of this test. It has been indicated there is a need to introduce more preplanning mechanisms, and point out ‘no-go’, ‘may-be-go’ and ‘go’ areas. This approach has many parallels to the Norwegian Master Plan.

Managing Catchments for Hydropower Generation Another mechanism that will probably inﬂuence the design of new hydropower plans is the market for ‘green’ electricity certiﬁcates. 11.5.2 Impacts of production, design and running due to climate change The implications on hydropower production from excepted climate changes have been addressed earlier. A study on the impacts of climate change on renewable energy in the Nordic countries (Fenger 2007) shows that the Box 11.8

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annual electricity production from existing hydropower stations may increase from 190 TWh in the period 1961–1990 to 240 TWh for the period 2071–2100. Most of this increase will come from Norway (Fig. 11.12). The study also indicates that the hydropower companies may expect more stress on dams from higher frequency of ﬂoods. Stricter rules for dam design and dam operations must, therefore, be introduced. On the global scale, the changes in precipitation pattern are more complicated. From the IPCC assessment

New insights and other user interests in the Alta River

Against a difﬁcult background, operational rules were set to be temporary for 5-year periods. Moreover, a comprehensive programme for follow-up studies was set up, mainly focusing on impacts on Atlantic salmon. The operational objectives are to: • obtain satisfying conditions for the salmon and good ecological potential; • maintain stable ice cover on the river for recreational users; • avoid dangerous ice runs; • reduce dangerous ﬂoods; • reach 10% reservoir storage at the start of the spring ﬂood; • obtain satisfying economic results. In order to help the daily operator to achieve these objectives, an advisory committee was established, representing the local salmon ﬁshing stakeholders (named ALI), the Finnmark county administration and one ice expert from the national Water Resources Authority. The committee is headed by the state-owned power company Statkraft Energy. So far, the use of this advisory committee has been a very successful experience. The hydropower plant operators base their decisions for the next operation period on: • hydrological forecasts based on weather forecasts and snow situation; • the ice conditions in the river; • the water reserve in the reservoir; • special events (dog-sledge race, vacations); • the electricity demand; • input from the Manoeuvering Advisory Committee. Statkraft has the responsibility to manage the water resources by respecting as much as possible all the different interests which are sometimes conﬂicting. For the municipality and for ALI, the Advisory Committee is of great importance and they want this experience to be continued as a permanent institution. Their main argument is that the operational decisions that Statkraft make are taken several hundred kilometres away from Alta and by personnel who frequently changes jobs. ALI argues that they represent the continuity and the practical know-how of the river. There is also a growing need to communicate to the public the status for the salmon in the Alta River. In order to meet this need, Statkraft publishes an Environmental Status sheet (see ﬁgure following). The sheet is updated every second year as new results from the ongoing research are known.

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Environmental Status - River Alta Last updated April 2008 Length of river: 230 km Anadrome stretch: 47 km Catchment area: 7,389 km2 Mean water discharge: 88 m3 /s (annual average) Characterisation according to the Water Framework Directive: 20 km has been defined as Heavily Modified Power plants: Alta Power Plant, opened 1987 Power production: 655 GWh, corresponds to the electricity consumption of approx. 32,750 households. Reservoir: Virdnejavri

Virdnejavri reservoir seen from the south.

Salmon and sea trout catches 600

30

500 400

20

300

10

200 100 2002

1998

1994

1990

1986

1982

1974 700000 600000 500000 400000 300000 200000

1989 1991 1996 1997 1999 2000 2001 2002 2003

Spawning pits were counted by means of helicopter observations and diving, and are considered minimum estimates. The yellow light is due to the number in the Sautso area remaining rather low.

100000 0 2004

Water discharge at Kista

2005 2006 Source: Literature list pts. 4

Water temperature at Sautso

600

Regulation has led to higher winter water flow and reduced spring floods. The annual volume is unchanged.

Average before regulation

1-Dec

1-Nov

1-Oct

1-Sep

1-Jan

1-Dec

1-Oct

1-Nov

1-Sep

1-Jul

1-Aug

1-Jun

1-Apr

1-May

1-Mar

1-Jan

1-Feb

0

1-Aug

2 –1

1-Jul

100

5

1-Jun

200

8

1-May

300

11

1-Apr

400

14

1-Feb

Average before regulation Average after regulation

500

Water temp [°C]

17

1-Mar

No. of spawning pits per km river

1978

Source: Literature list pts. 4

No. of salmonsmolt migrating to sea

No. of spawning pits

Water flow [m3/s]

0

0

Populations with densities above the blue line are not considered to be threatened or vulnerable (directorate for nature management) Source: Literature list pts. 4

90 80 70 60 50 40 30 20 10 0

Total river catch Norway

40

Total catch [tonnes]

Alta river Catch Alta [tonnes]

90 80 70 60 50 40 30 20 10 0

1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003

No. of fish per 100 m2

Density of juvenile salmon older than 1 year

Average after regulation

Regulation has led to a slight rise in winter temperatures, lower summer temperatures and higher autumn temperatures.

Data has been collected from the Norwegian Water Resources and Energy Directorate (NVE), Statistics Norway, the County Governor’s Office, Alta Salmon Fishing Association and Statkraft. The assessments shown as ’traffic lights’ have been carried out by SINTEF Energiforskning. The meaning of the traffic lights is as follows:

Good and stable status

Acceptable and/or variable status

Information sheet. Environmental status – River Alta.

Poor and unstable status

Managing Catchments for Hydropower Generation

Box 11.9

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Developing the power of osmosis

Osmotic power production is based on the different salinity of freshwater and seawater. This new technology does not require the damming of a river and hence has very little impacts on the environment. Fresh water and salt water are diverted into a plant located at the shore of an estuary, where a semi-permeable membrane (i.e. a membrane that retains the salt ions but allows water to ﬂow through) is installed between reservoirs containing fresh water and sea water respectively. As a consequence, a net ﬂow of fresh water (80–90%) towards the sea water side will be observed. Provided the salt water compartment has a ﬁxed volume, the pressure will reach up to 26 bars. This pressure is equivalent to a column of water which is 270 m high and can be used to drive a conventional turbine. Statkraft has developed this new technology together with several research partners and a ﬁrst small-scale pilot project for osmotic power generation will be built in Norway close to Oslo in 2008. The potential to generate electricity from osmosis is very high. Studies have shown that for Norway 20% of its current electricity production could eventually come from osmotic power, whereas for Europe an annual production of 180 TWh is indicated in Energy21.

(Fig. 11.13) it is clear that areas which are already wet will get wetter and the already dry areas will get drier. It is expected that mountainous areas with storage capacity will have an increased role to play to meet Europe’s peak demands. This ‘swing’ operation of those facilities will have negative impacts through increased variations in discharge which will have to be managed thoroughly. Increased autumn/winter precipitation and

increased glacier melting will increase the water availability in Norway’s and Iceland’s reservoirs. On the contrary, in Southern Europe, more conﬂicts on the use of declining water resources are anticipated. Norway has large hydropower plants that can produce ﬂexible ‘peaking power’ without severe ecological impacts, as many power stations have direct outlets into a fjord. In the future it is expected that Norway will be able to export this

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Fig. 11.12 Changes in HP production in the Nordic countries due to expected climate change (Fenger 2007).

Fig. 11.13 Changes in precipitation pattern (International Panel on Climate Change (IPCC) 2007).

‘peaking power’ southwards and add ﬂexibility to the inﬂexible base load electricity produced by the thermal power facilities (Fig. 11.14) of continental Europe. In order to achieve this, however, there is a need for more transmission lines. A step in this direction is the new 700 kV transmission line between Norway and the Netherlands ofﬁcially opened in June 2008. It should be stressed, however, that these developments will need large ﬁnancial investments and this will have to be reﬂected in electricity prices. Increasing transmission capacity also means that market prices in Norway will move up towards the European price level.

It is predicted that the price for electricity in Europe will increase in the coming years as a consequence of the EU energy and climate policy towards the year 2020 (see an illustration of the energy shortfall and demand needed up to the year 2030 in Fig. 11.15). 11.5.3 Increased need to co-ordinate licensing, national power plans and national/international energy planning The different challenges in the years to come quite clearly point towards a need to co-ordinate the energy and environment planning and poli-

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Fig. 11.14 Importance of hydropower production as peak load regulator (Taugbøl 2007).

Fig. 11.15 The need in GW for new installations to balance out the growth demand and to replace old installations in the EU15 countries. (Source: EPPSA – European Power Plant Suppliers’ Association and ABB.)

cies on an international level. In the European countries, the WFD will be an important tool in co-ordinating national licensing and environmental standards in surface waters. This probably will have an inﬂuence on the development of new hydropower and the rehabilitation of older hydropower developments. Harmonization of criteria and methods must be undertaken to avoid discriminations in the European market for electricity from hydropower. The EU Flood Directive was ﬁnished in October 2007 and shall be implemented before November 2009. The Flood Directive is connected to the WFD and the text points out how the two directives shall be

combined. However, the Flood Directive seems to give some opportunities for combining new hydropower schemes with ﬂood mitigation measures through the exceptions in the WFD. It should also be discussed if there can be stronger international co-ordination in the policies of watercourse protections. Today it seems that most countries make their national protection plans without looking too much at their neighbours. However, watersheds and energy markets are not ending at political boundaries. A challenge should be to ensure that all the different types of watercourses are represented in the sum of the protection plans, i.e on the

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European level. A part of such an evaluation should also include the hydropower potentials in the watercourses. The deregulation of the electricity market in the EU is moving on slowly and some countries are even showing signs of a more nationalistic policy. This may be a threat to the EU energy policy pointed out in section 11.3.2. Eurelectric – the Union of the Electricity Industry in Europe – has published a study with different scenarios on the possibilities to reach the energy demands in EU towards 2050 (Eurelectric 2007). The study recommends following a scenario which envisages the use of all options towards a low-carbon energy system. The policy recommendations, in brief, are: • unleash the potential of energy efﬁciency; • develop a low-carbon electricity system by using all available options; • intelligent electriﬁcations of the economy; • consistent deployment and a market-oriented approach; • global co-operation on global issues. One of the sub-recommendations is to replace much of the traditional oil consuming road transport with hybrid cars which can use renewable electricity as fuel. In order to put these recommendations into action, the study states that strong international policy co-ordination is needed and that more attention must be given to market mechanisms. References Ecologic (2007) Water Framework Directive and Hydropower. Common Implementation Strategy Workshop, Berlin 4–5 June 2007. Key Conclusions and Summary Report. Ecologic, Berlin. http://www. ecologic-events.de/hydropower/ (last veriﬁed 22 October 2008). Eurelectric – Union of the Electricity Industry (2007) The Role of Electricity. A new path to secure, competitive energy in a carbon-constrained world. Eurelectric – Union of the Electricity Industry, Brussels. European Union (EU) (2007) Water Framework Directive and Hydropower. Common Imple-

mentation Strategy, Berlin 4–5 June 2007. Key conclusions. Fenger, J. (ed.) (2007) Impact of Climate Changes on Renewable Energy Sources: their role in the Nordic energy system. Nordic Council, Copenhagen. International Energy Agency (IEA) (2005) Energy Statistics www.iea.org/textbase/stats/index.asp (last veriﬁed 30 October 2008). International Hydropower Association (IHA) (2003a) The Role of Hydropower in Sustainable Development. IHA, Sutton, Surrey, UK. International Hydropower Association (IHA) (2003b) First International Summit on Sustainable Use of Water for Energy. Country Report. IHA, Norway. International Hydropower Association (IHA) (2004). Sustainability Guidelines. IHA, Norway. International Hydropower Association (IHA) (2006) Compliance Protocol and Sustainability Guidelines. IHA, Norway. International Panel of Climate Change (IPCC) (2007) Climate Change 2007: the physical science basis. Working Group I Contribution to the Fourth Assessment Report. Cambridge University Press, Cambridge, UK. Price Waterhouse Coopers/Enerpresse. (2003) Price Waterhouse Coopers/Enerpresse, November 2003, European Carbon Factors. http://www.pwc.fr/fr/ pwc_pdf/pwc_carbon_factor_uk.pdf (last veriﬁed 13 October 2008). Renewable Energy (2007) Joint edition: Norwegian Water Resources and Energy Directorate, Enova/ Norwegian Research Council, Innovation Norway (2007) Renewable Energy. www.renewableenergy (last veriﬁed 13 October 2008). Statkraft (2002) Statkraft, Environmental Declaration of Trollheim Power Station. Statkraft, Oslo. United Nations (2002) World Summit on Sustainable Development: political declaration and Johannesburg Plan of Implementation. http://www.un.org/esa/ sustdev/documents/WSSD_POI_PD/English/WSSD_ PlanImpl.pdf (last veriﬁed 30 October 2008). United Nations (UN) (2006) Water – a shared responsibility. The UN World Water Development Report 2. World Water Assessment Programme. UNESCO (2006) Water: a shared responsibility. United Nations World Water Development Report 2. Berghan Books, New York. World Commission on Dams (2000) Dams and Development: a new framework for decisionmaking. Earthscan Publications, London.

Image facing chapter title page: Courtesy of the Norwegian Institute for Water Research.

The Danube River – the Most International River Basin

12

PHILIP WELLER1 1

International Commission for the Protection of the Danube River (ICPDR), Vienna, Austria

12.1

Introduction

The activities of the International Commission for the Protection of the Danube River (ICPDR) and Danube River Basin countries towards implementation of the Danube River Protection Convention (1994), as well as the EU Water Framework Directive (European Commission 2000), has created the basis for future-oriented water management addressing key water management issues in this large river basin. This approach has succeeded in the Danube River Basin, with its diverse historical, social economic and political circumstances, and therefore it could be applicable in other river basins. The Danube is the most international river basin in the world. Nineteen countries share the territory of the basin and 81 million people live within the basin. The region is geographically, politically and economically diverse. Despite the differences the countries of the Danube are co-operating together on water management issues and have signed the Danube River Protection Convention in June 1994. The Convention came into force in 1998. The Convention is a legal framework for cooperation in water management and assures protection of water and ecological resources and their sustainable use. Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

The main elements of the Convention are: • to strengthen international co-operation in water management; • to ensure conservation, improvement and rationale use of water; • to reduce inputs of nutrients and hazardous substances and their impact on the Black Sea, and control ﬂoods. To make the Danube River Protection Convention a living tool, the International Commission for the Protection of the Danube River Basin (ICPDR) has been created. All countries with major territory in the Danube are contracting parties to the Convention (except Montenegro which has now ratiﬁed the Convention and will be formally a member shortly) as is the European Community. The work of the Commission is co-ordinated by a small secretariat but undertaken by representatives of the countries who all actively participate in the work of the ICPDR.

12.2 The Danube River Basin The Danube River Basin is the second largest river basin of Europe, covering 801,463 km2. It is located to the west of the Black Sea in Central and South-eastern Europe. To the west and northwest the Danube River Basin borders on the Rhine River Basin, in the north on the Weser, Elbe, Odra and Vistula River Basins, in the northeast on the Dnjestr, and in the south on the catchments of the rivers ﬂowing into the Adriatic

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philip weller

Sea and the Aegean Sea (International Commission for the Protection of the Danube River 2005a). The Danube can be divided into three main parts based on its geological and geographical conditions. • The Upper Danube Basin stretches from the sources in the Black Forest Mountains to the Gate of Devín, east of Vienna, where the foothills of the Alps, the Small Carpathians and the Leitha Mountains meet. The area covers in the north the Swabian and Frankonian Alb, parts of the Oberpfälzer, the Bavarian and the Bohemian Forests, the Austrian Mühl- and Waldviertel, and the Bohemian-Moravian Uplands. South of the Danube lie the Swabian-Bavarian-Austrian Alpine Foothills as well as large parts of the Alps up to the water divide in the crystalline Central Alps.

Box 12.1

• The Middle Danube Basin covers a large area reaching from the Gate of Devín to the impressive Iron Gates gorge, which divides the Southern Carpathian Mountains in the north and the Balkan Mountains in the south. The Carpathians in the north and the east, and Karnic Alps and the Karawankas, the Julian Alps and the Dinaric Mountains in the west and south conﬁne the Middle Danube Basin. This circle of mountains embraces the Pannonian Plains and the Transsylvanian Uplands. • The Lower Danube Basin covers the RomanianBulgarian Danube sub-basin downstream of Cazane Gorge and the sub-basins of the Siret and Prut River. It is conﬁned by the Carpathians in the north, by the Bessarabian Upland Plateau in the east, and by the Dobrogea and Balkan Mountains in the south.

The Trans National Monitoring Network

In June 1994, the Convention on co-operation for the protection and sustainable use of the Danube River (DRPC) was signed in Soﬁa, coming into force in October 1998 with the main objectives of achieving sustainable and equitable water management, including the conservation, improvement and rational use of surface and ground waters in the Danube catchment area. DRPC also emphasizes that the Contracting Parties shall co-operate in the ﬁeld of monitoring and assessment. In this respect, the operation of the Trans National Monitoring Network (TNMN) in the Danube River Basin aims to contribute to the implementation of the DRPC. The TNMN has been in operation since 1996, but the ﬁrst steps towards it were taken about 10 years earlier. In December 1985 the governments of the Danube riparian countries signed the Bucharest Declaration. The Declaration had as one of its objectives to observe the development of the water quality of the Danube, and in order to comply with this objective a monitoring programme containing 11 cross-sections on the Danube River was established. The objective of the TNMN is to provide a structured and well-balanced overall view of the status and long-term development of quality and loads in terms of relevant constituents in the major rivers of the Danube Basin in an international context. The TNMN builds on national surface water monitoring networks. To select monitoring locations for the purposes of international monitoring network in the Danube River Basin, the following selection criteria for monitoring location had been set up: • located just upstream/downstream of an international border; • located upstream of conﬂuences between the Danube and main tributaries or main tributaries and larger sub-tributaries (mass balances); • located downstream of the biggest point sources; • located according to control of water use for drinking water supply. Monitoring location included in TNMN should meet at least one of the selection criteria. The ﬁnal TNMN contains 78 monitoring locations (78 sampling stations).

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The Danube River – the Most International River Basin Table 12.1 Coverage of the states in the Danube River Basin (DRB) and estimated population (data source: competent authorities in the DRB unless marked otherwisea

State Albania Austria Bosnia and Herzegovina Bulgaria Croatia Czech Republic Germany Hungary Italyb Macedonia Moldova Poland Romania Serbia and Montenegroc Slovak Republic Slovenia Switzerland Ukraine Total

Code AL AT BA BG HR CZ DE HU IT MK MD PL RO CS SK SI CH UA

Ofﬁcial coverage in DRB (km2)

Digitally determined coverage in DRB (km2)a 126 80,423 36,636 47,413 34,965

21,688 56,184 93,030 565 109 12,834 430 232,193 88,635 47,084 16,422 1,809 30,520 (801,463)

Percentage of DRB (%)

Percentage of DRB in state (%)

Population in DRB (Mio.)

Percent of population in DRB (%)

<0.1 10.0 4.6 5.9 4.4 2.9 7.0 11.6 <0.1 <0.1 1.6 <0.1 29.0 11.1 5.9 2.0 0.2 3.8 100

0.01 96.1 74.9 43.0 62.5 27.5 16.8 100.0 0.2 0.2 35.6 0.1 97.4 90.0 96.0 81.0 4.3 5.4

<0.01 7.7 2.9 3.5 3.1 2.8 9.4 10.1 0.02 <0.01 1.1 0.04 21.7 9.0 5.2 1.7 0.02 2.7 81.00

<0.01 9.51 3.58 4.32 3.83 3.46 11.60 12.47 0.02 <0.01 1.36 0.05 26.79 11.11 6.42 2.10 0.02 3.33 100

a For the purpose of comparison the coverage of the states was calculated using GIS based on the DRBD overview map. These values differ slightly from the ofﬁcial data of some countries, since other methods of calculation have been used. b Data source: Autonomous Province of Bozen – South Tyrol. c According to the 2002 census the population in Serbia and Montenegro without the provinces of Kosovo and Metohia is 7668,000 inhabitants. On the territory of Kosovo and Metohia the last census was in 1981. On the basis of this census and OEBS data the estimated population of Kosovo and Metohia in the Danube river basin today is about 1300,000 inhabitants. The information is relevant to the time when Serbia and Montenegro still belonged to the same state.

Due to the richness in landscape the Danube River Basin has a tremendous diversity of habitats through which rivers ﬂow, including glaciated high-gradient mountains, forested midland mountains and hills, upland plateaus and through plains and wet lowlands near sea level. The population residing within the Danube basin is 81 million Table 12.1 provides a listing of the countries (before the separation of Serbia and Montenegro) that have territory in the basin and their population as well as the percentage of the population living within the basin. In addition to being a major river basin itself, the Danube also has a number of sub-basins that

are equal to or greater than the size of other important international river basins in Europe. The Tisza River Basin, for example, is the largest sub-basin in the Danube River Basin (157,186 km2) and is slightly larger than the Elbe River Basin (148,268 km2). Five countries (Ukraine, Romania, Hungary, Slovakia and Serbia and Montenegro) have territory within the Tisza River Basin. In addition, the Sava River is the largest Danube tributary by discharge (average 1564 m3 s−1) and the second largest by catchment area (95,419 km2). The Sava River Basin includes territories of Slovenia, Croatia, Bosnia and Herzegovina, Serbia and Montenegro.

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The Danube River Basin District has been deﬁned in the frame of the work of the ICPDR. It covers: (i) the Danube River Basin; (ii) the Black Sea coastal catchments on Romanian territory; and (iii) the Black Sea coastal waters along the Romanian and partly the Ukrainian coast. The geographical coverage of the Danube River Basin District as well as the competent authorities is shown in Figure 12.1 (International Commission for the Protection of the Danube River 2005a). The outer boundary of the Danube River Basin District was deﬁned taking into consideration the hydrological boundaries of the surface waters and groundwater. The Danube River, however, is not only impressive because of its size. It contains highly diverse geographical regions and has a rich human history. It is also characterized by major socioeconomic differences among the countries. There is a strong gradient from the upper to the lower Danube for indicators of economic strength. Germany and Austria have economies among the strongest of Europe, while downstream in Moldova and Ukraine the per capita income is less than 1000 Euro per annum. This fact greatly inﬂuences the basin-wide management of water resources and restoration and pollution reduction efforts. For example, the percentage of population connected to public water supply as well as to sewage disposal shows great differences and standards from West to East.

12.3 Co-operative Management: the ICPDR and its History Progressively during the past three decades the countries of the region began co-operating in response to the reduction of the quality of Danube waters. Even before the collapse of the Iron Curtain attempts were made to jointly exchange data on water quality. The political changes that altered the region at the end of the 1980s and early 1990s, however, provided both an impetus and incentive to strengthen the co-operation. Efforts began to formalize the co-operation

through the development of a convention that was eventually signed in Soﬁa, Bulgaria on 29 June 1994. The ‘Convention on Cooperation for the Protection and Sustainable Use of the Danube River’ (Danube River Protection Convention, DRPC) forms the overall legal instrument for cooperation and transboundary water management in the Danube River Basin. The main objective of the Convention is the sustainable and equitable use of surface waters and groundwater and includes the conservation and restoration of ecosystems. The Contracting Parties co-operate on fundamental water management issues and take all appropriate legal, administrative and technical measures, to maintain and improve the quality of the Danube River and its environment. Austria, Bosnia and Herzegovina, Bulgaria, Croatia, the Czech Republic, Germany, Hungary, Moldova, Montenegro, Romania, Slovakia, Slovenia, Serbia, Ukraine and the European Union are presently Contracting Parties to the DRPC. To facilitate the implementation of the DRPC, the Danube countries agreed to establish the International Commission for the Protection of the Danube River (ICPDR). The ICPDR set up a Secretariat based in Vienna, which co-ordinates the work of the countries under the Convention and the work of the Expert Groups in particular. Expert Groups for River Basin Management (RBM), Pressures and Measures (PM), Monitoring and Assessment (MA), Flood Protection (FP), Information Management and GIS (IM&GIS) and Public Participation (PP), and a Strategic Expert Group (SEG) have also been created. The organization chart of the ICPDR can be seen in Figure 12.2. Each work group is composed of at least one expert from each country and meets twice or perhaps three times a year to undertake the work needed. Of importance, it is the experts from the countries who do most of the work needed in each of the work groups. The groups report regularly to the ICPDR on their work progress and/or seek guidance from the ICPDR on issues of policy.

The Danube River – the Most International River Basin

Fig. 12.1 Danube River basin district.

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Fig. 12.2 Organization chart of ICPDR.

In the 12 years since the signing of the Convention, the International Commission for the Protection of the Danube River (ICPDR) has been established and matured as the forum for co-operation among the Danube countries. All the countries of the Danube have been actively participating in the work groups of the Commission and achieved important progress in their joint efforts to manage this shared river system. As ICPDR President Catherine Day stated during the celebration of the 10-year anniversary of the signing of the Convention, ‘We have a number of successes to celebrate’. She went on to note, however, that we have a ‘number of challenges still ahead of us’. The biggest challenge faced by the ICPDR has been the development of an international river basin management plan as required by the EU Water Framework Directive.

12.4 Applying the EU Water Framework Directive (WFD) on an International River Basin With the coming into force of the EU Water Framework Directive in December 2000, the

countries of the Danube committed to use this visionary legislation to assist in meeting the goals of the Convention. The commitment to use the methods and meet the goals of the Directive was made by all countries – those in the EU, Accession countries and countries not in the EU (such as Serbia or Moldova). The ICPDR plays a co-ordinating role in ensuring that a river basin management plan for the entire basin is prepared. What makes the implementation process in the Danube River Basin a particular challenge is the fact that only some countries are EU Members and therefore obliged to fulﬁll the EU WFD (European Commission 2000). Besides Austria and Germany, four additional Danube countries became EU Members States on 1 May 2004. Two other Danube countries joined the EU in January 2007 and Croatia is in the process of accession to conform with the complete body of EU legislation in order to become EU Members. Others have not initiated a formal process to join the EU. The Danube River Basin falls into the third category of rivers addressed in Article 13.3 of the EU WFD – an international river basin extending beyond the boundaries of the European Community. The EU WFD requests the Member

The Danube River – the Most International River Basin State or Member States concerned to ‘endeavour to establish appropriate co-ordination with the relevant non-Member States, with the aim of achieving the objectives of this Directive throughout the river basin district’. Therefore, appropriate co-ordination mechanisms need to be introduced. The ICPDR has been utilized to secure this co-ordination. As the EU WFD has speciﬁed: In the case of an international river basin district extending beyond the boundaries of the Community, Member States shall endeavour to produce a single river basin management plan, and, where this is not possible, the plan shall at least cover the portion of the international river basin district lying within the territory of the Member State concerned.

In November 2000 all Contracting Parties of the Danube River Protection Convention stated their commitment to implement the EU WFD within their jurisdiction and to co-operate in the framework of the ICPDR to achieve a single, basin wide co-ordinated Danube River Basin Management Plan. For the states with territories of less than 2000 km2 in the Danube River Basin (Albania, Macedonia, Italy, Poland, Switzerland), the ICPDR has attempted to establish appropriate bilateral co-ordination. In view of the size and number of states that have territories in the Danube River Basin, coordination is required on different levels in order to fulﬁll ‘the environmental objectives established under Article 4, and in particular all programmes of measures’. For issues of basin-wide importance the ICPDR serves as the platform for co-ordination in the implementation of the EU WFD in the Danube River Basin District. Transboundary issues not covered by the ICPDR are solved at the appropriate level of co-operation, e.g. in the frame of bilateral/multilateral river commissions. Local issues remain a national task. Generally, co-ordination will take place at the lowest level possible so that the co-ordination via the ICPDR can be limited to those issues necessary on the basin wide level.

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At its 3rd Ordinary Meeting on 27–28 November 2000 in Soﬁa, Bulgaria the ICPDR adopted the following resolutions: • The ICPDR will provide the platform for the co-ordination necessary to develop and establish the River Basin Management Plan for the Danube River Basin. • The Contracting Parties ensure to make all efforts to arrive at a co-ordinated international River Basin Management Plan for the Danube River Basin. The important ﬁrst step in the process was completed in March 2005. The Danube Analysis report involving inputs from all the countries was submitted to the EU along with national reports from member states. This report responds to reporting obligations of the Water Framework Directive 2000/60/ EC (WFD) under Article 5, Annex II and Annex III regarding the ﬁrst characterization and analysis of the Danube River Basin District (DRBD). In addition, information is given on progress related to Article 6 and Annex IV for setting up an inventory of protected areas in the river basin district, as well as on progress related to Article 14 regarding public information and consultation. The report gives the basin-wide overview on issues requiring reporting under WFD. It provides information on the main surface waters and the important transboundary groundwaters. It includes, in particular, an overview of the main pressures in the DRBD and the related impacts exerted on the environment. The overview includes effects on the coastal waters of the Black Sea as far as they are part of the DRBD, since their status could be a reason for designating the whole DRBD as a sensitive area. The analysis is based on available data resulting from past and ongoing programmes and projects in the region and collected by the countries. The issues referred to in the basin-wide overview will be the basis for the preparation of the Danube River Basin Management Plan by the end of 2009.

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Fig. 12.3 Risk classiﬁcation of the Danube, disaggregated into risk categories.

Box 12.2

Pollution sources – point and diffuse sources (Moneris model)

MONERIS – MOdelling Nutrient Emissions into the RIver Systems – was successfully applied for the modelling of the nutrient inputs within the Danube River Basin. MONERIS is a semi-static emission model for point and diffuse sources of nutrients and after adaptation also for heavy metals and some priority substances (e.g. lindane). The results are calculated as yearly emissions into surface waters. The results are validated after inclusion of retention – processes with measured riverine loads. For MONERIS a harmonized database was established for all Danube countries. The results have been used for the analysis of the distribution of the point and diffuse nutrient inputs within the Danube and its sub-catchments in the development of Roof Report 2004 for the DRB. The need for further development of MONERIS as a management tool was conceived in recognition of some of the limitations of current approaches for management activities and decisions involving signiﬁcant pressures, major environmental costs and/or signiﬁcant environmental consequences/impacts, as required by the Water Framework Directive (WFD). The objective is to develop a Pollution Control Decision Support Tool based on MONERIS for implementation and evaluation of programmes of measures to achieve a good water status. The goal is to determine if the implemented measure or packages of measures meet WFD and DRPC targets in one speciﬁc sub-river basin or the whole catchment. This is seen as a support for the ICPDR in giving policy advice to governments on the need to invest in nutrient reduction projects, or implement speciﬁc measures in response to EU Directives. The tool will facilitate the decision process based on comparison of the effects of various measures implemented in different sectors, countries, regions or group of countries in the DRB.

The Danube River – the Most International River Basin 12.5

12.5.1

Critical River Basin Management Problems Signiﬁcant water management issues

The Danube Analysis Report (International Commission for the Protection of the Danube River 2005a) identiﬁed four signiﬁcant water management issues which needed to be addressed at a basin wide level. These are: • organic pollution; • nutrient pollution; • pollution from hazardous substances; • hydromorphological alterations. These problems are issues which if no action or improvements are made will result in a failure to meet the WFD goal of good water quality status by the year 2015. Figure 12.3 shows the risk classiﬁcation of the main stem of the Danube, disaggregated into risk categories (International Commission for the Protection of the Danube River 2005a). These four signiﬁcant water management issues (organic pollution, nutrient pollution, pollution resulting from hazardous substances, and hydromorphological alterations) will be in focus for the further management steps as part of the WFD implementation. Measure to address these problems will make up the ﬁnal Danube River Basin Management Plan that is expected by 2010. Future measures within the Danube River Basin will built on these four identiﬁed management issues and for each of them a relevant strategy will be developed to enable the achievement of good ecological status in all affected surface waters. Therefore, for each of the above mentioned water management concerns, issue papers are drafted and under preparation. A draft document on Signiﬁcant Water Management Issues (SWMI) was prepared based on the four issue papers. The document addresses the four issues and groundwater at a basin-wide scale to ensure a target-oriented DRBM Plan and an appropriate Joint Programme of Measures (JPM) (International Commission for the Protection of the Danube River 2007a). The document includes visions and operational management objectives for each SWMI,

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which provide guidance to the Danube countries towards realizing a commonly agreed DRBM Plan. A major surprise of the Danube analysis report (2005) was the conclusion that hydromorphological alterations had been so prevelant and were such a threat to the achievement of good ecological status. The reasons for this were the channelization of rivers for ﬂood control and navigation and the damming of rivers for hydropower production. An expert group speciﬁcally examining this issue has been formed. Since completion of the Analysis Report (or characterization report) the countries of the Danube have taken the next step and developed an integrated monitoring system to meet the requirements of WFD. This monitoring report – the next important milestone – was submitted to the Commission in March 2007. The summary (Roof) report on monitoring programmes in the Danube designed under Article 8 WFD has two integral parts describing monitoring networks for surface and ground waters. The description of the surface water monitoring network consists of: • Strategy for the development of WFD compliant monitoring programmes for the Danube River Basin District. • Annexes to the Strategy paper containing a list of monitoring sites, parameters, frequencies and analytical methods for the surface water monitoring networks. • Map of the surface water monitoring network in the Danube River Basin District. The description of the groundwater monitoring network consists of: • Groundwater Monitoring Report (status report). • Annexes to the report containing the description of monitoring sites, parameters, frequencies and analytical methods for the groundwater monitoring networks as well as a detailed characterization of the monitoring programmes and the network design in particular groundwater bodies. • Maps of the groundwater monitoring network in the Danube River Basin District. The basic consideration of the new monitoring system was to maintain the overview of water quality provided by the previous transnational

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monitoring system and to add to this the monitoring needed to fulﬁll the EU WFD. 12.5.2

Flood prevention and management

In addition to problems of water quality, the Danube River Basin has experienced a number of

Box 12.3

severe ﬂoods in recent years. A major ﬂood in 2002 triggered the formation of a ﬂood management expert group that has developed a basinwide Flood Action Programme. This programme calls for ﬂood action plans for the sub-basins and is being implemented parallel to the work under the WFD (International Commission for the Protection of the Danube River 2004)

Flooding in the Danube River Basin

The disastrous ﬂood events of the past 10 years have caused numerous human casualties and the material damage reached unprecedented dimensions. The prediction of climate change, which has all but become a certainty, has been another driving force for preventive action. Climate change is expected to further aggravate the situation, leading to an increased risk of damaging ﬂood events. Floods are the result of meteorological processes and are thus natural events and part of the natural water cycle. Massive damage is caused where humans increase the risk of ﬂooding through inappropriate land use in high-risk areas or through serious interference in natural processes. In November 2002, the International Commission for the Protection of the Danube River decided to establish the long-term Action Programme for Sustainable Flood Prevention in the Danube River Basin. The Action Programme is based on the sustainable ﬂood protection programmes developed in the various Danube riparian countries, and on networking existing structures and using the futureoriented knowledge base accumulated through a wide range of activities over the past decade. The overall goal of the Action Programme is to achieve a long-term and sustainable approach for managing the risks of ﬂoods to protect human life and property, while encouraging conservation and improvement of water-related ecosystems. Given that the Danube River Basin is the second largest among European river basins, and the most international worldwide, with the marked differences regarding sociological and topographic structures, the Action Programme represents an overall framework which needs to be speciﬁed in further detail for sub-basins. These must be consistent with the areas deﬁned for the enforcement of the WFD, in order to bring both planning processes together at an appropriate stage. The framework Action Programme deﬁnes the underlying principles and objectives for sustainable ﬂood protection for the entire basin of the Danube River together with a timeframe. In a ﬁrst stage, it deﬁnes a set of general objectives (e.g. the need to network existing national ﬂood reporting and forecasting systems) and sets out several categories of measures likely to reduce the risk of ﬂooding. These objectives and action plans need further speciﬁcation in the various sub-basins. The Action Programme also contains information on monetary and organizational mechanisms for realization. In the future it will be crucial to rapidly advance the planning process as well as to elaborate speciﬁc action plans for the various sub-basins of the River Danube, in order to be able to assemble from these components an overall programme by 2009. During this planning stage, it must be ensured that a harmonious development process ultimately leads to a consistent ﬂood action programme for the entire Danube River Basin, incorporating the future developments of the EU initiative on ﬂood risk management planning where possible.

The Danube River – the Most International River Basin This framework Action Programme is based on the sustainable ﬂood protection programmes developed in the various Danube riparian countries, and on networking existing structures and using the future oriented knowledge base accumulated through a wide range of activities over the past decade. The overall goal of the Action Programme is to achieve a long-term and sustainable approach for managing the risks of ﬂoods to protect human life and property, while encouraging conservation and improvement of waterrelated ecosystems. 12.5.3 Developing a programme of measures Following the completion of the characterization report and the identiﬁcation of the signiﬁcant water management issues the EU Water Framework Directive calls for the development of a programme of measures – the actions necessary to ensure the goal of good water quality status by 2015. The ICPDR is co-ordinating the development of the River Basin Management Plan (RBMP) at the basin-wide level and acting as a forum for exchange of information on measures and strategies for all countries. Steps to develop the RBMP by 2009 have been put in place. As a ﬁrst step, a short data collection was organized to identify the national planning of RBMP and Programme of Measures (PoM) developments. It appeared that most countries are closely following the deadlines in the WFD and will have limited information available that can be compiled on a basin-wide level. As a consequence, the ICPDR RBM Executive Group proposes to amend the timetable for preparation of the Danube RBMP because it will otherwise not be possible to ensure maximum consistency between basin-wide and national level. The revised timetable would involve collection of data until the end of 2008 and after that the preparation of a draft Danube RBMP for endorsement in June 2009. Thereafter, a shortened public consultation phase would take place up to October 2009 with an international public

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hearing organized by the ICPDR in Sept 2009. The ﬁnal Danube RBMP would be compiled by the end of 2009, which would make it necessary to organize the Ministerial Conference which approves the Danube RBMP in January or February 2010. This would time with submission of the River Basin Management Plan to the European Commission in March 2010. A road map for reaching this goal had been produced and the ICPDR is preparing issue papers on the signiﬁcant water management issues and preparing or engaging in stakeholder consultation processes, particularly with those stakeholders who need to undertake actions in support of reaching the good status (International Commission for the Protection of the Danube River 2007b). The measures under consideration and development are based upon a historical action programme in the region – the Joint Action Programme. This included a number of commitments for waste water treatment (WWT) plant construction, actions to reduce pollution, wetland restoration and policies to improve water quality. Important progress was made but additional efforts are needed (International Commission for the Protection of the Danube River 2005b).

12.6 Future Developments Future developments within the Danube River Basin will focus on: • monitoring the ecosystem; • analysing pressures; • generating political will. Although signiﬁcant efforts have been undertaken to put in place the structures and processes to bring long-term improvement to the management of waters of the Danube, future challenges remain. The most important issue is the need to develop a Joint Programme of Measures and to ﬁnd the political commitment to introduce and ﬁnance the needed measures. This applies as well to the necessary measures outlined in the Flood Action Programme.

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Under the ICPDR Joint Action Programme over 4 billion dollars of investment have been made in municipal wastewater treatment which have reduced Danube pollution (International Commission for the Protection of the Danube River 2003). Despite substantial progress there is the requirement for additional investment in the building of sewage treatment plants where they do not exist, the upgrading where necessary, and ensuring adoption of pollution reduction technologies in general. Further funding will be required to bring funding authorities and the country representatives together to ensure that the needed resources are available. In addition there is a need to adopt policies and more simple actions to reduce pollution. The Danube Regional Project of UNDP/GEF was helpful in identifying actions that could be beneﬁcial involving speciﬁc changes in agricultural practices, the reduction of phosphates in detergents through a ban, the protection of wetlands as a mechanism to ensure the reduction or pollution and the better management of nutrient pollution in particular (United Nations Development Programme–Global Environment Facility 2006). There needs also to be a basin-wide modelling of pollution and keeping track of national actions in a basin-wide context. For this purpose the Danube GIS and decision support tool (called MONERIS) for analysing actions related to pollution in the Danube River Basin are under the development. The upcoming Joint Danube Survey 2 should help provide information on the state of the Danube water resources and where additional actions are needed. In recent years the ICPDR has also taken important steps to engage a variety of stakeholders in the efforts to improve the conditions in the Danube River Basin. Of particular importance have been efforts to work with stakeholders of all kinds in carrying out the activities (International Commission for the Protection of the Danube River 2005c). It is also important to ﬁnd additional mechanisms for ensuring input from stakeholders. In

this context, this year the ICPDR has initiated a very important initiative to work with the navigation section to ensure a dialogue on development and maintenance of navigation and the environment.

Box 12.4 Involving the public – Danube Day The Danube is the most international river basin in the world. On 29 June 1994, Danube countries and the European Commission signed the Danube River Protection Convention committing themselves to cooperation in water resource management in the Danube basin – a commitment to work together to provide good quality and safe water for the 81 million people of this region. Strong efforts have been made by the ICPDR and the countries of the region to improve public understanding and involvement in issues related to water quality in the Danube. Ten years later, in 2004, Danube Day was celebrated for the ﬁrst time. Nowadays, 29 June is Danube Day and it is celebrated in all parts of the basin, creating solidarity among people. Danube Day is a powerful tool for enhancing the ‘Danubian identity’ of people living in the basin, demonstrating that in spite of their different cultures and histories they have shared responsibility to protect their precious resources. Danube Day is about people: getting involved and thinking about how their actions impact neighbours downstream. The Danube Day message is spread through discussion forums, river adventures, a Danube schools competition, a crossborder cycle tour, conservation tasks and simply by encouraging people to visit their own river. Danube Day is an occasion to celebrate the Danube River. It is also a day to acknowledge the challenges ahead and to mobilize the energies and resources for meeting those challenges

The Danube River – the Most International River Basin The positive results of the work that has been done to promote Danube solidarity through the convening of Danube Day are also an important part of the long-term efforts needed. The enthusiasm and interest shown by people throughout the Danube River Basin in participating and organizing activities in connection with 29 of June has been remarkable (www.danubeday. org). Hundreds of events have taken place in all Danube countries including the convening of the Danube Art Masters Competition. These actions have been very successful in strengthening public interest and involvement in the work of the ICPDR and in efforts to improve water quality. The initiation of activities at a sub-basin level have also helped make more concrete some of the actions needed and helped to deﬁne locally appropriate solutions. The active work of country representatives in the Tisza River Basin, Sava River Basin and in future in the Prut and Danube Delta have proved very beneﬁcial. Of importance for the ICPDR is the connection between the Black Sea and Danube. The Danube countries have signiﬁcantly reduced the pollution entering the Black Sea. The pollution loads of nitrogen and phosphorous have been reduced by 20% and 50% respectively (United Nations Development Programme–Global Environment Facility 2006). There is, however, a need for clear signals and data feedback from the Black Sea on the consequences of the work in the Danube to reduce pollution. Those indicators that exist show that there are ecological improvements but these need to be more systematically evaluated. To ensure this the Memorandum of Understanding (MOU) on co-operation between the two Commissions needs the support of all Danube and Black Sea countries. It is also necessary for actions to reduce inputs of pollutants similar to those on the Danube in other Black Sea river basins. For the future it will also be necessary for the ICPDR to acknowledge and act with an awareness of the implications of climate change. The recent problems of ﬂooding and drought in the Danube River are cause for concern and we will

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have to ensure that strategies that we develop are climate change resilient. Above all the work of the ICPDR has shown that the problems of water cannot be resolved by ministries of water and environment alone but there is a need for co-operation with other ministries (agriculture, ﬁnance and industry). In the future ways must be found to strengthen this cooperation at the national level. The ICPDR has also shown itself to be an effective forum to address transboundary conﬂicts and the continued role in this respect will require ongoing political support. The Danube is the most international river basin in the world and the challenge of co-operatively managing the basin is large. Nonetheless important progress has been made and important elements in achieving success have been realized.

Acknowledgements The author works for the International Commission for the Protection of the Danube River (ICPDR) and much of the text and information is based on reports and documents of the ICPDR.

References European Commission (2000) Directive 2000/60/EC of the European Parliament and of the Council of 23 October 2000 Establishing a Framework for Community Action in the Field of Water Policy. http://ec.europa.eu/environment/water/waterframework/index_en.html (last veriﬁed 13 October 2008). International Commission for the Protection of the Danube River (1994) Convention on Cooperation for the Protection and Sustainable Use of the Danube River (Danube River Protection Convention). ICPDR, Vienna. http://www.icpdr.org/icpdr-pages/drpc.htm (last veriﬁed 13 October 2008). International Commission for the Protection of the Danube River (2003) Policy and Legal Reforms and Implementation of Investment Projects Related to the ICPDR Joint Action Programme 2001–2005,

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Implementation Report, Reporting Period 2001– 2003. ICPDR, Vienna. International Commission for the Protection of the Danube River – Flood Protection Expert Group (2004) Action Programme for Sustainable Flood Protection in the Danube River Basin. ICPDR document IC/082. ICPDR, Vienna. International Commission for the Protection of the Danube River (ICPDR) (2005a) The Danube River Basin District, Roof Report. ICPDR, Vienna. International Commission for the Protection of the Danube River (ICPDR) (2005b) Joint Action Programme, January 2001–December 2005. ICPDR, Vienna. International Commission for the Protection of the Danube River (ICPDR) (2005c) Danube River Basin

Stakeholders Conference Final Report, 28th–29th June 2005. ICPDR, Vienna. International Commission for the Protection of the Danube River (ICPDR) (2007a) Document of Signiﬁcant Water Management Issues, vol. 3. ICPDR, Vienna. International Commission for the Protection of the Danube River (ICPDR) (2007b) Road Map for the Development of the Danube River Basin District Management Plan 2005–2010 (including a Work Plan, and an Operational Plan for Public Participation). ICPDR Vienna. United Nations Development Programme–Global Environment Facility (2006) 15 years of Managing the Danube River Basin, 1991–2006. ICPDR, Vienna.

Image facing chapter title page: Courtesy of ICPDR/Victor Mello.

Murray-Darling Basin – Integrated Management in a Large, Dry and Thirsty Basin

13

SARAH RYAN1 1

13.1

CSIRO Sustainable Ecosystems, Canberra, Australia

Introduction

The Murray-Darling Basin is the largest river basin in settled Australia, spreading over about 1 million square kilometres or 14% of the continent (Fig. 13.1). Its land and waters support signiﬁcant rural livelihoods but the impacts of tree clearing, broadscale cultivated agriculture and diversion of water for irrigation have been environmentally costly. The Basin has not been managed as a whole, for historical and political reasons, nor has its full hydrological cycle been well taken into account. Pathways to a more sustainable future for the land, water, biodiversity and people of the Basin, especially in the face of a drying climate, absolutely depend on managing the Basin and its contributing catchments as integrated systems. Policy makers are just beginning to demand this knowledge and this chapter outlines the current status and research directions needed to underpin more integrated Basin management. The chapter focuses on water because it is such a major driver of outcomes in this particular Basin, and because it is the connector that crosses catchments, from the very top end of the Basin to the very bottom.

Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

13.2 Nature of the Basin 13.2.1 Water and ecosystems The major rivers in the Basin rise in the western slopes of the Great Dividing Range which runs for 3700 km parallel to the east and south-east coasts of Australia (Fig. 13.2). After leaving the slopes, the rivers cross wide expanses of very ﬂat land until they discharge through an estuary system in the south of the continent. This geography results in a hydrological system in which only about 5% of the rainfall reaches the river system (Kirby et al. 2006). The ﬂat geography also means that the upper catchments have a disproportionate importance in terms of harvesting water. For example, the catchments above the major dams on the Murrumbidgee River represent 1.2% of the area of the Basin but contribute some 12% of Basin runoff (Commonwealth Scientiﬁc and Industrial Research Organization 2008). A further important driver of this Basin’s hydrological system is the variability of the rainfall and hence river ﬂows. Australian rainfall is amongst the most variable in the world (Table 13.1). The Basin is currently experiencing a prolonged drought, distinct warming and record low ﬂows into the major dams. These impacts are consistent with the trends expected in this part of Australia with climate change and are of grave concern to the nation.

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Fig. 13.1 The Murray-Darling Basin.

Table 13.1 Extreme variability of annual ﬂows in the Murray and Darling Rivers (from Chartres 2006) River Rhine Yangtze Potomac Murray Darling

Country

Ratio of max. and min. annual ﬂows

Switzerland China USA Australia Australia

1.9 2.0 3.9 15.5 4705.2

The low relief of the land through which the rivers ﬂow for most of their journey also produces highly meandering river systems (Fig. 13.3). Coupled with the huge variability in rainfall and river ﬂows, the rivers naturally ﬂooded periodically and were fringed by ﬂoodplains and wetlands that absorbed the larger ﬂows. These ﬂoodplains and wetlands supported vegetation that was diverse in species, structure and spatial

Fig. 13.2 MacIntyre River, in the north of the Murray-Darling Basin.

patterns, as well as a diversity of ﬁsh, birds and other mammals. Plants and animals were adapted to periodic wetting and drying in highly irregular cycles.

Murray-Darling Basin

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Fig. 13.4 Dead trees and reed beds at Macquarie Marshes, a consequence of altered channels, drought and moderately high abstractions upstream for irrigation. The Marshes are internationally recognized as an important waterbird breeding site. Fig. 13.3 The meandering Murray River, supporting redgum forests on the ﬂoodplain and irrigated crops and pastures on higher ground.

Over the last 120 years, people have signiﬁcantly altered the river system by constructing dams, weirs, levees, barrages and on-farm storages that alleviate ﬂooding and capture water for irrigation use. The capture of water in the uplands between autumn and spring, and its release in summer for irrigation, has altered the pattern of ﬂows so that peak ﬂows now occur in summer rather than in spring, particularly in the southern part of the Basin. The impact on the biota of the river, ﬂoodplains, wetlands and estuary has been substantial and the sustainability of many of the native ecological communities is in question. For example, native ﬁsh populations are estimated to be at only 10% of their pre-European settlement levels. Floodplains and some wetlands have lost connectivity with the river, ﬂoodplains have been grazed, and pests and weeds inadvertently introduced. Other wetlands are now inundated during the dry summer, due to the high river levels maintained to deliver irrigation water, and the species composition and structure has changed.

Towards the end of the river system in particular, where groundwater inﬂows are saline and there have been few ﬂushing ﬂoods in recent years, ﬂoodplain and wetland vegetation is dying and fauna are disappearing (Fig. 13.4). Drying lakes and wetlands are exposing sulphur-rich sediments to the air and the sulphuric acid formed is threatening ﬁsh and vegetation. In the Coorong, a 140-km-long lagoon at the mouth of the Murray River, the lack of fresh ﬂows has resulted in high rises in salinity and loss of habitat and resources that used to maintain waterbird populations recognized as internationally signiﬁcant through the Ramsar Convention. Flow through the mouth used to cease once in every 20 years before regulation, and once every 2 years on average since regulation. But no ﬂows have reached the mouth in the last 6 years and the mouth is now continuously dredged to keep the connection with the ocean open. The current drought is not the cause of most of these environmental impacts. The cause is the high level of abstraction and associated system changes that have been accumulating over time. The current drought is exacerbating this impact.

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Fig. 13.5 Originally a wetland near Mildura, only tree stumps remain. Salinization is often caused in this situation by rising groundwater levels bringing natural salt to the surface, a result of applying excess irrigation water nearby.

The groundwater systems underlying the Basin are also important as an additional source of water and as agents in moving salt down the system. Widespread clearing of woody vegetation has reduced transpiration and increased drainage to groundwater and rivers. Where the water quality is sufﬁcient, groundwater use has increased signiﬁcantly in the last 15 years and groundwater levels have been falling. Where the quality is poor, and usage low, shallow groundwater levels have been rising. In wetter periods, substantial areas of low-lying ground are at risk of becoming salinized (Fig. 13.5). The viability of many irrigation areas now depends on pumping of saline groundwater and storage in evaporation ponds. Many wetter ecosystems within the Basin also depend on groundwater, either wholly or in periods when surface water supply is short. The surface water and groundwater systems are managed separately, despite their hydrological connection. 13.2.2

Water quality

Apart from the impacts of salinity, the challenges of nutrient, sediment and chemical pollution

that dominate many other basins around the world, like the Great Barrier Reef catchments in Australia (Chapter 15), Chesapeake Bay in the USA (Chapter 18) and many European basins, are secondary issues in this Basin. This is partly a reﬂection of the generally low relief, forested uplands that contribute most of the runoff into the major dams, and relatively low fertilizer input in the dryland broadacre agricultural systems. There is negligible industrial activity in the Basin and the few point sources of pollution are well controlled by regulation. Nevertheless, sediment and nutrient levels are generally well above natural levels and this does have detrimental impacts on river and ecosystem health. Reinstating stream and riverine vegetation to improve ecological outcomes is in its infancy. 13.2.3 People and institutions The Basin is home to some 2 million people. Dryland and irrigated agriculture are the principal rural economic activities and apart from the 350,000 people in Canberra, Australia’s national capital, people live on farms and in rural settlements and towns. An additional 1 million people in South Australia heavily depend on water piped out of the Basin for their domestic use. The main dryland agricultural products are wheat and other grains, beef and sheep meat, and wool. Irrigated products are mainly grapes, rice, cereals, cotton and fruit and vegetables, as well as beef and dairy products grown on irrigated pastures. Agriculture in the Basin contributes about AUD$15 billion (34%) of Australia’s total revenue from agriculture, 4% of its export income and 2% of its GDP. While agriculture is low in importance relative to other sectors of the economy, it remains highly signiﬁcant for underpinning the economic activity of rural communities. Irrigation contributes disproportionately to this economic base. Occupying only 1.4% of the total land area of the Basin, it accounts for around 36% of the total proﬁt from agriculture in the Basin (Bryan and Marvenek 2004). Revenue per unit area is 13 times higher in irrigated areas compared to adjacent dryland areas (Meyer 2005).

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Fig. 13.7 Across the basin, water released from storages has roughly a 44% chance of being diverted for irrigation, a 32% chance of evapotranspiring in a wetland or ﬂoodplain, a 12% chance of evaporating in the river system, and a 12% chance of reaching the ocean (based on 1995–2003 data, from Kirby et al. 2006). Diversions are even higher in some catchments. Fig. 13.6 Irrigation in the southern basin provides jobs and incomes at 13 times the intensity of dryland agriculture (Meyer 2005).

This reﬂects the value that water adds to agriculture in this dry Basin, and explains why such high levels of river regulation and water abstraction have arisen. Across the Basin some 10,500 GL/yr (44% of runoff, and 95% of all diversions) is dedicated to irrigation (Kirby et al. 2006). These proportions, but not volumes, are very high by world standards. Communities and economies in the Basin have become very dependent on irrigation (Fig. 13.6). Practices on dryland farms in the Basin (a large 77% of the area) are of additional importance for their inﬂuence on hydrological processes and water quality. Widespread clearing of trees has now ceased, but expansion in the number of farm dams, and conversion of old pasture land to plantation forests, is reducing local stream ﬂows. Sediment and nutrient ﬂows to rivers are well above natural levels, but are secondary threats to water quality compared to salinity. The socio-political system is a key feature of the Basin. The Basin crosses the boundaries

of four states and a territory (Fig. 13.1). Constitutionally, state governments rather than the Commonwealth government have the power to make decisions about Australia’s natural resources. Decisions about water that is generated in one state and used in another therefore have to be made jointly. Formal collaborative decision-making has been in place for nearly 100 years but so inﬂuential has the nature of those arrangements been on outcomes for the Basin that it is described in more detail below. 13.2.4 Today’s challenges The key challenge in the Murray-Darling Basin today is to adjust the uses of water so that ecological assets and processes are better protected and viable irrigation industries maintained (Fig. 13.7). The feasibility of achieving both is yet to be demonstrated, especially in an era of great uncertainty about future climate with reductions in water supply highly likely. New scientiﬁc knowledge will play a role, but good leadership, effective governance and willingness and ability

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of water users to change will be equally important. There is a sense of urgency about the need for change. The south-east of Australia is in its sixth year of drought and inﬂows to the Murray River are at record lows. Many irrigators received unprecedentedly small or zero allocations of water last summer. Moreover, while drought is characteristic of the climate of southern Australia, temperatures are now clearly increasing as well, and the combination takes the Basin out of its previously recorded experience. There is great uncertainty about both short- and long-term futures for water supply.

13.3 The Past Shapes Potential for the Future The Murray-Darling Basin of today has been highly shaped by people’s actions in the past. The cultural forces that led to those actions will continue to shape the Basin of the future. Two such cultural forces stand out: the original goals of the irrigation settlements coupled with the subsequent scale of their development and domination of water use; and the relationships between the state and Commonwealth Governments (see Connell 2007 for a detailed history). Each of these has played a role in delaying the application of integrated catchment management (ICM) principles at Basin scale. Development of the Basin began in the 1850s as paddle-steamers developed lucrative markets for transporting wool out of the inland and back loading supplies for rural settlements. At this time, the states were independent colonies of England and ﬁercely competitive; part of the attraction of the river trade for South Australia was the customs duties it could collect on goods passing down the river. Federation of the states into the Commonwealth of Australia happened in 1901, only after long and difﬁcult debates about the powers that would pass to the Commonwealth Government. Management of land and water resources was one of many powers which remained with the states.

Box 13.1 We shall have a large and prosperous population obtaining its wealth by the surest possible means, that is from the soil, delivered from the risks of the natural rainfall, not dependent upon the chances of clouds, but able to secure a shower when it is needed.

Speech to the Victorian Parliament by Alfred Deakin, often referred to as the father of irrigation in Australia (Deakin 1886)

Even before federation the states had begun to investigate and invest in irrigation settlements along the major tributaries and the River Murray. Providing economic opportunities for individual land holders has been an important historical theme in the development of Australia, not only in irrigation settlements (Box 13.1). For over a hundred years farmers’ economic value to the nation and political power has restricted the range of policy options available to governments for inﬂuencing decisions farmers make about land and water use on their own farms. Markets, incentives, extension and voluntary schemes are therefore strongly preferred over regulation for inducing land and water use change. More recently, the granting of property rights over water to individuals and irrigator groups has further reduced the inﬂuence of governments. These cultural preferences have been a challenge to ICM in Australia because actions that deal with externalities are needed on-farm, while beneﬁts accrue to the wider catchment community. By the time of federation in 1901, irrigation settlements were already being independently established in South Australia, Victoria and New South Wales. Realizing that management of the water of the Murray River, which forms the border between Victoria and New South Wales, needed agreement between those states, and that supply of water to South Australia depended on actions in the two upriver states, moves to formalize joint decision-making were begun around

Murray-Darling Basin

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Fig. 13.8 History of dam building, growth in diversions and major decisions made under the two intergovernmental agreements. (1) Dam dates are initial construction. Some were subsequently enlarged. Dams with state afﬁliations were built by states, others under the Murray Water Agreement. Shaded events represent major changes in intergovernmental arrangements. Diversion data are from Close (1990) and Murray-Darling Basin Commission (2007).

the time of federation. It then took 14 years for member governments to agree and sign the River Murray Waters Agreement and to pass the requisite parallel legislation in their jurisdictions. A key characteristic of decision-making under this agreement (and its successor that persists until today) was that decisions would be unanimous. This allowed NSW, for example, to veto taking action over salinity for many years (Clark 2002). The scope of the River Murray Waters Agreement of 1915 was limited to the function and waters of the River Murray itself. Major activities under the Agreement were focussed on dams and weirs for securing water supply (Fig. 13.8). In parallel, the states continued developing storages on the major tributaries and gravity-fed channel distribution systems to deliver the water to irrigation areas. The impact on diversions is clear. In the 50 years between 1920 and 1970 water diverted for irrigation grew fourfold (Figs 13.8, 13.9). By the

1970s, declining water quality due to high salt levels began to be recognized as a serious issue. This was due to the combined effect of increased salt reaching the river as a result of over 100 years of land clearing in the catchments, and the impact of irrigation. The high salinity levels had already damaged irrigated crops in the lower part of the Murray in drier years, and the trend suggested further increases were likely without remedial action. This was the ﬁrst emergence of the need to consider much wider catchment issues to achieve desired outcomes in the river. Reform of the governance arrangements to allow Basin-wide decision-making about salinity and other catchment and environmental issues, took another 15 years. The very strength and success of the ﬁrst agreement in terms of building supply and sharing water had made it highly resistant to change (Doyle and Kellow 1995). The new Murray-Darling Basin Agreement was ﬁnally signed in 1987. Key bodies in the new

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Fig. 13.9 This book, published in 1945, both reﬂected and reinforced a non-scientiﬁc attitude to the beneﬁts of dam building and irrigation.

had been addressed, the ‘principle of comity was forgotten and the ogre of sovereignty unleashed once more’ (Clark 2002). The long-term consequences are signiﬁcant. For example, agreement was reached in the Basin Salinity Management Strategy (BSMS) to reduce levels of salt in the river, which was the immediate issue, but the subsequent storage of salt in the landscape is degrading ﬂoodplains and the land around salt disposal basins. The policy has not been adjusted. Despite sometimes slow progress, it must be acknowledged that agreements over sharing waters of the Murray River itself have been honoured all this time. Basin governance has faced up to major issues – water sharing (1915), salinity (1989), the cap (1995) and environmental ﬂows (2003), although always well behind scientiﬁc evidence of the consequences of not acting and only when unacceptable thresholds appear very close. In the same vein, the MDBC has been very slow to do preparatory work around future risks. The ﬁrst quantitative summary of the impact of six identiﬁed future risks was not published until 2006 (van Dijk et al. 2006). Australia as a whole has been slow to acknowledge and plan for its highly variable climate (Box 13.2), let alone the risk of climate change. It also needs to be acknowledged that the governance arrangements

Box 13.2 arrangement were a Ministerial Council of ministers from individual jurisdictions, the MurrayDarling Basin Commission (MDBC, a committee of senior bureaucrats from member jurisdictions and an administrative arm) and a Community Advisory Committee. A raft of new policies and strategies were developed in the following 15 years (Fig. 13.8). Most signiﬁcant in the ﬁrst 10 years were agreements about managing salinity and the cap on diversions. But a consequence of the divided loyalties of ministers and Commission members, and the expanded range of issues the new agreement covered meant that after the pressing issues

However, as they were admittedly quite abnormal years, it would be fair to discard them from consideration in either point of view. Economic engineering can seldom afford to cope with exceptional and abnormal conditions.

Australia has struggled to accept and adapt to its climate variability. The legacy of this quote (Monash 1904), about building weirs on the Murrumbidgee River, is still apparent today in the Australian Government’s ‘Exceptional Circumstances’ programme to alleviate the economic impact of drought.

Murray-Darling Basin for the Basin are not the only explanation for the slow progress towards a more integrated view of Basin management. Lack of integration of land and water issues has also been a challenge within states, and at times, intergovernmental political trading over unrelated issues has intervened (Doyle and Kellow 1995). Frustration with the inability of the Basin governance structure to deliver speedier solutions to major challenges led to a proposal from the Commonwealth government in December 2006 to reform the governance arrangements. ‘Existing arrangements centring on the MDB Agreement and the MDB Ministerial Council are unwieldy and not capable of yielding the best possible Basin-wide outcomes.’ (Australian Government 2006). Initially the Commonwealth sought, just for the Basin, a referral of the state powers to make decisions over water but one state adamantly resisted despite the signiﬁcant ﬁnancial incentive offered. A subsequent change of Commonwealth Government has now led to inprinciple agreement to form a new MurrayDarling Basin Authority that will incorporate the functions of the MDBC and in addition be responsible for developing a Basin Plan. State powers over decision-making at the Basin level only will be transferred to the Commonwealth government but the Ministerial Council will remain as an advisory group. Two theoretical frameworks help in understanding this historical pattern, and where opportunities and pitfalls might lie in the future. The ﬁrst of these is that being developed around water scarcity. Three types of water scarcity are deﬁned by Wolfe and Brooks (2003). First order scarcity refers to shortage of water supply and policy responses are dominated by engineering works. Second order scarcity refers to increasing the efﬁciency of use, and policy responses rely heavily on economic mechanisms (Fig. 13.10). In third order scarcity, the focus is on why water is used in this way at all and policy responses depend much more on new options and reallocation, for which social capacity is the most important. In the Murray-Darling Basin, we dealt with ﬁrst order scarcity until around 1995 (Fig. 13.8), our

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Fig. 13.10 Spray irrigation of pasture in the Murrumbidgee catchment.

current focus on water markets and increasing irrigation efﬁciency (‘We must adapt by reducing the use of water, and use what we have more efﬁciently’: Australian Government 2006) reﬂects second order responses, and our inability to secure environmental water indicates little progress towards dealing with third order responses. Molle (2003) makes a similar point in reviewing trajectories of ‘closing’ river basins around the world. ‘As the basin nears closure, sectoral allocation becomes a point of tension. Efforts are directed at allocating water towards the most economically valuable uses and new institutions evolve to address inter-sectoral competition and manage river-basin resources in an integrated fashion.’ This is a clear pointer to where research needs to be headed for the Murray-Darling Basin. The second theoretical framework is the adaptive cycle of complex systems theory (Walker and Salt 2006). The theory has been applied to ecological, social or linked socio-ecological systems, like the Murray-Darling Basin. The now thricerepeated pattern of reform in governance arrangements, followed by a ﬂurry of new policies, then increasing difﬁculty in dealing with new issues, matches many characteristics of this theory (Table 13.2). The theory may not predict the details of the next cycle, but it does provide the following warnings, especially at the release and

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Table 13.2 Adaptive cycles and the Murray-Darling Basin Phase in adaptive cycle Rapid growth

Conservation

Release

Reorganization

Events Characteristics

1st cycle 1890s to 1914

2nd cycle 1914–1985

Innovative exploitation of new opportunities Efﬁciency increases, capital grows, inter-connections strengthen, system becomes rigid, resilience declines

First irrigation settlements under state auspices

Major dam and weir construction, irrigation expansion Technological innovation in irrigation drives diversions to very high levels, irrigators develop strong political inﬂuence Long trend of increasing salinity ﬁnally resulting in damage to irrigated crops 1985 Murray-Darling Basin Agreement signed

Disturbance breaks the connections, regulatory controls weaken, capital is released Starting conditions set for next cycle

New Australian constitution (1901) locks in strong independent state powers over water

Pressure for reform built from inability to plan for the use of shared water under the constitution 1914 River Murray Waters Agreement signed

reorganization phases, where we likely are now. A release phase (in this case triggered by the lowest inﬂows on record) is generally short but the dynamics are chaotic. Certainly the consequences for irrigation communities and for the ﬂoodplains and wetlands of record low water allocations will be drastic. In the reorganization phase, small, chance events can have a big impact. Events here will determine whether the next cycle is a simple repetition of the previous one, or one that is more novel, or one that leads to new degradation. The water scarcity framework would indicate a need to shift to new institutional arrangements that grapple with the allocation challenge.

13.4

Integrated Catchment Management

The current Basin policy that directly addresses ICM is its Integrated Catchment Management Policy 2001: ‘We the community and governments of the Murray-Darling Basin commit ourselves to do all that needs to be done to

3rd cycle 1985–2006 Major new environmental policies developed Early success with salinity strategy, and introduction of Cap (1995) but then slow decision-making leads to lags in environmental ﬂows policy and dealing with future risks 2002–2007 drought, lowest inﬂows on record. Prime Minister announces plan for new Basin governance 2008 New Murray-Darling Basin Authority established

manage and use the resources of the Basin in a way that is ecologically sustainable’ (MurrayDarling Basin Commission 2001). The policy ‘seeks to respond to issues which require joint government action … or individual State but which could have implications for integrated resource management across the Basin.’ The policy is strong on aspirations but coy about how the change would happen, especially commitments to investing in the needed changes. And despite the integrative aspiration, the strong managerial culture in MDBC, fanned by vigilant attitudes of member states, has made it difﬁcult to move beyond a ‘one issue one policy’ approach. Rather, it has been developments outside the MDB governance that have fostered signiﬁcant progress in ICM practices in the Basin. Independently in the state of Victoria, formal catchment management agencies were formed in 1994 and planning and delivery of land and water programmes began to be more integrated and matched to catchment boundaries. Other states were beginning similar processes through the

Murray-Darling Basin 1990s when the Australian government, in 2001, offered the states large funds for landscape repair, contingent upon them forming ‘regional bodies’ to plan and implement integrated strategies. The terminology varies by state but the word ‘catchment’ remains in about half the names, e.g. the Murrumbidgee Catchment Management Authority. The regional bodies are governed by independent community-based boards, some statutory, some not. All are responsible for developing an integrated natural resource management plan for their region (‘iNRM’ plan), that has to be accredited by both state and Australian governments, and then forming diverse partnerships to deliver to the plan. The basic model is consistent across the 56 ‘regions’ deﬁned nationally but with allowance for some differences negotiated on a state by state basis. This results in anomalies like regions not crossing jurisdictional boundaries, hence the 16 regions in the MurrayDarling Basin include six that border the River Murray but do not cross it. They may be genuine political catchments but they are not genuine water catchments. Nevertheless, the regional bodies are becoming so instrumental in delivery of land, water and biodiversity policies that they are fast becoming the ICM groups envisaged in the ICM policy of the Murray-Darling Basin – although via a policy initiative outside the MDB governance and with a ‘NRM’ label rather than an ‘ICM’ label. Perhaps this suits Australia, which is a big country with sparse human resources, relatively dry catchments and where land and biodiversity issues can be as locally pressing as water issues. The integrated NRM plans are strongly driven by hierarchies of goals, targets for improving the condition of land, water, and biodiversity assets, and management actions to achieve them. There are regular reviews of progress and thus the adaptive management cycle is well institutionalized, even if the complexity of landscape change makes it difﬁcult to evaluate progress on short timescales. Community involvement, both in the governing boards and as delivery agents, means a core principle of ICM is recognized.

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The Australian government recently commissioned a review of the new regional delivery model (Keogh et al. 2006). Based on widespread consultation, the reviewers found ‘overwhelming support for the regional delivery of NRM in Australia’. Despite this progress, not all the plans involving water are made under this ICM umbrella, nor under the MDB agreement. A completely separate agreement between the state and Commonwealth Governments under the National Water Initiative (National Water Commission 2004) means that water sharing plans are developed under a different framework, and these are the critical plans that determine the relative share of water between irrigation and the environment. The development of these plans has been slow and their effectiveness has varied (Hamstead et al. 2008). This is not surprising as this is Australia’s ﬁrst step in dealing with the ‘third order scarcity’ of Woolfe and Brooks (2003), but it is frustrating to those who are watching the progressive decline in health of the Basin’s water ecosystems. Further institutional development is needed to integrate these plans with other plans involving water. Returning to the Murray-Darling Basin, it is clear that a set of catchment plans and actions are now in place that will inﬂuence outcomes for the Basin as a whole. What is missing is an integrated higher-level plan and regular monitoring reports at Basin scale. Very recent developments recognize this gap. Legislation in 2007, and the new intergovernmental agreement to establish the Murray-Darling Basin Authority (Council of Australian Governments 2008), commit member governments to developing a Basin Plan by 2011. The Plan will include new, sustainable limits on diversions and an environmental watering plan for the whole Basin as well as individual water resources. It will also pick up existing plans like the Basin Salinity Management Strategy and the programme that monitors river health consistently across the whole Basin – the Sustainable Rivers Audit. The new plans are inching Australia forward in integrated Basin management; the issue is whether this will be fast enough to

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deal with large disruptions to ecosystems and communities already evident in this warmer drier climate. Planning at Basin scale as if the past will return will not represent an improvement.

13.5 The Basin Knowledge System Australia has an internationally high rate of investment in publicly funded R&D relative to GDP but this does not translate to a high intensity of environmental data due to a small population and large continent. The system is also characterized by fragmentation and lack of coordination of both research funding and research providers. Consequently, while there is reasonably good knowledge about land, water and ecosystem assets and their processes in general, it is often scattered, in disciplinary and institutional silos, and not amenable to harnessing in spatially predictive tools at catchment and basin management scale (Fig. 13.11). Fragmented research funding particularly counts against developing the capacity to address large-scale knowledge challenges, as do promotion systems for scientists which are strongly weighted to individual disciplinary success. In the Basin, existing knowledge also shows bias towards those aspects that Australians have valued most in the past. For example, data and understanding about runoff and surface water ﬂows in regulated catchments, needed for reliable supply of irrigation water, are far more developed than knowledge about unregulated catchments or groundwater ﬂows or ecological processes in the rivers, ﬂoodplains and wetlands. The latter were not within the scope of the ﬁrst Basin agreement or even within the scope of state government concerns until the 1970s. Since then knowledge rushes have been stimulated in response to concerns about the next looming ecosystem threshold: salinity in the 1980s, then health of ﬂoodplain, wetland and estuary ecosystems in the 1990s, and then groundwater use. This is not a long scientiﬁc history for such an extensive, variable and complex Basin.

Fig. 13.11 Measuring ﬂow in Broken Creek, a tributary of the Murray River.

Consistent with the domination of state interests in governing the Basin, state processes for data collection and knowledge generation have also dominated knowledge development about the Basin as a whole. This has resulted in lack of data consistency and unwieldy mechanisms for more integrative studies. For example, New South Wales and Victoria each have different river models for their own tributaries and the Murray-Darling Basin Commission runs a different model again for the River Murray itself. Vegetation surveys have also been place-based or state-based, often using different vegetation community deﬁnitions, and this has made it very difﬁcult to collate for Basin-wide assessments. However good progress is being made now on

Murray-Darling Basin Basin-wide indicators of river health through the MDBC Sustainable Rivers Audit. The states have also agreed on a common monitoring programme for ﬁsh, macro-invertebrates and hydrological regime (Fig. 13.12). This falls short of more integrated, Basin-wide monitoring, but it is an evolutionary step in the right direction. At the national level, a new framework has just been established for assessing river health at broad scale (Norris et al. 2007). Murray-Darling Basin Commission investments in knowledge are usually focused on the rapid harnessing of existing knowledge to support new policy development, once an issue has built sufﬁcient steam to justify political action, and then further technical development to underpin implementation within the constraints of the agreed policy. Urgency and state domination of the committees that steer speciﬁc policy developments sometimes means that the wider scientiﬁc community is less involved than it could be. Scientiﬁc advisory committees, like those serving most Basin Commissions in the USA, have not evolved for the Murray-Darling Basin. The scientiﬁc underpinning of the Living Murray Initiative, however, was an excellent example of how best available expert knowledge could be quickly combined and structured in a model (Murray Flows Assessment Tool) to evaluate the ecological outcomes of returning different volumes of environmental ﬂows to the river system (MurrayDarling Basin Commission 2004). At times the Commission has also invested in more strategic research programmes, like the National Dryland Salinity Program from 1993 to 2004, and in partnered research groups like the Murray-Darling Basin Freshwater Research Centre and the e-Water Co-operative Research Centre, which are both current investments. The knowledge system about the Basin includes numerous local studies carried out by consultants and local organizations to answer speciﬁc questions about speciﬁc places. Drawing on the formal scientiﬁc literature for principles these studies can be inﬂuential in local and catchment scale decisions, but it is a concern that they are often not well scientiﬁcally

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Fig. 13.12 Assessment of the condition of the ﬁsh populations at 487 sites across the Basin (MurrayDarling Basin Commission 2008). Fish condition integrates two indicators, ‘expectedness’ and ‘nativeness’. ‘Good condition’ indicates the presence of many of the expected native species and fewer alien species; poorer ratings indicate progressively fewer of the expected native species and more alien species. No valleys merited better than moderate rating.

reviewed. A recent evaluation of the knowledge base used to determine environmental watering plans at one of the icons sites along the Murray revealed that only one of many publications (scientiﬁc and ‘grey’ literature) about the site satisﬁed criteria of scientiﬁc rigour and relevance (Abel et al. 2006). Developing the technical capacity of catchment management agencies across Australia was a priority recommendation of the Keogh report discussed earlier (Keogh et al. 2006).

13.6 Key Knowledge Gaps for Integrated Basin Management Finally, this section outlines where key knowledge gaps still lie for sound development and

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implementation of nested Basin and catchment policies and plans. This section draws in particular from Chartres (2006), who, writing in terms of the national agenda for water reform, recognizes the role of science: ‘Water planning, regulation and trading in an environmentally sustainable framework, the key tenets of the National Water Initiative, require robust scientiﬁc underpinning if the overall water reform process is to be transparent, credible and evidence-based’ (my emphasis). 13.6.1

The water resource

There are three major gaps in knowledge that are limiting better management of the water resources of the Basin: methods for consistent and accurate measuring, monitoring, accounting and reporting of Basin water resources (surface and ground water); improved knowledge about the interconnections between surface and ground water so that water resources are not double counted or double allocated; and methods of predicting change in water resources (both surface and ground water) with changes in climate or land use. These need to be set in the context of the whole Basin (Box 13.3). Major new science initiatives have recently been launched to address these. For example, the Water Resources

Box 13.3 • Balances inputs and outputs. • Explains connections between catchments and between surface and groundwater. • Explains poorly connected or disconnected hydrological elements. • Improves water accounting. • Indicates the relevant scale or scales for addressing speciﬁc questions. • Ensures numerical models include all relevant ﬂows and storages. A whole-of-basin approach (Kirby et al. 2006).

Observation Network (WRON) is developing distributed, web-based integrated data management tools which will allow end-users to access data quickly and easily, with the knowledge that the data has been subjected to a strict set of standards. The South East Australia Climate Initiative (SEACI) is developing new methods for downscaling climate projections and linking them with hydrological models to improve predictions of stream ﬂow across the Basin. And most recently, the Murray-Darling Basin Sustainable Yields project is assessing the implications of drought and climate change on water availability in every catchment of the Basin, as well as the Basin as a whole. A recent result (Fig. 13.13) for one of the

Fig. 13.13 Changes in rainfall, water availability and diversions (under current water sharing plans) in the Murray-Darling Basin in 2030 using the median outcome from 15 global climate models and three warming scenarios (Commonwealth Scientiﬁc and Industrial Research Organization 2008). Error bars show the second wettest and second driest outcomes. The north and south of the basin are shown separately because they are inﬂuenced by separate weather systems. The impact of a substantial reduction in runoff is not fully reﬂected in decreased diversions because current water sharing plans protect those diversions in favour of the environment. This emphasizes the need to squarely face the third order of water scarcity in this basin: reallocation, especially in the south where the majority of models predict a drier climate.

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Murray-Darling Basin major catchments in the Basin illustrates the gravity of concern about climate change, especially in the higher warming scenario. A reduction of 30% in runoff will have major consequences for irrigation and wetlands along this river. 13.6.2

Irrigation water use and impacts

Key research gaps are: understanding water, salt and nutrient ﬂows at local and regional level in irrigation areas so that the impact of interventions to reduce irrigated water use or of new irrigation developments can be assessed (e.g. impact on return ﬂows and their water quality); and designing farming systems and support tools, such as longer range rainfall forecasts, so that irrigation farmers can be more adaptive to variability in water availability 13.6.3

Ecology

A major summary of largely qualitative knowledge about the river ecosystems of the Basin was published in 2001 (Young 2001) and further built on in development of the Murray Flows Assessment Tool (Murray-Darling Basin Commission 2004). A notable shortcoming of this tool was the inability to quantitatively link water regime to ecological outcomes. On the hydrological side, methods for spatial modelling of water ﬂows and volumes over very large areas of low relief need further development, as well as signiﬁcant quantities of new data to populate them. On the ecology side, quantitative data on the growth and relationships between key functional species in rivers, wetlands, ﬂoodplains and the estuary – and how they respond to water regime – are still very sparse. The lack of historical data, naturally variable water regimes and ecosystems that are still in transition from natural to modiﬁed states make the task of teasing out robust relationships even more challenging. Interaction between ﬂow relationships and salinity and their combined impact on water ecosystems is a further knowledge gap. The water ecosystems of the Darling system are under-studied compared to the Murray system.

Box 13.4 The Barmah Forest … is a place where I come to revitalize, to re-energize and to refresh my spirit. … I sit at my favourite sand bar and swim in the Murray … sit quietly by one of the many lakes, wetlands and billabongs … wander through the many box ridges and gaze at the beauty of the wild ﬂowers and listen to the song of the birds and watch the tracks of the animals.

Comment by participant in non-market valuation survey (Dyack et al. 2007).

13.6.4 Socio-economics Four key knowledge challenges in this area are: (i) articulating and valuing the non-use values of water (e.g. Box 13.4); (ii) developing predictive capacity for social response to change and integrating it with economic analyses to evaluate policy options for dealing with changes in the major drivers of land and water use; (iii) developing market-based approaches to managing the externalities of using natural resources; and (iv) design of partnership structures and processes that harness and integrate initiative across Basin, region and community scales. 13.6.5 Integration and whole Basin Filling these sectoral knowledge gaps and having the knowledge used will only get part way towards more integrated management of the Basin. Post-hoc collation of multi-disciplinary knowledge misses the interactions between sectors that may be just as critical in determining Basin outcomes. Key knowledge challenges are in: describing and analysing the Basin as a socioecological system; developing complex systems science to gain insights into thresholds, tradeoffs and longer-term resilience; and understanding when and how to iterate between the system view and component knowledge. As well as technical challenges here, there are structural and

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attitudinal challenges – for water managers, policy makers and scientists – because of our strongly held reductionist world view (Fig. 13.13). 13.6.6 Creating impact from science As well as gaps in biophysical and socioeconomic knowledge about the Basin itself, there are knowledge gaps in how science is best drawn on by decision-makers. Key gaps include: quantifying and expressing the uncertainty involved in modelled data and in modelling complex systems; and understanding processes of interaction between science and decision-making in a context where science outputs and policy options are contestable, including the development of methods and tools to support deliberation, e.g. ‘translation goods’ (Jackson 2006). 13.6.7

Longer-term future

Finally, and already clearly on the horizon, there is research to be done on the interaction of energy, land and water under a changing climate. Making decisions about the use of Basin resources that does not include consideration of possible big changes in the nature and cost of energy is short-sighted, especially if proposals involve substitution of energy for another resource. For example, encouraging irrigation farmers to shift from high volume, gravity-fed systems to lower volume, pressurized systems will save water but entail higher energy costs. The salt interception schemes that pump saline water away from the River Murray to control water quality are also a highly energy-dependent intervention. The challenge in building the knowledge system in this Basin right now is keeping one eye on a future that isn’t here yet, while delivering good science to governments and communities who are under great strain due to the current drought and past over-allocation of water (Fig. 13.14). The second order responses discussed earlier (economic mechanisms that increase efﬁciency) are dominating now but they do not replace the need to be ready for third order responses (new options and reallocation). Many

Fig. 13.14 Decisions about water use in the basin have long-term impact. In 30 years’ time what will this Australian think of the decisions we are making today?

scientiﬁc challenges, of very large practical beneﬁt, remain.

Acknowledgements I acknowledge with thanks the six project leaders who have worked with me in an integrated Basin knowledge programme for the last four years, and taught me much about the Basin: Shahbaz Khan, Glen Walker, Nick Abel, Rod Oliver, Sebastien Lamontagne and Lu Zhang. Mac Kirby, Albert van Dijk and many others in the project teams also contributed to my learning. The Murray-Darling Basin Commission hosted me for two years. Nick Abel, Daniel Colloff, Ian Overton, Marcus Anastassiou and Peter Grieg kindly provided constructive comments on this manuscript.

References Abel, N., Roberts, J., Reid, J. et al. (2006) Barmah Forest: a review of its values, management objectives and knowledge base. Report to the Goulburn Broken Catchment Management Authority. CSIRO, Division of Sustainable Ecosystems, Canberra.

Murray-Darling Basin Australian Government (2006) National Plan for Water Security. http://www.pm.gov.au/docs/national _plan_water_security.pdf (last veriﬁed 8 May 2007). Bryan, B. and Marvenek, S. (2004) Quantifying and Valuing Land Use Change for Integrated Catchment Management Evaluation in the Murray-Darling Basin 1996/97–2000/01. Stage 2 Report to the Murray-Darling Basin Commission. CSIRO Land and Water Client Report, Glen Osmond, South Australia. Chartres, C.J. (2006) A Strategic Science Framework for the National Water Commission. National Water Commission, Australian Government, Canberra, Australia. Clark, S.D. (2002) Divided power, co-operative solutions. In: Connell, D. (ed.), Uncharted Waters. Murray-Darling Basin Commission, Canberra. Close, A. (1990) The impact of man on the natural ﬂow regime. In: Mackay, N. and Eastburn, D. (eds), The Murray. Murray-Darling Basin Commission, Canberra, Australia. Commonwealth Scientiﬁc and Industrial Research Organization (2008) Water Availability in the Murrumbidgee. A Report to the Australian Government from the CSIRO Murray-Darling Basin Sustainable Yields Project. CSIRO, Australia. Connell, D. (2007) Water Politics in the Murray-Darling Basin. The Federation Press, Sydney, Australia. Council of Australian Governments (2008) MurrayDarling Basin Reform. http://www.coag.gov.au/ coag_meeting_outcomes/2008-07-03/docs/Murray_ Darling_IGA.pdf (last veriﬁed 20 November 2008). Deakin, A. (1886) Water Supply and Irrigation: Speech of the Honorable A. Deakin, Chief Secretary of Victoria, in submitting to the Legislative Assembly a Bill to make better provision for the supply of water for irrigation, 24 June, 1886. http://nla.gov.au/nla. ms-ms1540-10-372 (last veriﬁed 13 October 2008). Doyle, T. and Kellow, A. (1995) Environmental Politics and Policy Making in Australia. Macmillan Education, Melbourne. Dyack, B., Rolfe, J., Harvey, J., O’Connell, D., Abel, N. and Ryan S. (2007) Valuing Recreation in the Murray: an assessment of the non-market recreation values at Barmah Forest and the Coorong. CSIRO Water for a Healthy Country, Canberra, Australia. Hamstead, M., Baldwin, C. and O’Keefe, V. (2008) Waterlines – Water allocation planning in Australia – current practices and lessons learned. National Water Commission, 2008. Canberra. http://pandora. nla.gov.au/tep/83966 (last veriﬁed 12 November 2008).

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Jackson, S. (2006) Water models and water politics: design, deliberation, and virtual accountability. In: José, A. Fortes, B. and Macintosh, A. (eds), Proceedings of the 7th Annual International Conference on Digital Government Research, DG.O 2006, San Diego, California, USA, 21–24 May 2006. ACM International Conference Proceedings Series. ACM, New York, pp. 95–104. Keogh, K., Chant, D. and Frazer, B. (2006) Review of Arrangements for Regional Delivery of Natural Resource Management Programmes. http://www.nrm.gov.au/publications/books/ regional-delivery-review.html (last veriﬁed 20 November 2008). Kirby, M., Evans, R., Walker, G. et al. (2006) The Shared Water Resources of the Murray-Darling Basin. Murray-Darling Basin Commission, Canberra, Australia. Meyer, W.S. (2005) The Irrigation Industry in the Murray and Murrumbidgee Basins. Technical Report No. 03/05. CRC for Irrigation Futures, Queensland, Australia. Molle, F. (2003) The ‘Closure’ of River Basins: Trajectories and Social Responses. 3rd Conference of the International Water History Association, Alexandria, Egypt, 11-14 December 2003. http:// www.iwmi.cgiar.org/assessment/FILES/pdf/ publications/ConferencePapers/Alexandrie%20 MOLLE.pdf (last veriﬁed 20 November 2008). Monash, J. (1904) Report on the Murrumbidgee Northern Water Supply and Irrigation Bill. Lower Murrumbidgee River Locking League, Balranald, New South Wales. Murray-Darling Basin Commission (2001) Integrated Catchment Management in the Murray-Darling Basin 2001–2010. Murray-Darling Basin Commission, Canberra, Australia. Murray-Darling Basin Commission (2004) Inside MFAT – the Murray Flows Assessment Tool. http://www. mdbc.gov.au/subs/information/mfat/index.htm (last veriﬁed 20 November 2008). Murray-Darling Basin Commission (2007) Water Audit Monitoring Report 2005/6. Murray-Darling Basin Commission, Canberra, Australia. Murray-Darling Basin Commission. (2008) Sustainable Rivers Audit. Murray–Darling Basin Rivers: Ecosystem Health Check, 2004–2007. MurrayDarling Basin Commission, Canberra, Australia. National Water Commission (2004) Intergovernmental Agreement on a National Water Initiative. http://www.nwc.gov.au/resources/documents/

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Intergovernmental-Agreement-on-a-national-waterinitiative.pdf (last veriﬁed 20 November 2008). Norris, R.H., Dyer, F., Hairsine, P. et al. (2007) Assessment of River and Wetland Health: a framework for comparative assessment of the ecological condition of Australian rivers and wetlands. Australian Water Resources 2005. A baseline assessment of water resources for the National Water Initiative. Level 2 Assessment. River and Health Theme. National Water Commission, Canberra, Australia. Van Dijk, A., Evans, R., Hairsine, P. et al. (2006) Risks to the Shared Water Resources of the Murray-Darling

Basin. Murray-Darling Basin Commission, Canberra, Australia. Walker, B. and Salt, D. (2006) Resilience Thinking: sustaining ecosystems in a changing world. Island Press, Washington DC. Woolfe, S. and Brooks, D. (2003) Water scarcity: an alternative view and its implications for policy and capacity building. Natural Resources Forum, 27, 99–107. Young, W.J. (ed.) (2001) Rivers as Ecological Systems: the Murray-Darling Basin. Murray-Darling Basin Commission, Canberra, Australia.

Image facing chapter title page and all images within the chapter: Courtesy of Sarah Ryan.

14

Water Resources in South East England – a Dilemma in Sustainable Development JOHN C. RODDA1

1

Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire and Hydro-GIS Ltd, Chalgrove, Oxfordshire, UK

14.1

The Dilemma

Because South East England is one of the driest parts of the UK and one of the most densely populated, the volume of water available per head of population is comparatively low. In fact, it is little more than half the ﬁgure of 1000 cubic metres per person per annum which has been employed by the World Bank, using concepts introduced by Falkenmark (1997), to indicate countries in the developing world which are suffering water stress. Continuing population growth, climate change and the accompanying sea level rise are projected to place even greater strains on South East (SE) England’s water resources. There are further constraints imposed by legislation such as the EU Water Framework Directive, the Habitats Directive, the Ramsar Convention and like international agreements. The SE Plan, which has been formulated by the SE England Regional Assembly at the behest of the UK Government, proposes to increase the population of the region by more than a million by 2026, principally by building upwards of 30,000 new houses a year to accommodate them. The viability of the Plan hinges on a number of factors, especially on the need to upgrade the infrastructure and particularly on the paucity of the region’s water resources and whether or not Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

they are adequate to permit the development to proceed. If they can assure the sustainability of the development proposed in the Plan, then there is the even more difﬁcult question of their ability to cope with continuing growth after 2026. This same dilemma is facing the politicians, policy makers and stakeholders in many regions of the world where the aim is to achieve sustainable development in the face of rapid growth. The answer in the past has been to enlarge the footprint of the city or region concerned. But is this still a viable option at this time in an already crowded country? This chapter examines the different aspects of water resources in the SE Region and their management, within a unit deﬁned by government decision rather than by physical or other factors, a unit which has come into existence quite recently.

14.2 Approaching River Basin Management The SE England Region is not a river basin but a number of basins, including the major part of the Thames Basin. Nevertheless the region shares many of the problems facing river basin management, including the management of water resources. Worldwide there are c. 260 international river basins shared by two or more nations, few of the boundaries of these basins, namely their watersheds, coincide with political and

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administrative boundaries. The Danube Basin is an example, the most international of river basins containing 19 nation states within its catchment of about 800,000 km2. While 100% of Hungary and 97.4% of Romania lie within the Basin (International Commission for the Protection of the Danube River 2004), major parts of the territories of the other nations extend beyond the Danube, for instance, some 90% of Austria and 99.8% of Switzerland. Indeed in a number of instances, the Danube and its tributaries form the borders between the riverine states. A similar situation exists within nation states. Internal boundaries between the provinces, prefectures, counties and other units of regional and local government rarely coincide with river basins. Rather than physical features, factors such as, history, religion, language and culture have often played important parts in the locations of boundaries, Then there is the fact that in many countries of the new world straight lines divide the nation’s territory into states and provinces, completely ignoring physical geography. Effective and efﬁcient water resources management has to ﬁnd ways of overcoming the impediments mentioned above and to move away from the over-used sectoral methodology, adopting a holistic attitude instead. Increasingly, integrated water resources management (IWRM) is advocated as the most sound and effective approach, because it engages both water resources and human activities, embracing ecological and socio-economic issues within a river basin based ecosystem approach (UNESCO 2006). IWRM aims to stimulate cross-sectoral co-operation and the co-ordinated management of surface and ground water resources together with land management, in order to optimize the beneﬁts to society, ecosystems and the economy in a sustainable fashion. IWRM may be simple to conceive and to advocate, but it is extremely difﬁcult to put into practice. It has to assimilate a wide range of complex and competing issues, as well as dealing with ineffective institutions, weak governance, lack of regulatory mechanisms and low levels of stakeholder participation. There have been signiﬁcant

steps towards IWRM, however, in many development projects, such as through the application of the four Dublin Principles (World Meteorological Organization 1992) (Box 14.1). Perhaps of greater consequence is the Water Framework Directive (Council of European Communities 1997), which came into force across the European Union in 2000. It requires member states to develop and enact integrated management in all river basins by 2015, in order to achieve good or high ecological status in all water bodies. Also important is the European Commission’s Habitats Directive (1992) which aims to protect over 200 sites across the EU and over 1000 species, many involving water. Of course the accurate and reliable assessment of the quantity and quality of the water available for use is essential to the development and management of water resources. Without it IWRM is an impossibility (World Meteorological Organization/UNESCO 1997)! Watersheds and political boundaries do not often coincide in the UK and within its relatively small land area, approximately 244,000 km2, there are few cases where watersheds currently deﬁne the political units as well as the hydrological ones. An exception was from 1973 to 1989 when, as a result of the 1973 Water Act, 10 regional water authorities and a National Water Council were established covering England and Wales. These authorities combining both regulatory and operational roles were based on the major river basins. They planned and controlled all uses of water, being responsible for water resources assessment, water supply, sewerage and sewage disposal, as well as for pollution control, ﬁsheries, ﬂood protection, recreation, environmental conservation and navigation. More recently, certain functions of central government have been devolved to the Parliament in Scotland and to the Assemblies in Wales and Northern Ireland. In addition the English counties, which historically have made up the second tier of government in England, have been grouped together in eight regions in an attempt by the national government at Westminster to promote ‘regional’ government. London is a separate ninth region. This is a result of the enact-

Water Resources in South East England

325

Box 14.1 The Dublin Principles Principle No. 1 – Fresh water is a ﬁnite and vulnerable resource Since water sustains life, effective management of water resources demands a holistic approach, linking social and economic development with protection of natural ecosystems. Effective management links land and water uses across the whole of a catchment area or groundwater aquifer. Principle No. 2 – Water development and management should be based on a participatory approach, involving users, planners and policy makers at all levels The participatory approach involves raising awareness of the importance of water amongst policy makers and the general public. It means that decisions are taken at the lowest level, with full public consultation and involvement of users in the planning and implementation of water projects. Principle No. 3 – Women play a central part in the provision, management and safeguarding of water This pivotal role of women as providers and users of water and guardians of the living environment has seldom been reﬂected in institutional arrangements for the development and management of water resources. Acceptance and implementation of this principle requires positive policies to address women’s speciﬁc needs and to equip and empower women to participate at all levels in water resources programmes, including decision-making and implementation, in ways deﬁned by them. Principle No. 4 – Water has an economic value in all its competing uses and should be recognized as an economic good Within this principle, it is vital to recognize ﬁrst the basic right of human beings to have access to clean water and sanitation at an affordable price. Past failure to recognize the economic value of water has led to wasteful and environmentally damaging uses of the resource. Managing water as an economic good is an important way of achieving efﬁcient and equitable use, and of encouraging conservation and protection of water resources.

ment of the Regional Development Agencies Act of 1998.

14.3 The South East England Plan South East England is one of the nine English regions (Figs 14.1, 14.2). This new level of regional government at present consists of an appointed Regional Assembly and a Development Agency. One of the main tasks of the Assembly is to produce a spatial plan for the region as required by The Planning and Compulsory Purchase Act of 2004. This Plan, the South East Plan, was pub-

lished initially in January 2005 (South East England Regional Assembly 2005a) and again in July 2006 (South East England Regional Assembly 2005b) as drafts for consultation. A third draft was published in March 2006 (South East England Regional Assembly 2006) for submission to government. This version of the Plan was the subject of a lengthy examination in public held between November 2006 and May 2007, before consideration by central government in late 2007 and further consultation before proceeding with the Plan. The Plan covers the period from 2006 to 2026 and consists of several development options at

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john c. rodda

Fig. 14.1 The South East England Region – physical features.

the regional and sub-regional scales. It deals with housing, the environment including water resources and ﬂooding, transport, the economy, waste and recycling, as well as social and cultural issues and a number of other matters. Its stated aim is to produce a healthierr, a more sustainable pattern of development and a more robust economy. In reality, however, the Plan is focused on how to accommodate a million extra people in the region by 2026 through the construction of 32,000 or more houses a year, a ﬁgure determined by central government. It is important to recall that while the Plan does not include London, London has an enormous inﬂuence on the region. It can be argued that this inﬂuence, based on proximity, is the only factor which brings any sense of homogeneity to the region. Indeed the counties of Kent, Surrey, East and West Sussex, Hampshire and the Isle of Wight, Berkshire, Oxfordshire and Buckinghamshire, which are covered in the Plan, can claim little cohesion and a sense of common identity, except possibly between those sharing

the Thames Basin. The disparity is greatest between the southern and eastern coastal counties and those inland to the west and north of the region which look westwards towards Wessex and to the Midlands. London’s inﬂuence on the region also affects its water resources because much of London’s water supply is abstracted from the Thames at Staines and passed through the reservoirs near Heathrow Airport. Consequently developments in the Thames Basin can affect the city’s supplies. Similarly the increase in London’s population from 7.3 million in 2003 to 8.1 million in 2016, as foreseen in the London Plan (2004), will require a greater volume of water to be taken from the Thames. Swindon, population 155,000, which is located immediately to the west of the region, draws much of its water supply from the Thames Basin and will require even more in the future, as it is due for further expansion under the South West of England Plan. Swindon’s treated efﬂuent is returned to the upper Thames so that about 90% of its used water is available for re-

Water Resources in South East England

327

Fig. 14.2 The South East England Region – political.

abstraction at Farmoor Reservoir just upstream of Oxford.

14.4 Characteristics of the South East England Region The SE England Region covers an area of 19,090 km2 containing a population estimated to be 8122,200 in 2004 (South East England Regional Assembly 2007). This is a population density of 425 per square kilometre, compared to a density of 240 for the UK as a whole and 377 for the Netherlands – the most densely populated country in Europe (The Times 2000). Excluding London, it is the most prosperous of the UK regions, it has the largest population and it gener-

ates the highest gross added value per head (The Economist 2007). The region is an international gateway, but its road and rail systems are very heavily congested, its airports are often stretched to capacity and its seaports deal with large volumes of trafﬁc, particularly Dover and Southampton. It has a high value landscape containing the highest proportion of land protected by national and international designations in the UK and it has the greatest proportion of woodland of any English region. Its attractiveness, proximity to London, its cross Channel routes and economic success make it subject to substantial development pressures, both national and international (South East England Regional Assembly 2005a). Currently these pressures give rise to serious doubts over whether the region’s

328 Table 14.1

Location

john c. rodda Selected climate data 1971 to 2000 averages (source: UK Met Ofﬁce website) July mean maximum temperature (°C)

February mean minimum temperature (°C)

Annual sunshine (hours)

Annual precipitation (mm)

17.8 22.5 22.3

3.2 1.5 1.7

1848.6 1534.7 1537.4

789.7 647.1 642.0

Eastbourne Wisley Oxford

infrastructure has the capacity to cope with present demands. They cause even more anxiety about the future. 14.4.1

Physical geography

Like the rest of the UK, the region’s physical geography is quite varied. Its coastline, which stretches some 580 km from the Thames Estuary to the point where the Hampshire Avon enters the sea at Christchurch, deﬁnes the limits of the region to the east and south and for part of the north (Fig. 14.1). Its western edge partly follows the course of this river northwards, but by and large, its western and northern limits are not based on physical features, except where its boundary follows the lower course of the Thames eastwards from London into the Thames Estuary. These physical features result, in the main, from earth movements during the Alpine orogeny in the Tertiary, which warped the Jurassic and Cretaceous sedimentary measures covering the south and east of England. Over the next 50 million years or so, erosion and deposition took place with water and ice as the main agents. These processes became dramatic during the Quaternary Ice Ages but they have slowed subsequently. Together they have created a gently rolling landscape with a small amplitude of relief, but one with gradients of up to 20% on certain scarp slopes. The climate of the British Isles is, in general, characterized by a small range of temperature between summer and winter, a fairly even distribution of precipitation from month to month and a freedom from extremes of one sort or another (Rodda et al. 1976). The South East is no excep-

tion although its summers tend to be a little warmer and its winters slightly cooler than the rest of the country. A February mean minimum temperature of about 4°C and a July mean maximum of 17°C would be typical of much of the region (Table 14.1). Climate change threatens to alter a number of the region’s climatic and hydrological variables. Projections from the Meteorological Ofﬁce indicate a rise in mean UK temperatures of between 2 and 3.5°C with the South East experiencing a possible increase of up to 5°C by 2080. Winters are likely to be milder with 30% more rain, much of this falling in more intense storms. Summers are anticipated to be warmer and drier with many more days when the temperature exceeds 36°C. A 60% decline in summer precipitation is considered very likely, a greater proportion occurring in more intense falls, although there is less certainty about the future summer rainfall than that for the winter. Evaporation rates are expected to be higher and soil moisture deﬁcits greater, both affecting the volume and distribution of runoff and of groundwater recharge. A sea level rise of between 19 and 79 cm will have a substantial impact on coasts, estuaries and wetlands. Kent is reputed to be the ‘Garden of England’ with numerous orchards and hop ﬁelds, in contrast to the more arable counties of Oxfordshire and parts of Berkshire where there are large areas of wheat, barley and oil seed rape (Table 14.2). Excluding the beech woods of the Chilterns, the amount of woodland increases from north-west to south-east, with the Weald being the most forested area. The region contains one National Park, the New Forest, and one proposed National Park, the South Downs. In addition there are nine

Water Resources in South East England Table 14.2 Land use in the South East England region (data provided by the Centre for Ecology and Hydrology) Area Category Broadleaf forest Coniferous forest Arable Improved grassland Semi-natural grass Upland Open water Built up Coastal Total

ha

%

286,717 46,404 646,585 409,412 222,530 35,612 8,926 249,189 19,965 1,925,340

14.9 2.4 33.6 21.3 11.5 1.3 0.5 12.9 1.1 100

areas classiﬁed as areas of outstanding natural beauty and more than 700 sites of special scientiﬁc interest. There are nearly 100 towns and cities in the region with populations in excess of 10,000, including nine with populations over 100,000, Southampton (234,224) and Reading (232,662) being the largest. Many of these urban areas are clustered close to the south and east of London, while others are strung along the South Coast (Fig. 14.2). 14.4.2

Hydrology and water resources

The Thames is by far the longest river in the SE Region with a total length of 338 km, although it ﬂows through the London Region for the lower part of its course. Its discharge is also by far the largest. Its average daily ﬂow at Kingston, at present the lowest gauging point, is 78.1 m3 s−1 from a basin of 9948 km2, the other main rivers having relatively small average ﬂows (Table 14.3). Many drainage basins contain a mixture of permeable and impermeable formations, so the response to rainfall can be very rapid in certain areas and very delayed in others. Depending on the lithology, groundwater can contribute larger or smaller volumes to the total ﬂow. Taking the Thames Basin as a whole, Andrews (1962) estimated that about 57% of the Basin contributed

329

to groundwater discharge and this produced on average a contribution of about 40 m3 s−1 to the daily mean discharge of the Thames at its lowest gauging point. The principal aquifer is the Chalk (Fig. 14.3), the most important in Britain (Rawson 2006), but Lower Greensand, Corallian Beds and Great Oolite Limestone are also noteworthy, as are the Folkestone Beds and the Hythe Beds in the Weald. Knowledge of the hydrology and water resources of the region comes from the network of observing stations which regularly monitors the changing state of the aquatic environment; it also comes from the results of various surveys, from satellite imagery, the weather radar network and from a number of other sources. The latter includes research studies of certain catchments, such as those carried out recently on the Pang and Lambourne (Wheater et al. 2006). SE England is possibly that part of the world where more environmental variables – physical, chemical and biological – have been recorded for longer and more intensively than for any other. At some stations, the Radcliffe Observatory, Oxford, for example, records of the weather and other phenomena date from early in the nineteenth century. The earliest systematic records of river ﬂows were started in 1883 on the Thames at Teddington, rather later than for a number of rivers in mainland Europe, while one of the oldest continuous record of groundwater levels is for Chilgrove in the South Downs which commenced in 1836. Now, as part of the national network reporting to the Meteorological Ofﬁce, there are 185 daily read raingauges in the region, 75 automatic rain recorders and 24 climate stations, together with 15 synoptic stations reporting hourly. In addition there are four weather radars that cover the region, forming a component of the national weather forecasting system. As parts of the respective UK networks, river ﬂows are gauged at about 150 sites in the region, river water quality at 10 sites and groundwater levels at about 50 observation wells. There is, however, a large number of other wells and boreholes where quality is monitored which are not part of the national network, many being pumped

330

john c. rodda

Fig. 14.3 The South East England Region – geology.

331

Water Resources in South East England

Table 14.3 Data for selected rivers in the South East England region (data supplied by the National River Flow Archive, Centre for Ecology and Hydrology) River Gauging station Basin area (km2) Start of record Mean ﬂow (m3 s−1) Maximum ﬂow (m3 s−1) Date 7-day minimum ﬂow Date (middle day)

Thames

Thames

Medway

Great Stour

Arun

Itchen

Test

Teddington/ Kingston 9948 1883 78.1 800

Days Weir

Teston

Horton

Alfolden

Allbrook

Broadlands

3444 1938 28.3 349

1256 1956 11.0 359

345 1964 3.18 33.1

139 1970 1.72 Under review

360 1958 5.38 20.5

1040 1957 11.1 >40

18/11/1894 0.001

17/3/1947 0,15

16/6/1968 0.43

8/2/2001 0.62

0.07

13/2/2000 2.21

13/2/2000 3.85

29/10/1934

9/7/1976

22/5/1976

1/10/1997

24/8/1976

12/8/1976

8/7/1976

for water supply. The same applies to surface water quality: regular samples are taken for analyses at hundreds of sites, some equipped with automatic water quality monitors. These are located particularly where water is abstracted for public water supply from rivers and boreholes and downstream from discharges of efﬂuent. Over the past 40 years a river pollution survey has been undertaken every 10 years, based on a combination of chemical and biological assessments. Now they are mounted more frequently and, at different times, a number of other biological surveys of the region’s rivers, ponds, lakes and reservoirs have been carried out. Much of these weather and water data are collected and archived by the Meteorological Ofﬁce, the Environment Agency, the water companies and the Natural Environment Research Council, but private individuals, local authorities non-governmental organizations and other bodies are also involved. The water resources problems facing the region hinge on the spatial and temporal variations in the different hydrological variables, in terms of both quantity and quality, the principal driver being, however, the volume and distribution of the precipitation. The average annual rainfall is approximately 740 mm (T.J. Marsh 2007, personal communication), while the annual actual evaporation is estimated to be about 480 mm (Rodda et al. 1976), giving a ﬁgure of

260 mm for the effective rainfall. This is close to the value of 280 mm which Marsh (2007) found was the mean annual runoff from 13 SE Region catchments. This 260 mm represents the volume of water which, in the average year, renews the surface and ground water resources, and is available for human use and for maintenance of the aquatic environment. It is approximately 610 cubic metres per annum for each inhabitant of the SE Region or 1670 litres per day. This ﬁgure is of course much less in a drought year. For example in the year to September 2006, the rainfall over the region was about 80% of the average, providing 224 mm for all the different uses and users, equivalent to 526 cubic metres per annum or 1440 litres per day. Floods are usually generated by intense rain storms, in winter associated with frontal movement across the region and in summer by more localized convectional storms. Some of the highest ﬂows occurred in March 1947 due to a combination of heavy rain and rapid snow melt. Some of these ﬂows, however, were matched in the upper Thames in July 2007, when on the 19 and 20 July a moist subtropical airmass stagnated over central England and produced extreme rainfalls over a range of durations (Hydrological Summary 2007). At Maidenhead 51 mm was recorded in 63 minutes, while at Pershore just over the north-west edge of the basin, 145 mm

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john c. rodda

was registered in 25 hours, a fall of about a 1000year return period. Oxford, Abingdon and a number of smaller towns and villages in the upper part of the Thames basin suffered from serious ﬂooding from river water and also from excess surface water not being able to drain away. Another feature of ﬂoods in recent very wet winters has been the generation of groundwater ﬂoods in the usually dry valleys in the Chalk. Following the winter of 2000/2001, they lasted for several months in a number of areas. Flood forecasts are provided to warn communities, which normally are well protected from ﬂoods with a low recurrence interval by drainage systems and structures of various types. Flood warnings are disseminated and emergency services are well organized to help protect the 208,000 properties at risk from ﬂooding in the region from a 100-year ﬂood or ones with a lesser return period (South East England Regional Assembly 2006). The region beneﬁts from an aquatic environment largely free from pollution, one which normally encourages fauna and ﬂora to ﬂourish. Surveys of the biological and chemical health of over 4000 km of the region’s river conducted in 2006 (Table 14.4) (Department for the Environment, Food and Rural Affairs 2007a) showed that, on average during the year, only about 2% were classiﬁed as poor or bad in biological terms and about 6% in terms of water chemistry. However, accidental spillages and releases take place from time to time and storm overﬂows occur occasionally, while farming practices cause high levels of nitrate, phosphate and other deleterious materials to exist in surface and ground water. In addition, because the mean ﬂow in many of the region’s rivers is small, they often contain a large percentage of treated sewage efﬂuent, a percentage which increases as the natural ﬂow decreases. The treatment capacity of sewage treatment works at a number of locations in the region is limited (Birks 2004) and increased volumes of sewage will demand larger sewers and more energy spent on pumping. Some works are operating at the boundaries of current treatment technology. To prevent degradation of receiving

Table 14.4 Results of the river pollution survey in 2006 (Department for the Environment, Food and Rural Affairs 2007b) (percentage of total length of river surveyed in the four classes) Type of survey

Good

Fair

Poor

Bad

Biological Chemical

77.3 65.0

20.8 29.1

1.9 5.5

0 0.3

waters consent conditions will have to be tightened and technology pushed into more and more expensive methods. More exacting standards of treatment may not be attainable, resulting in decreasing river water quality with possible impacts on biodiversity, ﬁsheries and human health (Box 14.2). Indeed the capacity of a number of rivers to take an increased load of efﬂuent limits the region’s ability to accommodate further growth. Box 14.2 Case study – Basingstoke (from Birks 2004) Basingstoke, population 90,000, is located in the headwaters of the River Loddon, a tributary of the Thames. It has expanded rapidly in recent years and an additional 10,000 houses are proposed to be built by 2011. However, the sewage treatment works is already operating to extremely stringent consent conditions on its efﬂuent. Should the development take place, conditions will have to be tightened further to protect the quality of the Loddon. Alternatives solutions are not feasible. Even if the new conditions are met, they only apply to BOD, suspended solids and ammonia. Other pollutants such as copper, zinc and endocrine disrupters will increase in volume as they are not controlled. 14.5 Water Resources Management and Use of Water 14.5.1 Governance Water governance in the region, as in much of the remainder of the UK, can at best be described

Water Resources in South East England as fragmented. Indeed it appears to be far from the ideal espoused by IWRM. This is mainly because of the large number of bodies which share interests in and responsibilities for the various processes of planning, assessing, managing and utilizing the resource and because of the limited amount of coordination that exists between them. One of the conclusions of a recent report of the House of Lords Science and Technology Committee (House of Lords 2006a) sums up the situation: Responsibility for water management is dispersed and unclear. We need clearer lines of responsibility, greater accountability and more effective funding procedures. Water management should be a partnership in which water companies, the regulators, Government and the consumer can all engage in a constructive dialogue. Stakeholder engagement requires transparency, accountability and a mutual respect for the interests of all participants.

It is only fair to say that this fragmentation is a situation that has pertained for the majority of the time since society and economy in the UK began to be organized, perhaps 300 years ago. Despite these disadvantages the population of the region generally enjoys a very satisfactory water service with some 99.9% of consumers being connected to a public water supply and a little fewer to public sewers. The present structure of water governance in England and Wales stems from a long series of acts of Parliament dealing with various aspects of water, dating back to the middle of the nineteenth century, one of the most inﬂuential at the present time being the 1989 Water Act. This Act privatized the 10 regional water authorities set up in 1974, creating from them 10 water private water companies owned by shareholders, which became responsible for water supply and waste water across England and Wales. Thames Water, for example, supplies 8.1 million customers through an infrastructure which includes 23 surface water reservoirs, 98 water treatment works and 31,000 km of water mains (Cook 2006). In certain parts of the country water supplies

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continued to be provided by the 15 private water companies that were in existence prior to 1974, some being established in the nineteenth century (Box 14.3). Together these 25 companies became accountable to the newly created National Rivers Authority (NRA), the Ofﬁce of Water Services (Ofwat) and the Drinking Water Inspectorate (DWI) of the Department for the Environment, Food and Rural Affairs (Defra). These three bodies regulate environmental, ﬁnancial and economic matters, together with drinking water quality and safety. With the passage of the Environment Act 1995, in 1996 the NRA became part of the Environment Agency (EA). In 2005, the Consumer Council for Water (CC Water) was established and in 2006, Ofwat became the Water Services Regulation Authority (WSRA).

Box 14.3 Proﬁle of the Mid Kent Water Company (from the Mid Kent website: http://www.midkentwater.co.uk (last veriﬁed 30 November 2008)) The Company was established in 1888 and in 2007 supplied nearly 600,000 customers in an area of 2050 km2, with an average of 165 million litres of high quality drinking water a day. The main towns supplied are Ashford, Canterbury and Maidstone. Some 88% of the water comes from 100 wells and boreholes and is distributed from 34 source stations and treatment works and more than 100 water towers and service reservoirs through 4205 km of pipelines.

The water service companies operate under long-term licensing arrangements overseen by Ofwat which sets price caps on charges to customers on a 5-year cycle (Heather and Bridgeman 2007). Together these companies possess a complex asset base with a replacement value of some £231 billion at 2002–2003 prices. As regards water, the EA is responsible for water resources management, including the control of abstractions and discharges and water quality. It is also

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john c. rodda

responsible for navigation, ﬁsheries and ﬂood risk management. Overall, its objective is to protect and enhance the environment so as to promote the achievement of sustainable development. CC Water was set up to protect the interests of water and sewage customers and to enable their collective voice to be heard in the national water debate. Each of these public bodies has regional arms with one or more units covering the SE England Region, Ofwat and DWI excepted. In the case of the EA, the responsibilities are shared by the Thames Region and the Southern Region, while for CC Water the division is the same. Water

supply and waste water services are provided by Thames Water and Southern Water, but there are currently eight companies which supply only water within the region (Box 14.4). To beneﬁt consumers, the water companies are supposedly in competition through the prices they are required to publish. One unspoken problem is the inability of one company to trust that its neighbour will honour a bulk supply agreement between them when the design drought or worse occurs. The water resources strategy for the region advocated by government, the regulators and the water industry is based on a twin track approach,

Box 14.4 The main bodies and institutions with responsibilities for aspects of water resources management in the South East England region National Department for Food and Rural Affairs (DEFRA) (EU Directives, policy, regulation, land drainage and ﬂood protection, monitoring) Department for Communities and Local Government (CLG) (planning, housing building and the environment) Government Ofﬁce for the South East (GOSE) Environment Agency (EA) (environmental regulation) Ofwat (economic regulation) CC Water (customer care) DWI (drinking water quality) British Waterways (amenity, recreation navigation) Natural England (EU Directives, policy, biodiversity, SSSIs) Regional and local (planning, resources, drainage, ﬂood protection) South East England Regional Assembly County Councils of Kent, East Sussex, West Sussex, Surrey, Hampshire and the Isle of Wight, Oxfordshire, Buckinghamshire and the Berkshire Unitary Authorities, together with the District Councils within the Counties Public companies (water resources, water supply, sewerage, sewage treatment, pollution control) Thames Water, Southern Water Public companies (water supply) Dover & Folkestone, Mid Kent, South East Water, Portsmouth, Cholderton and District, Sutton & East Surrey, Bournemouth and West Hampshire, Sutton and East Surrey

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Water Resources in South East England namely increased demand management and the development of new sources. Protection of the surface and ground water quality is an important component of this approach. The water companies’ water resources management plans and the drought plans and the EA’s Catchment Abstraction Management Strategies (CAMS) are amongst the other important elements.

Table 14.5 Household water use (per capita consumption in litres per day) (House of Lords 2006) Number of occupants 1 2 3 4 5 6

Thames Water

Portsmouth Water

201

222.7 181.7 135.7 128.7 120.8 84.7

124

14.5.2 Water consumption Some 70% of the water supplied in the region is derived from groundwater – the highest proportion in the UK. South East Water found (Environment Agency 2007a) that the cost of abstracting and treating groundwater for distribution to consumers is £35 per million litres, compared to £50 for river water and £450 for desalinated water. Currently the average charge to domestic consumers in England and Wales is 0.17 pence per litre for water, supplied and taken away. Assuming that consumers are supplied by groundwater, the difference between these two ﬁgures would provide a sum of £1665 per million litres to fund the remainder of the water service and to pay a dividend to shareholders. Because only 28% of households are metered and the remaining households are charged on the size of their property, this ﬁgure and others relating to consumption are open to debate. Indeed, that it is difﬁcult to make a reliable estimate of average domestic per capita consumption was a key observation in the House of Lords Report (House of Lords 2006b), which contained estimates ranging from 125 litres per head per day to 178 litres. This Report also displays (Table 14.5) a relationship between water demand and the number of people in a household, including the statement that three people living in separate households use as much water as four people living together. Figures collected by Ofwat (2007) indicate that metered customers in Thames Water and Southern Water pay an average of £165 and £210 per year for water supply and waste water disposal. For customers without meters, in 2007 Thames Water is introducing an average household charge of £245.

Table 14.6 Domestic water consumption in all households in the Thames and Southern Water Areas (litres per person per day) (Department for the Environment, Food and Rural Affairs 2007a) Year 2000–1 2001–2 2002–3 2003–4 2004–5

Thames

Southern

165 159 162 162 157

155 161 159 162 157

A consumption of 160 litres per day per person would seem a reasonable ﬁgure to apply to the whole of the SE Region (Table 14.6), a ﬁgure which has risen from 140 litres per day in the 1980s in the Thames Water area which also includes London. Of course the volume of leakage from the distribution system has to be added to this ﬁgure, which, as a percentage, has been reduced in recent years. Some of this leakage may contribute to stream ﬂows and percolate to aquifers. (Table 14.7). Thames Water’s leakage rates at 33% have been widely criticized as unacceptably high but much of the loss takes place within London rather than outside it. There, a large part of the distribution system dates from the nineteenth century and there are considerable obstacles to repairing and renewing it. For that part of Thames Water supplying the SE Region, a more appropriate ﬁgure would probably be one similar to that for Southern Water, namely 16%, a ﬁgure which is typical of most of the water supply companies in the region. Adding the leakage to the estimate of consumption would give a current per capita demand of 185 litres per day for the region.

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Table 14.7 Leakage as a proportion of water supplied (Ofwat)

Company Thames Southern Folkstone & Dover Mid Kent Portsmouth South East Sutton & East Surrey

Total lLeakage (ml day−1)

Distribution input (ml day−1)

Leakage per property (L day−1)

Percentage

Target total leakage for 2009–10 (ml day−1)

915 92 8 29 30 69 24

2809 586 46 163 180 391 161

261 89 114 116 102 116 90

33 16 17 18 17 18 15

690 92 8 27 30 69 25

There are, of course, a number of other users of water within the region, many abstracting directly from surface and groundwater sources, with much of this water being returned or reused. Reuse is a characteristic of the region: indeed it is said that the ﬂow of the Thames transits many kidneys between the source of the river and its mouth. Jamieson and Nicolson (1984) state that on average about 12% of the water abstracted for supply purposes from the Thames and its tributaries is efﬂuent, but that at certain locations in dry summers this ﬁgure rises to 70%. According to the Environment Agency (House of Lords 2006b), in 1997 abstractions in England and Wales were used for: public water supply, 45.4%; electricity generation, 32.1%; ﬁsh/cress farming, 11.4%; industry, 7.7%; spray irrigation, 0.8%; mineral washing, 0.8%; private water supply, 0.3%; other agricultural uses, 0.3%; and other uses, 1.1% (Table 14.8). The region has suffered from drought in recent years, usually as a result of two dry winters in succession and summers with high temperatures and low amounts of rainfall. The drought of 1975 and 1976 is one outstanding example, but there were others in the 1980s and 1990s. Most recently the low winter rainfalls of 2005 and 2006 sparked a notable drought in the summer of 2006. Under these conditions some water companies in the region initially attempted to manage demand by introducing the ﬁrst level of restrictions on the use of water by consumers, namely hosepipe

Table 14.8 Sectoral demand for water in South East England (in ML day−1) (House of Lords 2006) Sector Direct abstraction Primary industry Spray irrigation Power generation Public water supply Household use Non-household Leakage Total

Thames

Southern

130 20 105

180 30 2

1930 860 1150 4195

660 240 210 1322

bans. These prohibit garden watering, car washing and some other non-essential uses. Several applied for the next level, namely the imposition of drought orders. These resulted in a considerable amount of media attention being given to the effects of drought; consumers were urged to shower with a friend and to adopt other methods of reducing water use. It was reported in the press that this attention succeeded in lowering demand by between 8% and 11% during the summer of 2006 without detriment to public health. One company, Folkestone and Dover Water, successfully applied for ‘water scarcity status’ in March 2006, which allows it to compulsorily ﬁt meters to the premises of all customers, Should there be a change in the conventional design drought standard of 1 in 50 years to say 1 in 200 years?

Water Resources in South East England 14.5.3

Water efﬁciency

The threat posed by drought in recent years and the restrictions on water supplies to counter its effects highlighted the need for greater efﬁciency in the use of water. Public perception, however, is that leakage takes a considerable volume of the water that could otherwise have been used by the consumer. Consequently there is a feeling widespread in the region and beyond that water companies need to reduce leakage before taking other demand management measures. Indeed there is also resentment that water charges have risen considerably since or because of the privatization of water. Critics claim that water companies had not invested sufﬁciently in the systems they operate and are more intent on providing fat proﬁts for shareholders. Despite these sentiments, water efﬁciency is given considerable prominence by Government, the EA, Ofwat and the water industry. Water efﬁciency, the use of less water, can be achieved by changing human behaviour coupled with the use of appliances which function on smaller volumes of water. Most attention in the SE Region is directed at reducing domestic water consumption through the installation of low ﬂow showers, taps and toilets, together with washing machines and dishwashers which run on smaller volumes of water. Also important is the more judicious use of these appliances. Savings of up to 30% can be expected in new properties equipped with these devices and perhaps similar amounts by retroﬁtting old ones: savings of some 21 litres per person per day is a quoted ﬁgure (Environment Agency 2007b). The widespread use of water meters is also seen as a means of reducing use. Many new houses, however, are now being equipped with three, four or more bathrooms often with power showers while swimming pools are becoming more common, along with automatic garden irrigation systems. Use of smart meters which can be programmed to charge more during times of shortage have also been proposed. A more commonly employed device for saving water is the reduction in mains pressure.

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The costs of introducing water saving devices were found to range from 0.4% to 6.2% of the cost of a new dwelling, depending on the techniques used and the type of house (Environment Agency 2007a). Many of the water companies promote higher efﬁciency through educational programmes, leaﬂets, provision of bags to install in toilet cisterns to reduce their capacity, free drought tolerant garden seeds and other means. Water efﬁciency has been discussed in Parliament, the regulators are involved in it and it is the subject for several government committees, newsletters and websites. It also attracts the attention of pressure groups active in the environmental ﬁeld. The urgency of the problem usually disappears, however, as soon as a drought is terminated. A strong and continuous Government-coordinated campaign is needed to lower demand: In the long term the educational system will be crucial in ingraining water efﬁcient behaviour patterns in the minds of consumers. We urge the water companies to maximize their collaboration with schools in this regard. We also recommend that the Government make water efﬁciency – and the rationale behind it – a compulsory part of the citizenship syllabus (House of Lords 2006a).

Changing human attitudes to water and the use of water is of course the key to water efﬁciency, but such changes are very difﬁcult to make, particularly when there are few or no incentives, especially for consumers without meters. Even for those with meters, the charge for water is so low that it forms but a small part of the household budget, which is barely impacted by reducing the use of water. Some studies have shown that meters can reduce consumption by 10% or more. Consequently more metering is seen by Government, the regulators and the industry as a way of lowering consumption along with the use of water bills which are designed to highlight cost, so that the consumer is more aware of charges. Even more ‘artistic’ water bills, however, are likely to have little effect on

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john c. rodda

Fig. 14.4 Forecast demand – 2029/2030 (Source: Enviroment Agency 2007b).

demand, which will only be reduced in this context by moving meters from the inaccessible places where they are now located to kitchens where they can be visible all the time to all those in the household. Rainwater harvesting is also proclaimed as a way of increasing efﬁciency. Householders are encouraged to ﬁt water butts for garden watering and to use grey water from roofs for ﬂushing toilets. But they are also encouraged to use more water by means of the hosepipes, sprinklers and garden irrigation systems which are widely available. If government and the industry wish to be taken seriously about water efﬁciency, the ‘twin track’ approach which is currently advocated needs to be replaced by a ‘three lane’ approach. In the third lane, a concerted and continued Government-coordinated campaign in public awareness and education about water scarcity would parallel the thrusts

on resource development and demand management. Howarth (2004) lists a number of the steps that can be taken: mobile visitor centres, school activities such as plays, low cost sales of water butts and plants for dry gardens, web-based information, a concerted publicity programme for the media. The consequences of these different efﬁciency measures on the use of water are rather difﬁcult to understand and prediction of their impacts would seem to be even more problematical. For the future it might be reasonable to assume that the consumption per head of population will not be changed either upwards or downwards on the grounds that the volume of water saved by some will be balanced by the increase in demand by others. The Environment Agency (2007b) forecasts (Fig. 14.4) that demand will continue to rise in the region and that by 2029/30 use will vary

Water Resources in South East England

339

Fig. 14.5 Availability of surface water resources in the winter (Source: National Statistics Great Britain 2006).

from 150 litres per head per day to over 190. The Agency also forecasts (2007b) no change in leakage rates, so assuming the average use in the region in 2029/30 is some 180 litres per head per day with 16% leakage, this gives a total forecast consumption of about 210 litres per head per day at that time.

14.6 Finding Water Resources for Growth The task of ﬁnding new resources in areas facing rapid population growth has become increasingly complex in many parts of the world. Environmental pressures have largely curtailed the scope for new

reservoirs and overexploitation of groundwater has limited expansion of this source (Hanak 2007), particularly in the developed world. The SE Region is no exception. Indeed the EA has concluded that surface water resources suffer from unsustainable or unacceptable abstraction regimes over much of the region in both winter and summer. In other parts there is no additional water available (Figs 14.5, 14.6). Groundwater sources are in a similar state (National Statistics Great Britian 2006). Despite this knowledge, the SE Plan (SEP 2006) proposes to enlarge four existing reservoirs namely: Bewl, Broad Oak, Clay Hill and Havant Thicket. Thirty-ﬁve pumped storage reservoirs exist in the Thames Basin, most

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Fig. 14.6 Availability of surface water resources in the summer (Source: National Statistics Great Britain 2006).

located near London (Jamieson and Nicolson 1984). Together they contain a useful storage of 220 million m3, some 3 month’s water supply. There is also a proposal by Thames Water (2007) to supplement them by building a large reservoir some 6.7 km2 in area to the south-west of Abingdon to be ﬁlled by pumping from the River Thames mainly during the winter (Thames Water 2007). Its main purpose is to release water during dry periods to maintain the ﬂow of the Thames, so that water can continue to be abstracted downstream to supply London. This reservoir will also feed into the local system to provide water for consumers in and around Oxford and in Swindon.

While these new and enlarged reservoirs will increase the volume of storage in the region, they do not increase the resources of the South East. Another possible method of increasing storage is by artiﬁcial groundwater recharge, whereby surplus water is stored in aquifers for use when supplies are less plentiful. With extensive aquifer systems in the region, artiﬁcial recharge would seem an attractive proposition, but apart from a few pilot studies, this technique has not been employed operationally, as it is to a small extent in the London Region. Using treated efﬂuent to recharge aquifers seems an attractive proposition,

Water Resources in South East England but this could cause adverse public reaction, aquifers may become clogged and suffer water quality problems. 14.6.1

Increasing the resource

To increase the volume of the resource above that currently available, additional water needs to be brought into the SE Region. Among the possibilities are: transfer of water in tankers by sea, cloud seeding to promote rainfall and desalination of brackish and sea water, methods currently being practiced in various parts of the world. These would increase the volume of the resource, but none has seriously been considered for the region hitherto. In 2006, however, Thames Water submitted a planning application to build and operate a desalination plant at Beckton in East London, with a capacity of 140 million litres a day – sufﬁcient to supply about 1 million people. Planning permission was granted for this proposal in July 2007, but it is a contentious scheme. Although desalination has become a more efﬁcient process in recent years, it still consumes large amounts of energy and energy costs are rising rapidly. It also produces brine which is difﬁcult to dispose of to the aquatic environment without adverse impacts. A more realistic option for augmenting the resource is through water transfer. The Water Resources Board (Water Resources Board Great Britain 1973) devised a strategy for water resources development for England and Wales for the year 2000ad. This involved enlarging a number of reservoirs and building new ones, the conjunctive use of surface and ground water and the transfer of water from the humid north and west to the drier south and east of the country. In the main, rivers were to be used as carriers, but some aqueducts were to be constructed to move water between them. To derive the maximum efﬁciency, the use of the system was to be integrated. 14.6.2 Transfer plans and proposals The Water Resources Board (WRB) strategy included a transfer between the lower Severn and

341

the upper Thames, a new reservoir near Oxford and the enlargement of existing reservoirs at the head of the Severn and Wye. Few if any parts of the Board’s strategy were implemented, the Severn/Thames transfer being opposed on ground of cost, because of the ecological problems that were thought to result from the mixing of Severn waters with those of the Thames and for other reasons. Nevertheless in 1994 the National Rivers Authority (1994) set out the possibilities for the development of strategic resource options and transfer links nationally, including the Severn/Thames transfer (Fig. 14.7). One idea was to convey water directly to the London reservoirs from the Severn, using a 90-km-long pipeline at a capital cost of £117 million (1994 prices) for a 200 megalitres per day transfer. For the river to river option, bankside storage was foreseen alongside the Severn at the abstraction point for settlement of suspended solids and to act as a buffer against pollution incidents. Similarly storage would be provided in gravel pits adjacent to the Thames end of the transfer for blending and control. A pipeline would connect the two ends of the transfer, but there was also the possibility of using the Thames–Severn canal for part of its course. Water would be abstracted from the Severn only when the ﬂow was above a prescribed level. To augment the ﬂow of the Severn, there were proposals for enlarging certain of the reservoirs in its headwaters and redeploying some for river regulation. Brown and Askew (1993) reviewed three of the water transfer schemes which were parts of the NRA development strategy, including the linkage of the Severn and Thames. Taylor (1993) looked at the possibility of using part of the canal system operated by British Waterways to move 200 Ml/d per day to the south east from the north west. Lower capital and running costs, a shorter lead time and fewer environmental problems were considered the advantages of canals over a pipeline. Arguing that the 1995 drought highlighted the need to strike the correct balance between the availability of resources and the demands placed on them, Sheriff et al. (1996) examined the NRA water resources strategy afresh. They

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john c. rodda

Fig. 14.7 NRA proposals for transferring water from the Severn to the Thames (National Rivers Authority 1994).

focussed on the demand scenarios that would bring deﬁcits to the south and east of the country, some 629 Ml/d for the Thames and 152 Ml/d for the southern areas in the worst cases, together with the options for meeting these deﬁcits. These included developing more storage in the southeast and transferring water from the North of England or Wales. One element of the strategy was again a new reservoir in south-west Oxfordshire with or without a Severn/Thames transfer supporting it. This reservoir would cover 14 square kilometres, provide storage for about 150 million cubic metres and produce a reliable yield of 350 Ml/d (Table 14.9). Writing in the New Civil Engineer, Redfern (2006) reported a paper from Baker and Scott which was to be presented at the Institution of Civil Engineers Conference on Sustainable Water Resources. With the title ‘Don’t rule out a national water grid, warn UK engineers’ Redfern quoted costs and yields for a number of transfer schemes (Table 14.10) proposed by Baker and

Scott. These were very similar to previous proposals with some updated ﬁgures. They showed that while the operating costs for the four proposals were all in the region of £30 to £50 million per year, the capital costs ranged from £52 million for the Severn/Thames transfer and abstraction near London, to £560 million for the Oxfordshire reservoir. This suggests that transfer schemes are more cost effective than reservoir construction. It is salutary, however, that apart from the possibility of additional water supplies for Oxford and its hinterland and the proposal for Broad Oak and other reservoirs, the bulk of the water captured by these schemes is intended to augment supplies to London and to Swindon and is not for the SE Region. Implementing one or other of them would still leave a shortfall over much of the region outside the Thames Basin within the time horizon of the SE Plan. Some water transfers into these areas may be possible but artiﬁcial groundwater recharge of some aquifers within them may be a more realistic option.

343

Water Resources in South East England Table 14.9 Alternative yields and costs for Severn/Thames water transfer options (1995 prices) (from Sheriff et al.1996) Option −1

Severn/Thames (200 ML day ) Severn/Thames (400 ML day−1) South West Oxfordshire Reservoir

Potential yield (ML day−1)

Capital cost (£million)

Pumping costs (£thousand per year)

Time to promote

Supported 249 Supported 425 350

57 92 400

31.5 31.5 23.5

Medium High High

Table 14.10 Costs and yields for a Severn/Thames Transfer and Reservoir Scheme (from Redfern 2006, quoting Baker and Scott)

Scheme Augmentation of Severn by redeployment of Vyrnwy Reservoir with transfer to Thames for abstraction at Staines Regulate Severn using Craig Goch reservoir, transfer to Thames for abstraction at Staines Regulate Wye and Severn using Craig Goch reservoir with pipeline supplying water direct to London reservoirs South-west Oxfordshire reservoir Enlarge Broad Oak Reservoir

14.7 Confronting the Future 14.7.1

Conﬂicting views

In the past, with the notable exception of the concepts and plans advocated by the WRB, the national water policy promoted by central government has been river basin based. Little consideration seems to have been given to a holistic national approach. This parallels the government attitude at present to planning: each of the nine English regions has been required to produce its own development plan, but without a national plan as a basis, or in which to subsume them. Indeed this avoidance of considering the nation’s water needs as a whole seems to have been the current practice even when the National Water Council (NWC) existed from 1974 to 1988. For example, the Foreword to the 1978 Water Industry Review (National Water Council 1978) states

Capital cost (£million)

Annual operating cost (£million)

Unit cost (£ ML day−1)

Yield (ML day−1)

52

36.3

3.05

150

276

44.1

1.92

400

397

56

2.56

400

560 67

32.9 1.4

2.65 2.05

350 40

that ‘The Council has therefore not sought at this stage to adapt the review to be a formal statement of strategy or to include an itemized check list for further action.’ This was despite the existence at that time of the short-lived Central Water Planning Unit, which attempted to continue the role of the WRB. The existence of the ten ‘powerful’ regional water authorities must have been an important deterrent to the formulation of a national strategy. Since those times pressures on water resources have been mounting rapidly. Now a steep rise in population is underway. There are more than 60 million people in the UK and an increasing number of immigrants is expected. A population of 75 million is being predicted for 10–15 years hence and the pundits say that 3 million new dwellings will be needed. Well before the ﬁgure of 75 million is attained, the country’s infrastructure, especially its water services, will be vastly

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over-stretched. This will be most intense where the bulk of the increase in population will be focussed, namely in London and the SE Region. Consequently it is a very opportune moment to consider a national water policy afresh and the place of the SE Region within it. Water transfers to the south-east of the country must be the cornerstone of this new policy. The EA in its draft ‘Developing our Water Resources Strategy for England and Wales’ (Environment Agency 2007b) demonstrates very clearly (Fig. 14.8) that serious water stress is being experienced at present over the whole of south-east England. Nevertheless the Agency seems to deny that the most obvious solution and the most practical one is really necessary (Environment Agency 2007c). This is to promote water transfers. Quoting from a detailed report the EA website (www. environment-agency.gov.uk/subjects/waterres) proclaims that: We conclude that there is no new evidence of a need for large scale transfers of water to south east England from the north of England or Wales. Water companies’ existing plans provide for water supply in south east England to 2025 without the need for large scale water transfers. Such transfers are more expensive and environmentally damaging than the measures in the water companies’ water resources plans. It would be possible instead to build large pipelines to move water to south east England. The feasibility of such a scheme is not in question. It would be worth building a water grid only if: • The demand for water in south east England exceeds the available supply. • There is no better, cheaper option locally. The water companies’ estimates of future water demand in south east England allows for two million more people and 8% more water use by each person by 2030.

In this document there is an admission by the EA that beyond the 2020s water transfers may be necessary. This appears to conﬂict with the EA’s stance, because of the long lead times for such schemes. Even if plans for water transfers were to be commenced now, they would probably

become operational only at that time. Indeed the SE Plan (South East England Regional Assembly 2006) recognizes the length of time required for new schemes to make a contribution. But by focussing on the period between 2020 and 2030, neither the EA nor South East England Regional Assembly (SEERA) seem to acknowledge that if there is agreement on a strategy now which includes transfers, they will have a have a life of 100 years or more. The Council for the Protection of Rural England (CPRE) published ‘A Water Resource Strategy for South East England’ in July 2007 (Warren 2007). This extensive and well researched report, while recognizing that the region faces a very serious water problem, does not seem to offer a realistic solution. Giving more attention to the aquatic environment, changing the mandate of Ofwat, reducing leakage and improving water efﬁciency are amongst its recommendations. The paradox is easy to recognize. One part of the EA maintains that sufﬁcient resources exist in the region to meet demands until 2026. Another part of the Agency and other parties claim they are insufﬁcient and that the deﬁcit must in part be met by water transferred into the region. But circumstances are changing rapidly. In the average year now there are approximately 610 cubic metres per annum for each inhabitant of the SE Region or 1670 litres per day. This compares to a ﬁgure of 1012 litres per head per day for the Thames Basin as a whole (Cook 2006), of which the total licensed abstractions require 55% or 557 litres. When climate change is taken into account these average ﬁgures worsen. Assuming winter rainfall increases by 10% and summer rainfall decreases by 20% between now and 2026, then in the average year the region’s total rainfall will be about 660 mm. Actual evaporation can be expected to rise from the present 480 mm a year to an estimated 500 mm – probably a conservative ﬁgure. This leaves 160 mm, or 611 cubic metres per person per year, or 1030 litres per day. Adding 1 million more to the region’s population brings a further reduction to about 350 cubic metres or 960 litres under average climatic conditions. During future droughts as

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Water Resources in South East England 1. Anglian Water 2. Bournemouth and West Hampshire Water 3. Bristol Water 4. Cambridge Water 5. Essex and Suffolk Water 6. Folkestone and Dover Water 7. Mid Kent Water 8. Northumbrian Water 9. Portsmouth Water 10. Severn Trent Water 11. South East Water 12. South Staffordshire Water 13. South West Water 14. Southern Water 15. Sutton and East Surrey Water 16. Tendring Hundred Water 17. Thames Water 18. Three Valleys Water 19. United Utilities 20. Wessex Water 21. Yorkshire Water 22. Anglian Water (formerly Hartlepool Water)

8 22

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1 5

4 16

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Levels of water stress Not assessed

Low

Moderate

Serious

Fig. 14.8 Water stressed areas (Source: Environment Agency 2007b).

serious as that experienced in 2006, a further reduction of some 20% is likely in the annual rainfall total, together with an increase in the evaporation. This would reduce the available water resource even further, say by 20% to a ﬁgure of about 770 litres per head per day. There

are some additional factors which have to be taken into account. One is sea level rise and its effects on aquifers close to the coast where overpumping may contribute to salt water intrusion. Another is the Habitats Directive and the requirement to reduce the rate of abstraction from some

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sources and close others in order to comply. All these factors point to the water resources in the region coming under increasing stress in the future. Would the normal pattern of life be able to continue under these conditions? 14.7.2 Scenarios for the South East Region As 2026 approaches three scenarios are envisaged for the region (Rodda 2006): 1 Business as usual. Little or no recognition that most of the region’s water resources are already overstretched, that increasing the population of the region by 1 million will create a bigger demand for water and that there is over-reliance on improving water efﬁciency to meet it. The expected consequences will be: hosepipe bans 7 years out of 10, drought orders 5 years out of 10 and water rationing 3 years out of 10 over part of the region for part of the year. 2 Twin track approach. Recognition that population growth and climate change will impact the demand and the resource, along with small increases in the resource and in water efﬁciency and limited public awareness, participation and education initiatives. The expected consequences will be: hosepipe bans 5 years out of 10, drought orders 3 years out of 10 and water rationing 1 year out of 10 over part of the region during the summer. 3 Three lane approach. The resource will be increased by a Severn/Thames transfer, water efﬁciency and demand management will be practised effectively and a continuous, robust campaign mounted for public awareness, participation and education. The expected consequences will be: summer hosepipe bans over part of the region for 3 years out of 10.

14.8

Recommendations for Catchment Management

1 Develop, agree and enact legislation which provides a sound basis for catchment-based water resources management 2 Avoid convoluted governance structures for water resources management, by establishing clear and transparent lines of responsibility based on a small number of bodies, each with an agreed and well deﬁned role. 3 Maximize the co-operation between the institutions involved, such as by designating a lead body, preferably a department of government and by using a formal committee structure. 4 Make a clear separation between the operational and regulatory duties of these institutions with a single body being responsible for the latter. 5 Develop a basin-wide water resources plan setting out aims and objectives and how to attain them in the short, medium and long term. 6 Base this Plan and decision-making generally on reliable and representative data and information collected throughout each basin in question. 7 Ensure that the management and development of water resources is based on a full participatory approach, involving users, planners and policymakers at all levels (taken from the Dublin Principles WMO 1992). 8 Meter the water supplied to each consumer and charge for this water in a fair manner. Acknowledgements Thanks are due for their valuable help and assistance in the preparation of this chapter to: Frank Law, Terry Marsh, David Rodda OBE, Harvey Rodda, Nick Robbins, Colin Neal, Geoff Smith, Adrian Smith and Dee Galiford. The Met Ofﬁce Customer Centre and the Ofwat Library supplied some of the data used in the preparation of this chapter and their assistance is gratefully acknowledged. References

From this study of water resources in South East England, it is possible to make a number of observations and recommendations which may assist practice in other parts of the World:

Andrews, F.M. (1962) Some aspects of the hydrology of the Thames Basin. Proceedings of the Institute of Civil Engineers, 21, 55–90.

Water Resources in South East England Birks, C. (2004) The need for environmental infrastructure. Proceedings of a Conference, Planning for Water: A Major Challenge for the Government’s New Sustainable Communities. One Birdcage Walk, London, 21 October 2004. Brown, R.P.C. and Askew, T.E.A. (1993) Bulk water transfer systems. In: Proceedings of the 4th National Hydrology Symposium, British Hydrological Society, Cardiff, September 1993. Institute of Hydrology, Wallingford, Oxon., 1.93–1.99. Cook, D. (2006) Water Services in Thames Water. Proceedings, Centre for the Environment Conference, The Future for Oxfordshire’s Water Resources, Oxford University, 22 March 2006. Council of European Communities (1997) Proposal for a Council Directive establishing a Framework for Community Action in the Field of Water Policy, Brussels 26 February 2007. Department for the Environment, Food and Rural Affairs (2007a) Sustainable Development Indicators. http://www.defra.gov.uk/sustainable/government/ (last veriﬁed 10 October 2008). Department for the Environment, Food and Rural Affairs (2007b) River Pollution Survey for England. http://www.defra.gov.uk/environment/statistics/ inlwater/iwquality.htm (last veriﬁed 10 October 2008). The Economist (2007) A special report on Britain – living by their wits. 382, 3 February 2007. Environment Agency (2007a) Water Demand Management Bulletin. Issue 83, June. Environment Agency (2007b) Water for People and the Environment: developing our water resources strategy for England and Wales. Consultation document. Environment Agency, Bristol http://publications. environment-agency.gov.uk/pdf/GEHO0707BMXQe-e.pdf (last veriﬁed 8 October 2008). Environment Agency (2007c) Do We Need Large-Scale Water Transfers for South East England? http://www. environment-agency.gov.uk/subjects/waterres/9814 41/1464447/?lang=_e (last veriﬁed 8 October 2008). European Commission (1992) Council Directive 92/43/ EEC on the Conservation of Natural Habitats and of Wild Fauna and Flora: the Habitats Directive. European Commission, Brussels. Falkenmark, M. (1997) Meeting water requirements of an expanding world population. Philosophical Transactions of the Royal Society of London – Series B, 352, 929–936. Hanak, E. (2007) Finding water for growth: new sources, new tools, new challenges. Journal of the

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American Water Resources Association, 43, 1024–1036. Heather, A.I.J. and Bridgeman, J. (2007) Water industry asset management: a proposed service-performance model for investment. Water and Environment Journal, 21, 127–132. House of Lords (2006a) Water Management, Volume 1, House of Lords, Science and Technology Committee, 8th Report of Session 2005-06. The Stationary Ofﬁce Ltd., London, 160pp. House of Lords (2006b) House of Lords Science and Technology Committee (2006) Water Management, vol. 1. Report HL 191-I, 8th Report of Session 2005– 2006. The Stationery Ofﬁce Ltd, London, tables 14.5, 14.8. Howarth, D. (2004) Demand management of water in a regulatory environment. In: Cabrera, E. and Cobacho, R. (eds), Challenges of the New Water Policies for the 21st Century, Proceedings of the Seminar on Challenges of the New Water Policies for the 21st Century, Valencia, Spain, 29–31 October 2002. Balkema Publishers, Abington, pp. 141–170. Hydrological Summary (2007) Hydrological Summary for the United Kingdom, July 2007. Institute of Hydrology/British Geological Survey, Oxford. International Commission for the Protection of the Danube River (2004) The Danube Basin Analysis (WFD Roof Report 2004) Part A – basin-wide overview summary. International Commission for the Protection of the Danube River, Vienna, Austria. http://www.euvki.hu/euwfd/content (last veriﬁed 8 October 2008). Jamieson, D.G, and Nicolson, N.J. (1984) Water resources of the Thames Basin: quantitative and qualitative aspects. Proceedings of the Institute of Civil Engineers, 42, 379–391. London Plan (2004) The London Plan: a summary. Highlights from the Mayor’s Spatial Development Strategy for Greater London. Greater London Authority. National Rivers Authority (1994) Water: nature’s precious resource, an environmentally sustainable water resources strategy for England and Wales. National Rivers Authority, Bristol. National Statistics Great Britain (2006) Sustainable Development Indicators in your Pocket 2006: an update of the UK Government Strategy Indicators/ National Statistics. DEFRA, London. National Water Council (1978) Water Industry Review. National Water Council, London.

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Rawson, P.F. (2006) Cretaceous sea levels peak as the North Atlantic opens. In: Brenchley, P.J. & Rawson, P.F. (eds), The Geology of England and Wales, 2nd edn. The Geological Society, London, pp. 365–393. Redfern, B. (2006) Don’t rule out a national water grid, warn UK engineers. New Civil Engineer, 29 June 2006, p. 6. Rodda, J.C. (2006) Sustaining water resources in South East England. Atmospheric Science Letters, 7, 75–77. Rodda, J.C., Downing R. and Law, F.M., (1976) Systematic Hydrology. Newnes-Butterworth., London. Sheriff, J.D., Lawson, J.D. and Askew, T.E.A. (1996) Strategic resource development options in England and Wales. Journal of the Chartered Institute of Water and Environmental Management, 10, 160–169. South East England Regional Assembly (2005a) A Clear Vision for the South East: The South East Plan, draft for public consultation. South East England Regional Assembly, Guildford. South East England Regional Assembly (2005b) A Clear Vision for the South East: Draft South East Plan, Handed to Government. South East England Regional Assembly, Guildford. South East England Regional Assembly (2006) A Clear Vision for the South East: The South East Plan, draft for submission to Government March 2006. South East England Regional Assembly, Guildford. South East England Regional Assembly (2007) South East England Regional Assembly Website. http:// www.southeast-ra.gov.uk/ (last veriﬁed 8 October 2008). Taylor, J. (1993) Regional water transfers. In: Proceedings of the 4th National Hydrology Symposium, British

Hydrological Society, Cardiff, September 1993. Institute of Hydrology, Wallingford, Oxon., 1.101–1.106. Thames Water (2007) The Upper Thames Major Resource Development: Stage 2 Preferred Scheme and Design Options Report, vol. 1, 114. The Times (2000) The Times Concise Atlas of the World. Times Books, London. United Nations Educational, Scientiﬁc and Cultural Organization (2006) Water: a Shared Responsibility, United Nations World Water Development Report 2. World Water Assessment Programme, UNESCO/ Berghann Books, Paris/New York. Warren, G. (2007) A Water Resources Strategy for the South East of England. Campaign to Protect Rural England South East Report. CPRE, Kent. http://www. cprese.org.uk/campaigns/water/water_strategy_for_ the_southeast.htm (last veriﬁed 20 November 2008). Water Resources Board Great Britain (1973) Water Resources in England and Wales, Water Resources Board Publication no. 22–23, HMSO, London. Wheater, H.S., Neal, C. & Peach, D. (2006) Hydroecological functioning of the Pang and Lambourne catchments, UK: an introduction to the special issue. Journal of Hydrology, 330, 1–9. World Meteorological Organization (1992) International Conference on Water and the Environment: development issues for the 21st century, Dublin, Ireland, 26–31 January 1992. Keynote Papers. ICWE Secretariat, Geneva. World Meteorological Organization/United Nations Educational, Scientiﬁc and Cultural Organization (1997) Water Resources Assessment: handbook for review of national capabilities. UNESCO and the WMO Hydrology and Water Resources Department, Geneva, Switzerland.

Image facing chapter title page: Courtesy of John Rodda.

15

Managing the Catchments of the Great Barrier Reef

J A N E W A T E R H O U S E 1, M I K E G R U N D Y 2, I A I N G O R D O N 3, J O N B R O D I E 4, R A C H E L E B E R H A R D 5 A N D HUGH YORKSTON6 1

CSIRO Water for a Healthy Country Flagship and Reef Water Quality Partnership, Aitkenvale, Queensland, Australia 2 CSIRO Land and Water, St Lucia, Queensland, Australia 3 CSIRO Sustainable Ecosystems, Aitkenvale, Queensland, Australia 4 Australian Centre for Tropical Freshwater Research, James Cook University, Douglas, Queensland, Australia and CSIRO Water for a Healthy Country Flagship, Aitkenvale, Queensland, Australia 5 Eberhard Consulting Pty Ltd, Dutton Park, Queensland, Australia 6 Great Barrier Reef Marine Park Authority, Townsville, Queensland, Australia

15.1

Introduction

The Great Barrier Reef (GBR), Australia, is a remarkable structure – both for its abundant biodiversity and its extent. It fringes the north-east Australian coast and the state of Queensland for approximately 2000 km and comprises over 3200 coral reefs embedded in an ecosystem that includes mangrove forests, coastal wetlands and estuaries, seagrass meadows, deep shoals, continental shelf margin and slope. It is the world’s largest World Heritage Area. Like many reefs around the world, the GBR is under stress from three main inﬂuences: overharvesting of resources, climate change and terrestrial runoff of contaminants (Pandolﬁ et al. 2003). The GBR was declared a Marine Park in 1975. It is managed as a zoned, multi-use area with over 30% of the area of all of its 70 bioregions protected in high conservation, no take zones incorporated into a Park-wide zoning plan (Day et al. 2004). Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

The catchments which adjoin the GBR are equally remarkable in many ways and astonishing in their diversity, but have unfortunately received much less protection. Through runoff of contaminants though, the health of the GBR is intimately connected to the environmental management of these catchments (Haynes et al. 2007), especially with the projected impacts of climate change (Johnson and Marshall 2007). This chapter identiﬁes and examines the challenges to improving management of these catchments for their long-term protection and maintenance of GBR resilience. These challenges range from understanding the complex biophysical interactions between land and GBR ecosystems through to assessment of the economic and social realities, and the costs and beneﬁts of the change needed for better protection. A substantial start has been made with a Reef Water Quality Protection Plan (the Reef Plan – The State of Queensland and Commonwealth of Australia 2003) and should be accelerated through new Australian Government initiatives including a Reef Rescue Program introduced by the incoming Government in late 2007. Together these two plans will provide a

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framework and resources for supporting community-based regional Natural Resource Management (NRM) in target setting and investment in management change. New Australian and State (Queensland) Government investment is focussed on these initiatives and on the delivery of supporting science. The response is in essence a complex and broad scale adaptive management exercise. Much of this chapter explores this experience so far. 15.2 Nature of the System The catchments abutting the GBR are made up of 35 major river catchments (Furnas 2003), which include two of the largest river catchments on Australia’s eastern coast – the Fitzroy and Burdekin catchments (Fig. 15.1). Collectively, the mainland drainage basins have an area close to 424,000 km2 (Table 15.1) of which 68% of the total catchment area and 23% of total freshwater runoff is from the largest two basins. The rivers of the Wet Tropics in far north Queensland occupy only 5% of the GBR catchment land area but supply about 27% of the ﬂow (Furnas 2003). The climate ranges from humid tropical (Wet Tropics and Mackay-Whitsunday catchments) to semi-arid with signiﬁcant ranges in variability. Variability is high by Australian and world standards, especially in the drier catchments (the Burdekin catchments have a range of variability index from 0.75 to 1.25; the least variable regions are in the south and north with values from 0.5 to 0.75*). In the wetter catchments, there can be extreme short range gradients – the rainfall from Innisfail on the north-east coast to Mt Garnet some 98 km inland declines by 35 mm per kilometre, the sharpest rainfall gradient in Australia (Gentilli 1971). The topography of the northern catchments and the Mackay area is dominated by high nearcoastal ranges; the orographic effect of these ranges produces the ‘Wet Tropics’ of the GBR * The rainfall variability index used by the Australian Bureau of Meteorology is the ratio of difference between the 90 and 10 percentiles over the 50 percentile.

catchments. In the drier central and southern catchments, especially the Burdekin and Fitzroy catchments, there is subdued relief and the east– west catchment dividing range extends substantially to the west. The interactions between soils, topography and land use lead to substantial differences in the water quality coming from wet-catchment rivers (rainfall generally > 1500 mm) compared with dry-catchment rivers (rainfall generally < 1500 mm). In most cases, higher sediment and nutrient concentrations are found in rivers draining dry catchments (Furnas 2003) but the wetter catchments often have a higher proportion of bioavailable nutrients (Brodie and Mitchell 2005). Approximately 8% of sediment transport but up to 17% of the nitrogen and phosphorus load are estimated to emerge from these Wet Tropics catchments. Here, the higher rainfall supports more intensive agricultural activities with intensive fertilizer application. By contrast, the larger dry catchments are characterized by erosive landscapes and the generation of substantial loads of sediments (Brodie and Mitchell 2006). The largest agricultural land use by area in the catchments is beef cattle grazing (Table 15.1). Most of this land use is dominated by cattle husbandry; however, cattle husbandry in the drier areas is increasingly integrated with cattle fattening in either the wetter areas or where cropping produces grain for feedlots. High-rainfall areas are dominated by sugar cane and high-value horticultural crops while coastal irrigation areas are characterized by sugar cane and a smaller range of horticultural tree crops. Inland portions of the Fitzroy catchment support large areas of summer and winter rain-fed grain crops and are an important irrigation area for cotton and horticultural crops. Whilst mining is a major industry in the catchments, it occupies a relatively small area (Table 15.1). However mining operations can be a signiﬁcant local water user and old unrehabilitated sites continue to be a source of leachates including heavy metals and acidic waters. The total human population of GBR catchments is around 1 million, approximately 20% of Queensland’s population. Low densities in the

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Fig. 15.1 Drainage basins discharging into the GBR and the associated community-based NRM regions. Topography of the GBR catchments and bathymetry of the adjacent seaﬂoor adapted from a Geoscience Australia image.

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Table 15.1 Land use summary statistics for Great Barrier Reef catchments: 1999. (Data supplied by Queensland Department of Natural Resources and Water) Land use class Conservation and natural environments (without defence areas) Defence Forestry Grazing natural vegetation Sugar Other cropping Horticulture and tree crops Intensive animal production (mostly dairy) Urban and rural residential Mining Water and wetland Other (transport, utilities, etc.) Total

Area (ha)

% Area

5,589,891

13.3

428,217 2,155,925 31,426,854 568,970 1,165,062 63,041 103,350

1.0 5.1 74.0 1.3 2.7 0.1 0.2

261,417 74,826 618,236 11,908 42,467,697

0.6 0.2 1.5 0.03

order of 2.2 persons per square kilometre apply over much of the area but the urban centres of Townsville, Cairns, Mackay, Rockhampton, Gladstone and Bundaberg and adjacent coastal areas are growing rapidly and absorbing much of the 1.2% average population increase for the region (Gilbert and Brodie 2001; Hug and Larson 2006). There has also been signiﬁcant population growth around hinterland mining developments. Industries in the GBR catchments and activities within the GBR generate a signiﬁcant portion of Queensland’s regional economy (Productivity Commission 2003). In 2003 the gross value of production for mining, tourism and agricultural industries in the GBR region (the catchments and the GBR) was over AUD$14 billion (Fig. 15.2). In addition, the GBR region contributed 62% of Queensland’s exports from ports. Tourism has by far the largest growth rate for any industry and was predicted to nearly double in value within 20

Fig. 15.2 Projected gross value of production in the GBR catchment and lagoon (Productivity Commission 2003). (Source: ABARE projections.)

Managing the Catchments of the Great Barrier Reef years. The total (direct plus indirect) economic contribution of tourism, commercial ﬁshing, and cultural and recreational activity in the GBR and its catchments to the Queensland economy is AUD$5.4 billion per annum (gross product) and employs about 56,000 persons (Access Economics 2007; 2005–06 estimates). Tourism is the major contributor to these values (AUD$3.8 billion and 46,000 people). Indirect ecosystem services provided by the GBR include shoreline protection, maintenance of biological diversity, waste assimilation and reception, visual amenity, and lifestyle values, existence and bequest values. Land use is projected to remain relatively stable in the region (Marsden Jacob Associates 2008). In addition, the gross value of production of the major agricultural industries in total (sugar, beef and horticulture) is projected to remain substantial, with each industry growing by 1% or 2% per annum over the next 13 years.

15.3 The Catchments and the Great Barrier Reef The development of the GBR catchments since European settlement has resulted in water quality changes into GBR waters from increasing sediment, nutrient, pesticides and other contaminant loss from the land, and signiﬁcant alterations to the hydrodynamic regime of the ﬂoodplain (freshwater, estuarine and marine). The primary land uses driving water quality change in the GBR catchments are grazing, intensive cropping including sugarcane and horticultural crops, aquaculture and urban settlements (Gilbert and Brodie 2001). The change of extensive to more intensive uses (e.g. more intensive cattle production, or moving from cattle production to sugar cane) can be a major trigger of changes in loads impacting on water quality (Mitchell et al. 2001). The potential impacts of declining water quality on ecosystems in the GBR lagoon have been synthesized and reviewed in recent years (e.g. Haynes et al. 2001, 2007; Great Barrier Reef Protection Interdepartmental Committee Science Panel 2003; Hutchings et al. 2005; Brodie et al.

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2009a, 2009b); present knowledge is summarized below. Separate strands of evidence indicate that there has been at least a sixfold increase in delivery of sediments and nutrients from the GBR catchments to the GBR since European settlement (Brodie, et al. 2003). Signals of these increases in the GBR ecosystem have been detected in coral records through the analysis of coral cores. For example, the ratio of barium to calcium in corals offshore from the Burdekin River indicated a ﬁve- to tenfold increase in suspended sediment loads following European settlement of the Burdekin catchment (∼1860). This has been interpreted as indicating a large increase in erosion and delivery of suspended sediment to the mouth of the Burdekin River where the barium adsorbed onto the sediment desorbs and is taken up by the coral. The barium replaces calcium in the coral structure, resulting in an altered ratio of Ba/Ca indicative of a land-based inﬂuence (McCulloch et al. 2003; Lewis et al. 2007). Other metals including yttrium and manganese which are used as indicators of erosion and land settlement also show changed concentration in coral cores after 1860 (Lewis et al. 2007). The introduction of nitrogen fertilizer use in sugar cane from 1950 is evident in cores sampled in the Mackay-Whitsunday catchments (Marion 2007). Long-term time series from water quality monitoring in GBR lagoon waters also show increasing trends (Furnas et al. 2005; Brodie et al. 2007a). Agricultural land use delivers most of these loads to the GBR; sewage discharges contribute less than 3% to the overall nutrient load to the GBR lagoon (Brodie 1997). Increasing amounts of pesticides are also delivered to the GBR, especially in high river ﬂow conditions (Prange et al. 2007). Water discharged from rivers in ﬂood extends, as a plume, over the inshore areas of the GBR lagoon, and may reach mid- and outer-shelf reefs depending on the weather conditions (Devlin et al. 2001). The concentrations of contaminants, such as dissolved inorganic nitrogen, in these river plumes are typically 10–50 times the ambient concentrations; they may in the long

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term be producing a ‘pollutant enriched’ lagoon (Shaw and Müller 2005; Brodie et al. 2007a). The risk to marine organisms in the GBR from reduced water quality and the apparent biodiversity loss of inshore reefs adjacent to catchments with intensive agriculture has been described in recent years (see Haynes et al. 2001; Pandolﬁ et al. 2003; Fabricius 2005; DeVantier et al. 2006; Bruno and Selig 2007). For example, a correlation between water quality parameters and reef condition is evident throughout water quality gradients in the Whitsunday Island Group (Fabricius and De’ath 2004; Fabricius et al. 2005); observed changes include variations in the cover, composition and relative abundance of macroalgae, hard Box 15.1

corals and soft corals, the recruitment of young hard corals and the abundance of coral bioeroders. In addition, several studies have demonstrated the risks of increased nutrients such as increased frequency of crown-of-thorns starﬁsh outbreaks (Brodie et al. 2005); and laboratory studies have shown the high toxicity of several commonly used pesticides in the GBR catchments on marine organisms (e.g. Haynes et al. 2000; Negri et al. 2005; Markey et al. 2007). Residues of these pesticides (particularly the herbicides diuron and atrazine) are now ubiquitous in GBR lagoon waters adjacent to catchments with signiﬁcant pesticide use (Haynes et al. 2000; Shaw and Mueller 2005; Rohde et al. 2006; Prange et al. 2007).

Crown-of-thorns starﬁsh outbreaks

Coral-eating crown-of-thorns starﬁsh (Acanthaster planci) have caused widespread damage to many coral reefs in the Indo-Paciﬁc over the past four decades (Birkeland and Lucas 1990). On the Great Barrier Reef (GBR) a third observed outbreak cycle is now in progress. The cycles have occurred from 1962 to 1976, 1979 to 1991 and 1993 to present. Outbreaks may have occurred before 1962 but were not observed or reported. Each cycle appears to have begun in areas adjacent to the Wet Tropics region in the northern GBR. The impact of outbreaks on the GBR is a major concern to the multi-billion dollar tourism industry. Over a number of years, there was an outbreak on reefs between Cairns and the Whitsundays which was estimated to cost tourism operators, and the Queensland and Australian Governments about $3 million a year for control measures. The cause (or causes) of the outbreaks has been a controversial issue and there are two opposing groups of views as to the origin of the outbreaks. The ﬁrst one, initially proposed by Vine (1973) and partially by Potts (1981), postulates that population outbreaks are a ‘natural phenomenon’ due to natural ﬂuctuations in temperature, salinity or food for the planktonic larvae which could all increase larval survival. However, other evidence including feeding scars on massive corals indicates increased frequency of outbreaks over time. The second view states that outbreaks are due to anthropogenic changes to the environment of the starﬁsh. There are various possible anthropogenic causal hypotheses, however the two primary causes are believed to be: reduced predation on juvenile and adult starﬁsh from ﬁshing (removal of ﬁsh and gastropods) (Endean 1977; Sweatman 2008); and larval (phyto-plankon) food supply enhancement from nutrient enriched terrestrial runoff (Birkeland 1982; Lucas 1982; Brodie 1992). Current evidence provides increased support for terrestrial runoff (Brodie et al. 2005) triggering outbreaks, though recent monitoring of the re-zoning of the Great Barrier Reef Marine Park is highlighting that protecting predator populations can still have a signiﬁcant effect in reducing the populations of starﬁsh following outbreaks (Sweatman, 2008). Nutrient discharges from rivers have increased at least fourfold in the central GBR over the last century, and concentrations of phyto-plankton (>2 µm) off the inshore central GBR shelf in the wet season when A. planci larvae develop is double that at other places and times. In nutrient-enriched conditions phyto-plankton populations are also believed to shift from smaller species which are less desirable to crown-of-thorns starﬁsh larvae as food, to larger phyto-plankton which the starﬁsh larvae will favour. Larval development, growth and survival increase almost tenfold with doubled

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concentrations of large phyto-plankton. This and other lines of evidence suggest that frequent A. planci outbreaks on the GBR may be a result of increased nutrient delivery from the land. There are few options to manage outbreaks of crown-of-thorns starﬁsh once they have started, apart from having in place a comprehensive regime that maintains healthy and resilient reef ecosystems and their full suite of predatory species. Several techniques for speciﬁc site controls have been developed to reduce starﬁsh numbers through direct interventions, however they are labour-intensive and expensive, and are only practical in small areas. Because starﬁsh can quickly move from one area to another, control of a speciﬁc area must be an ongoing effort and may be required on a daily basis. The recommended control method involves trained divers injecting sodium bisulfate solution (dry acid which is non-toxic to other marine life) into the starﬁsh, which kills them within a few days. During active outbreaks, operators may need to inject 200–500 starﬁsh every day in an effort to keep selected sites free of starﬁsh (http://www.reef.crc.org.au/discover/ plantsanimals/cots/cotscontrol.html). Scientists will continue to search for the cause or causes of crown-of-thorns starﬁsh outbreak, monitor crown-of-thorns populations, investigate links between terrestrial runoff and crown-ofthorns starﬁsh outbreaks and investigate more cost-effective methods to control crown-of-thorns starﬁsh. However, as a ‘no regrets’ action, management to reduce terrestrial nutrient runoff is obviously an attractive option as it ﬁts with other imperatives to reduce nutrient pollution of the GBR. Photographs courtesy of the Great Barrier Reef Marine Park Authority for and on behalf of the Commonwealth of Australia.

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The catchment and the marine ecosystems of the GBR are interdependent. The GBR ecosystem has a biophysical dependence on the catchments; the catchments derive social and economic beneﬁts from the GBR. Nonetheless there are important distinctions. The health of the catchments seems to be largely inﬂuenced by ambient water quality conditions among other factors while the health of coastal waters is mostly inﬂuenced by the discharge from large river ﬂow events (Brodie et al. 2007b). However, as water quality monitoring has been driven by concerns about the health of the GBR, monitoring effort has been focused at the ends of rivers rather than catchment water quality or instream health. Consequently, we now know a substantial amount about end-ofriver contaminant discharges (Prange et al. 2007), loads during ﬂoods (Furnas 2003) and some implications for the health of the GBR (Fabricius et al. 2005; De Vantier et al. 2006), but we know much less about water quality and its effects on ecosystems within catchments. The GBR catchments support high biodiversity and many endemic species of freshwater ﬁsh (Pusey et al. 2004), some of the highest diversity of freshwater invertebrates in the world (Pearson et al. 1986; Vinson and Hawkins 2003; Pearson 2005) and many species of aquatic plants (Mackay et al. 2007). However, various estimates of loss of freshwater wetlands in developed catchments along the GBR coast range between 70% and 90% (Environmental Protection Agency 1999) while the condition of the remaining 10–30% range from moderate to no value as ﬁsheries resources (Veitch and Sawynok 2005). The most signiﬁcant reason for the reduction in the value of remaining wetlands to ﬁsheries is changed catchment hydrology resulting in loss of connectivity, habitat modiﬁcation, poor water quality and poor habitat quality. Wetland loss affects species whose lifecycle includes a marine phase (Veitch and Sawynok 2005). While the diversity of information on catchment freshwater health is increasing, much work is unpublished. Recent reviews on within-catchment water quality and ecosystem health have focussed on the wet tropics (e.g. Faithful et al.

2006; Connolly et al. 2007, 2008; Pearson and Stork 2008); there has been little work on the dry tropics, highlighting an important information gap. There are limitations in the current understanding of the transport, fate and impact of landbased materials from the catchment to the marine receiving environment. This capacity is critical to enable the development of meaningful water quality performance measures (targets), to facilitate the evaluation of the Reef Plan and Reef Rescue Program actions and regional water quality activities in the GBR catchments, and to guide future investment, such as the Reef Rescue Program, into areas of high risk. Whilst present knowledge enables a quantitative assessment of the relationship between the catchment and the GBR ecosystems, there is signiﬁcant uncertainty about the impact of land management interventions on the end-of-river contaminant loads, and even less conﬁdence in the relationship between loads and GBR ecosystem health. The complexity and variable nature of the system highlights the need for a co-operative management approach to address water quality issues for the GBR (Gordon 2007). The range of land and water management systems in place to tackle this challenge is described below. 15.4 Responding to the Water Quality Challenge 15.4.1 The plans and actions The evidence for the relationship between land use, water quality and declining GBR ecosystem health (described in Section 15.3) led to a national policy response in 2003 with Australian and State Government endorsement based around the Reef Plan. The Reef Plan built on existing Government policies, industry and community initiatives with a list of strategies and actions to be implemented by government, industry and community groups. Its goal is to halt and reverse the decline in water quality entering the GBR by 2013 (i.e. within 10 years); both for the overall protection of the ecological health of the GBR as well as for

Managing the Catchments of the Great Barrier Reef the health of its adjacent catchments and waterways. There is a focus on incentives, partnerships and enhanced planning processes to support the adoption of improved industry best management practices, essentially relying on voluntary approaches. The Plan has provided a substantial challenge for the delivery of widespread changes in land use practices and community attitudes through integrated government and community action. However, in the ﬁrst 5 years of implementation, the Reef Plan lacked a dedicated funding source. More recently the Australian Government has announced a AUD$200 million Reef Rescue Program (2008–2013) with the majority of funds (AUD$146 million) targeted at supporting onground actions to improve water quality. The other components of the Reef Rescue Plan are focused on improved catchment monitoring of water quality and management actions, targeted research and development to identify the most effective management actions, improved partnerships to engage the key stakeholders and development of a Reef Water Quality Report Card. The Reef Rescue Program also provides dedicated funding for better engagement of Traditional Owners in actions to support GBR protection and management. The focus for Reef Plan implementation has been on the development of Water Quality Improvement Plans in a number of priority GBR catchments and the associated development of speciﬁc Great Barrier Reef Water Quality Guidelines (Great Barrier Reef Marine Park Authority 2008) aimed at maintaining the health of key GBR ecosystems. These plans require the development of management objectives and targets for management practices, water quality and marine ecosystem health, and are facilitated by regional NRM bodies. 15.4.2 Priorities for intervention A priority for recent research has been identiﬁcation of the primary sources of terrestrial contaminant runoff to the GBR through more detailed regional assessments. For example, in most of the

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Burdekin catchment where grazing is the predominant land use, sediment is identiﬁed as the contaminant of greatest concern, whilst in the Tully catchment within the Wet Tropics, where intensive sugar cane and banana cropping is dominant, nitrogen and pesticides are the key concern. In these areas, management techniques have reduced sediment generation, although there remain relatively minor sediment issues in horticultural crops such as bananas. The focus for sediment control and reduction of erosion associated with rangeland grazing is in the Burdekin and Fitzroy catchments (Fig. 15.1) and to a lesser extent catchments in the Cape York and Burnett regions, and upper parts of the Wet Tropics catchments (Brodie et al. 2003; Furnas 2003; Cogle et al. 2006). Hillslope, streambank and gully erosion dominate sediment delivery processes, although further studies are required to demonstrate predominant erosion mechanisms in the catchments. Particulate nutrients are also sourced from soil erosion in grazing lands and can therefore be managed through soil erosion control. Dissolved inorganic nutrients are largely associated with fertilizer application in intensive cropping industries. Management of the majority of nutrient losses is essentially about managing sugar cane and to a lesser extent, horticulture, appropriately. For example, in the Tully catchment (a Wet Tropics catchment) where sugar cane production makes up only 13% of the catchment land use, 76% of the dissolved inorganic nitrogen discharged from the Tully River comes from sugar fertilizer losses; whilst some 85% is generated from sugar and bananas combined (Armour et al. 2007). Herbicides are mostly derived from sugar cane applications (Rohde et al. 2006) with some contributions from grazing, forestry and cotton production in the Fitzroy and other catchments (Packett et al. 2005; Prange et al 2007). 15.4.3 Management systems In the historical agricultural development of the GBR catchments, the evolution of land manage-

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ment practices was based around the resolution of the limitations posed by climate, soils, vegetation and distance against the opportunities available in local or export agricultural markets. So the early development of land management practices had steps such as: • Change from Bos taurus to Bos indicus cattle breeds with better adaptation to the climate, pest and disease loads, and increases in infrastructure such as watering points and fencing. • Breeding and selection programmes to adapt agricultural systems to the region’s limitations. • Development of irrigation systems and highvalue crops or the use of supplementary irrigation to increase reliability. • Improved pest control regimes. • Innovative work on soil chemistry and fertilizer practices (especially for the low activity soils of the tropics). The ﬁrst widespread interest in conservative land management practices arose from concerns about soil erosion on cropping lands with, somewhat later, concerns about pasture degradation and land cover deterioration on grazing lands (e.g. see McKeon et al. (2004) for discussion of the cycle of degradation and recovery in grazing systems). The immediate concerns around costs and prices then had to incorporate a longer-term view around the sustainability of the enterprise and the farm’s soil and vegetation resources. Nevertheless, adoption of soil and land conservation measures into management practices has had mixed success. Signiﬁcant and widespread change in management practice has been observed where a range of drivers concentrate. An impressive example of such adoption has been the widespread replacement of the annual burning of sugar cane before harvest with harvesting of green cane and subsequent application of the cane trash as a ‘trash blanket’, a form of stubble retention. This leaves a deep cover on the soil for all of the 4 years of ratoon crops (the new cane which grows from the stubble left behind after harvesting), and has become the dominant practice in many regions (Wrigley 2006). The adoption met many of the grower’s needs for productivity outcomes in addition to delivering

a substantial soil conservation impact (Rayment 2003). A broader interest in environmental outcomes came with the introduction of speciﬁc environmental legislation raising the issue of ‘environmental duty of care’ and community recognition of the importance of good environmental management, manifested by the growth of Landcare and related programs. This then led to an interest in ‘best management practice’ or BMP (Measham et al. 2007). The discussion by Measham et al. of BMPs illustrates both the ambition for the concept and the extraordinary complexity of the outcomes which are desired from their application. Many industries have advocated a version of BMP as a means to provide a template for effective multiple outcomes. For example, the sugarcane industry in Queensland has embraced a concept of Farm Management Systems (FMS) which incorporates ‘best practice’: ‘ … the FMS support tools will be a set of easy to use voluntary support tools which sugarcane growers can use to improve cane farming practices and farm profitability and to address the industries environmental duty of care …’. (Queensland Farmers Federation 2005). An in-depth overview of current management practices applied by the key sectors in the GBR catchment is provided in the research report Industries, Land Use and Water Quality in the GBR Catchment (Productivity Commission 2003). 15.4.4 Applied management practices Effective methods for managing sediment, nutrient and pesticide generation are already known and being implemented variably across the GBR catchments. Examples of methods designed for targeting speciﬁc problems are provided below. There are several schemes available for managing fertilizer application in the intensive cropping industries. Examples include ‘6 Easy Steps’ (Schroeder et al. 2005) where cane farmers are encouraged to follow a series of steps that tailor the fertilizer application rate to the plant and soil requirements, and has beneﬁts for productivity such as reduced fertilizer application. Thorburn

Managing the Catchments of the Great Barrier Reef et al. (2003) developed the N-replacement system and this system has been trialled in several sugar areas within the GBR catchments (Thorburn et al. 2007; Webster et al. 2008). Field trials have been positive, suggesting the assumptions behind the system are often valid. The main assumption is that soil nitrogen stores can buffer the difference between the amount of nitrogen needed by the crop and the amount of nitrogen fertilizer applied. For example, if the yield of the coming crop was larger than that of the previous crop, additional nitrogen requirements would be supplied from soil nitrogen stores. Conversely these nitrogen stores would be ‘topped up’ when a small crop followed a large one. This assumption means the concept of a ‘target yield’, e.g. used in programmes such as 6 Easy Steps, is no longer necessary in determining fertilizer rates. Target yields are generally related to possible production, not actual production, and so can be a signiﬁcant driver of high fertilizer application rates (relative to actual production) and high fertilizer and high fertilizer surpluses (Beaudoin et al. 2005). The success of the N-replacement system is a potential saving in nitrogen fertilizer applications of up to 40 %, and a reduction in the overall nitrogen surplus across the whole sugarcane industry of up to 60 %. The N-replacement research is currently in the ‘proof-of-concept’ phase and plans are underway for developing the concept into a practical management system. Sediment control in the grazing industry is guided by an industry-led initiative, Grazing Land Management, or ‘GLM’. This initiative has developed regionally speciﬁc BMPs. Sediment control in these areas requires increased vegetation cover, as well as improved pasture condition and soil health to retain water, sediments and nutrients on the land (Neldner 2006; Gordon 2007). In principle this means reductions in utilization rates of vegetation through reductions in stocking rate (particularly in regard to rainfall variability), wet season spelling to improve pasture condition, forage budgeting to ensure cover levels are adequate from year to year and preventing selective overgrazing of preferred areas in the landscape (Chilcott et al. 2003;

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Gordon and Nelson 2007). However, recent unpublished work by Bartley et al. (2007a) suggests that the majority of the sediments ﬂowing into the creeks and rivers comes from streambank and gully erosion which will sometimes need engineering solutions like contour banks or ripping (as opposed to retaining walls or sediment traps) and fencing riparian and gullied areas to provide reduced grazing pressure, rather than changes in grazing land management. Current practices largely address hillslope erosion and further work is required on management and restoration techniques for gully erosion. There are also other means of addressing sediment runoff in grazing lands. Maintaining soil health, for example through reduced stocking pressure, is also identiﬁed as an important contribution to improving soil inﬁltration, and therefore reducing surface water runoff and sediment loss (Dawes-Gromadzki 2005). Herbicide management is focussed on better and more effective delivery techniques which reduce losses and integrated pest management programmes focussed on reducing use. Current herbicide practices include zonal application and the replacement of residual herbicides, for example diuron, by other herbicides such as glyphosate that have shorter environmental half-lives. The response of the water quality system to these land management changes is signiﬁcantly inﬂuenced by system dynamics, depending on the contaminant and catchment systems. The major inﬂuences are summarized below: • Sediment control mechanisms and targets in large catchments like the Burdekin are likely to encounter long lag times to impact at the end of the system, depending on soil and ﬂow characteristics, and the time for riparian vegetation to grow. However, ﬁne sediments such as clays, which present the highest risk to GBR ecosystems, experience the least system lags in transport. In addition, the variability in the systems in terms of hydrology (decadal events) and climate mean that responses in the system are likely to be in decadal time-scales (Bartley et al. 2007b; Lewis et al. 2007).

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• Fertilizer management in intensively cropped areas such as the Wet Tropics and the MackayWhitsunday Region will also experience signiﬁcant lag times in system response because of the long sugar crop cycles (6–7 years) and storage in soils and groundwater. Responses are likely to be multiple years, i.e. 5–10 years. Lags in groundwater transport and sub-surface transport (Armour et al. 2006; Rasiah et al. 2007) and ﬂoodplain trapping (McJannet 2007; Wallace et al. 2007) also have a signiﬁcant inﬂuence on the system. • Herbicide management is expected to be characterized by limited time lags because most of the herbicides of concern have half-lives of less than 1 year (e.g. 50–160 days). Signiﬁcant reductions in loads are expected to be evident within 2 years of practice change involving reductions in use and loss. Changes in the presence of herbicides due to improved practices are also easier to identify in the system as they are not present naturally, generating a clearer signal related to practice change. A major limitation in detecting improvements in practices and measurable outcomes in GBR ecosystem health is the ability to detect the signal of change in the system. Noise in the signal is due to system variability, natural occurrence of sediments and nutrients in the system and limitations of the capacity to monitor and model material transport and fate. A good example of demonstrated time lags in system response to management changes is recorded in the Tully catchment. The Tully catchment is the least variable river in the GBR catchments, and yet very large changes in fertilizer use (increases) took 14 years to be robustly manifest as increasing nitrogen levels in the lower Tully River (Mitchell et al. 2001, 2006). This highlights the importance of the need for innovative monitoring and modelling techniques, and an improved understanding of the system dynamics to inform management decisions. 15.4.5 Land management systems at the enterprise level The catchment management challenge in the GBR catchments is essentially to reduce diffuse

pollution that is an accumulation of outputs from many individually managed farm enterprises over a large area. There are opportunities for intervention to address accumulated pollution but the primary accent has to be on change at the enterprise level. Farm sizes vary across industries and regions. For example, the average farm size in the Queensland cane industry is 77 hectares but farms range from 30 to more than 250 hectares; average yield is 98.9 t ha−1 of cane (http://www.canegrowers.com.au/informationcentre/about-the-industry/index.aspx). By contrast, grazing properties are substantially larger with an average size of 30,000 hectares (Productivity Commission 2003). These differences have signiﬁcant implications for the adoption and costs of environmentally sustainable management systems. Agricultural and more generally rural management of the GBR catchments is in the hands of private agents – traditionally farm families but in recent times there are many cases of larger farm businesses with local managers. These landholders manage either freehold farms or, in the case of much of the grazing lands, land leased from the State Government. While in the latter case there are constraints in terms of deﬁned land management expectations, landholders occupying freehold lands are generally free agents in choosing the land management systems used in agricultural production. However, there are signiﬁcant indirect constraints. Arguably, the most powerful is economic – the realities of markets, net returns and access – but there are also signiﬁcant social and biophysical constraints. Understanding of, and then adoption of, environmentally conservative land management practices is therefore complex. Pannell et al. (2006) demonstrates that adoption depends on the relevance to the individual’s economic, social and environmental goals. In the case where the aim is protection of an environmental asset such as the GBR which is removed in time, space and observable connection, then the inﬂuence on land managers for adoption or GBR protection measures is considerably diluted. Increasingly society in general has an interest in how these farmed lands are managed but has

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Fig. 15.3 An illustration of the inﬂuences on management practice. Better off-site outcomes would be achieved by a move to the right, however the interaction of vectors driving the development of the management practice may result in only minor change towards a better basin water quality outcome.

poorly deﬁned and variable pathways to inﬂuence management change (Allen et al. 2001). A key concept in GBR land management has been the use of industry BMPs. Here, an interesting difference in the desire to inﬂuence management practice, compared with other regions, is the apparently clear focus on the speciﬁc outcomes aimed at better water quality outcomes because of potential detrimental impacts on the signiﬁcant environmental, economic and social values in the GBR lagoon and the broader community’s perception of such an outcome. If BMPs are to be the vehicle for delivering sustainable land management with improved outcomes for water quality off-farm, then the development of these practices should optimize for these outcomes. However, these templates have multiple drivers; there is a tension between drivers of change in various directions – and some clear inertia against change at all. Figure 15.3 illustrates desired change for improving GBR outcomes as progression along a horizontal line. Other inﬂuences on land management choice are represented as vectors with variable impact on the GBR improvement direction. The relative impacts of these may change substantially; for example Roberts et al. (2008) illustrate the range of factors producing an historical improvement in the terms of trade for Australian agriculture. Rising food prices and the response to climate change are likely to inﬂuence land management choice in the GBR catchments over the next decade. Assessments that incorporate cost–beneﬁt and water quality outcomes of particular manage-

ment practices can assist in prioritizing the adoption of practices and predicting outcomes of a set of management actions. For example, Roebeling et al. (2007) analysed the water quality efﬁcacy and the economic dynamics of management practices in the Tully-Murray catchment. With soil, management, genetic and environmental factors interacting, this entailed some 1152 combinations for sugar cane alone. Based on the proposition that regional (i.e. private + social) beneﬁts are maximized where marginal private costs equal marginal social beneﬁts, the studies showed that the cost-structure around BMPs is such that a certain level of improvement in water quality can be made at no cost (e.g. for sugar – a 25–40% gain depending on adoption of new fertilizer practices) but beyond that point, costs for water quality improvement rise sharply. The current BMPs don’t yet balance both production and environmental goals. In addition, the spatial arrangement of industries in the catchment was not optimal for improved water quality outcomes. Subsidies, incentives and/or regulations will be needed to provide the business case for industries to develop and implement improved management practice settings and to guide spatial change. Consequently, it is clear that the perceived inﬂuence from GBR drivers to an individual landholder and the individual’s personal impact on these outcomes will not themselves drive the required changes. Moreover, there are still strong economic and social disincentives against sufﬁcient change. Consistent and effective change

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will depend on institutional consistency in support of the Reef Plan and Reef Rescue Plan to deliver targeted incentives. 15.4.6 Institutional complexities A fundamental institutional challenge in addressing the management of water quality within the lagoon of the GBR is to connect the management of the terrestrial and marine systems (Gordon 2007). State government agencies, local government, regional NRM bodies and individual land managers are responsible for the terrestrial components of the catchments, whilst the Great Barrier Reef Marine Park Authority is largely responsible for management of the Marine Park. The State Government has approached the management of water quality with regard to point source and non-point source pollution primarily through regulation, planning and development assessment instruments. Agriculture is not generally considered an environmentally relevant activity within the current regulatory process (unless it is intensive agriculture, e.g. feed lots, aquaculture), and as a result is not subject to local planning and development assessment mechanisms. Despite this limitation, a wide range of State legislative instruments have potential to be utilized to protect the values of the GBR. These include legislation that relates to the management of leasehold lands, water supplies, vegetation management, environmental protection and ﬁsheries management. Whilst a detailed overview of these potential management levers is not appropriate here, it is important to recognize that these mechanisms exist but are not currently utilized for this purpose. As a consequence, there are substantial inconsistencies in the way land uses are managed. For example, if an aluminium reﬁnery, aquaculture farm and tomato farm were proposing to operate in the same location and might contribute equally signiﬁcant water quality issues, the regulatory requirements would be markedly different – from a highly regulated approach with the former operations to little restriction on operation in the case of the tomato farm. Accordingly, landholders are

operating in a highly contrasted regulatory regime, leading to further complexity in the management regime. Regional NRM bodies have an increasingly important role in catchment management in Queensland and nationally. These groups lead the development of integrated Regional NRM plans which utilize the best-available science and involve strong community participation. Regional NRM plans are characterized by a hierarchy of targets that articulate management action and resource condition objectives for natural resource assets. Delivery is facilitated through partnerships between the regional NRM bodies and a range of other institutions (including all levels of government) and individual land managers. The Regional NRM plans are required to be consistent with other planning processes, and in the GBR catchments the Reef Plan has placed additional expectations on the regional NRM process to address water quality issues. In turn, the current focus of activity implemented to address the intent of the Reef Plan is delivered through regional planning processes. As noted above, additional investment has been provided in the priority GBR catchments to accelerate the development of scientiﬁcally rigorous Water Quality Improvement Plans. These plans facilitate integration and co-ordination of activities and management agencies, and provide a level of regional water quality management for the GBR in the absence of a regulatory agricultural management regime. These plans are, however, constrained by the level of resources available for implementation, and their jurisdiction, i.e. that of a community-based voluntary approach (albeit with the potential for considerable co-investment in adoption of BMP). Mechanisms currently utilized to support the voluntary adoption of improved management practices include increased awareness and education, extension and decision-support, grants and a range of market-based instruments such as auctions, incentives, ecosystem services payments, discharge trading and offsets (Marsden Jacob Associates 2008). Additional complementary mechanisms that governments could utilize

Managing the Catchments of the Great Barrier Reef include levies, permits, contracts, planning approval processes and changes to property rights, and whilst these options are being investigated they are yet to be implemented across the GBR catchments. The current arrangements rely on policy decisions at the GBR-wide scale with implementation of voluntary approaches within existing systems occurring at the regional scale, leading to further complexity in catchment management. Peak agricultural industry bodies are a major partner operating at both scales. This arrangement appears adequate on the assumption that the outcomes of the implementation of regional plans will achieve Reef Plan objectives. This analysis has not been performed, but as the science behind the regional water quality targets is emerging, it appears less likely that this is the case unless very high levels of uptake of all agreed best practices are achieved across all industries in all catchments. This issue is demonstrated in a recent draft water quality plan in the Wet Tropics catchments which sets a target for 23% reduction in nutrient exports based upon accelerated adoption of current industry BMP standards. Using estimations of the relationship between the end of catchment load and a marine water quality guideline (Wooldridge et al. 2006), a nitrogen load reduction of at least 80% is required to meet the guideline value. Whilst there is signiﬁcant uncertainty in the system understanding to support these estimations, it is evident that the current BMPs are not likely to deliver sufﬁcient outcomes in the next 5 years, and that substantial improvements in one catchment alone will not deliver the necessary outcomes for the GBR. A cross-regional approach to water quality management will be essential to achieve the necessary improvements, and more substantial measures, such as land use change, require further investigation in the longer term.

15.5 Improving the System The previous sections highlighted the complexity of the natural and institutional environments

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related to catchment management and marine ecosystem health in the GBR, and outlined the current management approaches. As the Reef Plan moves beyond the 5-year half-way mark there is increasing recognition of the need to improve the effectiveness of its delivery, particularly through improved partnership arrangements and increased resourcing to support real practice changes. A discussion of the interaction of science and management in responding to this need is presented below.

15.5.1 Institutional roles and scales Whilst the Reef Plan provides an innovative policy approach, premised on the value of aligning existing activities and initiatives within and beyond government to deliver an ‘integrated’ approach, the implementation of the Reef Plan has lacked a detailed plan, comprehensive commitment from the State Government and a dedicated funding stream. This has limited the ability of the Reef Plan to strongly inﬂuence actions both within government and the wider stakeholder network. The recent provision of the Reef Rescue Plan funding, especially the grants supporting on-ground actions, monitoring and science co-ordination, could accelerate uptake of initiatives developed under the Reef Plan. The complexity of jurisdictional arrangements described above is another factor that challenges the delivery of a co-ordinated approach. In recognition of these, the Reef Water Quality Partnership (RWQP) was formed in 2006 to improve co-operation and collaboration between Australian and State Government agencies and the regional NRM bodies of the GBR, to address the common science needs for targets-setting, monitoring and reporting, and to facilitate cross-regional consistency in approaches to managing water quality issues for the GBR. The RWQP was initially established to address a number of Reef Plan actions that relate to the role of regional NRM bodies and monitoring outcomes. The new Reef Rescue Program arrangements will need to build on this partnership

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approach if it is to deliver the outcomes sought by governments.

15.5.2

Approaches to targets

Regional targets for the GBR are being set for adoption of new management practices, water quality outcomes within and at the end of catchments, and GBR ecosystem health, i.e. reefs and seagrass beds. In 2006 a consistent approach, reﬂected in Figure 15.4, was adopted across the regions and signiﬁcant advances have been made in the rigour of the target-setting process than earlier efforts (e.g. Brodie et al. 2001). However, as a result of the limitations in monitoring and modelling capacity referred to previously, water quality targets are presently largely driven by an understanding of what is achievable water quality change within current land use systems and practices. The environmental tradeoffs in this decision have not yet received much attention because of the low conﬁdence in our understanding of what is actually being discharged from catchments and how this relates to requirements to sustain healthy GBR ecosystems. The implications for ecosystems of not

meeting targets, and the lag time to change practices and realize water quality beneﬁts for the target adopted, have high levels of uncertainty. In developing the targets three critical information needs emerge: • deﬁnition of the water quality required to protect the GBR ecosystems; • the degree to which water quality change can be achieved; and • selection of the most effective management actions to facilitate the necessary improvements. Fundamental to these questions is an understanding of the conceptual and quantitative linkages between land use, water quality and ecosystem health. These information needs are slowly emerging from long-term research and monitoring programmes that aim to understand and quantify the implications of poor water quality on the health of the GBR (De’ath and Fabricius 2008). This will allow quantiﬁcation of the effectiveness of investment in improved land management in the catchments to enhancing biodiversity in the GBR, which in turn underpins its resilience to other pressures, especially climate change. Whilst many management decisions are based on the precautionary approach, progression to a quanti-

Fig. 15.4 Water quality target setting within the GBR and relevant tasks within the Water Quality Improvement Plans.

Managing the Catchments of the Great Barrier Reef tative assessment of the implications of catchment activities on the GBR marine environment is the desired situation. Coupled with information of the water quality beneﬁt and cost-effectiveness of management practices and an understanding of barriers to adoption of improved practices, this information will guide future investment into management ‘hot spots’, e.g. fertilizer management (Brodie 2007) or erosion control (Brodie et al. 2003; Cogle et al. 2006), and provide the combined monitoring/modelling methodology needed to manage over decadal timeframes. Correlation between a catchment activity with a single ecosystem response and thus biological end points poses signiﬁcant challenges. Further research is required to link responses from management actions to biological indicators, such as chlorophyll a which has demonstrated a cause and effect response to nutrient enrichment in the marine environment (Devlin and Brodie 2005). Understanding what change is possible involves engaging the science of agricultural systems and its impact on water quality. Leading growers and agricultural experts have contributed to understanding the effectiveness of different practices and their applicability to local farming systems within the regional NRM planning process. Scaling up the knowledge of outcomes at a plot or paddock scale to catchment water quality exports has been through the application of catchment models (principally SedNet and ANNEX), which have allowed regional NRM bodies to explore scenarios of change in land use and management practices on water quality outcomes (Brodie et al. 2003; Cogle et al. 2006). This information has enabled the regional NRM bodies to consider the relative beneﬁts of different investment strategies, e.g. across different practices, industries and sub-catchments, and provides the foundation for targets adopted for the region. In some areas these outcomes will be compared with marine water quality objectives (where models and data allow), however, there remains a tension in identifying targets between what is considered achievable in terms of improved management actions, and what is con-

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sidered necessary to protect the health of GBR ecosystems. Setting water quality targets at regional scales, integrating science across agricultural, catchment and marine environments and scales through modelling tools and expert knowledge obviously involves high levels of uncertainty. This is most apparent in the marine environment, where relationships between water quality parameters and the resilience of GBR ecosystems is just emerging (Brodie et al. 2005; Fabricius 2005; DeVantier et al. 2006; Wooldridge et al. 2006; Cooper and Fabricius 2007). Tropical systems behave distinctly, yet much of the available eco-toxicological data is from temperate systems. The paucity of hydrodynamic, geochemical and ecological modelling tools for these areas is a major constraint. The roles of ﬂoodplains and estuaries as potentially signiﬁcant elements in the linkages between land use and GBR ecosystems has received little attention and remains largely unknown at this point. Catchment models also have limitations that are receiving attention from research and government agencies (e.g. Sherman et al. 2007; Grayson 2007; Hateley et al. 2007; Post et al. 2007). A common challenge across the system is understanding the dynamics during high rainfall events (at the enterprise level), ﬂood events in rivers, and impacts of ﬂood plumes in marine ecosystems. 15.5.3 Managing within uncertainty, the adaptive management of GBR water quality outcomes Adaptive management is a structured process that facilitates ‘learning by doing’ in the face of complex systems and high uncertainty (Holling 1978; Walters 1986). Therefore it is recommended such an adaptive, whole-of-system approach to planning and management for the GBR be established and implemented, where the catchment is deﬁned from the source of the rivers to the drop off of the continental shelf, and where the socioeconomic system feeds back to natural resource managers further up the catchment.

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At present, a passive adaptive approach for managing GBR water quality has been established, with strategies adopted and the expectation that monitoring will provide feedback on the success or otherwise of these strategies. However, an assessment of the regional water quality management strategies (Eberhard et al. 2008) suggests that there are signiﬁcant gaps in the current capacity to support effective feedback mechanisms, based on the uncertainty in science linking end of catchment loads to GBR health objectives, and potential decadal time lags in measuring signiﬁcant changes in mean annual loads. There are several risks with this ‘passive’ management approach (Parslow et al. 2007): • Monitoring will not be able to determine whether the strategies have been successful within a manageable timeframe. • Monitoring may indicate that the management actions adopted by the regional water quality plans have not been successful. • Where implementation of regional water quality plans successfully achieve water quality beneﬁts, wider GBR health beneﬁts may not be realized. • Climate change scenarios indicate that the GBR will be under increasing climatic pressures within one or two decades. This dictates some urgency in addressing the compounding pressure of declining inshore water quality, particularly given the likely lag times in facilitating change and realizing the beneﬁts. The adaptive management approach must therefore rest on feedback within manageable timeframes – both within the individual plans and across scales (Eberhard et al. 2008). One of the challenges is to effectively link explicit management strategies at the regional and sub-regional scale (on a 3- to 5-year planning cycle) to the GBR policy scale (a 10-year cycle). Many of the foundation knowledge needs are common to both scales, across ecological, social and economic dimensions, and would enable evaluation of the distinct but complementary management strategies. Even by adopting an adaptive management approach, the challenges of dealing with the timeframes of system response, compared to rea-

sonable timeframes for policy and management decisions, will persist. Measuring GBR ecosystem outcomes in the timeframe of an adaptive management cycle that is useful to managers is unlikely. However, there are indicators of change at smaller scales, such as adoption of management practices and (validated) predictions of improvement in end of catchment loads that can be used with greater conﬁdence in evaluating performance against targets. This emphasizes the role of regional knowledge in informing GBR outcomes, and the need to prioritize science needs to inform management within planning and implementation cycles.

15.6 Future Challenges for the Partnership of Catchment Managers Substantial progress has been made in catchment management efforts over the last 5–10 years to address water quality issues in the GBR (Brodie et al. 2009b). However, it is proposed that signiﬁcant improvement can be achieved through better collaboration between catchment managers and research providers, guided by a rigorous strategic plan for future investment. The RWQP provides an example of a collaborative arrangement that emerged in response to the science needs to support regional NRM delivery of the Reef Plan. The change in focus provided by the Reef Rescue Program and the transition of the achievement of the Reef Plan, especially the partnerships, will be crucial to maintaining the momentum of change. As these partnerships mature, it is timely to reﬂect on the science, management partnerships and policy challenges that lie ahead. In the last 5 years, governments have invested substantial resources into science to support implementation of the Reef Plan, estimated to be in the order of AUD$100 million (Brodie et al. 2009a). Despite this investment, the scale and nature of the science challenge for Reef Plan and Reef Rescue Program remains signiﬁcant. It is recognized that even the best science support will not ensure change in on-ground practices,

Managing the Catchments of the Great Barrier Reef although the emphasis of community-based regional NRM solutions in the GBR catchments has resulted in regional ‘action research’ that has assisted in facilitating social change. While a diversity of approaches provides a productive source of learning opportunities, a more strategic approach is required to support a major investment for Reef Plan (Ferrier 2007). Recent assessments of the critical gaps in knowledge to support Reef Plan highlight that integration of the science is key to addressing the complexities and uncertainties of the GBR system. The present approach of delivering components of the knowledge, without an overarching effort to collate, synthesize and integrate this knowledge, is likely to continue to fail to meet management needs. Science integration is the key to informing management decisions, and requires additional skills in conceptual and quantitative design and interrogation that go beyond traditional ﬁelds of expertise. An integrated approach is required to understand the whole system that results in GBR water quality, and includes relationships: • within and across catchments to the GBR, so that the linkages between catchment actions and GBR health, and within the components of the system (e.g. between water quality and coral health), can be quantiﬁed; • between biophysical, social and economic dimensions of the system so that realistic targets and implementation strategies can be developed and assessed; and • across scales, so that the sum of catchment and regional activities can be assessed to determine whether the existing and proposed activities are sufﬁcient to achieve the Reef Plan goal. The Reef Rescue Program’s Research and Development Program will need to be used efﬁciently to address the integration issues. Importantly, a formal mechanism to facilitate science co-ordination and the development of an agreed science strategy will be critical to this programme. The scale and complexity of the systems in question will require innovative approaches to monitoring and modelling, data and information management, and synthesis,

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communication and dialogue with policy and management agencies are essential elements of science integration. Ultimately the scope of science to support catchment management should include climate change impacts as these are inherently related to land management practices, water quality and GBR ecosystem health. The RWQP was established in recognition of the issues highlighted above and has facilitated and co-ordinated a collaborative approach between governments and regional NRM bodies to support target-setting, monitoring and reporting for the Reef Plan and regional NRM plans (Eberhard 2006). In the 2 years since its inception, the RWQP has clearly scoped the need for the issues identiﬁed above to be addressed. Whether the RWQP itself will expand to take on these additional responsibilities or whether new structures will emerge to replace or complement existing arrangements is not yet clear. What is clear is that despite the signiﬁcant challenges, the need for a partnership arrangement to support a collaborative approach to the Reef Plan and Reef Rescue Program has been clearly demonstrated to all of the partners.

15.7 Conclusions Management of water quality issues in the GBR catchment pose signiﬁcant challenges for the future of the GBR. Whilst substantial progress has been made in the last 5–10 years, effective linkage between regional NRM plans (community-based, regional) and Reef Plan (policy, GBRwide), and collaborative partnerships between management agencies, industry groups and science providers is critical. The regional NRM bodies appear to have great potential to deliver NRM outcomes within the scope of their jurisdiction, and available resources, however, expectations need to be tempered by the scientiﬁc evidence of the likely outcomes of these plans, and whether they are adequate to support the GBR ecosystems. Strengthening the ability to answer this question by addressing key knowledge gaps such as marine ecosystem objectives,

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and improving the capacity of our science programmes to effectively integrate knowledge across scales and systems is one facet of this. Another facet is the strengthening paradigm of institutional partnerships for collaborative NRM solutions (at both the regional and GBR scale). The ﬁnal element necessary to complete this picture is the linking of the regional NRM and Reef Plan and Reef Rescue Program. Reef Plan has played a very important role in exploring policy support through tools such as regulation and land use planning that lie outside the remit of regional NRM bodies. This needs to be coordinated with, and sensitive to, the regional community-based agenda – poorly timed or judged policy interventions have great potential to undermine regional NRM processes. Facilitating an effective dialogue between these scales whilst respecting the different roles played by each is a careful boundary to negotiate. The Reef Rescue Program will, for the ﬁrst time, provide signiﬁcant ﬁnancial support to achieve best management practice change across the GBR catchment. For all that, resources still need to be targeted at the priority areas and actions to get the most cost-effective outcome and return on investment in halting and reversing water quality decline in the GBR.

Acknowledgements The authors of this chapter would like to thank Bob Ferrier for inviting us to contribute and for his efforts in science for management of water quality issues of the Great Barrier Reef.

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Australian Centre for Tropical Freshwater Research (ACTFR) Report. No. 06/10 for FNQNRM Tully CCI. ACTFR, James Cook University, Townsville. Negri, A., Vollhardt, C., Humphrey, C. et al. (2005) Effects of the herbicide diuron on the early life history stages of coral. Marine Pollution Bulletin, 51, 370–383. Neldner, J. (2006) Why is vegetation condition important to government? A case study from Queensland. Ecological Management and Restoration, 7, S5–S7. Packett, R., Ford, P.W. and Noble, B. (2005) Overview of the delivery of pesticides to the Fitzroy Estuary and persistent organic pollutants in the food chain. In: Noble, B., Bell, A., Verwey, P. and Tilden, J (eds), Fitzroy in Focus: coastal science for the Fitzroy Region. Coastal CRC for Coastal Zone, Estuary and Waterway Management. Indooroopilly, Brisbane. Pandolﬁ, J.M., Bradbury, R.H., Sala, E. et al. (2003) Global trajectories of the long-term decline of coral reef ecosystems. Science, 301, 955–958. Pannell, D.J., Marshall, G.R., Barr, N., Curtis, A., Vanclay, F. and Wilkinson, R. (2006) Understanding and promoting adoption of conservation practices by rural landholders. Australian Journal of Experimental Agriculture, 46, 1407–1424. Parslow, J., Grayson, R. and Hart, B. (2007) Adaptive Management and Integration for the Reef Water Quality Partnership. A discussion paper prepared as part of the development of a protocol for developing adaptive management strategies for Reef Water Quality Improvement Plans. Subgroup of the RWQP Scientiﬁc Advisory Panel, December 2007. Pearson, R.G. (2005) Biodiversity of the freshwater fauna of the Wet Tropics region of north-eastern Australia: patterns and possible determinants. In: Bermingham, E., Dick, C.W. and Moritz, C. (eds), Tropical Rain Forests: past, present and future. University of Chicago Press, Chicago and London, pp. 470–485. Pearson, R.G. and Stork, N.E. (2008) Catchment to reef: water quality and ecosystem health in tropical streams. In: Stork, N. and Turton, S. (eds), Living in a Dynamic Tropical Forest Landscape. WileyBlackwell, Oxford, chap. 45. Pearson, R.G., Benson, L.J. and Smith, R.E.W. (1986) Diversity and abundance of the fauna in Yuccabine Creek, a tropical rainforest stream. In: de Deckker, P. and Williams, W.D. (eds), Limnology in Australia. Monographiae Biologicae Series no. 61. Junk, Dordrecht, pp. 329–342.

Post, D.A., Waterhouse, J., Grundy, M. and Cook, F. (2007) The Past, Present and Future of Sediment and Nutrient Modelling in GBR Catchments. Commonwealth Scientiﬁc and Industrial Research Organisation Publication (CSIRO). Water for a Healthy Country Flagship. CSIRO, Canberra. Potts, D.C. (1981) Crown of thorns starﬁsh: maninduced pest or natural phenomenon? In: Kitching, R.L. and Jones, R.E. (eds), The Ecology of Pests. Commonwealth Scientiﬁc and Industrial Research Organization, Melbourne, pp. 55–86. Prange, J., Haynes, D., Schaffelke, B. and Waterhouse, J. (2007) Great Barrier Reef Water Quality Protection Plan (Reef Plan) Annual Marine Monitoring Report. Reporting on data available from December 2004 to April 2006. Great Barrier Reef Marine Park Authority, Townsville, Australia. Productivity Commission (2003) Industries, Land Use and Water Quality in the Great Barrier Reef Catchment. Research report, Australian Government Productivity Commission, Canberra. Pusey, B.J., Kennard, M.J. and Arthington, A.H. (2004) Freshwater Fishes of North-Eastern Australia. CSIRO Publishing, Melbourne. Queensland Farmers Federation (QFF) (2005) Farm Management Systems Framework. QFF, Australia. Rasiah, V., Armour, J.D. and Cogle, A.L. (2007). Statistical characterization of impact of system variables on temporal dynamics of groundwater in highly weathered regoliths. Hydrological Processes, 21, 2435–2446. Rayment, G.E. (2003) Water quality in sugar catchments in Queensland. Water Science and Technology, 48, 35–47. Roberts, I., Puangsumalee, P. and Penm, J. (2008) Agricultural trade: the changing environment for trade policy and agricultural prices. Australian Commodities, 15, 217–235. Roebeling, P.C., Webster, A.J., Biggs, J. and Thorburn, P.(2007) Financial-economic Analysis of Current Best-management-practices for Sugarcane, Horticulture, Grazing and Forestry Industries in the TullyMurray Catchment. Report to MTSRF and FNQNRM, CSIRO Sustainable Ecosystems, Townsville, Australia. Rohde, K., Masters, B., Brodie, J. et al. (2006) Fresh and Marine Water Quality in the Mackay Whitsunday Region 2004–2005. Mackay Whitsunday Natural Resource Management Group, Mackay, Australia. Schroeder, B., Wood, A., Moody, P. and Panitz, J. (2005) Nutrient management – delivering the message to

Managing the Catchments of the Great Barrier Reef the industry. BSES Bulletin http://www.srdc.gov.au/ ProjectReports/BSS268_M4_Nutrient_ Management.pdf (last veriﬁed 18 November 2008). Shaw, M. and Müller, J.F. (2005) Preliminary evaluation of the occurrence of herbicides and PAHs in the wet tropics region of the Great Barrier Reef, Australia, using passive samplers. Marine Pollution Bulletin, 51, 876–881. Sherman B., Brodie J., Cogle, L. and Carroll, C. (2007) Appropriate use of catchment models for water quality target setting and land-use management. In: Water Quality and Sediment Behaviour of the Future: predictions for the 21st century. Proceedings of Symposium HS2005 at IUGG2007, Perugia, July 2007. International Association of Hydrological Sciences Publication 314. IAHS Press, Wallingford. The State of Queensland and Commonwealth of Australia (2003) Reef Water Quality Protection Plan; for Catchments Adjacent to the Great Barrier Reef World Heritage Area, Queensland. Department of Premier and Cabinet, Brisbane. http://www.reefplan. qld.gov.au/library/pdf/reefplan.pdf (last veriﬁed 17 November 2008). Sweatman, H. (2008) No-take reserves protect coral reefs from predatory starﬁsh. Current Biology, 18, R598–R599 (21 July 2008). Thorburn, P.J., Park, S.E. and Biggs, I.M. (2003) Nitrogen fertilizer management in the Australian sugar industry: strategic opportunities for improved efﬁciency. Proceedings of the Australian Society of Sugar Cane Technologists, 25, (cd-rom). Thorburn, P.J., Webster, A.J., Biggs, I.M., Biggs, J.S., Park, S.E. and Spillman, M.F. (2007) Innovative management of nitrogen fertilizer for a sustainable sugar industry. Proceedings Australian Society Sugar Cane Technologists, 29, 85–96.

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Veitch, V. and Sawynok, B. (2005) Freshwater Wetlands and Fish: importance of freshwater wetlands to marine ﬁsheries resources in the Great Barrier Reef. Sunﬁsh Queensland. Great Barrier Reef Marine Park Authority (GBRMPA), Townsville, Australia. Vine, P.J. (1973) Crown of thorns (Acanthaster planci) plagues: the natural causes theory. Atoll Research Bulletin,166, 1–10. Vinson, M.R. and Hawkins, C.P. (2003) Broad-scale geographical patterns in local stream insect genera richness. Ecography, 26, 751–767. Wallace, J., Hawdon, A., Keen, R. and Stewart, L. (2007) Water Quality during Floods and their Contribution to Sediment and Nutrient Fluxes from the TullyMurray Catchments to the GBR Lagoon. Report to FNQNRM for the Tully WQIP. Commonwealth Scientiﬁc and Industrial Research Organisation Publication (CSIRO) Australia, Canberra. Walters, C.J. (1986) Adaptive Management of Renewable Resources. McMillan, New York. Webster, A.J., Thorburn, P.J., Biggs, J.S., Biggs, I.M., Spillman, M.F. and Park, S.E. (2008) N replacement: a simple approach to improving nitrogen use efﬁciency in sugarcane. Proceedings Australian Society Sugar Cane Technologists, 30, 355–356. Wooldridge, S., Brodie, J.and Furnas, M. (2006) Exposure of inner-shelf reefs to nutrient enriched runoff entering the Great Barrier Reef Lagoon: post-European changes and the design of water quality targets. Marine Pollution Bulletin, 52, 1467–1479. Wrigley, T. (2006) Canegrowers Public Environment Report 2005. http://www.canegrowers.com.au/ publications-centre/public-environment-report.aspx (last veriﬁed 18 November 2008).

Image facing chapter title page: Courtesy of the Great Barrier Reef Park Authority and on behalf of the Commonwealth of Australia.

16

Catchment Management Case Study – Senegal River MIKE C. ACREMAN1

1

Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire, UK

16.1

Background

The Sahel (from Arabic sahil, meaning border) forms the transition between the Sahara desert to the north and the more humid and fertile Sudan region to the south. The Sahel runs 4000 km from the Atlantic Ocean in the west to the Red Sea in the east, in a belt that varies from several hundred to a thousand kilometres in width, covering an area of some 3 million square kilometres. The climate of the Sahel is tropical, with strong seasonal variations in rainfall and temperature. The Sahel receives between 200 and 600 mm of rainfall a year which falls mostly in the May to September monsoon season. Rainfall is generally higher in the south, declining rapidly to the north, and is characterized by great variation from year to year and from decade to decade, determined by the movements of the InterTropical Convergence Zone (ITCZ). There is a strong correlation between rainfall in the Sahel region and intense hurricane activity in the Atlantic. Monthly mean temperatures vary from a maximum of 33 ° to 36 °C to a minimum of 18 ° to 21 °C. During the winter, hot, dry Harmattan winds off the Sahara can bring sand and dust storms. The topography of the Sahel is mainly of low relief lying predominantly between 200 and Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

400 m elevation. Several isolated plateaus and mountain ranges rise from the Sahel but are designated as separate eco-regions because their ﬂora and fauna are distinct from the surrounding lowlands. The region’s vegetation is dominated by semi-desert grassland in the north and Acacia savannah wooded grassland and deciduous thorny bushland further south (White 1983). During the recent geological past, the region experienced a series of wetter (pluvial) episodes, some coinciding with high latitude glaciations (Goudie 1977), that created large lakes such as present day Lake Chad and the Inner Niger Delta and recharged underlying regional aquifers that now contain fossil water. The existence of permanent and temporary wetlands in a surrounding semi-arid region with a long and pronounced dry season and permeable soils is extremely important for wildlife and agriculture. In particular, the ﬂoodplain wetlands of large rivers, including the Senegal, Niger, Yobe and Logone, have for centuries played a central role in the rural economy of the region providing fertile agricultural land which supports a large human population (Groupe d’Experts des Plaines d’Inondation Sahéliennes 2000). The ﬂood waters provide a breeding ground for large numbers of ﬁsh and bring essential moisture and nutrients to the soil. Water that soaks through the ﬂoodplain recharges the underground reservoirs which supply water to wells beyond the ﬂoodplain (Acreman 1996). As the ﬂood waters recede arable crops are grown although some soil moisture

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persists to the dry season and provides vegetation that is essential grazing for migrant herds. The ﬂoodplains also yield valuable supplies of ﬁsh, timber, medicines and other products and provide crucial habitats for wildlife, especially migratory birds (Drijver and van Wetten 1992). The annual migration of birds along the Afrotropical– Palaearctic ﬂyway, as well as the intra-African migration associated with the seasonal climate ﬂuctuations, is dominated by the temporary nature of the wetlands in the Sahelian zone (BirdLife International 2007). In recent decades, drought, increasing populations of people and livestock and rising poverty have combined to put increasing pressure on the wetlands and has led to over-exploitation of their resources. In the face of such pressure, the key to development has often been seen as the implementation of major river engineering schemes such as dams for hydroelectric power generation and for intensive cereal cultivation. Dams can provide an all year supply of water in a highly seasonal climate. Few of the schemes, however, in Africa have ever realized their full potential and many are facing serious technical, economic, administrative, social and health problems. Added to this, the reduction in ﬂoodplain inundation caused by the retention of ﬂood water behind the dams has had disastrous effects on the traditional rural economy. Thus, in some cases, these schemes have diminished rather than improved the living standards and economy of the region as a whole (Acreman and Hollis 1996).

16.2

The Senegal Catchment

The Senegal River catchment (290,000 km2), one of the largest in Africa and the second largest in the Sahel, is shared by Guinea, Mali, Mauritania and Senegal (Fig. 16.1). The river ﬂows over 1800 km in three distinct zones: the mountainous upper basin lying mainly in the Fouta Djallon mountains of Guinea and Mali; the river valley

and its associated ﬂoodplain that varies between 10 and 20 km wide and forms the border between Senegal and Mauritania; and the delta where the river ﬂows into the Atlantic Ocean. The natural river ﬂow has a distinct seasonal pattern with high ﬂows during the wet season between April and October and low ﬂows between November and March. The river is a vital source of water in an otherwise arid landscape on the southern margins of the Sahara desert. Effective management of the Senegal River catchment is essential to the economies of the four riparian countries whose people depend upon the river for their livelihood in the form of agriculture, animal husbandry and ﬁsheries.

16.3 Catchment Hydrology Water is the major limiting factor for development in the region and understanding of the hydrological regime is crucial to effective catchment management. Rainfall in the Senegal catchment occurs from April to October causing an annual ﬂood from July to October which reaches its peak at some time during the months of August, September and October. The ﬂood is almost entirely generated by rain occurring in the upper basin over the Fouta Djallon and there is negligible inﬂow downstream of the Bakel which is at the head of the main valley. Recorded hydrographs show how the ﬂows have reduced during the 1980s and 1990s to an average that is less than half the overall average of 711 m3 s−1 (Fig. 16.2). While the mean annual ﬂow volume is 20,903 Mm3 the minimum is 6695 Mm3 (1984) and the maximum recorded annual ﬂow volume is 41,769 Mm3. In terms of ﬂooding characteristics, the valley may be divided into four stretches: Gouina to Bakel (202 km), Bakel to Kaedi (262 km), Kaedi to Dagana (363 km) and the delta (169 km). In the upstream stretch from Gouina to Bakel, the river is steep with a series of rapids and lies within an incised valley without a signiﬁcant ﬂoodplain.

Catchment Management Case Study – Senegal River

Box 16.1 A

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Four periods in the development of Senegal River catchment

The period of ‘average ﬂow’ 1904–1972

During this period ﬂoods occurred on a fairly regular basis, the average potential amount of ﬂood recession cropping is estimated at 110,000 ha. In practice, the population may not have been large enough to cultivate such an area and it is likely to have developed in parallel to signiﬁcant population growth. This period would also have been characterized by an abundance of physical resources related to the prevalence of ﬂooding. Forestry along the valley surviving on the periodic wetting would have been plentiful as would pasture for grazing. Fish populations are likely to have been large and ﬂora and fauna generally to have been more proliﬁc than can be seen at the present time. B

The drought years 1972–1987 prior to commissioning of dams

The drought had a dramatic impact on the physical environment and the dependent population. The average amount of ﬂood recession cropping is estimated at 37,000 ha. In several years it was almost zero, particularly in 1983 and 1984 which were the driest years of the century. The area of land under irrigation during the wet season began to be developed and increased from a total of 11,584 ha in 1977 to 42,517 ha in 1987. The signiﬁcant reduction in the ﬂooded area over a sustained number of years resulted in a substantial loss of forestry, pasture for grazing (and hence in the size of the herds) and ﬁsheries. The shortage of water was dramatically illustrated in the driest years of 1983–1985 when the army built a temporary dam across the Senegal to provide Dakar with water. The lack of traditional resources forced many active family members to seek work outside of the valley either in towns or aboard. This increased the burden on women and children. C The period 1988–2001 when the Manantali Reservoir was able to release ﬂoods without major constraints on other possible uses The storage of water behind Manantali Dam prior to the installation of turbines was used to regulate river ﬂows all the year round. In conjunction with Diama Dam, the saline intrusion that used to reach some 300 km upstream was eliminated. This had a negative impact on ﬁsheries, in part compensated by the development of ﬁshing in the Manantali Reservoir, but allowed the expansion of irrigation and the introduction of dry season cropping. The area developed for irrigation tripled in 10 years to reach 131,000 ha in 1998. Managed ﬂood releases were used to sustain some ﬂood recession agriculture, with the inundated area averaging 58,000 ha. This was an increase over the drought period (1972–1987), at a time when ﬂows have continued to be below average, but less than pre1972, when conditions were wetter. D 2001 onwards following commissioning of the hydro-power station after which ﬂood releases were in conﬂict with other water uses After 2001, the release of large quantities of water for managed ﬂooding reduced the amount of water for power generation. Computer simulation using the hydrological record has shown that release of the minimum managed ﬂood reduces the mean annual energy from some 800 GWh by an amount varying from 140 to 190 GWh depending on the hydrological record used. Nevertheless the importance of continuing to make environmental ﬂood ﬂow releases was widely accepted. The expansion of irrigation was bringing it into conﬂict with power generation in terms of water allocation and with ﬂood recession cropping in terms of space.

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Fig. 16.1 Senegal River catchment.

Fig. 16.2 Discharge hydrograph at Bakel showing reduced ﬂows in the 1980s compared to earlier in the twentieth century.

Catchment Management Case Study – Senegal River Downstream of Bakel, the valley widens and the ﬂoodplain attains up to 10–15 km width under peak ﬂoods, the average slope being 3 cm km−1. Downstream of Kaedi, the river divides into two arms, the Doué on the left and the Senegal on the right, as it encircles the lle Amorphil which has an average width of over 235 km. Downstream of Dagana, the river returns to a single channel before opening up into the delta (Acreman et al. 2000). The ﬂoodplain on either side of the main river channel is made up of 72 large natural depressions or basins varying from 1000 ha to over 15,000 ha in size. These serve to trap ﬂood water and its sediment and provide fertile land. Much of the ﬂoodplain wetland areas would have been vegetated naturally by gallery forests where Acacia nilotica predominates. Many areas have been cleared, however, for agricultural exploitation.

16.4 People of the Catchment In 1988 the population within the Senegal catchment was estimated at 1550,000 people (767,000 Senegalese, 696,000 Mauritanians and about 80,000 Malians). The population settled along the Senegal River was 1100,000. The Senegal River is a contact zone between the Arab Berbers to the north and the Negro Africans to the south. It has traditionally been a place of cohabitation between ethnic groups but also of tension. The drought since 1973 and the changes brought about by the construction of the dams have brought about profound changes to the social fabric of the region, particularly with regard to land use. This resulted in extreme tensions and frontier incidents between Senegal and Mauritania in 1989–1990 during which 120,000 refugees were repatriated from Mauritania and some 70,000 dispersed within the valley. Within the ﬂoodplain area of the lower and middle valley, the Peuls and Toucouleurs are the most important ethnic group. The distribution and activities of these different groups is broadly as follows:

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• Toucouleurs represent more than 60% of the population on the Senegalese bank and are cultivators and ‘masters of the land’. Traditionally many thousands of them used also to cultivate land on the right bank which they were forced to abandon following the clashes in 1989–1990. • Peuls are traditionally nomadic pastoralists; they represent only 4% of the population on the left bank and are spread in camps along the length of the river. • Wolofs are mainly concentrated in the delta and lower valley on the left bank where they make up some 25% of the population. • Maures are mainly located on the Mauritanian right bank. They are made up of white Maures of Berber Arab origin and black Maures who were previously the servile labour force. The traditional activities are commerce and pastoralism but the river has always held an attraction due to its water resources and grazing and there has been an increasing afﬂux of Maures who have been settling by the river. The numbers increased from 315,000 in 1977 to 600,000 in 1988.

16.5 Transboundary Catchment Management The countries of the basin have planned the development of the Senegal River over many years through Le Comité Inter-Etats pour l’Aménagement du bassin du Fleuve Sénégal (1964–1968) and L’Organization des Etats Riverains du Fleuve Sénégal (1968–1972) to address catchment management, particularly as this region has high natural inter-annual variability in rainfall and in river ﬂows. During the 1970s and 1980s, however, the Sahelian region experienced an extended period of drought which led to a serious water deﬁcit. In response to this situation, Mauritania, Mali and Senegal signed a treaty in 1972 to establish the transboundary Senegal River Basin Authority (L’Organization pour la Mise en Valeur du Fleuve Sénégal – OMVS) with the mandate of ‘securing countries’

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economies and reducing the vulnerability of peoples’ livelihoods through water resources and energy development’. Guinea joined OMVS in 2006. The structure of OMVS has evolved over time as it has been reorganized to meet the changing requirements of, ﬁrstly, the construction of the infrastructure and then its management, operation and maintenance. The present structure is as follows: • Head of State Conference – ultimate authority which meets once a year. • Council of Ministers – organ which controls OMVS composed of a minister from each member state and which meets twice a year. • High Commission – executive body directed by the High Commission who is appointed for 4 years. Responsible for political co-ordination, planning of the development programme, its implementation and supervision of its management, operation and maintenance. • Manantali Holding Company (SOGEM) – the company which holds the assets of the Manantali project and is charged with overall responsibility for its management, operation and maintenance. • Diama Holding Company (SOGED) – holding company responsible for the Diama Dam and other river infrastructure as well as its management, operation and maintenance. • Consultative bodies – the most signiﬁcant of these is the Permanent Water Commission which comprises representatives from each member state and is charged with deﬁning the allocation of water between states and between sectors. The objectives of OMVS are to: (i) stabilize and improve the livelihoods of the inhabitants of the basin and adjacent areas; (ii) maintain an ecological balance within the basin and promote its sustainability within the sahelian zone; (iii) render the member states less vulnerable to climatic variations and external factors; (iv) accelerate the economic development of the member states through intensive regional cooperation.

16.6 Catchment Infrastructure A key element of the plan of OMVS includes the implementation of a regional infrastructure, including: (i) the Manantali storage reservoir and hydro-electric project; (ii) the Diama saline intrusion barrage; (iii) embankments along the lower river to create high water levels for gravity irrigation and navigation between Saint Louis and Kayes. The OMVS vision is to ‘implement a joint basin development program that reinforces regional integration, yields beneﬁts and sustains growth among the four associated riparian countries’. The distributional effects were not seriously considered (i.e. which communities beneﬁt or lose out), the objectives were focussed on national and regional economic growth. The Diama Dam was constructed at the mouth of the river between 1981 and 1986 to stop saline water entering the river, thus making the river a reservoir of freshwater for irrigation. About 375,000 ha of land was under irrigation. Embankments were built from the dam upstream on both banks in the early 1990s to store water in the river at an elevation that allowed gravity feed to the ﬂoodplains and provided sufﬁcient depth for year-round navigation along the river into Mali. The barrage also provided water supply for stock and for Dakar. The construction of the upstream storage reservoir at Manantali on the Baﬁng River was undertaken during the period 1982–1987. The reservoir, with a capacity of 11,270 million cubic metres, was ﬁlled over the three subsequent years. The purpose of the dam was to generate 800 GWh year−1 of electricity (90% certainty) and to provide water for irrigation. The dam site controls 50% of the ﬂow on average and 70% in dry years. The hydroelectric power station was postponed pending further studies which were carried out in the early 1990s. Under the original project, irrigation has been developed on about 100,000 ha of the potential 375,000 ha of land suited for irrigation. A ﬁshery has developed on the Manantali Reservoir, leading to settlement.

Catchment Management Case Study – Senegal River Table 16.1 Value of ﬂoodplain production under pre-dam conditions Activity Recession agriculture Fishing Grazing Total

Value (US$/ha) 56–136 140 70 266–345

16.7 Environmental Services and their Values Prior to construction of the Manantali Dam, natural inundation of the ﬂoodplain of the Senegal valley supported vital ecosystem services for local communities (Horowitz and SalemMurdock 1990) including up to 250,000 hectares of ﬂood recession agriculture, forests which provided fuelwood and construction timber, ﬁshing, grazing for livestock, recharge of groundwater (OMVS/ISTI 1990), wildlife habitat (Acreman 2003) and maintenance of wetlands in the Senegal River delta (Hamerlynck and Duvail 2003). Whilst there is considerable debate about the actual economic value of ﬂoodplain production (Table 16.1), the principle has been accepted that the ﬂoodplain ecosystem has a high value that was not recognized in initial planning of the river basin. There was a succession of drivers for provision of environmental ﬂows within the Senegal River basin. The initial driver was the realization by Senegal that the belated development of the dams would cause signiﬁcant economic and social disruption to the ﬂoodplain communities. The construction of the Manantali and Diama dams created signiﬁcant environmental and social impacts. A primarily impact was the loss of ﬂood recession agriculture, fuelwood and grazing on the ﬂoodplain. There was also a 90% drop in the productivity of the ﬁsheries of the Senegal Delta which relied on inputs of freshwater from upstream. Although a new ﬁshery was created in the Manantali Reservoir, it was less productive and attracted different ﬁshermen. Also the character of the vegetation in the Djoudj

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National Park, adjacent to the river, changed signiﬁcantly as the periods of saline water intrusion into the river, which used to occur during the dry season, were replaced by a regime of continuous freshwater. For example, the river channel became chocked with Typha australis, which had previously been controlled naturally by saline water, and led to an increased incidence of bilharzia and malaria. The Diama barrage and embankments along both sides of the river led to severe degradation of the delta and loss of biodiversity in the Diawling National Park in Mauritania, also in the Senegal River delta.

16.8 Environmental Flood Flows Because of the delay between dam construction and installation of turbines, from 1991 OMVS agreed that major environmental ﬂow releases should be made from Manantali. This would create a managed ﬂood for a transitional period of 10 years to restore the ecosystem services to people undertaking ﬂood recession agriculture, herding and ﬁshing. Technical studies were undertaken to assess the feasibility of ﬂood ﬂow releases and the economic implications (Sir Alexander Gibb and Partners 1987). The aim of the work was to maximize the beneﬁts to ﬂood recession cropping whilst seeking to retain sufﬁcient water in the Manantali Reservoir for the purposes of irrigation, power generation and navigation. The studies were based on an analysis of the characteristics of the natural ﬂoods together with a review of the areas ﬂooded in the past and the agricultural requirements. In particular, a model was prepared to calculate the area ﬂooded in the valley from the topographic data available on the ﬂoodplain and its adjacent basins together with the longitudinal proﬁles of the ﬂood as recorded at the different gauging stations. This was combined with survey data that existed for inventories of actual ﬂood recession cropping undertaken for 6 years during the 1970s to try to correlate the ﬂooded area with that cultivated.

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The factors in the conception of the hydrograph were: • Area to be cultivated – the 6 years of inventory from the 1970s showed ﬂood recession cropping to vary from as much as 103,100 ha in 1970/71 to as little as 13,700 ha in 1972–1973. For a number of years in the 1980s, which were amongst the driest of the century, there was negligible ﬂooding and hence little cropping. A review of the labour requirements concluded that the maximum that could be cultivated with the population as at the mid 1980s was around 110,000 ha. Three hydrographs were produced that were capable of permitting the cultivation of 50,000 ha, 75,000 ha and 100,000, respectively. Thus in a dry sequence of years, such as those experienced in the period 1970–2000, the net effect would be to increase the overall area cropped. • Area to be ﬂooded – the area that can be cultivated is broadly that which has been ﬂooded for more than 15 days, allowing the soil to absorb sufﬁcient water, but less than 45 days, as beyond this point the climatic conditions are likely to be less propitious by the time the ﬂood has receded. Correlation of areas ﬂooded within this period suggest that approximately half the area inundated is likely to be cultivated. • Flood peak to generate ﬂooded area – an analysis of historical ﬂoods that had been the most effective at generating the required ﬂooded areas concluded that a ﬂood peak of 2200 m3 s−1 maintained for 10 days would be required to ﬂood 100,000 ha and 3000 m3 s−1 maintained for 20 days to ﬂood 200,000 ha. • Rate of rise – a rapid rate of rise is desirable for agricultural purposes in order that maximum water absorption can take place through the

deep cracked soils that otherwise have a tendency to close. Using actual historical values, a rate of rise of 2000 m3 s−1 over 10 days was selected. • Rate of fall – the rate of fall was selected based on historical values that were efﬁcient in terms of water use and the concept that the fall should not be too rapid in order to allow the farmers the time to follow the recession before the ground dried out. Two rates of fall were identiﬁed from historical data and the values selected were a reduction from 2500 m3 s−1 to 1500 m3 s−1 over a 10-day period and a fall from 1500 m3 s−1 to 500 m3 s−1 over a 20-day period. • Timing of ﬂood – historically the ﬂood peaks occurred over the months of August to October. A key objective was to build managed ﬂood releases on top of the natural ﬂoods occurring on the unregulated tributaries which account for about 50% of the ﬂow. Simulation of historical years of data showed that it was possible to recognize the trigger point at which the ﬂood occurred and to combine the releases efﬁciently with the tributary ﬂows to minimize the use of water. This was achieved in real time using hydrometric data relayed by satellite, or by radio, which was used should the automatic stations not be functioning. The overall conclusion of the study was a set of ﬂood hydrographs (Table 16.2). During ﬁlling of the Manatali reservoir, and up until 1990, negligible ﬂooding took place. However, in the years following ﬁlling environmental ﬂow ﬂood releases were made with the precise scenario (A, B or C; Table 16.2) depending on the available water. The average area of ﬂoodplain inundated was around 58,000 ha. Figure 16.3 shows the ﬂood hydrograph scenario A superimposed on

Table 16.2 Design ﬂood releases for the Senegal River (after Sir Alexander Gibb and Partners 1987) Managed ﬂood A B C

Designed for the cultivation of:

Volume Aug–Oct (km3)

Peak ﬂow (m3 s−1)

Duration at 2000 m3 s−1 (days)

50,000 ha 75,000 ha 100,000 ha

7.5 8.5 10.0

2000 2750 3000

10 15 20

Catchment Management Case Study – Senegal River

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Fig. 16.3 Environmental ﬂood ﬂow releases (scenario A) compared with natural hydrographs for 1988 and 1990.

recorded ﬂow data for two years (1988 and 1990). It can be seen that the same ﬂood hydrograph released each year can be less than (as in 1988) or greater than the natural ﬂow conditions (as in 1990). Thus a managed release may produce a more reliable ﬂood for ﬂoodplain-dependent livelihoods. The very ﬂat hydraulic proﬁle of the Senegal River means that there is not a unique water level–ﬂow relationship and the same level can be produced at different ﬂows at a particular location depending on the rate of rise and fall. A second set of studies was undertaken to produce a statistical model relating ﬂow and level data and a hydraulic model that enabled the transmission of ﬂoods down the valley to be simulated. The use of satellite imagery also permitted the extent of ﬂooding and areas cropped to be surveyed and inventoried more accurately and quickly.

16.9 Water Allocation Trade-offs Although the environmental ﬂood ﬂow releases included in the plan inundated only around 58,000 ha (20% of the original area), they had signiﬁcant beneﬁts. Fishermen in the Senegal River at Mauritania saw their annual catch rise from 10 to 110 tonnes once the annual ﬂoods were re-established. After installation of the tur-

bines in 2001 the economics of environmental ﬂows changed. The release of water for managed ﬂoods reduced the amount of power that could be generated. Even the minimum environmental ﬂood ﬂow release reduces the mean annual energy from some 800 GWh by an amount varying from 140 to 190 GWh depending on the hydrological regime. Horowitz and Salem-Murdock (1990) concluded that a combination of environmental ﬂood ﬂow releases and generation of some hydro-power was the most efﬁcient economically. Hollis (1996) reviewed the original consultants’ reports and suggested that an environmental ﬂood ﬂow release could be made which would inundate 100,000 ha whilst retaining sufﬁcient water to generate 912 GWh of electricity with a 95% certainty. The trade-off between electricity generation and managed ﬂood releases has important political implications since electricity beneﬁts the urban elite (mainly Maures), commerce and industry (there being little rural electriﬁcation) whilst ﬂoods beneﬁt the rural poor (Wolofs, Peuls and Toucouleurs). Financial assistance from the World Bank for the hydropower turbines was conditional on the basis that OMVS retained the managed ﬂoods as a possible long-term option. A Plan for Mitigating and Monitoring Impacts on the Environment (Plan d’Atténuation et de suivi des Impacts sur l’Environnement – PASIE) was undertaken during

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the implementation of the Regional Hydropower Development Project to mitigate the environmental and social issues that had occurred as a result of the construction of the original dams. Amongst other activities, PASIE included a programme for optimizing management of the Manantali Reservoir, so as to restore some level of beneﬁt to downstream farmers. Further environmental ﬂow programmes were implemented to release water through the embankments to re-inundate the Diawling National Park in the north of the delta based on hydro-ecological studies (e.g. Duvail 2001). Such releases were possible from the water stored within the river embankments immediately behind the lower Diama dam and thus did not require additional water from Manantali and so did not impact on electricity generation. In addition, since development of intensive irrigation was considerably less than planned, the releases did not compromise the agricultural sector. The precise timing and volume of ﬂows was derived through participation of local resource users. The return of freshwater ﬂows revised the ﬁshery particularly mullet, birds, such as white pelicans, and the growth of grasses (Sporobolus robustus) which re-vitalized traditional handicrafts such as mats made by local women (their main source of income). These environmental ﬂows led to a major revival in ﬁsh stocks. The total costs of the restoration over 12 years was around $US100 per hectare, whilst the added value of annual natural resource production is US$65 per hectare (Hamerlynck et al. 2006). The GEF-funded Senegal River Valley Water and Environmental Management Project commenced in 2003 to develop a basin-wide framework to integrate national water resources activities within an environmental action programme. Along with strengthening monitoring networks and building technical capacity, the project will assess the impacts of the seasonal ﬂoods from changes to the ﬂow regime due to the Manantali Dam. In 2002, the Governments of Mali, Senegal and Mauritania signed a Water Charter that guarantees an annual managed ﬂood (Article 14) and

minimal environmental ﬂows (Article 6), except under extra-ordinary circumstances. The objective of the Water Charter is to ‘provide for efﬁcient allocation of the waters of the Senegal River among many different sectors, such as domestic uses, urban and water supply, irrigation and agriculture, hydropower production, navigation, ﬁsheries, while paying attention to minimum stream ﬂows and other ecosystem services’.

16.10 Stakeholder Participation OMVS was established as a top-down organization driven largely by the Council of Ministers and High Commission. OMVS operates consultative bodies, such as the Permanent Water Commission, but these also comprise government ofﬁcials from member states and are charged with allocation of resources. There is no forum for local communities’ participation in the policies or implementation work of OMVS. Local resource users only had inﬂuence on decisions through participation in studies and workshops run by outsiders, such as IUCN – The World Conservation Union. The Water Charter extended stakeholder involvement within the Senegal basin to include farmers and NGOs. Further stakeholder participation was stimulated by the Global Environment Facility (GEF) project which included participation in its design and implementation. Now local co-ordination committees exist throughout all countries of the basin.

16.11 Conclusions The development of the Senegal River catchment in Sahelian west Africa is limited primarily by water availability. The international catchment authority, OMVS, with its membership of basin countries, provides an institutional framework for managing water resources of the transboundary Senegal basin. The principle followed is one of sharing the beneﬁts (electricity, irrigable land and navigation) rather than the water itself. It allows for transboundary co-ordination, such as

Catchment Management Case Study – Senegal River releasing water from the Manantali Dam in Mali to generate electricity, support a navigable river, supply irrigation and maintain river ecosystem services in Senegal and Mauritania. Environmental economics was a key tool in water allocation decisions. This led to environmental ﬂood releases from Manantali Dam to maintain the ﬂoodplain ecosystem at least in the short term; without this information the Manantali basin infrastructure would have been operated purely to serve power generation, intensive irrigation and navigation sectors. The catchment has a complex ethnic mix and a wide range of communities characterized by an urban elite and rural poor. Decisions made in the best interests of the catchment as a whole have important distributional impacts, with signiﬁcant winners and losers. References Acreman, M.C. (1996) Environmental effects of hydroelectric power generation in Africa and the potential for artiﬁcial ﬂoods. Water and Environmental Management, 10, 429–434. Acreman, M.C. (2003) Case Studies of Managed Flood Releases. Environmental Flow Assessment Part III. World Bank Water Resources and Environmental Management Best Practice Brief No 8. World Bank, Washington DC. Acreman, M.C. and Hollis, G.E. (1996) Water Management and Wetlands in Sub-Saharan Africa. IUCN, Gland, Switzerland. Acreman, M.C., Farquharson, F.A.K., McCartney, M.P. et al. (2000) Guidelines for Artiﬁcial Flood Releases from Reservoirs to Restore and Maintain Downstream Wetland Ecosystems and their Dependent Livelihoods: Final Report to DFID. Centre for Ecology and Hydrology, Wallingford, UK. BirdLife International (2007) BirdLife’s Online World Bird Database: the site for bird conservation. Version 2.1. BirdLife International, Cambridge, UK. http://www.birdlife.org (last veriﬁed 8 October 2008). Drijver, C.A. and van Wetten, J.C.J. (1992) Sahel Wetlands 2020. Centre for Environmental Science, Leiden, The Netherlands.

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Duvail, S. (2001) Scénarios Hydrologiques et Modèles de Développement en Avl d’un Grand Barrage. Les Usages de L’eau et le Partage des Ressources Dans la delta Mauritanien du Fleuve Sénégal. Unpublished PhD Thesis. Louis Pasteur University, Strasbourg. Goudie, A.S. (1977) Environmental Change. Clarendon Press, Oxford. Groupe d’Experts des Plaines d’Inondation Sahéliennes (GEPIS) (2000) Vers Une Gestion Durable des Plaines D’inondation Sahéliennes. IUCN, Gland, Switzerland. Hamerlynck, O. and Duvail, S. (2003) The Rehabilitation of the Delta of the Senegal River in Mauritania. IUCN Wetlands and Water Programme Blue Series. IUCN, Gland, Switzerland. http://data.iucn.org/ dbtw-wpd/edocs/WTL-029.pdf (last veriﬁed 18 November 2008). Hamerlynck, O., Duvail, S., Messaoud, B. and Benmergui, M. (2006) The restoration of the lower delta of the Senegal River, Mauritania (1993–2004). In: Symoens, J-J. (ed.), Coastal Ecosystems of West Africa, Biological Diversity-Resources-Conservation. Foundation for the Promotion of Scientiﬁc Research in Africa, Brussels, pp. 195–210. Hollis, G.E. (1996) Hydrological inputs to management policy for the Senegal River and its ﬂoodplains. In: Acreman, M.C. and Hollis, G.E. (eds), Hydrological Management and Wetlands in Sub-Saharan Africa. IUCN, Gland, Switzerland. Horowitz, M. and Salem-Murdock, F. (1990) Senegal River Basin Monitoring Activity Synthesis. Institute for Development Anthropology, Binghampton, New York. (OMVS/ISTI) Organization Pour la Mise en Valeur de Fleuve Senegal (OMVS) and International Science and Technology Institute (ISTI) (1990) Groundwater Monitoring Project, vol. II, Hydrogeological synthesis of the Senegal River Delta. OMVS, Dakar, Senegal. Sir Alexander Gibb and Partners (1987) Etude da la gestion des ouvrages communs de l’OMVS: Rapports Phase 1, vol. 1B. Optimization de la crue artiﬁcielle (rapport déﬁnitif). Organization Pour la Mise en Valeur de Fleuve Senegal (OMVS), Dakar, Senegal. White, F. (1983) The Vegetation of Africa: a descriptive memoir to accompany the UNESCO/AETFAT/ UNSO Vegetation Map of Africa (3 Plates, Northwestern Africa, Northeastern Africa, and Southern Africa, 1 : 5000,000). UNESCO, Paris.

Image facing chapter title page: Courtesy of the Centre for Ecology and Hydrology.

Laguna De Bay – A Tropical Lake Under Pressure

17

MARIA VICTORIA O. ESPALDON1 1

School of Environmental Science and Management, University of the Philippines Los Baños, Laguna, Philippines

17.1

Introduction

17.1.1 A brief description of Laguna de Bay The Laguna de Bay is the largest lake in the Philippines and one of the largest in Southeast Asia. It is a classic model of a multiple resource system, being situated in the rapidly industrializing and urbanizing region of the Philippines. It provides a resource for ﬁsheries, irrigation and power generation. It is situated within latitudes of 13 °55′ to 14 °50′ N and longitudes of 120 °50′ to 121 °45′ E in Luzon Islands, Philippines (Fig. 17.1). The Laguna de Bay, also known as the Laguna Lake, has a total surface area of approximately 900 km2 and a 220-km shoreline. The lake has an average depth of 2.5 m and a maximum water holding capacity of about 2.9 billion m3. As such, it is seen as a potential source of potable water supply over the next 10 years, even though its present water quality is an issue. The lake is divided into West, East, South and Central Bay. The lake straddles two provinces of Laguna and Rizal, and its watershed spreads in three provinces including Cavite, 12 cities and 49 municipalities; or a total of 2656 barangays or villages. The whole Laguna de Bay watershed including the lake is referred to as the Laguna lake basin or region. Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

The Laguna de Bay region is characterized by two distinct seasons: dry from November to April; and wet from May to October. However, some areas experience evenly distributed rainfall throughout the year. It is also notable that the lowest air temperatures and highest wind velocities occur from December to February causing high water turbulence and turbidity. As a consequence ﬁsh growth is limited even with an abundant supply of nutrients. In the last three decades of Laguna Lake management, the importance of anthropogenic activities in the watershed has been highlighted and with the magnitude of the lake’s watershed, integrated management is a difﬁcult task. The total land area of the watershed is approximately 2903 km2. Of this, 52% is agricultural land, 14% is open grassland, 5% is forested and 20% is built-up areas or zoned for industry (Fig. 17.2). The Laguna de Bay region is home to about 9% of the total population of the Philippines (approximately 6.6 million in 2000). The more densely populated municipalities are those located close to Metro Manila or the West Bay. The rest of the Laguna de Bay (East, South and Central Bay) are generally dominated by agricultural communities. Urban sprawl caused by its proximity to the country’s business and commercial capital, Metro Manila, has resulted in the urbanization and industrialization of the cities and municipalities along the West Bay. Currently expansion is moving towards the southern portion of the Laguna de Bay region. While this

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Fig. 17.1 Location map of Laguna de Bay.

promotes opportunity for employment in the commercial, industrial and even in the agroindustrial sectors, it has also highlighted the environmental pressures affecting the lake.

17.2

Environmental Concerns

To appreciate water pollution and associated water resource use concerns of Laguna de Bay

requires an understanding of the lake’s hydrodynamics. Three major factors that drive the lake’s hydraulics are the inﬂows to the lake from the surrounding watershed, tidal forcing from Manila Bay through Pasig/Napindan River, and surface wind stresses (Fig. 17.3). The presence of structures such as ﬁshpens and ﬁsh cages, bottom friction or bed sheer stress, the occurrence of high turbidity, and the occasional proliferation of ﬂoating macrophytes

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Laguna De Bay – A Tropical Lake Under Pressure 121 °0 ′0 ″E

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Fig. 17.2 The Laguna de Bay region (Lasco et al. 2005).

(Eichhornia) are some of the factors that contribute to lake circulation patterns (Lasco et al. 2005). The many rivers that drain into the lake carry contaminants from land-based anthropogenic activity located throughout the catchment. The lake’s location, therefore, makes it a sink for various kinds of water pollutants many of which degrade water quality. The lake’s major sources of pollution are domestic, industrial and agricul-

tural. These pollutants enter the lake through the 24 major tributaries, including the backﬂows from Pasig River. Pasig River is the only outlet of the lake to the Manila Bay and cuts through the densely populated and heavily congested part of Metro Manila (Lasco et al. 2005). As early as 1973, the Societe Grenobloise d’Etudes et d’Application Hydrauliques (a French consulting ﬁrm; SOGREAH 1974) reported that the nutrient levels in the lake were extremely

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Fig. 17.3 Laguna Lake hydrodynamics (Lasco et al. 2005).

high, (Villadolid 1933; Lasco et al. 2005), which was attributed to the high ﬁsh yield during that time. In the 1970s, of the estimated 5000 tonnes of nitrogen that drained into the lake from the watershed, 36% was from livestock and poultry production, 26% from domestic sources, 22% from the Pasig River backﬂow, 11% from the leaking of agrochemicals and 5% from industrial sources. By 2000, the Laguna de Bay waste model (Zafaralla et al. 2005) estimated that the majority of the load came from domestic sources (79%), agricultural activities (16.5%), industry (4.5%) and other sources (0.5%). The trend is not surprising because of the rapidly increasing human population living in the watershed, increased urbanization and industrialization and the decreasing amount of agricultural land. Increased urbanization has not been mirrored through supporting waste water infrastructure and management. Presently, some cities and municipalities in the watershed do not have centralized sewerage systems and still utilize an open canal system where wastes are directly dumped into rivers. This issue is currently being addressed as a high priority by the Laguna Lake Development Authority (LLDA), the government entity created by virtue of Republic Act (RA) 4850 in 1966 and designated to manage the lake and its watershed (Oposa 2002). In many of the lakes’ tributaries total coliform counts are in excess of the allowable limits of

5000 MPN 100 ml−1 set for Class C rivers (Lasco et al. 2005).* Additionally, the lake has high bacterial counts, though the inﬂow of saltwater from Manila Bay in the summer months reduces the coliform burden in the central portion of the lake. The Millennium Ecosystem Assessment of 2005 indicated that, based on the exceedance mean pollutant thresholds and water quality trends during the decade 1990–1999, there is a worsening of the condition of the lake. The key parameters are ammonia, nitrate, total N, orthophosphate, total P, chemical oxygen demand (COD), turbidity, chloride and hardness. Other parameters that have exceeded thresholds or showed erratic trends include water clarity, dissolved oxygen (DO), pH and alkalinity. Lake waters are also contaminated with toxic and hazardous substances such as heavy metals and persistent organic pollutants, including pesticides arising from industrial and agricultural sectors. The heavy metals lead (Pb), chromium (Cr), cadmium (Cd), copper (Cu), arsenic (As) and mercury (Hg) have been found to be in concentrations exceeding the prescribed safe levels for Class C waters and also for lake sediments Results from various studies highlight the * A Class C banding indicates that the water is only ﬁt as a ﬁshery and not for domestic or potable use; MPN is most probably number – this is an estimate used to measure the concentration of a target microbe in a sample through serial dilution tests.

Laguna De Bay – A Tropical Lake Under Pressure previous lack of a structured sampling regime and large spatial and temporal variability. Bioaccumulation of metals into the lake’s biota is a serious concern, especially given the importance of aquaculture to the region. Rapid shoaling of the lake is a major concern due to sedimentation and siltation of inorganic deposition from the watershed via its tributaries. The rapid denudation of the watersheds following conversion to cropland and urban areas has contributed to the sedimentation of the lake. Primary production in the lake has declined since the 1990s (3.8 g C m−2 day−1) with a marked decrease in algal biomass. Zooplankton communities also declined especially when ﬁshpens were widely distributed throughout the lake. Concurrently, lake ﬁshery production has declined but this could also be attributed to the perennially turbid lake water which prevents light penetration, reducing primary production. Indeed, records of the Bureau of Fisheries and Aquatic Resources (BFAR) and other agencies including the LLDA identiﬁed a rapid decline in ﬁshery productivity. From 1980 to 1996 there was a 64% decline in the production levels (Palma et al. 2005) and a concomitant decline in species diversity. Historically 33 species were found in the lake of which nine were indigenous. Presently four of these indigenous species have been lost. Migratory species of ﬁsh have also disappeared, and the dominant ﬁsh are now the aquaculture species, Oreochromis nilotica (tilapia) and Chanos chanos (bangus). Fishpens and ﬁsh cages for commercial rearing were the saviour of the ﬁshery sector. When the open ﬁshery declined due to deteriorating water quality, the advent of ﬁsh pens and ﬁsh cages in the early 1980s allowed the ﬁshery sector to increase production through supplementary feeding. The raising of tilapia and bangus in the middle of the lake where salt water intrusion occurs beneﬁtted from higher background levels of primary production. However, the pen and ﬁsh cages considerably diminished the area for open ﬁshery which resulted in conﬂicts between larger capital enterprises and traditional ﬁshing communities.

393

17.3 Historical Background to Laguna de Bay Fishing is the most important traditional use of the lake. Fishermen along the coast of the lake even refer to this huge body of water as ‘dagat’, which is the local term for an ocean or sea. Fishery activities consist of both open water capture, and in the 1980s, aquaculture. Fishery production shows decreasing trends. The overall ﬁsh production has declined from about 90,000 t in 1980 to about 38,000 t in 1996, with positive trends in 1982–1984, and between 1988 and 1993 (Palma et al. 2005). However, after 1993, a continuous decline is observed (Fig. 17.4). BFAR attributed the short-term increases in ﬁshery productivity to the expansion of ﬁsh pen technology during this period. The use of traditional ﬁshing gear remains a primary practice among ﬁshermen, especially those without capital to invest in ﬁsh pen or cages. This situation renders them very vulnerable to competition over ﬁshing area rights and to the deteriorating lake water quality (Palma et al. 2005). Fish pen culture with goby and carp was ﬁrst introduced as early as 1965 by the then Philippine Fisheries Commission. Expansion was rapid following the establishment of a LLDA demonstration pen of 0.38 km2 in Looc, Cardona in 1970. The development of the pen technology was implemented through the Laguna de Bay Fish Pen Development Program Funded jointly by the Asian Development Bank (ADB) and Organization of Petroleum Exporting Countries (OPEC). These ﬁsh pens were stocked at a density of 30,000– 60,000 bangus ﬁngerlings per hectare. Fry are purchased from specialist nursery growers who operate in the catchment. Harvesting is year round; and ﬁsh initially fed naturally, but artiﬁcial feeding became a more common practice when lake productivity started to decline. There are concerns though that artiﬁcial feeding is becoming another source of pollution. Cage culture, introduced later in 1974 by LLDA using tilapia, uses small-scale net cages which are more affordable to local ﬁshermen with modest capital.

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Fig. 17.4 Total ﬁsh production, 1980–1996 (Palma et al. 2005).

In 1983, with the success of ﬁsh pen technology, the lake was covered with ﬁsh pens to about 31,000 hectares which was way above the then estimated carrying capacity of the lake (21,000 hectares) (Fig. 17.5). This has sparked intense conﬂict between the ﬁsh pen operators and local ﬁshermen, resulting in violence and even death among ﬁshermen and security guards of the ﬁsh pens.

17.4 Evolution of Policies

Fig. 17.5 Map of Laguna de Bay showing ﬁshpen area, 1984 (Palma et al. 2005).

The establishment of the Laguna Lake Development Authority (LLDA) was as a response to the growing consciousness of both the national and local communities in the lake region for the need to deal with the growing problem of resource over-exploitation and environmental degradation of Laguna de Bay. Environmental planning and

Laguna De Bay – A Tropical Lake Under Pressure management (EPM) in the Laguna Lake region represents the very ﬁrst example of this type of institutional collaboration in the Philippines (Cardenas et al. 1987). RA 4850* (1966) laid the legal basis for EPM in the region through the creation of the LLDA. RA 4850 mandated the LLDA to ‘lead, promote, and accelerate the development and balanced growth of the Laguna de Bay area’. Technical assistance from the United Nations Development Program (UNDP) provided a more stable basis for EPM in the region. Between 1967 and 1977, the UNDP undertook several studies including the most controversial study on the ‘Feasibility Survey for the Hydraulic Control of the Laguna de Bay Complex and Related Development Activity’. The study was later relegated to LLDA in 1970, which served as the start up for the newly established agency. The various studies conducted identiﬁed regional development possibilities in agricultural diversiﬁcation, industrial development, mining, aggregate production and land reclamation. Major constraints that were identiﬁed are the: • need for a comprehensive evaluation of water quality and related ecosystems; • determination of the effects of potential developments on these systems; • evaluation of institutional management concepts that could relate to social and environmental issues. The studies recommended close examination of the programmes for lake ﬁshery, lake water quality, water supply, industrial estate planning and irrigation. Already during that earlier period, the main concerns included the following: • outdated Philippine regulations on water quality, and a weak, underfunded national pollution control agency; • a rapidly increasing population around the lakeshore without corresponding development of sewerage facilities; * RA 4850 is the act creating the Laguna Lake Development Authority as the central management entity of the Laguna Lake region.

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• an expansion of industrial zones contributing pollution to the lake; • increased nutrient loss from agricultural land to the lake; • the consequences of geothermal ﬂows of undeﬁned magnitude from hot springs in the southern end of the lake (Cardenas et al. 1987). RA 4850 was further amended through Presidential Decree (PD) 813 on 17 October 1976 to make the LLDA more institutionally robust in dealing with the major issues, many of which involved more than one sector. While PD 813 strengthened the charter of LLDA, Executive Order (EO) 927 dated December 1983 further improved the institutional capabilities of the LLDA to rationalize the allocation of resources. This enactment allowed the LLDA to modify and improve its organizational structure, extend its scope of jurisdiction to cover the whole lake region and the power to issue standards, rules and regulations pertaining to water pollution control. Today, LLDA stands out as an example of an institution that encourages the development of an integrated approach to lake management; and is guided by strategic objectives designed to reinforce its role as an integrated watershed management institution (Laguna Lake Development Authority 2004). As such, it promotes the use of ecosystem-based planning and management of lake resources that adopts the basin/sub-basin as the planning unit. It also adheres to the concept of participatory and partnership approach in planning and implementation that appreciates the signiﬁcant roles of the government, civil society, business enterprises and the local community to preserve the lake and its resources. To update the antiquated water quality management mechanisms, the re-engineered LLDA is set to adopt broadened market-based instruments that encourage more meaningful participation and partnership using environmental users’ fees and charges. The LLDA pursues the elusive goal of improving the lake water quality to a level that satisﬁes the competing demands for water use and other ecosystem services.

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maria victoria o. espaldon 17.5 Current Solutions to Address Environmental and Socio-Economic Concerns

17.5.1 Lake Fishery Management Program Indiscriminate proliferation of ﬁsh pen and cage structures has placed tremendous pressure on the carrying capacity of the lake with a steady recent increase in aquacultural production (Fig. 17.6). It was also a socio-political problem as ﬁsh pen occupancy by corporations and individuals often exceeded set limits because of their patterns of ownership (Cardenas et al. 1987). To address the intense conﬂict, the LLDA spearheaded the lake zoning and management plan (ZOMAP) which represents a democratization mechanism for allocating lake resources. The ZOMAP demarcated areas into open water ﬁshery, aquaculture (ﬁsh pen and ﬁsh cage),

transport and navigation, ﬁsh breeding and propagation. The ﬁsh pens and cages zones are arranged in concentric layers, following the shorelines, which allows for open ﬁshing, navigation and transport between layers. The ZOMAP developed in 1983 was further reﬁned through continued consultation with local ﬁshermen, ﬁsh pen operators and the scientiﬁc community. The revised ZOMAP considered updates on the natural carrying capacity of the lake, and the enactment of the 1991 Local Government Code (Laguna Lake Development Authority 1984). The revised ZOMAP became an integral part of the Master Plan in 1996, and in 1999, LLDA passed a Board Resolution No. 5 to transfer the unproductive areas to more productive sites but maintaining 10,000 hectares for ﬁsh pen and 5000 hectares for ﬁsh cages (Fig. 17.7). The implementation of the ZOMAP however remained a tough challenge for LLDA administrators who found it difﬁcult to dismantle ﬁsh

Fig. 17.6 Fish production from aquaculture, 1980–2002 (Palma et al. 2005).

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Fig. 17.7 The Laguna Lake Zoning and Management Map (ZOMAP) (from Laguna Lake Development Authority, Philippines, the regulatory body of Laguna de Bay).

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pens outside the zones because of the political inﬂuence of ﬁsh pen operators during their initial establishment in the 1980s. During the 1990s the LLDA made headway in the implementation of ZOMAP and maximized the opportunity for rezoning following typhoon damage to the ﬁsh cage structures. 17.5.2 The Environmental Users’ Fee The LLDA, in line with updating the mechanisms for water quality management, has pioneered a water pollution charge system in the Philippines. The Environmental User Fee System (EUFS) is an integral part of the LLDA’s Environmental Management Program. The EUFS is an attempt to use economic incentives to address environmental problems. This system operates on the ‘Polluter Pays Principle’ and aims to encourage the companies to invest in and operate pollution prevention and/or abatement systems. This system effects direct accountability for damage inﬂicted on the Laguna de Bay. It requires enterprises to obtain a Discharge Permit (DP) – a clearance or legal authorization from the LLDA to discharge liquid waste of a declared or speciﬁed concentration and volume into Laguna de Bay or any of its tributaries. Only establishments that meet the required efﬂuent standards set by the Department of Environment and Natural Resources Administrative order No. 35 Series of 1990 are given Discharge Permits. A report showed that there is reduction of biological oxygen demand (BOD) loading after 6 years of EUFS implementation (Laguna Lake Development Authority 2004). The total BOD waste load into the lake in 2000 is about 74.7 MT yr−1. Of this, about 19% came from industry and 11.5% from agriculture and forestry. The domestic source remains to be the biggest BOD contributor (68.6%). Based on these data, it can be said that EUFS as a mechanism for water quality protection can be very effective in regulating discharges from industry. However, it still falls short in terms of dealing with pollution originating from the set-

tlements and catchment.

population

centres

in

the

17.5.3 River Rehabilitation Program and Community Participation While the ZOMAP or EUFS are considered landmark institutional responses to the various problems in lake management, the formation of river rehabilitation councils anchored on individual watersheds as a management unit is strengthening the role of local communities in the lakes’ management. LLDA General Manager Carlos Tomboc, who has a strong watershed and community orientation, launched a comprehensive program to rehabilitate the 24 river systems ﬂowing into the lake. The rehabilitation programme initiated: • a comprehensive survey of the river systems gathering baseline information; • an education and motivation campaign; • the creation of an environmental army called ‘Hukbong Pangkapaligiran’; • organization of river rehabilitation councils and foundations to ensure multisectoral partnerships and stakeholder involvement; • sustainability monitoring of the rehabilitation efforts. This programme addressed, for the ﬁrst time, the issue of pollution loading from non-point sources. The formation of the river rehabilitation councils was motivated by shared concern among stakeholders (community members, local government ofﬁcials, non-government organizations and civic organizations) to reduce the pollution of the lake through reducing the pollution of its tributary rivers. At present, about 20 river councils are operational. The LLDA Board of Directors issued ofﬁcial pronouncements to institutionalize and extend recognition and support to these Community Based Organizations (CBOs) (Espaldon et al. 2005). Presently, the conﬁguration of the river rehabilitation councils varies. Some river councils are primarily managed and facilitated by the industry sector in areas of the Laguna Lake

Laguna De Bay – A Tropical Lake Under Pressure region where industrial parks and sectors are located. Others are led by Non Government Organizations (NGOs) and the scientiﬁc community, whilst others are led by Local Government Units (LGUs) and local farmers/ﬁshermen organizations. Their experiences in reducing pollution also vary – while some are making substantial headway, others are still sub-optimal. For example, the BISIG CATA (located in the western portion of the Bay with many urban areas and industry) has already met targets in terms of decreasing the river’s pollutant loading and has agreed targets for the future. According to members, the activity galvanized the community into taking deﬁnitive shared action to enhance opportunities for sharing knowledge. The LLDA organizes an annual Laguna Lake Conference that is open to all river rehabilitation councils, NGOs, LGUs, researchers and the public to enable the different sectors to present their views.

17.5.4 Laguna de Bay Shoreland Management The shoreline of Laguna de Bay spans 220 km with an estimated total shoreland area of 13,634 km2 based on lake level water drawdown from an elevation of 12.5 m to 10.5 m (Espaldon et al. 2005). RA 4850 deﬁnes the coverage of the lake up to 12.5 m elevation. This shoreland is the strip of land along the perimeter of the lake, which is exposed when the water level is low (about 10.5 m), and submerged when the water level rises (12.5 m). This area has been for many years subjected to illegal reclamation by human settlements, quarrying, domestic and industrial garbage dump sites, and agriculture. To address the present and potential threats arising from the competing use of shorelands of Laguna de Bay, the LLDA enforced the shoreland management programme which is designed to set up proper management of the use and occupancy of shoreland areas. The programme covers the inventory of titleholders, occupants, owners and

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claimants, with a corresponding information, education and communication programme. It attempted to set up an administrative system that guarantees the execution of the rights of legitimate shoreland users before setting or conducting activities on public land. Delineation was done, but some opposition was registered in parts of the lake’s perimeter. Rice farmers along the eastern bay in Laguna Province had evolved a system of use rights over the many years of rice cultivation in these drawdown areas, including special rice growing practices. Hence, the need for an inventory of ‘legitimate’ owners was received with reservation. To date, the implementation of the programme has only progressed slowly. In the northwestern part of the lake the consequences of the construction of the ﬂood control structure along the Taguig-Napindan area, the proliferation of dumpsites and the establishment of squatter camps along the lake’s perimeter are all issues not fully resolved by the shoreline management plan.

17.6 New Scientiﬁc Insights 17.6.1 Chemical contaminants The different types of pollutants that enter the lake come from industry (heavy metals and other inorganics), human settlements and urban centres (coliforms and other disease causing bacteria) and agriculture (pesticide and fertilizer residues), and all pose serious threats to human health and well-being. However, health risk studies, and responses to these potential risks, are limited. Analysis of heavy metals in water, sediments and the biota of Laguna de Bay show levels of heavy metals exceeding prescribed criteria for Class C waters. In particular, for cadium (Cd), copper (Cu), and lead (Pb) (LLDA 1996–1998; Madamba et al. 1994; Lasco et al. 2005). Levels of other metals (Cr, Ni, Zn and AS) present in lake water were low but concerns about biomagniﬁcation exist (Lasco et al. 2005).

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In some locations the heavy metal concentration of the lake’s sediment is still considered low (Lasco et al. 2005). Other studies, however, have indicated that the concentration of heavy metals on the sediments can be about 1000 times greater than the concentration in the water column (Lasco et al. 2005). Cr, Cd, Cu and Zn were all found in the muscle tissue of different ﬁsh species caught in Laguna Lake, but their concentrations were lower than in structural and intestinal tissue. Cr, Cd and Pb concentrations were highest in organ and structural tissue. Cu and Zn levels were highest in the intestinal tissue. Polyaromatic hydrocarbons (PAH) pose another threat to the lake and persistent organic contaminants have been found (Barril et al. 1995). Other pollutants such as oil and grease have been conﬁrmed in sediments and other toxic and hazardous chemicals such as pesticides and other industrial persistent organic chemicals have been identiﬁed in ﬁsh and other biota. However, the true extent of organic pollutant distribution and abundance needs to be more thoroughly assessed. This has potentially serious implications on the health of the local communities and consumers of ﬁshery products. Lasco et al. (2005) notes that very little is known of the extent of pollution in terms of the levels of toxic and hazardous substances such as heavy metals, polychlorinated biphenyls (PCBs) and polyaromatic hydrocardons in the water, sediments and biota. In summary, a key challenge for the integrated management of the lake must be to enhance the background of environmental knowledge both spatially and temporally. Future issues that need to be addressed are: • What is the extent of pollution with heavy metals and other toxic and hazardous substances in the lake? • What is the geography of the different types of pollution? • What are the implications for human health, ﬁsheries, and to other uses of the lake? • Are the policies directed at maintaining water quality effective in addressing this issue?

17.6.2 Alien invasive species, or species substitution As noted in an earlier section, the ﬁsh species diversity of the Laguna Lake was severely affected by the declining water quality and human activities. An inventory of ﬁsh species in the lake reported a total of 33 species, consisting of 14 indigenous (ﬁve of which are anadronomous), and 19 exotic or introduced species. Anadronomous ﬁsh totally disappeared in the 1990s and this is attributed to the blocking of the migration path by the Napindan Hydraulic Control Structure and the pollution of the Pasig River. Nineteen exotic species were introduced in the lake through direct stocking and by accident. Aquaculture led to the introduction of bangus, four species of tilapia, four species of carp, three species of gouramy and mosquito ﬁsh into the open lake water as cage escapees. ‘Janitor ﬁsh’ (Pterygoplichthys), an exotic aquarium ﬁsh, was inadvertently introduced to the lake, and is now affecting the local ecology of the lake and some of the major rivers like the Marikina River. The janitor ﬁsh causes extensive damage to ﬁshing gear and aquaculture structures, generates instability through burrowing through substrate and riverbanks, increasing water turbidity, and is also a voracious eater of algae competing with other fresh water species. The uncontrolled proliferation of this invasive species is not only a threat to the livelihood of 28,000 small-scale ﬁshing families, but also to the lake’s biodiversity (World Bank 2007a). Fishermen report that the janitor ﬁsh now make up nearly of 75% of the daily open water catch, suggesting that this ﬁsh is fast becoming the dominant species in the lake. Mapagpala, an alliance of ﬁshing communities from 29 municipalities around Laguna de Bay, highlighted that without direct intervention, tilapia numbers will continue to decline and janitor ﬁsh will dominate all catches. Reports of janitor ﬁsh weighing as much as 30 kg are now common. Scientists from the Department of Science and Technology (DOST) also agreed that the problem

Laguna De Bay – A Tropical Lake Under Pressure caused by janitor ﬁsh in the lake is an environmental disaster. Guerrero III, Executive Director of the Philippine Council for Aquatic and Marine Research and Development, noted there is evidence that the janitor ﬁsh has also become invasive in other countries, having become established in tropical and semi-tropical regions of North America, Puerto Rico, Malaysia and Indonesia. This ﬁsh introduced in the country possibly by hobbyists has escaped into local fresh waters (GMA News TV 2007). Several attempts have been conducted to explore economic uses of the janitor ﬁsh. One initiative is converting the ﬁsh into feeds for ducks, chickens and other livestock, while another is exploring the use of the skin of janitor ﬁsh as source of cheap leather to make wallets, watch straps, billfolds and key chains. The government through the LLDA, has also launched a programme to purchase janitor ﬁsh for ten pesos per kilo in order to reduce the population of the invasive species in the rivers and in the lake (in particular mature adult ﬁsh). The most recent and celebrated project involved high school students who extracted biofuel from the ﬁsh. They were able to produce 0.5 kg of biofuel for every 12 kg of janitor ﬁsh. A recent study on janitor ﬁsh found concentrations of Pb ranging from 0.06 to 0.19 mgkg−1 (wet weight). As, Cd, Cr and Hg are also detected in ﬁsh ﬂesh at low concentration during the wet season (Chavez et al. 2006). Total bacterial counts in the ﬁsh ﬂesh were also higher during the wet season than in the dry season. 17.6.3 Carbon ﬁnancing for ‘microwatershed’ development action plans Concern about the consequences of future climate variability and change is now being factored into integrated lake management planning. Laguna de Bay is the largest lake in the Philippines and the second largest in Southeast Asia, and it is also where the most intensiﬁed land use change is taking place. Given this, lake management planning has also examined opportunities for carbon sequestration at regional and community levels through agroforestry.

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Of the total basin area, the area of land suitable for forests is 73,000 ha, of which only 19,000 ha are actually forested. The remaining land has been turned over to grass and annual crops. The consequence of this has been increased erosion and siltation of the lake. Siltation is roughly estimated to be about 1.5 million cubic metres per year (Lasco and Pulhin 2007). Under the Biocarbon Fund of the World Bank Carbon Finance programme, the LLDA is expected to implement the Laguna de Bay Community Watershed Rehabilitation Project. This project consists of a set of small-scale community-based watershed rehabilitation sub-projects in the Laguna de Bay watershed which includes; • streambank rehabilitation; • reforestation in upland areas near the headwaters of key river; • the establishment of an agroforestry carbon credit scheme. Agroforestry is expected to be directly implemented by communities and provide a source of income in addition to that from farming. LLDA acts as the umbrella institution initiating the scheme which reduces the potentially high transaction costs associated with developing smallscale carbon ﬁnance activities. The maximum emission offset of small scale afforestation in agroforestry projects is estimated to be 8000 t CO2 yr−1. This equates to around 0.03 Mt CO2 by 2012 and around 0.05 Mt CO2 by 2017 (World Bank 2007b). It is expected that the reforestation, aside from earning the communities LLDA carbon credits, will help address land degradation, reducing siltation and landslides. The reforestation scheme enhances land and water conservation, ﬂood mitigation and enhanced biodiversity. Socio-economic beneﬁts of the scheme include livelihood improvement (through agroforestry), a reduction of localized erosion problems, improved ﬁsheries and reduced costs of treating water for domestic use, reduced ﬂooding and landslides, increased groundwater recharge, and improved agricultural production in the long term. The primary areas targeted for implementation are those with simple land management and ownership issues.

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Issues and Challenges

17.7.1 Local community participation in lake management Local community participation is at the heart of the integrated natural resource management strategy for a system like the Laguna de Bay. It is characterized as a multiple resource–multiple use ecosystem, hence more dynamic and complex than simply a river or lake. While the LLDA is considered as the main agency planning and implementing a management strategy for the Laguna de Bay resources, the power or authority for environmental governance is not monopolized by a formal institution. Time and again, the local communities of ﬁshermen, the biggest sector and the most traditional users of the lake, assert their signiﬁcant role in the management of the lake. The formation of the Hukbong Pangkapaligiran or environmental army is a testament to the interest and determination of local people. Their experience highlights that substantive people pressure can be mobilized to effectively hault or curb pollution. Partnerships between various groups has also been an effective means of mobilizing the local resources of the region. Industry, local government units, non-government organizations and CBOs are all active actors in the performance and success of the river rehabilitation councils. Currently this modality of lake management has been recognized by various sectors at various levels (local, national, regional and global). In fact, the community participation and partnerships with industry and other sectors were highlighted as a signiﬁcant contributor to the adoption of the Laguna de Bay as the eighteenth member of the International Living Lakes Network in August 2001. Whilst it is recognized that the integrated management of Laguna de Bay has progressed positively, there is a realization that wider stakeholder engagement and involvement especially with local ﬁshing communities and lake shore residents would enhance future environmental

governance of the lake. The LLDA has a key role to play in this regard. 17.7.2 Strengthening the science base to support management decision-making A previous section highlighted the paucity of background information, especially relating to heavy metals and other toxic and hazardous substances found in the lake water, sediments and biota. Continuous and regular monitoring of the input river systems and the lake itself are needed. Additionally an appropriate reporting system must be in place to determine the success of environmental management programme implementation. Studies on the impacts of the different types of pollutants in the lake biota, especially on ﬁsh, are still limited and inconclusive. Risks to human health should also be a priority not only for LLDA but also for other relevant government and non-government agencies. Background baseline information is required to assess the potential emergence of health issues arising from the intensiﬁcation of some types of pollution. While studies highlight that diffuse agricultural pollution is a major concern, there is no concerted and deliberate effort to address this issue. The increase in the livestock sector is another issue that has to be addressed. A recent study highlighted the enormous contribution of the poultry and livestock industry in Laguna Province to the eutrophication burden of the lake (Espaldon et al. 2007). Future inventories of this sector including small-scale backyard and commercial operations is required, though recent studies have been initiated to quantify such sectoral production (Dyer 2008; Espaldon et al. 2008). 17.7.3 The watershed or lake basin approach to lake management One of the signiﬁcant contributions of Laguna de Bay management experience is the promotion of the watershed as a planning and management unit. Lake management has identiﬁed that water quality issues are directly associated with human

Laguna De Bay – A Tropical Lake Under Pressure activities in the sub-basins draining into the lake. The re-positioning of the LLDA under the administration of the Department of Environment and Natural Resources (DENR) has now strengthened the LLDA mandate – of leading the planning and management of the Laguna de Bay region as a watershed or a basin. While Laguna de Bay has been the centre of debate, action and struggles over who should manage and how, and for whom the beneﬁt from its various ecosystem services accrue, it has also been a showcase of various mechanisms of management and partnership. These are being continuously challenged, and to a certain extent inspire others towards a philosophy of ‘integrated’ watershed management with which the lake has become synonymous.

References Barril, C.R., Tumlos, E.T., Bato, R.C. and Madamba, L.S.P. (1995) Water quality assessment and management of a hypertrophic lake, Laguna de Bay, Philippines: problems and strategies. Proceedings of the 6th International Conference on the Conservation and Management of Lakes, 23–27 October, Tsukuba, Japan. Cardenas, M.L., Francisco, F.R., Nepomuceno, D.N. and Espaldon, M.V.O. (1987) Assessment of Environmental Planning in Laguna Lake Region. Regional Development Dialogue, vol. 8, no. 3. United Nations Centre for Regional Development, Nagoya, Japan, pp. 100–131. Chavez, H.M., Casao E.A, Villanueva, E.P., Paras, M.P., Guinto M.C. and Mosqueda, M.B. (2006) Heavy metal and microbial analyses of Janitor ﬁsh (Pterygoplichthys spp.) in Laguna de Bay, Philippines. Journal of Environmental Science and Management, 9, 31–40. Dyer, R. (2008) Livestock and Water Pollution in Asia: analysis and modeling of causes, impacts and policy options in the Philippines. PhD thesis, University of Aberdeen, Aberdeen, UK. Espaldon, M.V.O., Nepomuceno, D.N. and Dizon, J.T. (2005) Institutional arrangements, social conﬂicts and ecosystem trends. In: Lasco, R.D. and Espaldon, V.O. (eds), Ecosystems and People. The Philippines Millennium Ecosystem Assessment (MA) Sub Global Assessment. Environmental Forestry Programme,

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College of Forestry and Natural Resources, University of the Philippines Los Banos, Laguna, pp. 189–199. Espaldon, V.O., Alcantara, A.J., Valdez, C. et al. (2008) GIS-aided Animal Production Impacts Analysis on the Environment in Laguna Province. 11(2), 42–47. University of the Philippines Los Banos, College, Laguna Espaldon, M.V.O., Alcantara, A.J., Sevilla, C. S., Paraso, M. and Valdez, C. (2007) A GIS-aided study on Environmental Animal Health and Production in Laguna Province, Philippines. University of the Philippines Los Banos, Bureau of Animal Industry and Food and Agriculture Organization, Manila. GMA News TV (2007) Fishermen vs. ‘nuisance’ Janitor Fish, 25 November 2007. http://www.gmanews.tv/ story/64717/Fishermen-declare-war-vs-nuisancejanitor-ﬁsh (last veriﬁed 10 October 2008). Laguna Lake Development Authority (LLDA) (1984) http://www.llda.gov.ph/zomap.htm (last veriﬁed 30 November 2008). Laguna Lake Development Authority (1996–1998) http://www.llda.gov.ph/ (last veriﬁed 30 November 2008). Laguna Lake Development Authority (LLDA) (2004) Laguna de Ba’i: The Living Lake. LLDA, Pasig, Philippines. Lasco, R.D., Espaldon, V.O., Tapia, M. (eds) (2005) Ecosystems and People: The Philippines Millennium Ecosystem Assessment (MA) Sub Global Assessment. Environmental Forestry Programme, College of Forestry and Natural Resources, University of the Philippines Los Banos, Laguna, Philippines. Lasco, R.D. and Pulhin, F.M. (2007) Laguna Lake Basin and the Sierra Madre Community Forests, the Philippines, Case Study 8. World Agroforestry Centre, Laguna, Philippines. http://www.worldagroforestry.org/downloads/publications/PDFs/bc06222. pdf (last veriﬁed 22 October 2008). Madamba, L.S.P. and Pamulaklakin, M.A. (1994) Heavy metals in selected ﬁsh species collected from Laguna de Bay. Philippine Journal of Science, 123, 135–146. Millennium Ecosystem Assessment (MA) (2005) http:// www.millenniumassessment.org/en/index.aspx (last veriﬁed 8 October 2008). Oposa, A.A. (2002) A Legal Arsenal for the Philippine Environment. Batas Kalikasan, Philippines, pp. 445–460. Palma, A.L., Mercene E.C. and Goss, M. R. (2005) Fish. In: Lasco, R.D. and Espaldon, V.O. (eds), Ecosystems and People: The Philippines Millennium Ecosystem Assessment (MA) Sub Global Assessment.

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Environmental Forestry Programme, College of Forestry and Natural Resources, University of the Philippines Los Banos, Laguna, Philippines, pp. 117–132. Societe Grenobloise d’Etudes et d’Application Hydrauliques (SOGREAH) (1974) Laguna de Bay Water Resources Development Study. United Nations Development Programme and Asian Development Bank, Manila, Philippines. Villadolid, D. (1933) Some causes of depletion of certain ﬁshery resources of Laguna de Bay. National Applied Scientiﬁc Bulletin, 3, 251–256. World Bank (2007a) The State and Trends of Carbon Markets 2007. The World Bank Institute and

International Emission Trading Association, Washington DC. World Bank (2007b) Catalyzing Markets for Climate Protection and Sustainable Development. The World Bank Carbon Finance Unit. http://www.carbonﬁnance.org (last veriﬁed 8 October 2008). Zafaralla, M.T., Barril, C.B., Santos-Borja A.C. et al. (2005) Water resources. In: Lasco, R.D. and Espaldon, V.O. (eds), Ecosystems and People: The Philippines Millennium Ecosystem Assessment (MA) Sub Global Assessment. Environmental Forestry Programme, College of Forestry and Natural Resources, University of the Philippines Los Banos, Laguna, Philippines, pp. 65–114.

Image facing chapter title page: Courtesy of School of Environmental Science & Management/Maria Victoria Espaldon.

18 Chesapeake Bay Catchment Management – Lessons Learned from a Collaborative, Science-based Approach to Water Quality Restoration TOM SIMPSON1 1

18.1

Water Stewardship, Inc., Severn Avenue, Annapolis, Maryland, USA

Introduction

The Chesapeake Bay is the largest and, historically, most productive estuary in the USA. The bay is long and narrow being about 250 km long and 25–50 km wide. The estuary is very shallow with an average depth of only 7 m but it contains a narrow, deep trench up to 60 m deep that is the drowned river valley of the Susquehanna River. What is remarkable about the Chesapeake Bay is the very large size of its catchment area to its relatively small water volume. The catchment is approximately 165,000 km2 in size (Fig. 18.1). The bay has a catchment area to water volume ratio of about 2400 : 1, which is six times greater than the estuary with the next largest catchment area to water volume ratio, the Bay of Finland. The Chesapeake Bay has been in some stage of decline for more than 200 years with sedimentation, wetlands loss, ﬁsheries over-harvesting, and toxicant and pathogen pollution as principal causes of decline prior to 1950. However, during the 1960s and 1970s, widespread loss of submerged aquatic vegetation (SAV) and increased Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

evidence of hypoxic and anoxic zones in the summer became apparent. The US Congress authorized a major research effort in the mid 1970s to determine the cause of this accelerated decline in habitat and living resources in the bay. When this research was completed, it became evident that nutrient over-enrichment was the principal cause of the more recent systemic decline (United States Environment Protection Agency 1999). Nutrient over-enrichment, or eutrophication, causes excessive algal growth which results in areas of low to no oxygen in most deep waters and some shallow creeks and rivers from May through September. This effectively eliminates the cooler, deeper waters as warm weather habitat for ﬁnﬁsh and shellﬁsh and makes survival of benthic organisms difﬁcult. In shallow tidal rivers and creeks, low oxygen is responsible for many reported ﬁsh kills, particularly during spring. The excessive algal concentrations also impair clarity in shallow water, and along with suspended sediments, are responsible for turbidity that has resulted in the loss of most of the underwater grasses in the tidal shallows (0.5– 2.0 m depths). These grasses are critical habitat for many species of ﬁnﬁsh and crabs; serve as ﬁlters to improve water quality and clarity; and,

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Fig. 18.1 The Chesapeake Bay Watershed.

buffer shorelines from wave action (United States Environment Protection Agency 2003). Anthropogenic nitrogen and phosphorus loads to the Chesapeake Bay are estimated to be more than six times pre-settlement loads. Increased nutrient loads have resulted from all human activities in the watershed. The relative contribution of different sources to nutrient loads has been estimated using a model-based approach (Fig. 18.2). Agriculture, while occupying less than 25% of the land area, is estimated to contribute about 40% of the nitrogen and phosphorus. Some reductions in nutrient losses from agriculture have occurred due to implementation of pollution control measures, or Best Management Practices (BMPs). However, as discussed later, it is difﬁcult to assess the actual impact of these BMPs, so model estimates have been used. Wastewater discharges, primarily from sewage treatment plants, account for about 20% of the nitrogen and phosphorus reaching Chesapeake Bay annually. However, their phosphorus loads have been cut by more than 50% in the last 20 years, due to both a phosphate detergent ban and increased removal during wastewater treatment. Many sewage treatment plants in the catchment are being upgraded for nitrogen removal and, by 2015, most treatment plants will have advanced

nitrogen removal. Atmospheric deposition is an important source of nitrogen to the Chesapeake Bay. It is estimated that nearly 20% of the nitrogen delivered to the bay from the catchment is atmospheric nitrogen that passes through urban and forest land uses and that about 6% of the agricultural contribution is from atmospheric deposition, dominantly ammonia from animal production operations. Developed lands currently account for less than a quarter of the nitrogen and phosphorus reaching the bay. Unlike agriculture and sewage treatment plants, however, loads from developed lands are increasing over time as population and development increase rapidly within the catchment. The growth in population will also make it difﬁcult to maintain the reduced loads from sewage treatment plants over time as the plants will be operating at the current limits of technology for nutrient removal.

18.2 History of the Chesapeake Bay Program In 1983, the states of Virginia, Maryland and Pennsylvania, the District of Columbia, the US government and the tri-state Chesapeake Bay Legislative Commission signed the ﬁrst Chesapeake Bay Agreement (United States

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Natural 1%

Manure 16%

Air Deposition to Tidal Water 7%

Fertilizer (Agriculture) 15%

Manure & Fertilizer Air Deposition 6%

Animal Confinment 2%

Air Deposition to Watershed 19% Municipal & Industrial Wastewater 20%

Fertilizer (Nonagriculture) 10%

Phosphorus 2003 Natural 3%

Septic 4%

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Atmosheric Deposition to Tidal Water 7% Municipal & Industrial Wastewater 22%

Manure (Agriculture) 23% Chemical Fertilizer (Agriculture) 17% Fertilizer (Nonagriculture) and Stream Scouring 26%

Animal Confinrment 2%

Fig. 18.2 Nutrient sources to Chesapeake Bay. (Source: EPA, WSM 2004.)

Environment Protection Agency 1983), committing in very general terms to work together to reduce nutrient pollution through the state and federal partnership, the Chesapeake Bay Program (CBP). Since 1983, two additional agreements and one major amendment have been signed that guide bay restoration and management policies and activities. In 1987, the second Chesapeake Bay Agreement established the CBP’s goal to reduce the amount of nutrients, primarily nitrogen and phosphorous, that enter the bay by 40% from 1985 levels by the year 2000 (United States Environment Protection Agency 1987). The agreement continued cooperative efforts to restore and protect the bay, further committed partners to speciﬁc restoration actions and declared that implementation of these commitments would be reviewed annually with additional commitments developed as needed. The new agreement also contained goals and priority commitments for: living resources; water quality; population growth and development; public

access; governance; and public information, education and participation; but the 40% nutrient reduction goal dominated future activities. The ‘40%’ reduction was applied only to controllable sources from within signatory jurisdictions and did not initially include atmospheric deposition. As a result, the goal was actually a reduction of about 23% of the total nitrogen and phosphorus loads to the bay. In 1992, the 1987 Chesapeake Bay Agreement was amended to require jurisdictions to develop sub-basin-speciﬁc nutrient reduction strategies, called tributary strategies, designed as riverspeciﬁc clean-up plans for reducing nutrient levels in each sub-basin and the bay itself by 40%. The amendment also required tributaryspeciﬁc strategies to maintain the capped loading goals beyond 2000 in the face of growth and development. The state of Maryland used local agricultural tributary teams to develop agricultural strategies based on estimated levels and impacts of BMP

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implementation. A ‘technical options team’ developed BMP efﬁciency estimates based on a 1990 Chesapeake Bay Program BMP report, the scientiﬁc literature and expert consensus. These were used by the local agricultural tributary teams in each of Maryland’s ten sub-basins to develop strategies appropriate for, and implementable in, their basin. The agricultural strategies were combined with state developed stormwater and wastewater treatment plant strategies and were estimated, based on BMP efﬁciencies and model projections, to achieve the 40% reduction if fully implemented. No post2000 capped load management strategies were developed. The BMP deﬁnitions and efﬁciencies were adapted for use by all signatories to the bay agreement for estimating impacts of all nonpoint sources in 1995. The Agricultural Tributary Teams were very successful in engaging the local agricultural community in development of the river basin tributary strategies. As a result, local stakeholders took ownership of the strategies and maintained interest in their implementation. This approach was so successful that the state of Maryland decided to create Tributary Teams, modelled on the Agricultural Teams, in each of its tributary basins. The teams are composed of local government ofﬁcials, farmers, environmentalists, civic organization representatives, business and industry representatives and interested citizens. These teams have helped raise awareness of local water quality issues and served as advisors to the state on Tributary Strategy implementation in their basins for the last 12 years. In June 2000, CBP partners signed Chesapeake 2000, an agreement intended to guide restoration activities throughout the bay catchment for the future (United States Environment Protection Agency 2000). In addition to identifying key measures necessary to restore the bay, Chesapeake 2000 provided the opportunity for Delaware, New York, and West Virginia to become more involved in the bay programme partnership by working to reduce nutrients and sediment ﬂowing into rivers from their jurisdictions. In 2001, New York, Delaware and West Virginia committed to

the Chesapeake 2000 water quality goals. As a result, the entire catchment area became part of the restoration effort for the ﬁrst time, almost 20 years after the ﬁrst agreement. The new agreement committed the CBP to reducing total nutrient loads by nearly 50% from all sources in the entire watershed, or more than twice the reduction needed to achieve the ‘40%’ controllable load agreed upon in 1987. The 2000 agreement committed to reducing nutrient and sediment loads sufﬁciently to remove all related water quality impairments in the bay, as required by the US ‘Clean Water Act’, with a goal of accomplishing this by 2010. In 2003, CBP partners agreed to nutrient loading limits of no more than 80 million kilograms of nitrogen and 6 million kilograms of phosphorus to the bay annually (Fig. 18.3). Partners also agreed to reduce land-based sediment loads so that no more than 3.75 million metric tons would be delivered to the bay. These were considered capped loads that had to be maintained, once achieved, in the face of subsequent growth. These reductions in nutrients and sediment are expected to produce the improved water quality conditions necessary to support the living resources of the bay and needed to meet the requirements of the Clean Water Act. If sufﬁcient progress towards these goals is not achieved by 2011, a federally mandated ‘Total Maximum Daily Load’ (TMDL) and associated implementation plan will have to be developed. The Tributary Strategies that States have developed to achieve the Chesapeake 2000 goals will likely serve as the starting point for development of the TMDL implementation plan. At the agreed-upon reductions, the Chesapeake Bay water quality model predicts the existence of a bay similar to that in the 1950s. Water quality conditions necessary to protect aquatic living resources will be met in 96% of the bay within designated use requirements, and the remaining 4% will not fully achieve water quality conditions for only four months a year. Between 1985 and 2002, annual phosphorus and nitrogen loads delivered to the bay from the entire catchment were estimated to have declined

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350 300 250 200 150 100 50 0 1985

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2003

25 20

2010 Phosphorus Goal

15 10 5 0 1985

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7

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6 5 4 3 2 1

Source: CBP Phase 4.3 Watershed Model.

0 1985

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Fig. 18.3 The 2010 pollutant reduction goals are to reduce annual nitrogen loads to no more than 175 million pounds, annual phosphorus loads to no more than 12.8 million pounds and land-based sediment loads to no more than 4.15 million tons. (Source: CBP Phase 4.3 Watershed Model.)

by 3.5 million kilograms and 27.3 million kilograms, respectively, based on watershed model projections. Monitoring data at river input stations showed signiﬁcant reductions on most rivers but not as much as projected by modelling. Tidal monitoring in the bay’s mainstream showed no signiﬁcant change. The reductions obtained between 1985 and 2002 included offsetting signiﬁcant increases due to population growth. Maintaining reduced nutrient and sediment levels into the future will be a challenge because of the expected growth in human population and shifts in animal manure nutrients and the land available for application of those nutrients. In 2005, tributary strategies were completed by jurisdictions to implement the agreed reductions. Nutrient loads were allocated to nine major tributary basins that were further divided into 16 major tributary basins by jurisdiction, then into 37 jurisdictionally deﬁned tributary strategy sub-basins (Fig. 18.4). As in 1993, the 2003 tributary strategy process began with a strong emphasis on stakeholder involvement but these goals were much more challenging and many of the more easily implemented practices were in place. As a result, stakeholders could not agree on an implementation strategy that model results indicated could achieve the proposed

nutrient loading goals. Jurisdictions then developed strategies that were estimated to achieve the loading goals based on BMP efﬁciency estimates and watershed model output. These strategies required near complete implementation of many existing practices and incorporated numerous new BMPs. Nearly all jurisdictions also required major additional nutrient reductions from point-source discharges. The non-pointsource strategies focused heavily on agricultural BMPs since they are more cost effective than urban BMPs.

18.3 Goal Setting and Nutrient Reduction Load Allocations One of the hallmarks of the Chesapeake Bay Program has been a willingness to set challenging quantitative goals with deadlines for implementation of the practices, programmes and policies that will achieve these goals. The 1987 Bay Agreement committed to a 40% reduction in non-point source and point source loads from the signatory jurisdictions of Pennsylvania, Maryland, Virginia and the District of Columbia by 2000 from a 1985 base year. In 1992, this agreement was amended to require the development of

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Fig. 18.4 Nutrient loading allocation approach: (a) by nine major tributary basins; (b) by 16 major tributary basins by jurisdiction; and (c) by 37 state-deﬁned tributary strategy sub-basins.

basin-speciﬁc strategies to accomplish these goals in each of the major river basins leading to the Chesapeake Bay. While these goals were not achieved by 2000, signiﬁcant progress towards the 40% nutrient reduction goal was made through agricultural BMP implementation and urban point source controls (largely wastewater treatment plants). Progress was estimated based on reported practice implementation and associated reduction efﬁciencies as projected using the Chesapeake Bay Watershed Model (WSM), a state-of-the art model based on the Hydrologic Simulation Program – FORTRAN (HSPF) model. It was estimated that about 75% of the phosphorus goals would be met but that the nitrogen controls would only achieve about 50% of reductions needed to reach the target loads. Further, these were model estimates. Actual monitoring data at the fall line on major rivers showed some progress but not as much as the WSM projected. However, data in the tidal portions of the same rivers and the main stem of the bay showed little or no improvement. While the apparent discrepancies are still not

completely understood, nutrient transport, lag times from ground water, in-stream and tidal nutrient processes, model limitations, and overly optimistic assumptions regarding BMP performance are widely considered the principal causes of model versus monitoring discrepancies. The last two factors are discussed in detail later in this chapter. The 1987 goals were essential to provide direction and guidance to jurisdictional funding, programmes and policies needed in support of the strategies. The original ‘40%’ nutrient reduction goal was expected to reduce anoxic-volume days in the bay by about half. It was never intended to restore the bay to an unimpaired status but was considered a realistic and potentially achievable goal that would provide substantial improvements in water quality within the realm of political will and acceptability. The goal helped to focus Tributary Strategy development and implementation, as well as becoming a credible end point on which to garner public support, political will and implementation support and funding.

Chesapeake Bay Catchment Management By contrast, the Chesapeake 2000 Agreement committed to remove all nutrient impairments to the Chesapeake Bay to avoid a regulatory TMDL. It took 3 years to determine the loading cap needed to remove impairments. The load reductions required additional progress that was more than double what had been accomplished, based on model estimates, from 1987 to 2000. This was even more challenging since many of the more easily achieved reductions had been implemented to make the progress noted in 2000. The Chesapeake 2000 goal may have been so challenging that actually achieving it could require implementing control measures that require systems changes which could have substantial economic and cultural impacts. The 2005 state tributary strategies required higher levels of implementation of control measures than are likely to occur with current funding, political support and voluntary participation of major stakeholders, particularly the farm community. Nutrient and sediment goals were based on the best scientiﬁc estimates of loads needed to remove impairments to living resources. Lesser goals will not remove the nutrient impairments as required by the Clean Water Act. However, it has proved difﬁcult to maintain current rates of progress and implementation when there is relatively widespread recognition that the strategies designed to meet the goals cannot and will not be implemented as currently written. Efforts are underway to re-evaluate the goals and also to look at longer-term strategies to achieve the needed reductions that are economically and politically feasible. It is anticipated that the Tributary strategies will be adapted to become a TMDL implementation plan by or shortly after 2011. Goal setting has been critical to maintaining the commitment of the different jurisdictions to attempt to accomplish these challenges. Experience in the Chesapeake Bay catchment suggests that quantitative goals are an essential driver of public and private actions in any water quality restoration programme. In addition, allocating the loading caps or nutrient reductions to each of

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the signatory jurisdictions is essential to provide a base upon which they can develop their strategy. The allocation process was kept as technically based as possible but, of necessity, was largely driven by model-based estimates. These estimates were then presented to policy level managers for ﬁnal negotiation. In 1993, the three original Bay States and the District of Columbia each agreed to take on a certain load reduction. In 2003, all six Bay states and the District of Columbia reached agreement on individual loading cap allocations. While this was clearly a mixture of science and policy negotiation, assigning loads by jurisdiction and gaining the commitment of each jurisdiction to achieve their portion of the goal is essential to coming close to accomplishing the goal.

18.4 Tributary Strategies In 1992, and again in 2003, the signatories to the Chesapeake Bay Agreement committed to develop tributary speciﬁc strategies that would achieve their needed nutrient reduction based on agreed upon jurisdictional load allocations. These strategies required implementation of control measures that would reduce loads from sewage treatment plants, agriculture, urban stormwater, forest management and for 2003, atmospheric deposition of nitrogen. The strategies are based on prescribed levels of implementation of a wide range of speciﬁc control measures and BMPs in speciﬁc sub-basins of the Chesapeake catchment. Estimates of the effectiveness of various practices, systems and approaches in reducing nutrient pollution were developed for each of the control measures in the strategies. These were based on the best available scientiﬁc information. As discussed below, the reduction efﬁciency estimates developed during 1992 to 1993 were based on limited available research results and did not account for spatial or management variability that might occur when these control measures were applied at a watershed scale. The ramiﬁcations of this are discussed below. In

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Fig. 18.5 Progress in agricultural pollution control.

2003, reduction estimates were revised to reﬂect our increasing understanding that research level efﬁciencies likely overestimated watershed scale implementation of practices and that tracking, reporting, operation and maintenance of the control measures had also likely been overestimated. These changes created both a policy and public reaction since they reduced the model-estimated impact of the control measures that had been implemented through 2003. A reduction in model-estimated progress was reported for 2003 (Fig. 18.5). This was difﬁcult for policymakers to explain since they had annually been reporting on the positive progress based on model results. It was also difﬁcult for the public, particularly farmers, to have it suggested that the amount of progress due to their efforts was actually less than had been previously

reported. The lesson that was learned from this activity was that there had been an over-reliance on model output as an estimator of progress in water quality improvement with a concurrent under-reliance on actual monitoring data that must eventually show the real impacts of our actions.

18.5 Modelling and Monitoring to Support Watershed Restoration Monitoring and modelling are both essential to development, implementation and evaluation of catchment restoration strategies. The appropriate role and relative reliance on each should be determined early in the restoration programme. Modelling is essential to allow prediction of the impact of nutrient reduction strategies

Chesapeake Bay Catchment Management and actions on current and future land use, population and agricultural conditions. It allows for the development of ‘what if’ scenarios that can enhance development of cost-effective strategies using practices, programmes and approaches that are more politically, socially and economically acceptable. The Chesapeake Bay Program has chosen to rely on two very large coupled models to simulate conditions in the watershed and in the Bay. The estuarine-based Chesapeake Bay Water Quality Model estimates changes in water quality and living resources that would result from reductions in riverine nutrient loads to tidal waters. Riverine nutrient loads are delivered to the Water Quality Model using the Chesapeake Bay Watershed Model (WSM), which estimates nutrient loads from sub-basins and different land uses throughout the catchment under various land and nutrient management scenarios. The WSM is calibrated at the river input monitoring stations of all the major rivers and at other calibration stations within the watershed. As a result, the above-fall line nutrient loads are accurately reﬂected in the model; however, there are continual reﬁnements and upgrades to the model to better distribute those loads within the watersheds and to simulate the impact of management actions. The CBP relies on one very large watershed model that is based on the HSPF model. The HSPF model is widely used to estimate nutrient processes and loads typically from much smaller watersheds than the Chesapeake Bay. The HSPF has been continually reﬁned and updated to improve simulation of nutrient and sediment delivery from the catchment. Phase 5 of the WSM was calibrated for use in 2007. While it is clear that each new version of the model is a substantial improvement over the previous one, it remains challenging to accurately simulate land use conditions, changes and the impacts of control measures throughout a 165,000-km2 watershed using one model. The CBP concluded, and independent reviewers veriﬁed, that the current version of the WSM is as detailed and reﬁned a simulation of the watershed as is possible based on available

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data and our ability to deﬁne and simulate land uses in land-river segments using HSPF. There is much discussion about whether to continue to reﬁne the HSPF model for further application in the Chesapeake Bay Watershed. Some recommend looking for other types of models for the entire watershed that may be used in concert with HSPF to enhance our simulation capabilities, while others suggest different models that best simulate land uses, conditions and processes in different parts of the watershed to develop more accurate ‘representative’ simulated sub-watersheds. Caution must be exercised when using models to estimate restoration progress. The Chesapeake Bay Program began running annual ‘Tributary Strategy Implementation Progress Scenarios’ in the mid 1990s. These were used as estimators of progress to determine the anticipated future impact of control measures, since it was thought that lag times from surface and ground water processing and other factors were delaying seeing progress in monitored water quality. While there was some validity in lag time assumptions (and the lack of them in model output), overoptimistic assumptions about BMP implementation, operation and performance appears to have resulted in over-optimistic model-based progress estimates. However, for many years, annual progress in reducing nutrient pollution and restoring the Chesapeake Bay was reported by state and federal political ofﬁcials based on modelled BMP implementation impacts. In 2003, the Washington Post did a series of investigative articles entitled ‘Bay Progress Overstated’ that provided a rather sensationalized perspective on the misuse of model information in estimating progress in bay restoration. Bay Program management had been advised of these issues for several years. The media publicity resulted in several Congressional and Executive Branch investigations into progress estimation for the Chesapeake. What was found was that a sincere attempt to provide simplistic estimates of progress based on model output had unintentionally become perceived and used by ofﬁcials as if it were actual progress.

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The programme reviews also pointed out the need to separate ‘implementation’ progress from ‘restoration’ progress. Strategies, with implementation plans, were essential to successful catchment restoration. The ability to show progress in implementation of these strategies is equally critical to maintain public and political will. It is apparent that scientiﬁc uncertainties about BMP efﬁciencies, large spatial variability in BMP performance, management, and programmatic uncertainty about actual BMP implementation and operation and maintenance make estimation of the actual impact of practice implementation on water quality extremely difﬁcult. As a result, the CBP now uses model-based BMP impacts as estimates of ‘implementation progress’ and relies on monitoring to estimate water quality improvement. Model results are used to estimate what proportion of the potential nutrient reducing impacts of implementing an adopted strategy has been achieved based on the amount and robustness of practices implemented. This is no longer directly linked to suggested water quality response since it has been determined that monitoring is ultimately the only way to determine actual impact of practice, programme and policy implementation. The WSM continues to be a critical tool to develop strategies, evaluate new practices, programme and policies and predict the impact of future land use, population or agricultural changes. The WSM is no longer used as a substitute for monitoring in estimating actual progress in achieving bay water quality goals. The Chesapeake Bay Program implemented an extensive river-input and tidal water monitoring programme in the mid 1980s. This has provided critical information for calibration of both the water quality and watershed models and allows us to observe long-term trends in nutrients related both to control measures and climatic variation. Since 1985, many other monitoring stations have been added to enhance calibration capability for the watershed model. This was particularly important to allow calibration of the model at jurisdictional boundaries so that loads and load reduction allocations could be made. The monitoring network has provided extremely

important trend information over the last two decades; however, as stated above, from the mid 1990s until 2006 progress was rather assessed on modelled loads rather than on monitored loads. This was justiﬁed based on information that suggested signiﬁcant lag times in observing the impact of actions implemented on the land. It has become apparent that progress is being overestimated using the model and that monitoring offers the only real tool that can be used to determine whether or not water quality is really improving. Extreme inter-annual variability due to climate makes trends in water quality extremely difﬁcult to detect over anything but a very long sampling period. In an attempt to dampen the effect of inter-annual variability, the monitoring data has been converted to loads based on average hydrology. While this may be the only realistic way to compare loads over time given the huge variability, it can also result in monitoring information that does not seem logical to the public or policy-makers since they, and the living resources of the bay, see actual water quality conditions, not the average or ﬂowadjusted hydrology. Efforts are being made to use ﬂow-adjusted hydrology to determine long-term trends while reporting near real-time monitoring that reﬂects the actual climatic conditions that were observed during the past water year. Long-term monitoring data for the Chesapeake Bay indicate statistically signiﬁcant reductions in nitrogen and phosphorus have occurred at the majority of river-input stations (Fig. 18.6) but reductions are less than predicted by the WSM based on the estimated impact of control measure implementation. It is unclear why phosphorus shows some signs of continuing to increase in the lower rivers of the bay. Evaluation of long-term monitoring data also suggests that the rate of reduction may have slowed in recent years. It is unclear if this is due to increases in load due to population growth and development, higher loads due to high precipitation years in recent years or increasing difﬁculty in continuing progress and reducing nutrient pollution load sources. While some progress in reduction has been apparent at river-input sta-

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Fig. 18.6 Flow adjusted phosphorus concentration trends entering Chesapeake Bay.

tions, limited reductions have been observed at the mainstem tidal monitoring stations. It is unclear if this is simply an artifact of very slow processing of nutrients within the tidal system or if load increases from below the fall line (below the river-input stations) have been sufﬁcient to offset reductions observed at the river-input station. Even if progress at the river-input station is used as a gauge of success, however, it is far below what would be needed to remove the nutrient impairments to the bay and substantially below what has been estimated based on model projections of control measure implementation. Both modelling and monitoring are essential to the Chesapeake Bay Restoration Program. Modelling allows estimation of land use loads and projection of the impact of actions proposed in tributary strategies. It can also provide an estimate of progress in implementing the strategies. The CBP has concluded that modelling should be used primarily as a planning tool and for projecting future conditions. Monitoring, on the other hand, has no capability of projecting future conditions of the impact of proposed actions but is critical to evaluation of long-term trends in water quality due to land use changes and control measure implementation. In the short term,

monitoring is critical to understanding the impact of major climatic events on Chesapeake Bay water quality and living resources.

18.6 Current BMP Deﬁnitions, Efﬁciencies and Assumptions Much of the progress towards the 2000 nutrient reduction goals came from state reported implementation of agricultural BMPs between 1993 and 2000. Both structural BMPs, e.g. lagoons and manure storage sheds; and agronomic management practices, e.g. cover crops and nutrient management, are included in state tributary strategies. BMP deﬁnitions vary somewhat between states but are co-ordinated and made consistent by the Chesapeake Bay Program. Each practice has an approved deﬁnition, reduction efﬁciency and tracking/reporting procedure (United States Environment Protection Agency 1998). BMPs may be applied in three ways (Fig. 18.7). Some BMPs represent land use conversions such as the change from conventional tillage to conservation tillage or from crop land to forest or perennial grasses. Other practices are simulated in the model by a change in the rate, timing or

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Meteorological and Precipitation Time Series Inputs

Atmospheric Deposition Time Series Inputs

Final Output

Point Source Discharges

Opportunities for BMPs

External Transfer Module Opportunities for BMPs Land Variable File Management Practices Time Series Land Use Time Series Time Series from Land Simulation

River Variable File Time Series to River Simulation

method of nutrient inputs. Nutrient management, estimated to have been implementing about 1 million ha, is the most widely implemented practice that is simulated in the model by changing inputs and application timing. Practices that cannot be represented in the model either by a land use change or simulation are credited using a post processor with a standard reduction applied to the load from a speciﬁc land use based on the state reported implementation of the practice. The post process reduction efﬁciency for any one practice varies throughout the watershed based on soils, climate, topography, cropping systems, etc. There currently are about 40 BMPs or land use conversion practices that are approved for use by states in tributary strategies and are tracked, reported and represented in the WSM. There are also nearly 20 new practices (United States Environment Protection Agency 2007) awaiting peer review and approval by the Chesapeake Bay Program so they can be used (Table 18.1) The Chesapeake Bay Program appears to be the only programme nationally that has developed such a sophisticated quantitative BMP tracking and crediting system (i.e. assigning numerical load reduction efﬁciencies to various

Fig. 18.7 The Chesapeake Bay Watershed model opportunities for BMP.

practices for nitrogen and phosphorus). The fact that it has been in existence for 10 years, has undergone two self-imposed internal reevaluations and has openly identiﬁed a list of weaknesses and needed improvements is laudable. The scientiﬁc community, however, is somewhat skeptical of the quantitative use of model-based results such as bay-wide nutrient reductions based on reported BMP implementation and efﬁciency assumptions when they are used to shape policy. Progress is likely to have been overestimated. Evaluation of the BMP crediting system used in the CBP reveals both strengths and weaknesses. While practice deﬁnitions appear to be consistent within the watershed, the WSM is limited in its capacity to simulate different land uses and different crops, i.e. it only simulates a ‘composite crop’ for each watershed segment that may be many hundreds of square kilometres in size. As a result, it is a daunting task to develop an approach that adequately describes practice efﬁciency. The CBP has tried to account for this by limiting the acreage reported treated by a down gradient practice (e.g. buffers) or by limiting the potential for nutrient loss from certain conﬁned animal operations. Such efforts are

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Table 18.1 Chesapeake Bay Watershed Tributary Strategy BMPs (http://www.chesapeakebay.net/tribtools.htm) Approved BMPs

BMPs requiring peer review

Agricultural BMPs Riparian forest buffers Riparian grass buffers Wetland restoration Land retirement Tree planting Conservation-tillage Carbon sequestration/alternative crops Poultry phytase Poultry litter transport Nutrient management Enhanced nutrient management Conservation plans/soil conservation and water quality plan (SCWQP) Cover crops (early- and late-planting) Small grain enhancement (early- and late-planting) Off-stream watering w/ fencing Off-stream watering w/o fencing Off-stream watering w/ fencing and rotational grazing Animal waste management systems: livestock Barnyard runoff control/loaﬁng lot management Animal waste management systems: poultry

Agricultural BMPs Continuous no-till Dairy precision feeding/and forage management Swine phytase Ammonia emission reductions Precision agriculture Precision grazing Water control structures Stream restoration

Urban and mixed open BMPs Riparian forest buffers Wetland restoration Tree planting Urban growth reduction Wet ponds and wetlands Dry detention ponds and hydrodynamic structure Dry extended detention ponds Urban inﬁltration practices Urban ﬁltering practices Urban stream restoration Erosion and sediment control Urban and mixed open nutrient management

Urban and mixed open BMPs Riparian grass buffers Forest conservation Horse pasture management Abandoned mine reclamation Mixed open stream restoration Dirt and gravel road erosion and sediment control Urban street sweeping

Forest BMPs Forest harvesting practices

Forest BMPs Stream restoration Dirt and gravel road erosion and sediment control

Septic BMPs Septic connections Septic pumping Septic denitriﬁcation

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useful in trying to limit overestimation of nutrient reductions but overly optimistic BMP efﬁciency estimates may still occur. There are numerous causes for the optimistic BMP reduction estimates. Hydrological lag times may delay actual water quality improvements but delays in planning, implementation and practice maturity are also important factors, not usually considered, that are sources of lags in practice effects. Hydrological lag times are well documented. Dissolved nitrogen associated with groundwater may have a transport time of years to decades, with a median time of about 10 years (Lindsey et al. 2003). Nutrients associated with sediment can have much longer transport times (several decades) in the watershed because of their storage in soil and stream corridors, both of which are greatly inﬂuenced by annual rainfall. Planning, implementation and practice maturity lag times may be easier to estimate than hydrological lag times but are rarely considered. Other factors that caused overly optimistic assumptions about progress were BMP efﬁciency and application assumptions. Probable sources of error in estimating BMP impacts include limited data and/or ﬁeld observation; research and plotscale reduction efﬁciencies applied to catchmentscale implementation; extreme spatial variability in soils, hydrology and management; assuming proper implementation, operation and maintenance, function and replacement; and ﬁnally, optimistically reported implementation rates. These all appear to result in greater pollution control credits for practices than what is likely to occur (Simpson et al. 2004). As a result, in 2003, selected BMPs were evaluated and their efﬁciencies revised to better reﬂect current research and knowledge; the goal was to provide more realistic estimates of expected pollution reduction levels. Nutrient management plan efﬁciency estimates, for example, were reduced from 30% to 24%, although the initial technical proposal would have reduced them to 18%. Conservation plan and cover crop (applied to conservation tillage) efﬁciency and implementation assumptions were also substantially reduced. This impacted water-

shed model output by reducing apparent ‘progress’ and making it more difﬁcult to develop tributary strategies that would achieve loading goals. For both progress runs and strategy scenarios, the changes resulted in substantially lower modelled phosphorus reductions and somewhat lower nitrogen reductions that more closely related to observed conditions. Policy-makers and programme managers reacted to the proposed changes much more differently than did people in the scientiﬁc community. Individuals in the scientiﬁc and technical communities supported the changes. They agreed with the evidence that suggested BMP impacts were overestimated and that changes in reduced modelled ‘progress’ would bring model results closer to expected water quality. Agency managers responsible for explaining progress to stakeholders and policy-makers were put in the difﬁcult situation of having to explain the slower progress than previously reported as well as the increased difﬁculty in achieving already challenging loading goals. Stakeholders and policymakers were initially confused, which detracted from the accelerated implementation that was needed. It was apparent that the changes were needed and did in fact improve estimated impacts and, ultimately, CBP credibility. But neither managers nor scientists could afford to lose stakeholder or policy-maker support. With time, most people have recognized the need for, and value of, the changes, but gaining initial acceptance was difﬁcult. However, evaluation of a few BMPs in 2003 led to wider recognition of the need for a broad review of deﬁnitions and effectiveness estimates for all existing BMPs or new ones proposed for use in Tributary Strategies. This resulted in revision of existing effectiveness estimates and a more conservative approach to establishment of reductions associated with new practices. These revisions have further reduced modelled progress estimates but are considered to more closely represent actual water quality. The revision process was extremely challenging and both scientiﬁc and policy interests attempted to defend optimistic effectiveness estimates for BMPs of interest. Unlike in 2003, however,

Chesapeake Bay Catchment Management policy-level consensus was reached for all BMP effectiveness revisions proposed in 2007, as was agreement to an adaptive process in which BMPs would be reviewed every 3–5 years. Experience in the Chesapeake Bay helped identify approaches for conservatively predicting conservation effects and using adaptive management in water quality restoration programmes to avoid some of the difﬁculties encountered. Reﬁning efﬁciencies is an iterative, adaptive process that must continue while research identifying and adding new BMPs is being conducted. It was recognized that there is no ‘good time’ to change BMP assumptions but they must be made at recurring intervals. Changes should not be made continuously because policy cannot accommodate constant adjustments. It is essential to be conservative in setting and revising assigned efﬁciencies. It is always easier to gain public and policy acceptance for an increase in practice efﬁciencies than to reduce them. However, there are no known examples in the Chesapeake catchment, or elsewhere, where BMP efﬁciencies underestimate pollution reduction impacts. This further reinforces the need to use conservative efﬁciency estimates. Research- and demonstration-site derived efﬁciencies for catchment-scale implementation efforts do not reﬂect the spatial variability of an entire catchment area. Efforts should be made to assure that reported implementation is close to actual and to determine if farmer implementation and operation is as rigorous as speciﬁed in the practice. Model output and monitoring data must be consistent and used appropriately. Implementation progress, water quality restoration progress and habitat/living resource and restoration progress should be reported separately, using different data sources and approaches. This will more accurately convey the health of the ecosystem to stakeholders and is critical to maintaining the long-term momentum of any restoration effort. Better research or demonstrations and monitoring of individual BMPs, conservation systems and small catchment area conservation effects will increase conﬁdence in BMP effectiveness. Finally, managers, policy-makers and

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involved citizens must be made aware of potential implications of the iterative-adaptive BMP effectiveness approach so they understand the recurring need to change effectiveness estimates as knowledge advances (Simpson and Weammert 2007). States are responsible for collecting and reporting levels of practice implementation on an annual basis. Implementation rates vary widely between states and from year to year within states. This may be related to differences in staff support between states and political emphasis on the need for reporting. Methods for tracking and reporting to the Bay Program are consistent but data collection within states varies. There is concern that there is double counting of certain practices and possible optimism in reported implementation. Reporting is further complicated by the perceived need to always show progress. If a state ﬁnds errors in previous reporting, they must reduce reported implementation and run the risk of being further from goals than previously reported. While the revised numbers may be more accurate, they may create public and political perception concerns. The Chesapeake Bay Program accepts state-reported implementation rates without question except when implementation in the WSM model segment exceeds available acres. Third party review of annual implementation progress should be considered to ensure quality of reported implementation information. Reported progress is usually based on plans written or structures designed, not actual implementation. There is concern that this results in overestimation of implementation. A Rural Clean Water Project study in Pennsylvania found as low as a 22% implementation rate for some nutrient management plan and animal waste BMPs (United States Department of Agriculture– Agricultural Stabilization and Conservation Service 1992). A 1998 statewide survey of nutrient management plan implementation in Maryland supported reductions in nutrient inputs as a result of the plans. Depending on deﬁnition, however, 40–70% of farmers reported ‘following the plan’ and many of those had not reduced

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nutrient inputs as much as assumed in the nutrient management reduction efﬁciency. Another WSM implementation assumption is that the practice is implemented as deﬁned by the CBP. For example, cereal grain cover crops are an efﬁcient post-harvest nitrogen management practice with high reduction efﬁciencies. However, these efﬁciencies assume timely fall planting (before 1 October in the coastal plain). In some cases, cost-sharing standards for cover crops have allowed planting until after 1 November. The same data on which the cereal grain nitrogen reduction efﬁciencies were based showed a major decline in efﬁciency of nitrogen reduction after 1 October. There are numerous other cases where either state requirements or farmer practice are inconsistent with CBP BMP deﬁnitions. It has been suggested that the CBP should adjust implementation rates of practices based on available data. Additional research and surveys should be conducted to reﬁne actual rates of implementation and identify differences between practice deﬁnitions and what is being implemented (Simpson et al. 2004). All practices are assumed to be implemented and maintained as prescribed and to function at design efﬁciency over time and in all types of storm events. The CBP is working to improve assumptions related to BMP function, maintenance and design storm events but it is evident that these, along with optimistic efﬁciency and implementation assumptions, are likely to have resulted in substantial overestimation of nutrient reductions.

18.7 Systems Approaches to Conservation and Nutrient Balances The discussion above identiﬁes speciﬁc actions that can improve estimates of progress from BMP implementation. There are broader issues related to the overall approach to reducing agricultural nutrient losses using BMPs. Most BMPs are selected to address the current farm nutrient budget using the current crop and animal produc-

tion systems. It is also important to address farm nutrient imbalances and to look at feasible changes in production systems as part of a comprehensive farm planning process. Despite efforts for more comprehensive planning, BMP practices or systems are still frequently implemented individually without looking at how practices work together to address whole farm water quality concerns. Current focus is on tactical plans that prescribe speciﬁc conservation practices for current systems rather than strategic actions such as feed management or alternative crop rotations. Structural and pollution abatement practices are frequently considered equivalent to pollution prevention practices. Efforts should be made through feeding, waste storage, cropping systems and nutrient management to minimize the opportunities for nutrient losses prior to using pollution abatement practices to treat lost nutrients. The planning and implementation process needs to be an ongoing repetitive process that continually reassesses crop and animal management options at a farm, watershed and industrywide level to meet water quality objectives. The cost of BMP/farm system implementation is not currently part of production costs. As a result, farmers expect government cost-share to offset a large part of BMP expense. Market incentives or disincentives are needed to build environmental costs into recoverable production costs. Where cost share is used, it needs to be targeted and viewed in light of overall environmental objectives. Cost share continues to be viewed and designed as another form of support payment to farmers. The tendency, within government programmes, to pay for implementation not consistent with CBP practice deﬁnition for the sake of farmer convenience, should be minimized or reduction efﬁciencies should be adjusted accordingly. Cost share practices should support strategic, not just tactical, action and must ﬁt within the whole farm environmental management system. For example, most cost share standards call for 6 months maximum storage of animal waste which forces manure application in the fall, just prior to the period of most intense

Chesapeake Bay Catchment Management runoff and leaching. Care must be taken to assure that practices designed and/or cost-shared with public funds do not reduce losses from one pathway while increasing them from another (Simpson et al. 2004). Specialization, intensiﬁcation and concentration of agricultural production, particularly poultry and livestock, have created ﬁeld, farm gate and regional nutrient imbalances. Historically, farms producing animals and animal products depended heavily on the farm crops to feed the animals. In the last half of the twentieth century, farms became more specialized and livestock and poultry production became more regionalized with crops to feed the animals being produced in other areas. As a result, crops are often not consumed on the farms where they are produced but are exported to other farms and regions where intensive animal production is located. This new organization of production is typical for non-ruminant animals, such as poultry and hogs, and is becoming more important in the dairy industry. Although nutrients are removed from the farm where crops are produced and there is very little opportunity for them to return, the crop farms are sustained by the replacement of the exported nutrients primarily with fertilizer. In crop/feed importing regions, the ﬂow of nutrients to individual farms may far exceed the ﬂow of nutrients from the farm in animals, animal products or crops. This imbalance in ﬂows results in short- or long-term accumulations of nutrients that are susceptible to unintended or unexpected losses to surface and ground waters (Simpson et al. 2004). On farms or in production areas with substantial nutrient imbalances, the application of BMPs must be done in a manner to address the complete production system, from feed to animal to waste management to land application of manure (feed production). Comprehensive plans are essential in areas with nutrient imbalances but it may still not be possible to manage waste nutrients in an economically and environmentally acceptable manner. On such farms or in entire production regions, transport of manure to nutrient-deﬁcit areas and/or development of high

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value alternative uses may be necessary to supplement on-farm BMPs. While the Chesapeake region has several transport and alternative use programmes, none have yet proven to be economically viable without substantial public sector subsidy.

18.8 Observations and Lessons Learned in the Chesapeake Bay Watershed • The Chesapeake Bay restoration effort has been a unique partnership of six states, the District of Columbia and the Federal government. Only two of these states have signiﬁcant areas of the bay within the boundaries so the fact that others are active in the partnership and have remained so for 20 years is laudable. It is our conclusion that a formal partnership among catchment jurisdictions is an essential starting point for a long-term restoration programme. In addition, annual gatherings of the leadership of the jurisdictions demonstrating their political and executive support for the restoration effort are critical to maintaining both the partnership and restoration progress. While goals have proven more challenging than expected, there is no question that what progress has been made would not have occurred without the strong support at the executive level from the jurisdictions and the Federal government. • Quantitative nutrient reductions goals or loading caps with deadlines to achieve them are essential to a successful catchment restoration programme. In the Chesapeake Bay catchment, recent goals were set to completely remove impairments in an effort to meet the objectives of Federal regulatory requirements through a voluntary partnership. It appears that these goals may require such extraordinary reductions that they have made it difﬁcult to maintain the support and momentum for the restoration. While goals should be based on improvement in living resources, it may be necessary to establish challenging, but accomplishable, interim goals or milestones along the path to full restoration of living resources.

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• The ability to allocate loads to jurisdictions and allow those jurisdictions to develop control strategies consistent with their social, cultural, ﬁscal and political situation is also essential to the development of accomplishable strategies and long-term maintenance of a true catchment partnership. • Development of control measure nutrient reduction efﬁciencies allows jurisdictions to develop strategies and scenarios that allow them to see if, based on model estimates, needed reductions are estimated to be achieved. It is recognized that data supporting control measure efﬁciency is limited and these quantitative efﬁciencies should be viewed as relative estimates. However, these efﬁciency estimates are the foundation upon which any catchment management strategy must be built so efforts must continue to improve these estimates through additional research and small watershed monitoring during implementation of selected practices. • The appropriate balance between modelling and monitoring as evaluation tools must be established early in the restoration programme. As discussed above, over-reliance on model output to estimate water quality restoration progress resulted in overestimation of progress for the Chesapeake Bay and a substantial policy and public reaction when these efﬁciencies were corrected to more accurately reﬂect accurate progress. However, model simulations allow the projection of the impact of future land use, population and agricultural changes and evaluation of scenarios to determine the most effective and cost-efﬁcient control measure strategies to achieve needed nutrient reductions. • Monitoring must be used as the primary tool to determine actual progress in achieving needed nutrient reductions; however, extreme interannual variation in climate makes it very difﬁcult to see the relatively small incremental effects of control measure implementation. A combination of long-term trends based on average hydrology and near real-time evaluation of water quality must be used to evaluate actual progress. • The Chesapeake Bay Program developed best management practices for control measure efﬁ-

ciencies before the mid 1990s, nearly a decade before anyone else attempted to do this. It became apparent that these estimates assumed research level implementation of the practice, proper operation and maintenance and long-term functioning of the practices. It is apparent that all of these assumptions were optimistic and that wide-scale implementation under actual management operation and maintenance conditions results in lower reductions than observed at a research scale. This can be addressed through initially conservative efﬁciency estimates rather than adjusting existing optimistic efﬁciency estimates. Adjustments may make model estimates of impact closer to actual but give policy-makers and the public the appearance that progress towards goals is actually being reduced which can harm both public support and political will. • It is currently estimated that greater than a 50% reduction in nutrient loads, from a 1985 base, will be needed to remove the nutrient impairments to the Chesapeake Bay. Reductions comparable to this level may be needed in other coastal waters, such as the Gulf of Mexico/ Mississippi River Basin, to remove nutrient impairments. Limited data is beginning to indicate that the implementation of most current BMPs or control measures to non-point sources of nutrient pollution, particularly agriculture, may not be capable of achieving these levels of nutrient reduction. It is likely that removing nutrient impairments to an estuary like the Chesapeake Bay will require full and rigorous implementation of existing practices along with systems changes so that less polluting development and agricultural production systems are used. Experience in the Chesapeake Bay catchment suggests that full and rigorous implementation of existing practices to current production systems is very challenging and systems changes are even more challenging and frequently met with resistance by the agricultural community. • While much work remains before nutrient impairments are removed in the Chesapeake Bay and other coastal waters, focus should not solely be on reducing nutrient losses from existing sources. It is equally critical to minimize growth in load due to development or agricultural inten-

Chesapeake Bay Catchment Management siﬁcation, particularly for animal production or grain-based biofuel production. Perhaps the greatest challenge will occur if, or when, nutrient reduction loading goals are achieved, since all future growth in load must be completely offset to maintain the water quality that was so challenging and costly to achieve.

References Lindsey, B.D., Phillips, S.W., Donnelly, C.A. et al. (2003) Residence times and nitrate transport in ground water discharging to streams in the Chesapeake Bay watershed. Water Resources Investigations Report 03-4035. USGS, Baltimore, MD. Simpson, T.W. and Weammert, S.E. (2007) The Chesapeake Bay Experience: learning about adaptive management the hard way. In: Schnepf, M. and Cox, C. (eds), Managing Agricultural Landscapes for Environmental Quality: strengthening the science base. Soil and Water Conservation Society, Ankeny, Iowa, pp. 159–169. Simpson, T.W., Musgrove, C.A. and Korcak, R.F. (2004) Innovation in Agricultural Conservation for the Chesapeake Bay: evaluating progress and addressing future challenges. A White Paper from Scientiﬁc and Technical Advisory Committee (STAC) Chesapeake Bay Program, Publication 04-003, Annapolis, MD. http://www.chesapeakebay.net/ content/publications/cbp_13325.pdf (last veriﬁed 25 November 2008).

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United States Department of Agriculture–Agricultural Stabilization and Conservation Service (USDAASCS) (1992) Pennsylvania-Conestoga Headwaters Project, 10 Year Report 1981–1991. Annapolis, MD. United States Environment Protection Agency (US EPA) (1983) The Chesapeake Bay Agreement of 1983. Chesapeake Bay Program Ofﬁce, Annapolis, MD. United States Environment Protection Agency (US EPA) (1987) The 1987 Chesapeake Bay Agreement. Chesapeake Bay Program Ofﬁce, Annapolis, MD. United States Environment Protection Agency (US EPA) (1998) Chesapeake Bay Watershed Model Applications and Calculation of Nutrient and Sediment Loadings – Appendix H: tracking best management practice nutrient reductions in the Chesapeake Bay Program. US EPA 903-R-98-009. CBP/TRS 201/98. Annapolis, MD. United States Environment Protection Agency (US EPA) (1999) The State of Our Nation’s Waters. Environment Protection Agency, Washington, DC. United States Environment Protection Agency (US EPA) (2000) Chesapeake 2000. Chesapeake Bay Program Ofﬁce. Annapolis, MD. United States Environment Protection Agency (US EPA) (2003) Ambient Water Quality Criteria for Dissolved, Oxygen, Water Clarity and Chlorophyll a for the Chesapeake Bay and Its Tidal Tributaries. Annapolis, MD. United States Environment Protection Agency (US EPA) (2007) Tributary Strategy Tools. Annapolis, MD, USA. http://www.chesapeakebay.net/tribtools

Image facing chapter title page: Courtesy of Tom Simpson.

19

The Glasgow Strategic Drainage Plan J. BRYAN ELLIS1

1

Urban Pollution Research Centre, Middlesex University, Hendon, London, UK

19.1

Introduction

The city of Glasgow in Scotland has long suffered severe intra-urban ﬂooding with the 1:100 storm event of July 2002 for example, generating a maximum rainfall intensity of 95 mm per hour with a total of 75 mm falling in 10 hours and producing some 6000 m3 of overland ﬂows. Rail and road transport was seriously disrupted with the M8 motorway closed for a prolonged period as well as over 100 urban roads being rendered impassable. In addition, over 500 properties were ﬂooded causing some £100 million damage. The River Clyde is also associated with serious water quality problems due primarily to intermittent urban runoff discharges and the Class C or D river water quality is well below any EU WFD ‘good ecological status’ (GES) target classiﬁcation. In addition, the region is predicted to be particularly vulnerable to future climate change which will further exacerbate ﬂood risks and water quality problems (Macdonald and Jones 2006). These hydraulic and water quality problems are constraining development regeneration and severely restricting riparian and ecological improvements as well as deterring business investment. The problems of drainage infrastructure refer to both above-ground watercourses and below-ground sewer systems, both of which have Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

suffered from over a century of industrialization and urbanization and are clearly struggling to cope with present day demands. Hydraulic capacity to accommodate continuing regeneration and growth within the urban area is no longer available, with the environmental damage of combined ﬂooding and water pollution being beyond acceptable limits in terms of future EU legislation. The Glasgow City Council (GCC), Scottish Water (SW) and the Scottish Environment Protection Agency (SEPA) undertook a major review of drainage infrastructure provision following the 2002 ﬂooding events detailing the causes, consequences and potential mitigating measures (Glasgow City Council 2002). The review report also identiﬁed the problems of stormwater management resulting from divided and overlapping responsibilities as well as unclear legislation. It recognized that a joint, integrated partnership approach and long-term strategy for runoff control was required to successfully address the drainage problems in Glasgow and satisfy the interests and needs of a variety of stakeholders. At present the various stakeholders are required to progress drainage solutions through a range of separate legislative instruments, each of which has its speciﬁc procedures and timescales as well as often conﬂicting requirements. In addition, funding mechanisms, such as ﬂood protection grant assistance, are closely linked to the successful negotiation and conclusion of these separate procedures.

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These issues stimulated a multi-agency and private sector forum led by the GCC and SW to undertake a major and wide ranging examination and masterplanning of the sewerage and urban watercourse infrastructure. The Glasgow Strategic Drainage Plan (GSDP) will provide for a 1.1 million population and incorporates 27 discrete catchments in terms of local drainage area plans (DAPs) and four sewage treatment works (STWs) receiving wastewater from a 4-km2 catchment and possessing over 540 combined sewer overﬂows (CSOs). The key objectives of the GSDP masterplan are: • ﬂood risk reduction; • water quality improvement; • removal of development and carrying-capacity constraints; • habitat and watercourse improvement; • support for future integrated and optimized investment planning The successful delivery of sustainable drainage infrastructure solutions clearly requires a more streamlined and collaborative planning approach to deal with the complex, multi-layered issues associated with the control and management of impermeable overland ﬂows. The key to its success will lie in the effective integration of urban land use planning, building design, surface

water drainage, access and open space provision with sustainable drainage systems (SUDS) having a prime role in providing inter-linkage between these elements.

19.2 Drainage Masterplanning Given the number of stakeholders (Fig. 19.1) involved in drainage management within the greater Glasgow region, it was recognized that a joint holistic strategy was required to successfully address the issue of delivering an integrated, sustainable drainage infrastructure. Two organizations lead in the process (shown in bold in Fig 19.1), but other principal stakeholders and functional areas involved in the strategic surface water management for the GSDP emphasize the difﬁculty of achieving progressive, joined-up and motivated integrated thinking. The multi-agency Metropolitan Glasgow Strategic Drainage Partnership (MGSDP) is a unique collaborative stakeholder platform formed from the organizations most closely involved with the operation and management of the sewerage and drainage system with the lead agencies being SW (responsible for the sewerage network),

Fig. 19.1 Glasgow Strategic Drainage Partnership stakeholders and functional areas.

The Glasgow Strategic Drainage Plan GCC (responsible for roads drainage, watercourses and ﬂood risk) and SEPA (responsible for water quality and ﬂood advice). The MGSDP has also drawn in seven local authorities lying within the larger metropolitan region as well as consulting with the Scottish government on emerging legislative, ﬁscal and administrative issues. The masterplan was therefore required to address not only the variety of stakeholder concerns, but also performance drivers and future urban land use development horizons. At the core of the masterplanning process lies the ﬂows and loadings to the four STWs during wet weather events and the interactions of stormwater and combined sewer discharges with receiving water capacity and assimilation. The interactions between surface and below ground drainage systems have always been complex in Glasgow with some culverted watercourses serving as sewers, e.g the Gartloch ‘burn-sewer’. Drainage area plans (DAPs) undertaken by SW had never traditionally considered the ‘surface’ water system within their sewer asset modelling routines, so that excess wastewater frequently surcharged during extreme wet weather conditions. Such asset-based modelling approaches totally failed to consider the effects of excess overland ﬂows generated from impervious surfaces which forced their way into the foul/combined sewage system during extreme rainfall events. Greater Glasgow has four STWs located within the urban area and which serve over 1 million people (Fig. 19.2). Amalgamated network modelling for 27 hydraulically discrete catchments will underpin cross-catchment solutions to allow environmental impacts to be addressed at a regional level. It was considered that local subcatchment solutions, whilst perhaps individually robust, would be insufﬁcient to deliver collective strategic sustainability and to appropriately balance traditional ‘hard’ engineering controls with ‘soft’ watercourse and BMP (best management practice)/SUDS source control approaches. Masterplanning also offered opportunities for fuller stakeholder engagement and more potentially equitable environment quality-of-life solutions for sub-catchment communities.

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A staged approach of the master-planning process was agreed and adopted by the GSDP Steering Group (Fig. 19.3). Stage 1, which was completed in April 2004 at a cost of £120 million, investigated the overall urban drainage and ﬂooding issues as input for initial stakeholder information and debate, with the critical East End/ Dalmarnock sub-catchment drainage area modelling study providing a baseline context and pilot study for identifying the nature and scale of possible initial solutions. Some of the watercourses in the East End are culverted for up to 90% of their length and subject to increased ﬂow velocities, ﬂood surcharging and reduced attenuation opportunities and have a constant need for silt and debris removal. The macro-modelling approach therefore needed to analyse both sewer and watercourse systems together in order to fully evaluate their interactions (Tufail et al. 2003) and give a better understanding of the hydraulic processes operating. The mapping and modelling of drainage area plans (DAPs) by UK water companies such as Scottish Water has traditionally only been concerned with the sewer network (i.e. an asset survey), and thus an integrated sewer/ watercourse modelling approach represented an innovative advance. These solutions included consideration of both traditional ‘hard’ and alternative ‘soft’ source control approaches as well as SUDS retroﬁt and watercourse improvement, estimated as requiring an investment of some £120 million The sub-catchment modelling of the Light Burn showed that ﬂooding was caused by a lack of hydraulic capacity in both the sewer network and (frequently culverted) river channel, which was exacerbated during extreme storm events by overland ﬂow from nearby highways and the M8 motorway (exceedance overland ﬂows). Analysis of a 1 : 5 RI (storm return interval), 2-hour duration design storm event showed that >75% of the receiving water volume during higher and extreme ﬂows was composed of CSO spillage, and this has signiﬁcant impact on both ﬂow regime and water quality (Akornor and Page 2004). It was estimated that £1.5 million of

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WEST DUNBARTONSHIRE EAST DUNBARTONSHIRE Glasgow Dalmuir WwTW NORTH LANARKSHIRE

Dalmuir

CITY OF GLASGOW Dalmarnock Shieldhall WwTW Daldowie Dalmarnock WwTW Daldowie WwTW Shieldhall EAST RENFREWSHIRE East Kilbride

Hamilton

River Clyde

SOUTH LANARKSHIRE

Fig. 19.2 Sewage treatment works drainage areas within Greater Glasgow.

development potential was affected by these strategic drainage constraints within the subcatchment. Three ‘hard’ engineering solutions were proposed at the end of Stage 1 for the sub-catchment, constituting interceptor tunnel, off-line storage tanks and ﬂow separation options. These would all provide a 1 : 30 level of service protection from sewer ﬂooding and 1 : 100 for watercourse ﬂooding. The separation solution, involving over 13 km of sewer pipe, would incorporate source control BMPs at a total estimated cost of £13.6 million. The target levels of service (expressed as storm return interval; RI) would be achieved over the planning horizon to 2020 for a combination of sewer separation, source controls and river

channel management (Fig. 19.4). The sewer separation option would involve a step-change approach with annual separation targets (in the order of 10–15%) combined with in-stream dredging and/or aeration in addition to BMP/SUDS implementation. Stage 2, which is still ongoing, deﬁnes an initial strategic masterplan to facilitate meaningful stakeholder debate in terms of the scope, complexity, costs and solutions for achieving a sustainable outcome. It involves the completion of all drainage area investigations across the metropolitan Glasgow region and an urban pollution study of all discharges into the River Clyde. Of particular importance was the need for an understanding of the water quality of the River Clyde

The Glasgow Strategic Drainage Plan

Fig. 19.3 Stages of the Glasgow Strategic Development Plan.

Fig. 19.4 Glasgow Strategic Development Plan sewer separation option.

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with respect to wet weather CSO, surface water outfall (SWO) and STW discharges as a basis for achieving future target water quality compliance. In addition, the applicability of using the critical sub-catchment modelling as a basis for regional catchment up-scaling and cost-effective crosscatchment options is to be assessed. Both design rainfall events and annual time series have been considered in this modelling analysis and 2020 development horizons included, as well as the potential effects of future climate change. Local and strategic solutions have been developed from the analysis including local options essentially based on source control BMP/SUDS retroﬁtting as, for example, the Light Burn subcatchment (Fig. 19.5). A preferred strategic option of interception and transfer has been identiﬁed

however, which would cost some £2.7 billion and optimize existing sewerage and wastewater assets but intercept Formula A ﬂows at each STW and transfer these ﬂows to a new STW further downriver (Page and Fleming 2005). The alignment of the interception tunnel transferring the Edinburgh Road CSO discharges downstream is shown together with the principal culvert ‘daylighting’ along the Light Burn and adjacent riparian areas scheduled for re-development, landscaping and amenity use. In addition to such end-of-pipe strategic ‘hard’ engineering, there would be wide-scale local adoption and retroﬁtting of BMP source controls to minimize impermeable surface runoff to both separate and combined systems. Green roofs, street rain gardens (planters), storage attenuation

CSO to be abandoned

Interceptor Tunnel 1km long; 1600mm diameter

Roof disconnection to public buildings

Detention Basin Porous Surfacing

Historical flooding

Watercourse re-development de-culverting and landscaping

Fig. 19.5 Local retroﬁtting drainage solution for the Light Burn subcatchment.

Grass Swale

The Glasgow Strategic Drainage Plan in parks and open spaces as well as the development of green access corridors adjacent to existing watercourses (with de-culverting), would comprise the core elements for surface water management plans (SWMPs). By utilizing source control BMP techniques, these SWMPs will permit development to proceed on areas where the combined sewerage and/or watercourse system is already at capacity and will serve as drainage ‘blueprints’ for the regional catchment. In addition to providing a drainage solution, by linking up with broader environmental strategies embodied in the GCC Development Plan, there will be opportunities for well-designed schemes such as the Light Burn to generate additional environmental and community beneﬁts including: • enhanced ecological and wildlife habitats with expanded riparian corridors and new niche areas associated with re-opening of culverted watercourses; • landscape improvements and improved public access and recreation/amenity opportunities; • provision of green space and water space for local amenity beneﬁt; • protection and enhancement of existing environmental designation. A major component within the Stage 2 process has been the hydrodynamic and water quality modelling of the River Clyde to provide a better understanding of the existing impact of tributary CSO discharges. The water quality modelling predicts no major changes to river status even if all signiﬁcant inputs to the Clyde were removed (Page and Fleming 2005). Improvements would be essentially dependent on both increased and improved STW capacity, although a reduction in stormwater inputs and upstream diffuse runoff control would be a supportive measure. The urban pollution planning study conﬁrmed the conclusion that end-of-pipe improvements alone would be insufﬁcient to meet river target standards. Nevertheless, the MGDSP has progressed a range of measures since 2002 under Stages 1 and 2 of the GSDP including fast action ‘squads’ to clear blocked watercourses and sewers, replace-

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ment of critical sewer lengths, a £6 million initial investment to tackle intermittent CSO discharges and £16 million for watercourse ﬂood alleviation schemes as well as an improved ﬂood warning service. Almost 500 properties have been removed from the ‘At Risk’ ﬂood register since 2002. SEPA is also considering a request to retroﬁt SUDS schemes within the urban area to reduce surface water discharges to the foul sewage system thereby reducing the hydraulic load and potentially freeing up capacity in the foul sewers. Issues at Stage 2, however, include those concerned with the appropriate assessment of sewer performance, dealing with illicit cross-connections, surface water diversions and overland exceedance ﬂows as well as assessing the scale of potential inﬁltration capacity for SUDS schemes (Scottish Environment Protection Agency 2004). The initial masterplan developed within Stage 2 has identiﬁed upgrading works with estimated costs of nearly £1.5 billion which will need to be agreed with relevant stakeholders, in particular SW, GCC and other local authorities as well as with central government. Stage 3 of the GSDP will be concerned with detailed master planning of the GSDP catchment with development scenarios extended to 2025 and targeted surface and stormwater management plans on a catchment wide basis. Stakeholder consultation will enable preliminary designs to be developed for design and implementation in the ﬁnal Stage 4 which will be rolled out over the next decade. It is expected that Surface Water Management Plans (SWMPs) will be ﬁrmly founded on area SUDS implementation associated with the introduction of ‘green corridor’ subcatchment schemes together with more resilient building design, and the widespread introduction of rain gardens, street planters, water squares and attenuation ponds to cater for exceedance overland ﬂows (Macdonald and Jones 2006).

19.3 Some Key Issues There has been a tendency to view the drainage problems of Glasgow as a local or area issue but

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it has much wider regional and national signiﬁcance in terms of the population affected and economic impacts if drainage problems prevent or restrict successful re-generation. In addition, continued ﬂood and water quality infractions could have serious consequences under future EU WFD regulation. In this respect, SEPA has a distinct and separate role as the responsible monitoring and regulatory agency, and although an active stakeholder in the MGSDP, this partnership must not be at the expense of their regulatory responsibilities. The lessons learned from the Dalmarnock East End ﬂood review showed that ad hoc and unsustainable solutions to local drainage problems required a wider sub-catchment consideration of both up- and downstream actions to deliver speciﬁc local improvements. This means that in some cases it may be necessary to accept standstill or even short-term deterioration in environmental quality in order to achieve long-term improvements. This would be a controversial but perhaps pragmatically necessary issue for SEPA in the WFD river basin planning process. A clear mandate to develop and implement the GSDP on a city-wide basis has not yet been legally or administratively approved in terms of current regulations and legislation or funding alignments. There is a real possibility that a solution for one area could simply transfer to, or heighten the risk for, another area. Delivery of the GSDP must be achieved within, and be consistent with, the objectives of wider Planning and Strategic Environment Assessment (SEA) statements and plans as well as Local Plans and the Glasgow and Clyde Valley Structure Plan (GCVSP), all of which set the framework for planning within the GSDP site. The GCVSP provides goals and aspirations for economic development, improved social mix through co-location of high quality social and public housing, improved environmental quality and the means whereby these are to be achieved. Drainage infrastructure constraints could be prejudicial to these aspirations and given the scale and nature of the prevailing constraints, there will be a need to prioritize and time-stage any

solutions. Prioritizing expenditure within the masterplan is, however, constrained by allocated institutional funding and cost-beneﬁt criteria. Funding of the GSDP, however it may be prioritized, goes well beyond the capacity of any single partner agency to deliver and it is crucial that the central Scottish Executive be involved as the lead national public sector administration. Scottish Water funding for sewers does not include the level of strategic stormwater infrastructure investment identiﬁed by the Stage 2 initial master planning. In addition, GCC can only procure increased budgets for watercourse improvements through a ﬂood prevention order and developers can hardly be expected to fund rectiﬁcation of historical deﬁciencies. Under current legislative arrangements, however, a developer can use any spare capacity created or available in the sewer system regardless of whether or not they contributed to the creation of that capacity. Thus one developer’s contribution can be utilized to subsidize and unlock drainage capacity for another. This issue needs to be addressed if a signiﬁcant brake on private sector infrastructure investment is to be avoided. An appropriate mechanism for calculating and managing developer contributions towards infrastructure provision and improvements needs to be identiﬁed and agreed. It is also essential to ensure the continued participation of the private sector in the GSDP and, therefore, the public sector participants need to regulate in a fair, effective, proportionate and consistent manner. In addition, they need to provide not only leadership direction but also anticipate issues which might arise in the future as a result of legislative or institutional arrangements. It will also be necessary to recognize that the private sector will only make a full commitment to a development site when it is conﬁdent that it can be developed to give a reasonable return on investment and within a reasonable timescale. In this respect, drainage and transport infrastructure planning is a necessary pre-requisite for development. The river basin management (RBM) planning process required under the EU WFD timetable for

The Glasgow Strategic Drainage Plan 2010 will consider the Clyde catchment and its centrality to the management of water resources in the west of Scotland. The GSDP has commenced its planning and implementation ahead of this river basin planning process and there is a clear need to ensure consistency with the wider catchment RBM planning and the accompanying proposed mitigating Programme of Measures (PoMs). There will be many connections and synergies through mutual objectives, particularly related to achieving improvements or protection against deteriorating urban diffuse water quality and ecological status. Effective management linkages between the central and local planning processes urgently need to be established to enable water priorities and funding arrangements to be put in place. The continued operation of CSOs and illegal stormwater discharges also gives rise to risks of WFD infraction. Surface Water Management Plans are being currently developed by SW and GCC for the Clyde Gateway Project Area to overcome shortfalls in urban drainage and to facilitate future development. Potential solutions to manage ﬂood and pollution risks include detention and retention basins which have land-take implications and which will need to be considered in local development and structure plans as well as in the RBMP process. Glasgow City Council is committed to the development of SWMPs for the greater metropolitan region, and its participation in the Clyde catchment Flood Liason Advisory Group (FLAG) will provide a cross-local authority link with SEPA and SW to support strategic policy for the RBMP process. The strategic GSDP vision of sustainable drainage for Glasgow incorporates SUDS as an essential component to deliver environmental beneﬁts. Under the GCC City Plan, there is a requirement that all development proposals will be ‘required to make provision for SUDS’ wherever practicable. Such SUDS proposals will still be expected to be designed to accommodate a 1 : 30 event with an ability to deal with 1 : 200 events through safe ﬂood routing arrangements. In addition, some stakeholder reservations have been expressed about the risk of failure that may

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be associated with inﬁltrative SUDS devices (Jones and Macdonald 2006), either due to poor installation or maintenance. The pressure on land take required for surface attenuation facilities also prejudices developers against these SUDS options. Despite the legislative drivers, however, a ‘hard’ engineering culture is deeply engrained within regulatory agencies which favour structural control measures such as interceptor tunnels and storage tanks over non-structural options. This attitude is reinforced by local pressure groups who argue strongly for structural measures as well as a perception that there is no standard, agreed method of calculating efﬁciency and effectiveness for non-structural methods. Additionally current funding and grant appraisal systems are generally deemed not to be suitable for non-structural methods. In the short term, however, structural controls are difﬁcult to deliver, so ‘soft’ SUDS engineering and deculverting become more credible alternatives. Other control measures including weather warning systems, emergency response systems or watercourse maintenance are already being embedded in GDPS strategic infrastructure policy but mainly as second-order measures. At present SUDS introduction is largely dependent on opportunistic delivery, contingent on what developers happen to bring forward for planning approval. It is also important to raise public awareness of surface water issues in order to achieve community ‘buy-in’ to the GSDP vision.

19.4 Support Systems Decision support approaches have been applied for the identiﬁcation and location of appropriate BMP options for surface water control and treatment based on criteria of ﬂood potential, runoff quality, public acceptability, cost, ownership responsibility, site factors (soil type/depth, groundwater levels, etc.) and receiving water ecology and habitat (Scholtz 2005). This decision support approach has been tested at demonstration sites within the Gadburn area in the

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north-east of the city centre, although Jones and Macdonald (2006) have suggested that local councils remain concerned about space requirements as well as future maintenance and cost burdens. As much as 110,000 m2 might be taken out of the Gadburn development site for the construction of SUDS, which would represent a substantial capital loss to the developers. In a parallel study elsewhere in Glasgow, Scholtz et al. (2006) suggested the use of below-ground storage in combination with surface swales and attenuation ponds as an alternative drainage solution which would reduce the pressure on land take. The creation of strategic drainage plans will allow easier prioritization, planning and integration of all funding sources which should result in a more effective and efﬁcient use of available resources. Currently, however, no speciﬁc funding exists to establish, develop, manage or maintain this strategic approach. This will, therefore, necessarily depend on ‘top-slicing’ existing municipal funds to provide the required resources, although the implications of this have yet to be determined. A major potential driver for the adoption of BMP/SUDS solutions for future urban drainage in Scotland, however, is the legislation introduced under the 2005 Water Environment and Water Sewers Act whereby the ‘Controlled Activities’ Regulations (CAR) now require that surface water discharges should be via BMPs/ SUDS to ensure sustainable ﬂood management. Such management should provide the maximum possible social and economic resilience against ﬂooding by protecting and working with the environment to deliver a quartet of ﬂood ‘awareness’, ‘avoidance’, ‘alleviation’ and ‘assistance’. The 2007 technical guide ‘Sewers for Scotland’ also contains a clause recognizing the need for stakeholder co-operation and the development of SWMPs as well as vesting SUDS drainage to the responsibility of SW (Water Research Council 2007). This is the ﬁrst time a UK sewage utility has set out such comprehensive requirements and guidance for the design, construction and maintenance of public source control BMPs/ SUDS. The guidance also encourages Scottish local authorities to manage highway runoff by

providing appropriate attenuation and treatment prior to discharge. The preclusion of certain SUDS components, including exclusions for individual properties, does leave however some questions regarding overall responsibilities and liability for BMP/SUDS drainage infrastructure. Irrespective of these issues there is an increasing regulatory pressure from SEPA for the consideration and inclusion of SUDS in new and retroﬁt developments with supportive legislation for their adoption and maintenance by SW. In addition, the existence of an ofﬁcial design guidance manual (Construction Industry and Research Information Association 2000), the formation and outputs of a Sustainable Urban Drainage Scottish Working Party chaired by SEPA from the mid-1990s (SUDSWP 2005), planning guidance notes (PAN 61, 2001: Planning and SUDS, Scottish Executive, Edinburgh) and other literature (www.sepa.org.uk/suds), have all provided administrative and legislative support. At national level, Strategic Policy 9c(iv) of the Scottish Executive (2004) advice on building standards also includes a requirement for SUDS to be related to the relevant design strategy with national planning policy (SPP7) for ﬂood control, including a vested requirement for SUDS (Scottish Government 2004). The recent consultative document on future ﬂood risk management in Scotland recommends that SEPA should take the lead role in the implementation of the EU Floods Directive with the local authorities maintaining their responsibilities for surface water sewers and highway drainage (Scottish Government 2008). Area Flood Risk Management plans setting the strategic framework would be vested in SEPA with local authorities responsible for translating the objectives set out in the Area Plans into speciﬁc measures to address local ﬂood risk. This dual designation contains inherent conﬂicts of interest and responsibility which would only be overcome by close participative collaboration and inter-action between these primary stakeholders. There is currently a lack of integration of water industry infrastructure with other drainage and ﬂooding infrastructure in Scotland, although the MGSDP

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The Glasgow Strategic Drainage Plan is addressing this problem for the Glasgow region. Managing stormwater on the surface is the key to future sustainable urban drainage and this will involve the identiﬁcation and creation of designated ﬂood routes; either along green corridors or road surfaces. It would also involve the designation of open space areas within development sites which would be ‘reserved’ for exceedance ﬂows. These recommendations represent signiﬁcant planning issues for urban areas such as the metropolitan Glasgow region and the suggestion for deemed planning permission for certain ﬂood risk management proposals will be highly controversial (Scottish Government 2008). Nevertheless, the GSDP is now included in the Scottish Government’s National Planning Framework. Linking drainage plans to development plans will provide a measure of certainty for developers, allowing a more structured and controlled approach for investor contributions to be identiﬁed and established. Glasgow City Council has asserted the importance of connecting surface water drainage with green infrastructure, public access and good urban design to create regional SUDS schemes and strategic conveyance routes within their local development plans (LDPs). In their LDP for the East End region of Dalmarnock, surface water management will be delivered as an integral element of the green space network (Glasgow City Council 2008). The ﬁnal strategic drainage development plan will need to incorporate policies, legislation, standards and guidance to ensure that the physical and design requirements of dealing with overland ﬂow and both on- and off-site BMPs/SUDS, can be effectively incorporated into the long-term drainage infrastructure. The need for micro-management of new development (e.g. to ensure construction methods and materials, etc.) and retroﬁtted solutions will also have implications for permitted development rights and building control standards and there will be a need for a clearer deﬁnition of responsibilities. The importance of supporting public education and awareness campaigns has been recognized by Glasgow City Council and information on source control approaches has been included as part of ‘ﬂood

fairs’ held around the city. General public engagement in regional drainage infrastructure planning remains, however, an outstanding issue. Additionally, the inclusions of BMP technologies within the brownﬁeld sites of the Greater Glasgow area are not without their own problems. Direct stormwater inﬁltration to ground could lead to widespread sub-surface leachate mobilization and groundwater contamination as well as leaching out into the Clyde estuary. This means that many BMPs/SUDS devices including storage and inﬁltration facilities will have to be lined. The GSDP represents an innovative and challenging planning-led approach to deal with the complexity of issues associated with the control, treatment and management of urban drainage. The multi-agency approach will facilitate stakeholder integration which ultimately beneﬁts both the environment and the local population affected by drainage-related problems. The staged investigations have allowed process interactions to be quantiﬁed in order that the causes of ﬂooding and poor water quality can be better understood as a basis for developing more sustainable and cost-effective solutions. The new legislation supporting the introduction of BMP/SUDS technologies is forcing radical changes in thinking and approaches to urban land use planning, with surface runoff water becoming regarded as a resource rather than a waste efﬂuent. The introduction of dispersed source control approaches will, however, require a regional perspective of their distribution in terms of equilibrium ﬂow and quality regime effects accumulatively downstream. This analysis will also need to be accompanied by the introduction of ﬁxed rules, ﬁnancial incentives and/or general binding codes of practice for urban diffuse pollution to ensure that a satisfactory distribution of BMPs/SUDS is achieved over the long term. References Akornor, O. and Page, D.W. (2004) Glasgow strategic drainage plan Stage 1. In: Proceedings of WaPUG Conference, Dunblane, Scotland, 17 June 2004.

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http://www.wapug.org.uk (last veriﬁed 18 November 2008). Construction Industry Research and Information Association (CIRIA) (2000) Sustainable Urban Drainage Systems: Design Manual for Scotland and N. Ireland. Report C521. Construction Industry Research & Information Association, London. Glasgow City Council (2002) Flooding in the East End of Glasgow. Report to local members of the Scottish Parliament, 30 July 2008. SEPA, Scottish Water and Glasgow City Council, Glasgow. Glasgow City Council (2008) East End Local Development Strategy. Development & Regeneration Services, Glasgow City Council, Glasgow. Jones, P. and Macdonald, N. (2006) Making space for unruly water: sustainable drainage systems and the disciplining of surface runoff. Geoforum, 36, 534– 544. Macdonald, N. and Jones, P. (2006) The inclusion of sustainable drainage systems in ﬂood management in the post-industrial city: a case study of Glasgow. Scottish Geographical Journal, 122, 233–246. Page, D.W. & Fleming, N. (2005) Glasgow strategic drainage plan Stage 2. In: Proceedings of WaPUG Conference, Blackpool, November 2005. http:// www.wapug.org.uk (last veriﬁed 18 November 2008). Scholtz, M. (2005) The Glasgow sustainable urban drainage system management project. In: Proceedings 10th Interntional Conference on Urban Drainage,

Copenhagen, 21–26 August 2005. IWA Publishing, London. Scholtz, M., Corrigan, N.I. and Yazdi, S.K (2006) The Glasgow sustainable urban drainage system management project: case studies. Environmental Engineering Science, 23, 908–922. Scottish Environment Protection Agency (SEPA) (2004) Glasgow Drainage Infrastructure. Report WR 04/12. West Region Board, 15 September 2004. Scottish Environment Protection Agency, Stirling. Scottish Executive (2004) Planning and Building Standards Advice on Flooding. Scottish Executive, Edinburgh. Scottish Government (2004) Scottish Planning Policy SPP7: planning and ﬂooding. Development Department, Scottish Executive, Edinburgh. Scottish Government (2008) The Future of Flood Risk Management in Scotland. Scottish Government, Edinburgh. Sustainable Urban Drainage Scottish Working Party (SUDSWP) (2005) Drainage Assessment: a guide for Scotland. Scottish Environment Protection Agency, Edinburgh. Tufail, S., McKissock, G. and Adshead, H. (2003) Urban watercourses in the East End of Glasgow. In: Proceedings of WaPUG Conference. Blackpool, November 2003. http://www.wapug.org.uk (last veriﬁed 18 November 2008). Water Research Council (WRC) (2007) Sewers for Scotland, 2nd rev. edn. WRC Publications, Swindon.

Image facing chapter title page: Courtesy of the Urban Pollution Research Centre.

20 The Ruhr Catchment (Germany) – the Contribution of Reservoirs to Integrated River Basin Management GERD MORGENSCHWEIS1 1

20.1

Water Resources Department of the Ruhrverband, Essen, Germany

Introduction

The ‘EU Water Framework Directive’ put into force in 2002 created the legal background and the essential prerequisite for the holistic management of rivers within the borders of a natural catchment basin. The main goals of the WFD are the protection and improvement of the status of aquatic ecosystems as well as the promotion of the sustainable use of water resources. The ‘Ruhr River Association’ (Ruhrverband) has undertaken the integrated management of water quantity and water quality in a holistic way in the Ruhr River catchment basin since the end of the nineteenth century. Within the context of a supra-regional ‘division of labour’ in water resources management, the Ruhr river was assigned the role of supplying drinking water for the Ruhr district, the most densely populated and industrialized area in Europe (Fig. 20.1) To achieve the above-mentioned main goals of the WFD, river basin management plans with dedicated water quality measures were to be drawn up. For the effective development of such plans, computer models for the integrated quantitative and qualitative management of water resources are highly useful. Since none of the available models proved to be adaptable to the Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

Ruhr river basin, a model system for water quantity management has now been developed, as a ﬁrst step, by the Ruhrverband in co-operation with the Institute for Water Resources Management at the University of Karlsruhe in Germany. This model can be used as an integral part of operational river basin management on a daily basis. It has been adapted to the watershed of the Ruhr River in Germany, and was implemented in July 2002. During ﬂood events a dedicated ﬂood forecast model operating on an hourly data basis (Göppert et al. 1998) is utilized (see section 20.5).

20.2 The Ruhr Catchment The Ruhr River is situated in a range of low-lying mountains, with a length of 209 km, a catchment area of 4485 km2, and a long-term mean streamﬂow of merely 70 m3 s–1 with strong seasonal variation (Figs 20.2–20.4). Such characteristics do not ﬁt well with the objective of supplying more than ﬁve million people plus industries and businesses in the Ruhr district (Fig. 20.1) with drinking and process water. Consequently, supply deﬁcits have occurred especially during prolonged dry periods. Although these water-scarcity problems were particularly apparent in times of natural drought, their primary cause was the withdrawal of water for human use. Very substantial water

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Fig. 20.1 Ruhr district and Ruhr catchment basin with reservoirs and water catchment.

Fig. 20.2 Lower reach of the Ruhr River at Duisburg ﬂowing into the Rhine River. Photograph by Ruhrverband Essen.

Fig. 20.3 Middle reach of the Ruhr River at Essen with inﬁltration basins. Photograph by Ruhrverband Essen.

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Table 20.1 Water abstractions and water losses in the Ruhr catchment basin

Fig. 20.4 Upper reach of the Ruhr River. Photograph by Ruhrverband Essen.

abstractions have been common in the Ruhr River (Table 20.1). The annual abstraction exported to neighbouring catchment basins represents a ‘loss’ for the Ruhr River and has to be replaced by additional water supplied by reservoirs. It shows that, at times of the highest water demand (e.g. 1975), 1.35 B m3 of water was withdrawn from the catchment basin per year. Even today (water year 2007), about 515 M m3 of water is abstracted and about 225 M m3 is pumped into the neighbouring catchment basins of the Emscher and Lippe Rivers, where coal mining and heavy industry historically exerted an enormous impact on natural water supply structures (Table 20.1). On the other hand, low-ﬂow situations occurred periodically in the Ruhr River, especially during dry periods in the summer and autumn. Apart from creating water quantity shortages, such low-ﬂow periods caused tremendous public health problems: the region experienced frequent epidemics of malaria, typhoid and amoebic dysentery up to the end of the nineteenth century. To deal with these problems, a comprehensive water resource management system was established during the ﬁrst decades of the twentieth century. The basic objective of this system was to divide water management tasks among the

Year

Water abstraction (m3 a−1)

Water losses (m3 a−1)

1900 1925 1950 1975 2000

179 545 892 1031 527

128 287 384 397 272

Max. Actual

1353 515

474 225

different river basins. The Ruhr River was assigned the task of providing a supraregional supply of drinking water to a geographical region including areas located within neighbouring catchment basins (e.g. the Emscher and Lippe River basins).

20.3 Operational Water Quantity Management in the Ruhr River Basin To compensate for water deﬁcits caused by exports and for natural ﬂuctuations in runoff, a system of reservoirs (Fig. 20.1) – which currently provides a total storage capacity of 473.6 M m3 – was established on the upper reaches of the Ruhr River and its tributaries. The physiographical features of the Ruhr River basin are extraordinarily favourable for dam-building. For instance, there is relatively high precipitation in the mountainous areas in the eastern part; here land use is dominated by forests and the mountainous terrain slopes from east to west, enabling the river to transport water to the densely populated regions in the low-lying areas (Fig. 20.1). Given these geographical features, dams were built and reservoirs created. Besides several smaller reservoirs, there are two key groups of reservoirs: the so-called Northern Group comprising the Moehne Reservoir (134.5 M m3), the Sorpe Reservoir (70 M m3) and the Henne Reservoir (38.4 M m3), and the Southern Group with the Bigge Reservoir (171.7 M m3), the Verse

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Fig. 20.5 Aerial photograph of the Bigge Dam. Photograph by Ruhrverband Essen.

agement. In this way, the integration of water quality and water quantity within the borders of a natural catchment basin was guaranteed from the beginning of the Association’s history. The Ruhr River Association Act of 1990 created the present legal foundation for the management of the reservoirs in the Ruhr catchment area. In addition, there are regulations applicable to particular reservoirs; these are deﬁned in the plan approval documents for the individual reservoirs. On the basis of these statutory regulations – and taking account of other background conditions affecting reservoir operation and water quality management – primary and secondary criteria for the control of the reservoirs, especially during low-ﬂow periods, have been deﬁned.

Box 20.1 Primary control criteria for reservoir operation Primary control criteria – compliance with the threshold values for minimum ﬂow: 1

Villigst

5-day average: 8.4 m3 s−1 Lowest daily average: 7.5 m3 s−1

Fig. 20.6 The Moehne Dam. Photograph by Ruhrverband Essen.

Reservoir (32.8 M m3) and the Ennepe Reservoir (12.6 M m3) (Figs 20.1, 20.5, 20.6). Only after the Bigge Reservoir was put into operation in 1965 was the system able to meet the demands for abstraction. All of this was carried out under the aegis of two water authorities, which were founded by special acts of the Prussian legislature in 1913: the Ruhr Reservoir Association (Ruhrtalsperrenverein) responsible for water quantity management and the Ruhr River Association (Ruhrverband) responsible for water quality man-

2 Hattingen to Duisburg (where the Ruhr enters the Rhine) 5-day average: 15.0 m3 s−1 Lowest daily average: 13.0 m3 s−1 3

Henne – special control scheme

Threshold at Oeventrop: 2.5 m3 s−1

Thus water quantity management in the Ruhr River basin focuses on maintaining runoff threshold values during both low-ﬂow periods and ﬂood events. Because of the supraregional role of water supply placed on the reservoir system it is controlled from the Headquarters of the Ruhr River

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Fig. 20.7 Measuring network for control and ﬂow regulation of the reservoirs within the Ruhr catchment basin.

Association located at Essen, more than 100 km away from the individual reservoirs. Decisions on water releases from the individual reservoirs are taken in real time on the basis of a complex hydrological network – comprising streamﬂow gauges and precipitation measuring stations – combined with sophisticated long-distance data transmission equipment, a monitoring system, and the outputs of mathematical models. A simpliﬁed overview of the present network of streamﬂow gauging stations established to record surface runoff is shown in Figure 20.7. With a total of 96 stations, this network has an above-average density of gauging stations. The control sections at Villigst and Hattingen are critical locations as they are where a minimum runoff has to be maintained (in conformance with statutory regulations) during daily water management operations (Box 20.1). Parallel to the network of gauging stations, a precipitation measurement network, with a total of 94 stations, is

operated in collaboration with the German Weather Service. In summary, there is currently a closed information system, including facilities for long-distance data transmission, in existence in the catchment area of the Ruhr via which all relevant high-frequency system control data are conveyed in real time to the ﬁve local reservoir operation centres and Headquarters in Essen. The data are then processed at an Operation Centre (Fig. 20.8) with the aid of state-of-the-art visualization techniques and made available for the real-time control of the system (Morgenschweis 2001, 2003). As a result, information on the state of the system throughout the entire river basin is available in real time, allowing the Association to react immediately to any changes in the hydrological or meteorological conditions. In addition, a forecasting system was installed in the Operational Centre in 2002, continually updates

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gerd morgenschweis demand, Morgenschweis 1995) was developed using fuzzy logic techniques. On the basis of the quantities taken by the major waterworks along the Ruhr and the air temperature forecast of the German Weather Service, this model can predict withdrawals at the control river sections of Villigst and the mouth of the Ruhr at Duisburg for 1–5 days within a 95% conﬁdence interval. This model has been in daily use since 1994. The model has proved to be very useful in low-ﬂow situations and is thus an important component of the system set up to safeguard the water supply of one of the most densely populated regions of Germany.

Fig. 20.8 The operation centre of Ruhrriverband. Photograph by Ruhrverband Essen.

expected conditions in the entire basin for several days, hence increasing operational efﬁciency.

20.4

Operation of Reservoirs During Low-Flow Periods 20.4.1

Daily water demand

Water abstractions have a large impact on streamﬂow in the Ruhr River under low-ﬂow conditions. Water demand is highly variable, but since the mid 1970s, an overall declining trend in water demand has been observed. Weekly variation (periodicity) in water demand along with strong seasonal patterns (maximum withdrawals in June) is highlighted in Figure 20.9. Since the reaction time of the river after changes in water releases from the reservoirs may be between 3 and 5 days under low-ﬂow conditions, forecasts covering this period must be available for the control of the reservoir system. Moreover, since considerably larger water volumes are withdrawn from the Ruhr River during low-ﬂow periods to ensure water supply, it is necessary to estimate future withdrawals for up to 5 days in advance. For this purpose the forecasting model EZVOR (EntZiehungsVorhersage = forecast of daily water

20.4.2 Daily water quantity management The forecasted water demand data, the daily weather forecasts of the German Weather Service (DWD), and the transmitted measurement values of stations within the catchment are input data of a real-time river basin simulation model (RRM), which has been developed as a network ﬂow model consisting of a limited number of system elements such as river reaches, reservoirs, barrages, gauging stations, water withdrawals, water inﬂows and ﬂow control structures (e.g. pumps, weirs, water gates, etc.), which are combined in modular fashion to simulate the complex system. The model includes nodes and transport elements with different features which are interconnected to reproduce the topology of the river basin (Morgenschweis et al. 2002; Morgenschweis 2003). The calibration and validation of the model were carried out with historical data from the four water years 1992 to 1995 using the ‘splitsample test’. As is shown by the comparison of measured and calculated hydrographs at the Hattingen gauging station (Fig. 20.10) situated 60 km upstream of the mouth of the Ruhr River, the ﬂood-routing method is suitable for the calculation of variable ﬂow velocities. The model structure accommodates the various man-made structures upstream of this gauging station (e.g. reservoirs, barrages, sewage plants, withdrawals, inlets, etc.; Fig. 20.11), and displays a

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Fig. 20.9 Daily water exports (m3 s−1) in the 1996 catchment water year.

Fig. 20.10 Comparions of measured and calculated hydrographs at the ganging station Hattingen for different periods: (a) entire water years of 1992 and 1993; (b) low-ﬂow period in 1993; (c) ﬂood period in 1993.

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Fig. 20.11 Wastewater treatment plant at Essen. Photograph by Ruhrverband Essen.

high performance (Morgenschweis 2003) for annual and low-ﬂow conditions (Fig. 20.10a,b). On the other hand, however, the results of model calculations during ﬂood events have proved to be unsatisfactory as the daily time-step is not at a high enough resolution (Fig. 20.10c). For this reason, ﬂood forecasting is carried out using a dedicated model operating on an hourly data basis (see section 20.1). The real-time model enables engineers to view river basin management as an integrated entity and to make decisions that take account of the state of the system at any time and place. The software system ensures that solutions for individual management aspects – e.g. forecasts of water exports and real-time reservoir management in the basin of the Ruhr River – are connected and combined. This information can serve as the basis for advanced water quality modelling and water quality management in river basins (see section 20.6). Modelled and measured data for the streamﬂow gauging stations Villigst (at the mid reach of the Ruhr River), Hagen-Hohenlimburg (at the Lenne River), Hagen-Eckesey (at the Volme River) and Wetter (at the lower reach of the Ruhr River) are shown in Figure 20.12. The measured and simulated values should display a good congruency, e.g. as at the gauging station of HagenEckesey. Forecasted values for up to 6 days are

also shown; with forecasts for the ﬁrst 3–4 days corresponding extremely well with the measured values. In contrast to this, the forecasts for the ﬁfth and sixth days are unsatisfactory due to low precipitation forecasts, highlighting the importance of this latter factor in model performance. The experience gained during 6 years of use shows that the RRM model system is a powerful tool for operational water quantity management in the Ruhr catchment basin. It is able to simulate a structured watershed with strong anthropogenic impacts and to generate short-term forecasts for real-time system operation. The model system can be applied for prognostic long-term simulations as well as for operational short-term forecasts. The implemented ﬂood-routing method, a modiﬁed double cascade model, is able to reproduce the variable ﬂow velocities in river reaches, which have to be taken into particular consideration for low-ﬂow conditions. A comparison of measured and simulated runoff volumes during the testing period of the initial model is shown in Table 20.2. The differences at the most important control sections Wetter and Hattingen – which are situated in the lower reach of the Ruhr River (Fig. 20.7) and thus integrate most of the anthropogenic impacts on the entire catchment area – are highly correlated. Because the input and output data of the model system are closely related to the spatial structure of a modelled river basin, the management tool has been incorporated into a Geographic Information System (GIS). This facilitates spatial visualization of the system state of the entire river basin by the model users, who are water engineers with a broad scientiﬁc background.

20.5

Operation of Reservoirs During Flood Events

20.5.1 Problems and approaches to ﬂood management The catchment basin of the Lenne River, the main tributary of the Ruhr River (Fig. 20.1),

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Fig. 20.12 Real-time forecast at control gauging stations. Location of sampling sites shown in Fig. 20.7. The dotted line represents the time of forecasting.

Table 20.2 Comparison of measured and simulated runoff volumes as a measure of model quality Gauging station Villigst Hohenlimburg Eckesey Wetter Hattingen Werden Muelheim

Measured (mm3)

Stimulated (mm3)

Difference (%)

1001 893 311 2197 2292 2431 2463

927 899 291 2181 2261 2308 2332

7.35 0.69 6.31 0.71 1.35 5.05 5.30

covers an area of 1357 km2 and has a very distinct topography, with the contributing streams cutting deeply into the narrow gorges and with narrow ﬂood plains. The Lenne catchment area

is a typical European mountainous region covered mainly by forests. An analysis of the historical ﬂoods occurring during the period 1968–1993 revealed a signiﬁcantly higher occurrence of ﬂood events in the European winter season between October and March. A preliminary investigation showed that western circulation patterns, in particular of the type ‘west cyclonic’, are responsible for most ﬂoods. These are very stable weather conditions characterized by a rain event of several days’ duration and intensity (Bardossy 1993). With extreme events occurring predominantly during the winter months, the probability of temperatures falling below the freezing point – at least temporarily – is high. The analysis of historical ﬂoods revealed that approximately half of

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the major events were inﬂuenced by snow. It should be noted that the runoff during these winter events is determined mainly by a change in the form of precipitation, for instance from snow to rain, and not primarily by melting of snow (Göppert 1995). The characteristics of the catchment area, together with the spatial distribution of precipitation that is typical for the region, cause runoff conditions which are characterized by: • a quick response of sub-catchments to rainfall and steeply rising limbs of ﬂood hydrographs; • slow recession of ﬂoods after the end of precipitation in areas with distinct retention behaviour, especially wooded areas; • long-lasting, predominantly multi-peak events of 4–6 weeks’ duration. There are two reservoirs located in the Lenne catchment basin (Fig. 20.1): • the Bigge Reservoir with a storage capacity of 171.7 M m3 (including the Lister Dam); • the Verse Reservoir with a storage capacity of 32.8 M m3 (including the Fuerwigge Dam). Together these two reservoirs control the discharge of about 25% of the entire Lenne catchment basin. In accordance with legal requirements, the two reservoirs must be operated so as to maintain minimum runoff at certain control sections along the Ruhr River at all times and to reduce ﬂood damage downstream of the reservoir. The only reservoir with an explicitly deﬁned ﬂood protection function is the Bigge Reservoir, which has a maximum ﬂood protection capacity of 32 M m3 that may be used during the winter months. Two observations have been made repeatedly during ﬂood events in recent years: • The ﬂood-protection storage volume of the Bigge Reservoir is not sufﬁcient to retain the entire ﬂood volume of major events (Morgenschweis and Heitefuss 2005). • The present discharge capacity in the lower reaches of the Lenne River is not sufﬁcient to prevent ﬂood damage even during medium ﬂoods. Owing to the very limited possibilities for either improving the discharge capacities of the focal points along the Lenne River or for increasing the

Fig. 20.13 Flooded street beside the Lerne River. Photograph by Ruhrverband Essen.

ﬂood-protection storage volumes of the reservoirs, the existing system had to be optimized. The optimization strategy was based on the decision to use the reservoirs mainly for protection against minor and medium ﬂood events (Fig. 20.13). To put this philosophy into practice, three prerequisites were essential for ﬂood management of the reservoir system (cf. section 20.3): 1 a comprehensive information network; 2 an online ﬂood forecasting model; 3 a ﬂood warning system. With regard to the ﬁrst prerequisite, a comprehensive hydrological information network with automatic data transmission and data visualization has already been installed in an Operation Centre at the Association headquarters. The second prerequisite is the existence of an online forecasting model enabling the Association to develop optimal control strategies during a ﬂood event. To ensure improved ﬂood protection at critical points downstream of the reservoirs, the forecasting model, which was developed by the Institute for Water Resources and Hydrology of the University at Karlsruhe (IWK) in Germany, was based on a spatially distributed rainfall–runoff model. This model – the Vorhersagemodell or Flood Forecast Model (VMOD) – calculates the runoff at critical points downstream of the reservoirs with a high degree of accuracy (Göppert et al. 1998). It was implemented for the Lenne subsystem in 1995 and expanded to cover the entire catchment basin in 2002.

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Fig. 20.14 Flood forecasts for Altena during the October 1998 ﬂood.

The ﬂood forecasting model proved to be very useful for optimizing reservoir management. With this model, it has been possible to precisely forecast peak runoff at the critical focal point of Altena (Fig. 20.14). Figure 20.14 shows the results of various model calculations for 28 October 1998 compared with the measured values. The inﬂuence of the precipitation forecast on the ﬂood forecast is clear. The high performance of the forecast can be demonstrated by the small difference between the predicted discharge of 429 m3 s−1 and the measured value of 418 m3 s−1. This difference is within the range of model accuracy. The third prerequisite is an effective ﬂood alarm and warning system. In the Ruhr district there are two clearly demarcated roles; ﬂood forecasting is carried out by the Ruhr River Association and ﬂood warnings are issued by the State Environmental Agency in the city of Hagen. 20.5.2 Challenges and limits of the present system At the end of October 1998 a huge ﬂood occurred in the Ruhr catchment basin. It was caused by heavy rains throughout the entire catchment

basin from 24 to 31 October 1998 after a period of unusually high precipitation. Figure 20.15 shows the balance of inﬂow, discharge and storage volume for the Bigge Reservoir during the ﬂood period. Clearly the entire peak inﬂow of more than 200 m3 s−1 had been retained in the reservoir. Owing to the inadequate ﬂood storage volume, an intermediate discharge period was necessary during the decline of the ﬁrst ﬂood wave (Morgenschweis and Heitefuss 2007). Additional calculations performed with deterministic and stochastic hydrological models as well as hydrodynamic models have revealed that the ﬂood protection capacity of the existing reservoir system is limited to ﬂood events with a return period of about 50 years (HQ50). This means that more extreme ﬂood events will deﬁnitely cause major ﬂood damage along the Ruhr and Lenne Rivers. For this reason, risk assessment techniques have been employed to develop Flood Action Plans (State Environmental Agency Hagen 2002; State Environmental Agency Duisburg 2003). These Flood Action Plans have been developed by the State Environmental Agencies and Ruhrverband to perform two functions: 1 Assess the extent of the expected ﬂooding and the probable damage caused by ﬂoods with a

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Fig. 20.15 Operation of the Bigge Reservoir during the October 1998 ﬂood.

return period of 100, 200 and 1000 years, respectively. 2 Propose measures for reducing or preventing damage (e.g. local ﬂood protection, ﬂood prevention, disaster management plans, evacuation plans, etc.). The Flood Action Plans for the Ruhr River and its main tributary, the Lenne River, have been presented to the public and were put into force in 2002 and 2003 after discussion with local authorities, industry and other stakeholders (Fig. 20.16). As part of an ongoing process, the local authorities are now putting the proposed measures into action and the state Ministry of Environment is granting ﬁnancial aid to the affected communities.

Fig. 20.16 Flooded Ruhr valley near Essen. Photograph by Ruhrverband Essen.

The Ruhr Catchment (Germany)

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20.6 Integrated Water Quantity and Water Quality Management Linking water quality with water quantity management also presents a major challenge to integrated river basin management (Bode and Morgenschweis 2001). This is illustrated by the following two examples. 20.6.1 Reduction of ammonia nitrogen concentrations in the lower Ruhr by increasing the amounts of water released by the reservoirs Elevated ammonia concentrations in river water are very undesirable. Ammonia is released during the biochemical decomposition of nitrogenous substances (e.g. proteins, urea, etc.) and is thus present in certain concentrations in bodies of water. Elevated ammonium concentrations in water cause intensiﬁed chlorine consumption during water treatment processes. The chloramines formed during chlorination are responsible for the typical ‘chlorous’ smell of drinking water. The nitrogen in the ammonium ion is converted under aerobic conditions – both during sewage treatment and as a result of degradation by the microbes present in water – to nitrite and then to nitrate. During sewage treatment part of the nitrate thus formed can be denitriﬁed under anoxic conditions to form elemental nitrogen. Both processes, but especially nitriﬁcation, are temperature-dependent. Moreover, the activity and growth of the nitrifying micro-organisms are slower in colder water. This phenomenon is especially relevant for surface water, where water temperatures around freezing point can occur for several days during cold periods in the winter. Analysis of the ammonia nitrogen concentrations in the Ruhr at Essen and the associated water temperature during the period 1980–1990 conﬁrm this temperature dependency (Nusch 1997). This process is evidently reinforced by the simultaneous occurrence of low ﬂow in the Ruhr River. Detailed investigations have established

Fig. 20.17 Weir and Lake Hengsty. Photograph by Ruhrverband Essen.

a signiﬁcant correlation between high ammonia nitrogen concentrations and low ﬂow in the Ruhr. It is evident, therefore, that critical concentrations of ammonium can occur in the Ruhr when low ﬂow coincides with low water temperatures during longer periods of cold weather. The consequence for water quantity management is that larger amounts of water should be released from the reservoirs whenever: 1 The water temperature drops below 6 °C. 2 The ammonium concentration rises above 1.0 mg1−1. 3 Runoff falls below 10 m3 s−1 at the gauging station Villgst/Ruhr (Fig. 20.2). As soon as a trend of more than 3 days’ duration is observed, especially an increase in the ammonium concentration, it is necessary to release additional amounts of water (Fig. 20.17). However, a decision to release additional water must always be contingent upon having sufﬁcient water levels in the reservoirs for the particular time of year. This integrated approach to river quality management was successfully tested in the Ruhr catchment basin during the period from 1984 to 1991 and introduced in the winter of 1996/97 as a standard operational procedure (Bode and Morgenschweis 2001).

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gerd morgenschweis 20.6.2 Increased ﬁsh mortality in the Ruhr in the spring

In recent years increased ﬁsh mortality has been observed in the lower reach of the Ruhr River during the period from mid-May to mid-June. (‘Corpus Christi Day ﬁsh deaths’, Nusch 1997). Analysis of the water quality and water quantity indicators during the period 1989–1998 indicates that some quality indicators (e.g. chlorophyll a, pH and ammonia nitrogen) display signiﬁcantly higher values while runoff – measured at the Hattingen gauging station for example – is usually low. The low runoff in the middle and lower reaches of the Ruhr River in May and June can be explained by the operating rules of the reservoirs system, which are based exclusively on criteria used for water quantity management. After the winter months (November to April), when the main goal is to store as much water as possible in the reservoirs under normal conditions, it is necessary to release additional water from the reservoirs during the summer months (May to October) to maintain the minimum discharge of 15 m3 s−1 required by law in the lower reach of the Ruhr River (cf. chapter 20.3). With the exception of the year 1996, in which lower limit values were permitted because of the extremely low precipitation in the winter, discharge never fell below 20 m3 s−1 at the Hattingen gauging station during this period. Nevertheless, ﬂow conditions appear to be one of the main factors involved in the increased ﬁsh mortality observed in the spring. This is because of the nearly stagnant ﬂow conditions prevailing in the impounded part of the Ruhr River and the resulting increase in the concentrations of ecologically relevant contaminants. This can result in massive growth of plankton algae, pH change and the development of eutrophic conditions. Since the processes described above are time-dependent and temperature-dependent, the release of increased amounts of water from the reservoirs can reduce the frequency and intensity of ﬁsh mortality in the lower Ruhr during the spring months. However, this is critically dependent on annual variations in rainfall and the pres-

sures placed on available reservoir storage by water quantity demands. Taking account of all of these considerations, the Ruhrverband has been implementing the procedures described above since the spring of 1999. None of the water quality indicators alone, but rather the interaction of several abiotic factors (e.g. runoff, water temperature) and biotic factors (e.g. spawning stress, susceptibility to parasitic infection), is responsible for the increased ﬁsh mortality occurring in the spring. For this reason, the policy is to release water whenever an increased number of dead ﬁsh (i.e. > 50 ﬁsh day−1) are found at Lake Baldeney, situated near Essen in the lower reach of the Ruhr River. Water quality can be predicted on the basis of weather forecasts issued by the German Weather Service and in this way, the water quality managers at Ruhrverband gain valuable time in which to undertake effective corrective action. An example of this occurred in mid May 1999, when a water release from the Moeöhne Reservoir was increased by 6 m3 s−1, thereby reducing ﬁsh mortality downstream.

20.7 Summary and Future Issues For several decades now the reservoir system in the Ruhr catchment area has been contributing to river basin management in many different ways: 1 It has assured the security of the water supply in the densely populated Ruhr district by making controlled releases of water from the reservoirs. During the summer months of 2003, for example, the Ruhrverband was able to supply sufﬁcient water to the population and to industry despite extreme high temperatures and low water levels. 2 It has reduced peak runoff during ﬂood events via dynamic water retention in the reservoirs. This water management instrument has effectively prevented ﬂooding in endangered cities, in particular in the Lenne Valley, in recent years. 3 It has improved the water quality situation in the Ruhr via the controlled release of water from

The Ruhr Catchment (Germany) the reservoirs. For example, the problem posed by elevated ammonia nitrogen concentrations during cold periods in the winter and increased ﬁsh mortality during the spring months. The successful employment of the reservoir system depends on sophisticated communication technology and river basin modelling as well as on comprehensive limnological investigation programmes involving monitoring of water quality indicators. Next to models of water quantity management, which have been expanded in recent years to include the entire catchment area and the scope of water management activities, the integrated water resource management approach will be broadened to include an operational water quality model. The integrated and catchment-area-based approach to water management foreseen by the EU Water Framework Directive has already been applied successfully by the Ruhrverband for several decades. The management of the reservoirs systems in the Ruhr catchment area addresses the requirements for quantity, quality and ecological target laid down by the EU Water Framework Directive.

References Bardossy, A. (1993) Stochastische Modelle zur Beschreibung der raum-zeitlichen Variabilität des Niederschlages (Stochastic models to characterize the variability of precipitation in space and time). Universität Karlsruhe, Institut für Hydrologie und Wasserwirtschaft, Mitteilunge Heft 44 (in German). Bode, H. and Morgenschweis, G. (2001) Beitrag der Talsperren zum Flussgebietsmanagement im Einzugsgebiet der Ruhr. Schriftenreihe der Arbeitsgemeinschaft der Trinkwassertalsperren, 3, S.69–94. Göppert, H.G. (1995) Operationelle Hochwasservorhersage zur Steuerung von Talsperren (Operational ﬂood forecasting for the management of reservoirs). Universität Karlsruhe, Institut für Hydrologie und Wasserwirtschaft, Mitteilungen Heft 49 (in German).

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Göppert, H.G., Morgenschweis, G., Ihringer, J. and Plate, E.J. (1998) Flood forecast model for improved reservoir management in the Lenne catchment, Germany. Hydrological Sciences Journal, 43, 215– 242. Morgenschweis, G. (1995) Kurzfristige Vorhersage der Wasserentnahme aus einem Flußgebiet. Proceedings der 8. Wiss. Tagung des DVWK an der Ruhr-Universität Bochum, 15 S. Morgenschweis, G. (2001) Echtzeitbewirtschaftung eines Flussgebietes am Beispiel der Ruhr. Wasserwirtschaft, 91, 575–581. Morgenschweis, G. (2003) Real-time river basin management in the Ruhr catchment basin/Germany. In: Mostoﬁnejad, D. and Chamani, M.R. (eds), Proceedings of the 6th International Conference on Civil Engineering (ICCE). Water Resources, Geotechnics and Transportation Series 7. Isfahan University of Technology (IUT), Isfahan, Iran, pp. 1–9. Morgenschweis, G. and Heitefuss, Chr. (2005) Use of a reservoir system for ﬂood control in the Ruhr River basin in Germany: positive effects and limitations. In: Proceedings of the 73rd Annual Meeting of ICOLD. May 1–6, 2005 in Tehran/Iran. Theme S2, pp. 1–7. Morgenschweis, G. and Heitefuss, Chr. (2007) Operational Flood Protection Management in the Ruhr River Catchment Basin, Germany: challenges and limitations. IAHS publication no. 317. IAHS, USA, pp. 305–311. Morgenschweis, G., Brudy-Zippelius, T. and Ihringer, J. (2002) Operational water quantity management in a river basin. Water Science and Technology, 10, 111–118. Nusch, E.A. (1997) Water quantity and quality in lakes and reservoirs for human uses. In: Jorgensen, S.E. and Matusi, S. (eds), Guidelines of Lake Management, vol. 8. International Lake Environment Committee; United Nations Environment Programme, Shiga, Japan, pp. 121–146. State Environmental Agency Duisburg (2003) Hochwasseraktionsplan Ruhr (Flood Action Plan Ruhr) Duisburg. www2.brd.nrw.de/Dezernat 53 State Environmental Agency Hagen (2002) Hochwasseraktionsplan Lenne (Flood Action Plan Lenne) Hagen. www.stua-ha.nrw.de/map/p/hwlenne/main/ tr/Frame.html

Image facing chapter title page: Courtesy of Ruhrverband Essen.

21 Evolution of River Basin Management in the Okavango System, Southern Africa P I O T R W O L S K I 1, L A R S R A M B E R G 1, L A P O M A G O L E 1 AND DOMINIC MAZVIMAVI1 1

University of Botswana, Harry Oppenheimer Okavango Research Centre, Maun, Botswana

21.1

Introduction

The Okavango River basin is a large, hydrologically unique basin containing one of the world’s most magniﬁcent wetlands – the Okavango Delta. The basin is shared between three countries: Angola, Namibia and Botswana, with multiple users, multiple interests and, hence, multiple stakeholders present within each of them. The water resources in the basin are currently little utilized, and the basin and the Delta are in near pristine conditions. This situation is not a result of conscious management efforts as these have been very weak or non-existent. Rather, it is a result of conditions in the basin, both natural and socio-economic, that have not favoured development. Almost the whole basin is on Kalahari sand that is very nutrient poor, has low agricultural potential and thus the basin has always had low population density. In Angola, from where the majority of the basin’s water currently emanates, there was civil war from 1975 until 2002 and no development has taken place. In addition, the Okavango Delta in Botswana has been infested with tsetse ﬂies, which effectively kept livestock out and thus provided an exclusive environment for African wildlife. Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

With peace in Angola, population growth and an imperative to improve the livelihoods of the predominantly poor people inhabiting the basin, increased pressure on water resources in the basin is inevitable. From upstream to downstream, there is a clear gradient in development priorities. Angola is re-populating its part of the basin with agriculture as the economic cornerstone and has plans for irrigation and construction of a number of hydroelectric power plants. Namibia is putting emphasis predominantly on agriculture and irrigation. Botswana, in turn, has a thriving and economically very important tourism industry in the Okavango Delta which is based on wildlife and is dependent on the conservation of the sensitive ecosystem. There is also here, however, pressure on water resources for the growing population in the vicinity of the Delta. As a result, management of the basin and its water resources requires balancing of often conﬂicting needs and international co-operation and co-ordination in water resource utilization (Turton et al. 2003). The current management paradigm in the basin is that of sustainable development. Problems facing decision-makers in the basins are, therefore, centred on how to achieve a balanced development of local populations without compromising the integrity of the basin’s natural environment and thus without jeopardizing the basis for such development. To achieve this, a

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management framework has been established at the local and international levels. Here we describe how the current management institutions function, what challenges they face and what their achievements are against a background of social and natural features and processes.

21.2 Geography, Hydrology and Biology 21.2.1 Geology and geomorphology The Okavango system is a part of the endoreic Makgadikgadi basin located in central southern Africa (Fig. 21.1). Based on hydrological differences related to geology, topography and climate the Okavango system can be divided into three

Fig. 21.1 The Okavango Basin.

parts: the Okavango River catchment, the Okavango Delta and the Boteti sub-basin. The whole basin, apart from the headwaters of the Cubango where crystalline rocks outcrop, is covered by Kalahari sands. These are highly weathered, well sorted, ﬁne to medium sands, of mixed aeolian-ﬂuvial origin, the nature of which projects strongly on the character of the basin (Thomas and Shaw 1991). The headwaters of the main tributaries, the Cubango and the Cuito, occupy a topographically higher area (1500– 1800 m a.s.l.) where mean annual rainfall is over 1200 mm yr−1 and the drainage network and topographic relief are well developed. The mid reaches of the Okavango River, as well as the Delta and Boteti basin, occupy a ﬂat plateau (900–1100 m a.s.l) with topographic gradients smaller than 1 : 3000 (McCarthy et al. 1997). This, combined

Evolution of River Basin Management in the Okavango System, Southern Africa with the very permeable nature of the Kalahari sands, and low rainfall (less than 450 mm yr−1), makes the area essentially devoid of structured continuous drainage other than that of the Okavango River. Fossil valleys indicate, however, that in the (geological) past, rainfall in the region must have been much higher than present. The upper part of the Okavango Delta, the Panhandle, is a broad ﬂat-bottomed valley with anastomosing and meandering channels and a system of permanent and seasonal ﬂoodplains. The Panhandle broadens into the Okavango Delta proper, which geomorphologically is a large alluvial fan (McCarthy 1993). The upper part of the fan is occupied by a central swamp; a broad, featureless permanently inundated area. This separates into four main distributaries, each of which splits into a set of secondary distributaries. There is no consistent channel continuity within the Delta and a large part of the ﬂow takes place in the form of overland ﬂow through interconnected ﬂoodplains of breadth reaching 20 km. The distributaries terminate at a system of northeast to south-west trending faults, where ﬂows are taken over by rivers collinear with the faults, the Thamalakane and the Kunyere. Two topographic depressions are present at the north-east and south-west extensions of the faults, which at times form lakes, but tend to dry out in drier periods. The Boteti sub-basin comprises essentially a single river channel of ephemeral character, which terminates in a small alluvial fan in the Makgadikgadi pans. The Boteti River currently receives its water exclusively as a spillover from the Okavango Delta through the Thamalakane River. 21.2.2 Rainfall, runoff and ﬂooding The runoff from the Okavango River catchment amounts on average to 9306 Mm3 yr−1, which is approximately 7% of total average catchment rainfall (Fig. 21.1, Table 21.1). This adds to approximately 4800 Mm3 yr−1 falling over the Okavango Delta in the form of local rainfall. The ﬂow into Boteti averages only 20 Mm3 yr−1.

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Table 21.1 Variability of inputs to the Okavango Delta: local rainfall and ﬂow of the Okavango River

Annual Minimum Maximum Average 5-year period Minimum Maximum

Rainfall at Maun (mm yr−1)

Okavango River ﬂow at Mohembo (Mm3 yr−1)

150 (1994/95) 1,190 (1973/74) 448 (1922–2008)

5,266 (1996) 16,012 (1963) 9,300 (1934–2008)

314 (1925–1929) 700 (1973–1977)

5,909 (1993–1997) 12,614 (1961–1965)

Outﬂow from the Delta is, therefore, only 2% of the inputs (Fig. 21.2). The remainder is removed from the Delta by evapotranspiration. Importantly, however, a signiﬁcant part of the evapotranspiration takes place through terrestrial vegetation, supplied by ﬂood water inﬁltration and lateral groundwater ﬂows of local character (Ramberg et al. 2006b). The water balance indicates that this process uses up to 24% of the total inﬂow (Wolski et al. 2006). Rainfall in the basin is highly seasonal, with distinct wet (November–April) and dry (May– October) seasons. The ﬂow in the Okavango River is perennial, however, with an annual ﬂood event. Inter-annual variability in rainfall and ﬂows is moderate but there is evidence of dry and wet pluri-annual periods occurring at 9 and 40year time frames (Tyson et al. 2002; Mazvimavi and Wolski 2006). The ﬂooding in the Delta is predominantly caused by the seasonal ﬂood pulse from the Okavango River. This causes the inundation to expand along the Delta, a process that takes about 6 months, with annual minimum inundation occurring in February–March and maximum in August–September (McCarthy et al. 2004). Local rainfall contributes to inter-annual variation in ﬂood extent, but causes expansion of the inundated area only during high rainfall years (McCarthy et al. 2004). The hydro-period conditions vary throughout the Delta. The central swamp is permanently inundated and is surrounded by a system of seasonally ﬂooded plains,

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Fig. 21.2 Long-term discharge of the Okavango River at Mohembo.

but the frequency of ﬂooding here has large variations from regular-annual to once or twice in a 40-year period. These seasonal ﬂoodplains are not ﬁxed in space, but migrate, expand and shrink in response to inter-annual and pluri-annual variation in hydrological inputs. 21.2.3

Fate of solutes

The transport of waterborne solutes into the Delta, being in effect an endoreic system, would be expected to cause accumulation of precipitated salts at the surface. Yet the Okavango Delta is a lush, freshwater environment. This is due to the dominant role of inﬁltration in the system and chemical evolution of groundwater associated with Okavango Delta islands. In the model summarized by Ramberg and Wolski (2008), inﬁltrating water is drawn towards islands by the evapotranspirative uptake of riparian vegetation. This uptake causes a concentration of solutes that ultimately leads to precipitation of carbonates and silica to form the island’s soil matrix. The groundwater, chemically enriched mainly with sodium and chloride, is trapped under the island centres due to groundwater depression caused by vegetation transpiration. This groundwater then sinks occasionally to the deeper strata through density-driven ﬂow. The whole process leads to vertical and horizontal stratiﬁcation of

groundwater with freshwater localized to shallow aquifers in areas subject to inundation and saline groundwater occupying deeper parts of the aquifer and present within islands or larger land bodies. This system is kept in balance by the presence of riparian woody vegetation and continuous growth of islands. Degradation of the island vegetation belt can thus potentially lead to mobilization of saline groundwater lenses and destructive salinization of freshwater ﬂoodplains. 21.2.4 Sedimentation and abandonment of streams The stream channels in the Okavango Delta are banked by stands of papyrus and reeds and water can easily seep across banks. The water ﬂowing into the Delta carries little suspended sediment and the majority of sediment is carried in the form of bed load. This cannot be carried across vegetated banks and is trapped in channels, ultimately leading to channel aggradation, development of vegetation blockages and channel abandonment. This process causes major changes in the distribution of water within the Delta at timescales in the order of 100–150 years, but local changes in channel structure and ﬂooding patterns occur at shorter timescales as well. This has clear negative consequences for human activity, as it renders waterways non-navigable,

Evolution of River Basin Management in the Okavango System, Southern Africa jeopardizes infrastructure such as tourist camps and affects water supply schemes and sedentary communities whose livelihoods are based on utilization of wetland resources. It is, however, recognized as a very important process in the maintenance of the Okavango Delta ecosystem, leading to its dynamic vegetation successions and high biological productivity through release of nutrient pools during the drying phase, accumulated and stored before then within the permanently ﬂooded areas. The most dramatic example of the process is that of the Thaoge distributary, which dried out between 1890 and 1920 and turned several thousands square kilometres of wetlands into dry Kalahari savannah. Changes at smaller spatial scales have been observed since in several parts of the system (Wilson 1973; Wolski and Murray-Hudson 2006). As mentioned earlier, the Okavango Delta is located within a tectonically active zone. Tectonic movement is, therefore, another factor that can lead to abandonment and re-ﬂooding of large parts of the Delta. 21.2.5

Vegetation in the basin

The Okavango basin is covered by savannah woodlands with deciduous species in the south and evergreen species in the north. Acacia spp. and Colospermum mopane dominate the south, while Burkea dominates the central part of the basin. In the north, Brachystegia dominates, but the topographically highest parts are covered by a plan alto grassland. Wildlife densities are generally low with the exception of the Okavango Delta but the situation in Angola is largely unknown (Mendelsohn and El Obeid 2004). 21.2.6 Biodiversity in the Okavango Delta For the Okavango Delta area (c. 28.000 km2) the number of identiﬁed plant species is 1300, ﬁsh 71, amphibians 33, reptiles 64, birds 444 and mammals 122 (Ramberg et al. 2006a) (Table 21.2). In the Delta there are large variations in habitat patterns over small distances, although the Delta is very ﬂat and is made up of homoge-

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Table 21.2 Number of species in taxonomic groups of originally terrestrial origin observed in each major habitat in the Okavango Delta (modiﬁed from Ramberg et al. 2006) Taxonomic group

Total number of species

Aquatic/ perennial swamp

Wetland/ seasonal swamp

Dry land/ terrestrial

Plantsa Reptiles Birds Mammals

1061 64 444 122

205 7 112 3

519 5 57 21

704 52 275 110

a

From Snowy Mountains Engineering Corporation (1990) and therefore the total number of species differ from the ﬁgure in the text.

neous sand. Small differences in altitude of 1–2 m lead to large differences in the frequency and duration of ﬂooding, which creates habitat gradients; from permanent rivers and lagoons to permanent swamps with reeds and papyrus, to seasonally ﬂooded grasslands; occasionally ﬂooded grasslands; riverine woodlands and dry woodlands. Each of these ecotones has a distinct species composition not only of plants but also of reptiles, birds and mammals. In a worldwide biodiversity comparison (Junk et al. 2006) of six globally important wetlands located in the tropics and sub-tropics, the Okavango Delta had a low number of ﬁsh species, but the second highest number of plants and mammals, the third highest number of amphibians and the highest number of reptiles and birds. In particular, the number of large mammal species and their high abundance are outstanding in the Okavango Delta as compared with the other large wetlands. Wetlands, however, generally do not have high biodiversity in comparison with, for instance, sub-tropical and tropical forests, and the species richness in the Okavango Delta is about the same as that in savannah systems in Southern Africa. A total of 46 plant habitat types have been identiﬁed in the Okavango Delta. Each speciﬁc habitat fragment, however, is fairly small (c. 0.05 km2) and thus is repeated many times over the Delta landscape (Ramberg et al. 2006a). The

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highest habitat diversity (areas with exceptionally high vegetation heterogeneity) occur along the perimeter of the wet Delta, along the Panhandle and along the major ﬂow channels to the east and west. It is very likely that the total species diversity is also highest in these areas. The implications for the management of biodiversity in the Delta are immense since these areas are the ones that will be impacted ﬁrst, probably becoming drier, if the inﬂow in the Okavango River is reduced by upstream developments or climate change. 21.2.7 The importance of the ﬂood pulse The uni-modal ﬂood pulse sensu Junk et al. (1989) creates and maintains the seasonal ﬂoodplains. This ecotone forms perhaps the most important habitat in the Delta as it sustains the high abundance of grazing mammals and is very important for ﬁsh productivity. The biomass of large mammals in African savannah systems is correlated to rainfall and nutrient status (Fritz et al. 2002). A model estimate, based on a large number of data, predicts a biomass in the Okavango Delta of around 1200 kg km−2, while it is actually almost ten times higher and in the range of the more nutrient-rich savannah systems in Africa (Table 21.3). The prolonged period of favourable grazing caused by the ﬂood pulse arriving several months after the rains is one reason for this. The other is the relatively high nutrient levels in the ﬂoodplains. Concentrations of phosphorus in water of sea-

Table 21.3 The number of species, density and biomass of large herbivores in a 6966 km2 study area in the Okavango Delta (modiﬁed from Bonyongo 2004) Habitat category Grassland dependent Not grassland dependent Total a

Number of species

Density (no. km−2)

Biomass (kg km−2)

8 4 12

29.09 2.04 31.11

6,678 5,210a 11,888

The high biomass is caused by elephants with 4083 kg km−2.

sonal ﬂoodplains are typically 7–50 times higher than in the permanent channels, while the concentrations in dry ﬂoodplain sediments are 6–10 times higher than in the dry woodland soils (Ramberg et al. in preparation). This enrichment is probably caused by nutrient accumulation and storage in the dry sediment/soil matrix from year to year and release when ﬂooded. In African ﬂood plain systems there is a direct positive relation between ﬂood size and ﬁsh production (Welcomme 2001) and this is the case also for the Okavango Delta (K. Mosepele, personal communication, 2008). In particular, the high ﬂooding years seems to provide very favourable conditions for ﬁsh reproduction (Lindholm et al. 2007). The ﬂood pulsed seasonal ﬂoodplains are thus likely to be decisive for the total biological productivity in the Delta both of terrestrial grazing mammals and their large predators such as lions, leopards, cheetahs, hyenas and wild dogs, and of the aquatic ﬁsh fauna with dependent ﬁsh-eaters such as birds, otters and crocodiles. 21.2.8 The importance of river channel abandonment The large-scale switching of river channels caused by sedimentation and the resulting drying up of large areas and ﬂooding of others described above, cause ﬁrstly a dramatic succession of vegetation where for instance woodlands are turned into grasslands, while permanent swamps are turned into woodlands. It also creates a regional nutrient pattern where the areas under ﬂooding accumulate nutrients and the drying up areas release them with very high biological productivity as a result (Ellery and McCarthy 1994).

21.3 Historical Use and Management of the System The population density in the whole Okavango River basin has always been low, caused mainly by the nutrient poor and low productive Kalahari sandy soils and the low, erratic rainfall with

Evolution of River Basin Management in the Okavango System, Southern Africa fairly frequent periods of drought that can last 4–5 years. Small-scale farming has been and still is the dominant source of livelihood in the basin but the value of farming varies from the north to south following the gradient in rainfall (Mendelsohn and El Obeid 2004). Normally each household cultivates a few hectares and keeps small herds of sheep and goats. In the fringes of the Okavango Delta ﬂood recession agriculture is practised and ﬁshing provides an important supplement. Expansion of farmers and livestock into the Delta has been restricted by the presence of tsetse ﬂies. Under the traditional way of utilizing the Okavango Delta some management and water control activities such as reed burning and bunding for reclaiming of agricultural land, building of dams and water diversions for farming, clearing of vegetation blockages in channels for accessibility, were carried out (Wilson 1973). These actions had only local effects and had no inﬂuence on the system as a whole (Potten 1976). Importantly, people were moving in response to variable ﬂooding conditions. The capital of the Batawana, for instance, the largest group living in the Delta area, was moved three times during the 1800s (Wilson 1973). The early (1850–1920) European travellers in the region regarded the Okavango Swamp as a waste of water and suggested often unrealistic schemes for its ‘better’ utilization such as irrigation of the Kalahari with redirected ﬂows of the Zambezi and Okavango (Schwarz 1920). Such schemes did not, however, pass the economic scrutiny of hardnosed colonial governments. From about 1920, a number of smaller-scale water management projects were carried out in the Okavango Delta. Attempts were made to remove the extensive papyrus blockages and restore the water ﬂow in the failing Thaoge channel in the western part of the Delta. In the eastern distributaries, channels were cleared of vegetation and bunds constructed in order to reduce evapotranspiration and increase outﬂow (Wilson 1973). These projects had only short or no desired effects.

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With the arrival of independence in Botswana and with new enthusiastic Government ofﬁcers, and likewise sympathetic foreign aid organizations, a number of water development schemes were implemented in the Delta such as a largescale rice cultivation project (funded by China), a modernization of the traditional ﬂood recession agriculture (funded by Germany) and a construction of a bypass canal to substitute for the failed Thaoge channel (assisted by the Netherlands). All of these projects, in spite of considerable ﬁnancial input (several million US$ each), failed due to either a lack of knowledge about the nature of the Delta system or ignorance of the local political and socio-economic situation. Before 1975, the colonial authorities in Angola developed plans for utilization of Okavango water to the stage of feasibility studies. Due to the relatively steep gradient of the headwater rivers, 17 potential locations for hydroelectric dams were identiﬁed. Also, irrigation schemes of 54,000 ha, mostly along lower reaches of the Cubango and Cuito, were considered feasible (SWECO GRONER 2005). The independence in 1975 and subsequent civil war in Angola, however, prevented realization of these plans. In Namibia, the majority of colonial development (mainly large-scale cattle farming) took place in the central and southern parts of the country while the Okavango region was left to the indigenous population, so no large-scale plans were made there. In the period 1970–1980 both Botswana and Namibia made prognoses of population growth and economic development that indicated increased water shortage by the 1990s and so turned their attention towards the Okavango. In Botswana, projects were initiated with ﬁnancial support from international aid organizations to investigate feasibility and means of water abstraction from the Delta and its transfer towards the more populated eastern part of the country. In Namibia, a Water Development Plan was formulated in 1974, with plans to include the Okavango into the water supply system of the country, and building of a hydroelectric station

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at Popa Falls, just upstream from the border with Botswana. In Botswana, the situation was additionally complicated by the discovery of diamonds in the Orapa area located in the distal Boteti basin. In search of water to supply the constructed mine and mining town, the engineers turned to the Boteti and the Okavango Delta. A dam was constructed at the end of the Boteti to store water for the mine. In 1973, the Boro channel was dredged and bunded at a length of 17 km in order to reduce the loss of water into ﬂoodplains and increase ﬂows down the Boteti River. This reduced the seasonally ﬂooded area in the lower Boro, which was exacerbated by the general decline in rainfall and inﬂow during that period. This negatively affected the local ﬂood recession farming and lead ultimately to a ‘popular uprising’ and breaching of the bunds in 1981. In the mid 1980s, Botswana was hit by a major drought. Associated reduction of water supply to the Orapa diamond mine triggered the initiation of the Southern Okavango Integrated Water Development Project (SOIWDP) (Snowy Mountains Engineering Corporation 1990). The aims of the project were to supply water to the Orapa mine, to Maun and to the rural population along the Boteti. The proposed engineering works aimed to reduce channel overﬂow and ﬂooding along a 50-km stretch of lower Boro River and construction of several reservoirs along the Thamalakane and Boteti Rivers. In 1992 the project was approved. At this stage, however, the local community remembered the effects of the 1973 dredging and started to oppose the project. The tourism industry, already well established in the Delta, expressed its disagreement as well, and eventually international conservation groups started to lobby the Government. The Government decided to put the project on hold, and commissioned the International Union for Conservation of Nature (IUCN) to do an independent evaluation. The major conclusions of the assessment were that the beneﬁts were overestimated and too uncertain while the socioeconomic costs were underestimated (Scudder

et al. 1993). When the Government received the report in 1993 it called off the project immediately. As mentioned earlier, Namibia had considered plans to augment its interior water supply with Okavango water since the 1970s. The 1974 National Water Master Plan included a phased construction of a system of reservoirs and water carriers in the interior and ﬁnally linking it to the Okavango River. Four out of ﬁve phases were implemented by 1987. The last phase was to construct a pipeline to bring water from Rundu by the Okavango River to Grootfontein and into the already constructed system of canals. This part of the project was delayed, but after a several years’ long drought a crisis situation had evolved, with very little water left in the reservoirs on the interior Namibia highlands, and a forced implementation of the project started in 1996. This was done unilaterally (without consulting Botswana and Angola) with reference to the emergency situation and caused strong opposition in Botswana. Fortunately, the drought ended with better rains in beginning of 1997, and the construction of the pipeline was postponed (Ramberg 1997). These events coincided with, and probably contributed to, the Botswana Government unilaterally ratifying the Ramsar Convention in 1996 and declaring the Okavango Delta a Ramsar site: a wetland of international importance. This was the ultimate recognition that the Okavango Delta should be managed as a conservation area, to maintain the natural wetland system and its biodiversity, and to use its water and natural resources sustainably.

21.4 Current Situation in the Basin Estimates of the population living in the Okavango River catchment and within 20 km of surface water in the Delta and Boteti are: 350,000 in Angola, 163,000 in Namibia and 88,000 in Botswana (Mendelsohn and El Obeid 2004). The population densities are low, and the water resources of the Okavango Basin are relatively

Evolution of River Basin Management in the Okavango System, Southern Africa Table 21.4 Estimates of current water abstractions within the Okavango basin Water users Botswana

Domestic water supply

Namibia

Domestic and irrigation – state water scheme Irrigation – public and private water schemes Domestic

Angola

Water use (Mm3 yr−1) 3.84 (recorded in 2004, Department of Environmental Affairs 2008) 3.0 (recorded in 2007, NamWater personal communication) 13.2 (allocated, actual value unknown, NamWater personal communication) 13.8 (estimated, Ashton and Neal 2003)

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activity after the mineral industry. Tourism is also of some importance in Namibia’s part of the basin. In the Delta area, tourism generates more income than all other economic activities together such as agriculture, ﬁshing, etc. (Murray 2005) and spin-offs from tourism such as construction, transport, supplies, training, etc. form an engine of development in the urban centre of Maun. The tourism in the Okavango Delta relies on exclusivity (‘low numbers – high costs’) and is dependent on the large numbers of charismatic wildlife and the pristine nature of the environment. 21.4.2 Economic development potential

little utilized. The total water abstractions for human use do not exceed 33 Mm3 yr−1, which is less than 1% of mean annual discharge of the Okavango River (Table 21.4). The water supply schemes are mostly for domestic use and only in Namibia are there a few irrigation schemes. This low utilization of the water resource takes place in spite of the Okavango River being the only permanent water body within the borders of Namibia and Botswana and in spite of their growing populations. The low water abstraction in Angola is not surprising, however, considering that this country ﬁrst went through a 25-year liberation war against the colonial rulers that ended with independence in 1975, but then almost immediately turned into a civil war that did not end until 2002. The southern part of Angola, within the Okavango River basin, was at times a war zone during this period and became de-populated. 21.4.1

Sources of livelihood

The majority (77%) of the population in the basin is rural and their livelihood relies on small-scale agriculture and cattle, supplemented by ﬁshing and utilization of river-related and dry land veld products (Mendelsohn and El Obeid 2004). The Delta is the major tourism attraction in Botswana and tourism is the second largest GDP generating

The potential for economic development in the basin is limited mainly by its remoteness, lack of mineral resources, relatively poor soils and harsh climate. Some potential for irrigated agriculture exists, however, in parts of Angola and Namibia and is planned to be explored by both governments. There also exists potential for hydropower generation along the Okavango River in Angola and Namibia, and the feasibility of hydropower plants has been investigated there (Mendelsohn and El Obeid 2004). This might be an important direction of development in the basin considering rising energy prices and climate change concerns related to the use of fossil fuels. Ecotourism is a realistic development option, and plans are being considered for wildlife parks in Angola. Within the Botswana part of the basin, development is envisaged to go towards further expansion of wildlife-based tourism. 21.4.3 Emergence of water resource management initiatives in the basin Some of the developments described above have the potential to cause reduction of water ﬂows and change to the hydroperiod; changes in sediment transport; and to cause pollution of the river with nutrients and pesticides. This is further complicated by the upstream–downstream principal differences in the way the river water is perceived to be used. The two upstream countries

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are predominately planning for water withdrawals for domestic use and irrigation, while Botswana, as an end user and with its unique economic dependence on wetland tourism, emphasizes its need for water conservation. The necessity for collaboration and information exchange on water development between the Okavango basin states was realized in the early 1990s. The Permanent Okavango River Basin Water Commission (OKACOM) was established in 1994 with a protocol signed by the three countries which replaced earlier bilateral technical commissions. The OKACOM is essentially a communication platform with no decisionmaking power; however, it effectively acts as a management body, as it is mandated to advise and present recommendations to respective governments. The OKACOM recognized the need for a uniﬁed management plan for the basin, and started working towards development of such in the mid 1990s. By 1995 a proposal for a project tasked with development of such a plan was prepared and submitted to UNDP for funding (Okavango Basin Water Commission 1995). This was approved and a preparatory project was completed during 1997–1999 resulting in a full planning proposal. Due to lack of consensus between delegations and with UNDP-GEF, the project was not approved until 2001. For the same reasons, it did not start until 2005 as ‘Environmental Protection and Sustainable Management of the Okavango River Basin (EPSMO)’. It was, unfortunately, hit by serious problems with red tape and poor governance. The project manager resigned and the project was on hold for more than a year and started again only in 2007. In the meantime, in Botswana, the ﬁrst discussions on the development of a management plan for the Okavango Delta took place in 1998. The need to formulate such a plan stems directly from the recognition of the Okavango Delta as a Ramsar site, and the fulﬁlment of Article 3.1 of the 1971 Ramsar Convention which states that ‘The Contracting Parties shall formulate and implement their planning so as to promote the conservation of the wetlands included in the List,

and as far as possible the wise use of wetlands in their territory’. The ﬁrst conceptual paper outlining the need and content of such a plan was produced in 1999 by the then National Conservation Strategy Agency (later the Department of Environmental Affairs) Government of Botswana, IUCN and HOORC (see section 21.5.2 for deﬁnition of HOORC). This proposal was funded by the Ramsar Bureau and in the course of subsequent years, the Okavango Delta Management Plan Project (ODMP) got funding in two more steps for more detailed planning. The third comprehensive phase started in 2003, ended in April 2007 and produced a management plan that remains to be implemented. So although it would have been ideal to have the basin-wide plan in place and then align the Delta plan to that, it happened the other way around (Box 21.1).

Box 21.1 The progress of OKACOM and ODMP Permanent Okavango River Basin Commission (OKACOM) 1994 – Three nations (Angola, Namibia, Botswana) agreement signed 1997 – Preparatory planning of management proposal started and ﬁrst funding received 2001 – Funding for substantial planning project received 2005 – Planning project (EPSMO) started 2007 – Planning project re-started Okavango Delta Management Plan (ODMP) 1996 – Ramsar Convention signed by Botswana 1997 – Okavango Delta designated as a Wetland of International Importance 1999 – Preparatory planning of management plan proposal started and ﬁrst funding received 2003 – 2007 Planning project (ODMP) 2008 – Implementation started

Evolution of River Basin Management in the Okavango System, Southern Africa 21.5 The Okavango Delta Management Plan 21.5.1

Guiding principles for the ODMP

The main activities of the ODMP project (2003– 2007) aimed ‘to develop a comprehensive, integrated management plan for the conservation and sustainable use of the Okavango Delta and surrounding areas’ (Department of Environmental Affairs 2008). It is not a water management plan, although it encompasses elements of such. The approach to development of the plan was borrowed from Ramsar Planning Guidelines (Ramsar Convention Secretariat 2007) and IUCN’s Ecosystem Approach to wetland management (Shepherd 2004). It is integrative and adaptive in nature and based on participation of stakeholders in the plan development and implementation process. It has to be realized that the integrated approach to management is not strongly supported within the regulatory framework of Botswana. The 1968 Water Act regulating water resources planning and management is based on a centralized approach with minimal participation of water users in decision-making. Only more recent regulations, such as the National Wetlands Policy and Strategy (not approved yet), and the National Water Master Plan (1992, currently being revised), advocate integrative approaches to resource management. Additionally, the requirement to achieve integrated management of water and environmental resources was a formidable challenge to the existing fragmented planning in the Delta, which had been sub-divided in about 20 Wildlife Management Areas (WMA) and one Game Reserve each with its own management plan. In addition there were sectoral plans by largely unco-ordinated government departments with a mandate to manage different resources within the Delta. Lack of communication often led to confusion, duplication and omission of roles. This led to avoidable mistakes such as construction of roads on seemingly dry land that later, during high ﬂood years, came under water and thus became useless. It is also believed that this fragmentation of manage-

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ment made matters worse when it came to natural resources use conﬂicts in the Delta (Magole 2008) and these conﬂicts are still rampant. The main friction is between the traditional resource users and the modern and expanding tourism industry which, because it is based on nature and wildlife, also ﬁnds natural allies in conservation groups. Local people in general complain about large land allocations for tourism or other commercial purposes that cut them out from their subsistence sources such as ﬁrewood, reeds, grass and ﬁshing. Farmers in particular complain about increased crop damage and predation by wild animals. Elephants are the worst problem as they have increased dramatically during the last decades and are protected. Sectoral planning and management is well illustrated by the actions of Department of Animal Health. To separate livestock from wildlife that carries diseases such as foot-and-mouth and to stop its transmission to cattle, a fence was constructed in the 1980s to surround the Delta on the south, east and north. That this also cuts off the Kalahari wildlife populations from their natural fall-back areas in the Delta was not considered. Similarly in 2001–2002, spraying against tsetse ﬂies was undertaken in the Okavango Delta, with little or no involvement of the Departments of Tourism, Environment and Wildlife or other stakeholders. Similarly, the increasing human populations around the Delta, urbanization and demands for higher living standards have caused stress on the water supply systems and in particular Maun has had long periods of water shortage. A number of additional groundwater sources have been identiﬁed by Department of Water Affairs that, in typical fashion, narrowed down the alternatives to one based on technical and economic grounds before consultations with other departments and stakeholders were done. 21.5.2

Organization and work of ODMP

Initially, two possible organizations for the project were discussed. The ﬁrst was to establish

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a strong planning team with special mandates outside of the existing management structures to ensure integration and efﬁciency in the planning phase. The drawback, however, would come in the implementation phase as it would require major changes in legislations to transform the planning team into an implementing body. This was regarded as unrealistic and it was concluded that it would not be possible to separate the existing government departments from the tasks they had been given by law and they would be unlikely to have much sympathy for what had been planned without their involvement. An alternative approach was followed whereby new structures and functions were put in place that linked the existing government departments together by frequent meetings and improved communications, but basically the responsibility for development and implementation of the integrated ODMP was vested in them. ODMP was initially divided into 12 components each representing a prominent resource use sector; such as water, ﬁsh, land, plants and range, wild animals, livestock, tourism, and the less conventional planning sectors of stakeholder participation, research strategy and data management. In most planning processes the latter are considered cross-cutting issues and hence inherent and to be dealt with within the other sectors/ components. In the ODMP process, however, these together formed a substantive sector which was complete with a budget and allocated to an equally unconventional but appropriate planning partner, University of Botswana’s Harry Oppenheimer Okavango Research Centre (HOORC). The sectors were tasked with situation analysis. They had to answer questions such as: what is the state of the resource? Who is using the resource? Who is managing the resource? Who else has interest in the resource? Are there any conﬂicts in the use of the resource? To achieve integration, sectors were organized into task forces with members from all the professional areas involved in the use and management of the resources within the Delta. Thus, multidisciplinary teams were formed in

accordance with the requirements of the integrated water resources management (IWRM) (Al Radif 1999). These allowed new opportunities for the sectors to communicate their views and plans, learn from other sectors and to negotiate their stake in other sectors, e.g. the tourism sector was able to negotiate its land and other resources requirements from the land board and wildlife sectors. Communities were also brought in through community leaders and resource use groups representatives as well as community focal persons who became members of task forces and acted as the intermediary between the ODMP planning team and the communities. A large number of traditional community meetings (Kgotlas) were organized for people to contribute their perspective on the situation of the resources, their suggested solutions and to appraise solutions suggested by the technical planning team. The planning process was led and co-ordinated by the Department of Environmental Affairs (DEA) under the Ministry of Wildlife, Environment and Tourism. The co-ordinators (DEA) reported to a National Steering Committee of Permanent Secretaries and Directors of involved departments. At the district level in Maun the co-ordinators and the planning sectors were supposed to report to the Okavango Delta Wetland Management Committee made up of all district level stakeholders and task forces of experts, but this was not implemented. All this led nevertheless to a large number of meetings and acrosssector interactions that gradually broke down barriers between all these ofﬁcers from different sectors. Last but not least, the co-ordinators and the sectors had to report every milestone achievement to the communities in village Kgotlas. This gave the ODMP an unprecedented stakeholder involvement. 21.5.3 Problems faced by ODMP The ODMP project became seriously delayed right from the outset (Table 21.5). It took a long time to mobilize project ofﬁcers and to organize the project secretariat, partly related to the com-

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Table 21.5 The planned and actual time frame for the ODMP project phases Milestones

Proposed duration (months)

Actual duration (months)

7 6 12 14 12 12–24 39

20 12 6 6 Left for implementation phase Left for implementation phase 44

1 Inception report 2 Framework management plan 3 Draft management plan 4 Final management plan 5 Studies and researcha 6 Pilot projectsa Total a

Activities 5 and 6 were planned to be done in parallel with activities 3 and 4.

plexity and bureaucratic nature of the procurement process in the public sector. This led in the end to a forced ﬁnalization of the project and important elements were left out from the ﬁnal product: an integrated management plan for the Okavango Delta. Most government ofﬁcers seconded to Maun were at a basic or intermediate level of competence. Many in charge of project components had a limited capacity to either deliver, or ensure delivery of project output of adequate quality. Additionally, as most government ofﬁcers only stayed in Maun for a limited period, by the end of the project in 2006 about 80% of the government ofﬁcers who took part in the start of the project in 2003 had been transferred. Again, basic training and introduction had to be given to the new replacements on the principles of IWRM and the particular Delta environment. Institutional memory could never be developed. For HOORC, the participation in the ODMP project became a mixed blessing. Because of the deep involvement of its staff in all task forces it became automatically disqualiﬁed in tendering for research projects when these were advertised. Instead, external consultants were used, often from outside Botswana, and some of them relied heavily on HOORC resources and know-how. The opportunity for local and national capacity building within the project was thus not effectively used. The integrated nature of the ODMP was supported by the prevailing culture of cross-sectoral

integration at District level. However, the governance in Botswana is centralized and most important decisions are taken on higher levels in Gaborone, 1000 km from Maun and the Okavango Delta. Sometimes the Secretariat in Maun had problems with the collaborating partners that could be traced back to lack of buy-in for the project at a senior level. This is where the ODMP started to suffer the consequences of operating the integrated approach within a sectoral institutional setting. The ODMP Secretariat was not in control of the collaborating ofﬁcers’ time and budget and the participating departments had the liberty to make spending and priority decisions ‘almost’ independent of the Secretariat. A case in point was the water component; the Department of Water Affairs chose to do its work for ODMP in Gaborone and used most of its budget for development of a hydrological model at the expense of the integrative and stakeholder driven process promoted by the ODMP. Negotiations for a better approach proved futile because the ODMP Secretariat could not force the preferred management approach on a department which had the legal mandate to make decisions on water management. The lack of buy-in for the project at top level in ministries and their departments also became clear in their participation in the National Steering Committee, which was the forum for heads of departments, usually directors, who rarely participated but sent fairly low-level representatives. However, the same was also the

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case with representatives for major donors: IUCN, Swedish International Development Agency (SIDA), Danish International Development Agency (DANIDA) and German Development Service (DED). The overall governing committee for the project became, therefore, fairly insigniﬁcant.

21.6 Management Framework at the International Level – OKACOM The Permanent Okavango River Basin Water Commission (OKACOM) was established in 1994 essentially as an advisory body with a mandate to deal with aspects of water management in the basin such as measures and arrangements to determine the long-term safe yield of water in the basin, the reasonable demand for water, the criteria to be adopted in the conservation of the natural environment, equitable allocation and sustainable utilization of water resources, the investigations related to the development of any water resources, prevention of pollution of water resources and control of aquatic weeds, and measures to alleviate shortterm water shortages (Okavango Basin Water Commission 1994). Co-operation of the three basin states over the shared basin is also regulated by the Southern Africa Development Community (SADC) Protocol on Shared Waters (South African Development Corporation 2001), to which all three countries, as members of SADC, subscribe. The protocol is based on the principles in the UN Convention on the Law of the NonNavigational Uses of International Watercourses. According to the protocol, countries sharing a river basin are expected to harmonize their planning, development and use of water resources, and avoid situations leading to conﬂicts, maintain an appropriate balance between water resources development and use, and environmental conservation, utilize water in an equitable and reasonable manner and prevent causing signiﬁcant harm to other riparian states, and mitigating any harm caused.

The Ramsar Convention, ratiﬁed by Namibia and Botswana, is another international agreement that is guiding interactions of the three countries over the Okavango basin issues. The Okavango Delta, a declared Ramsar site, is not within the borders of Namibia, however, that country as a signatory to the Convention agrees to participate in the conservation and wise use of the wetland. It does not directly affect management of water resources upstream, particularly in Angola which is not a signatory to the Convention. The framework of international agreements guiding the process of basin management is thus almost as good as it gets. It is, however, important to realize that these agreements are lacking compulsory jurisdiction and enforcement, and rely instead on the opinion of the wider world community and the will to maintain good relationships with the neighbouring countries. This seems, however, not to be a strong deterrent as illustrated by the example of Namibia in 1996 when they decided unilaterally to build a pipeline from the Okavango River to supply water to Windhoek. With reference to the ‘emergency situation’, provided for in the Helsinki Rules (International Law Association 1966), they did not have to seek permission from the other riparian states. Luckily good rains fell just before the construction project was about to start and it was shelved without the need to further test the international water management framework (Ramberg 1997). 21.6.1 Organization of OKACOM The OKACOM is made up of a Commission, a Steering Committee and a Secretariat. The Commission is composed of three representatives from each country; senior civil servants usually with expertise in water management, law and environmental issues. The Okavango River Basin Steering Committee (OBSC) comprises technical experts from usually the ministries responsible for water and environment, and other invited experts. The OBSC provides technical advice to the commission and has formed task

Evolution of River Basin Management in the Okavango System, Southern Africa forces dealing with hydrology, biodiversity and institutional issues. A Secretariat (sponsored by USAID and SIDA) based in Maun, Botswana was established in 2008 to be the administrative arm of OKACOM and facilitate information exchange among the basin states. The Commission is consulting with a number of stakeholders in the basin such as the independent Basin Wide Forum. This Forum was originally organized by the SIDA sponsored project ‘Every River has its People’ and executed by the national conservation organizations in the three countries. It comprises ten representatives of local communities from each country and provides an opportunity for local communities to share experiences about uses of the basin and problems encountered. 21.6.2

The work of OKACOM

In contrast to the Okavango Delta, where there are a number of acute conﬂicts, the problems in the Okavango River basin are mainly potential and perceived. This is a lucky situation because OKACOM has moved very slowly. It has functioned in an erratic way and has had problems in starting even a fairly straightforward planning project similar to ODMP. The slow progress was caused by a cluster of factors that seem to be typical for international resource sharing projects. The three countries have all been colonies or protectorates under different foreign powers that have given them different administrative cultures, with Angola from Portugal, Namibia from Germany and South Africa, while Botswana got it from Britain. They also have different ofﬁcial languages and in particular the Angolan delegation needs to rely on translators. During the 1990s, the deliberations at the meetings were remarkably antagonistic and the delegations did not show much trust in each other. In addition, the communications between the three delegations functioned in a haphazard way, and venue and time for meetings were frequently changed often at the last minute. There is now a remarkable improvement in the relations between the delegations, probably because there have been very few changes of del-

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egates and eventually good personal relations have evolved, but the delegates have also been trained in conﬂict resolution through an initiative by USAID. This organization also funded the Okavango Integrated River Basin Management Project (IRBM), a 4-year (2004–2008) initiative that provided support to the Commission’s institutional development. It improved the ability of OKACOM to function as a multinational planning and consensus building institution to effectively manage and co-ordinate the use of river basin resources; targeted development of a formalized information system; and supported local governance of natural resources. The OKACOM Secretariat that has been established in Maun is now evolving into the much needed communication hub. The OKACOM is becoming a clearinghouse for research- and management-oriented projects in the river basin and has established working relations with ODMP and has HOORC staff in two of its technical task forces. Importantly, OKACOM, initially concerned primarily with allocation of water, has embraced the integrated, participatory approach to water resource management, and started looking at the basin as a whole. OKACOM’s own planning project, EPSMO, that restarted in 2007, is now, among other things, collaborating with HOORC to determine the environmental ﬂow requirements for the Okavango basin.

21.7 Conclusions The contradiction between development policy and conservation policy is not grasped when it comes to the Okavango Delta and is a major challenge. The strong trend towards a sedentary lifestyle with ambitions to develop permanent infrastructures such as roads, electrical power supplies and water channels for communication and irrigation, are in direct conﬂict with the commitment to the Ramsar convention where the requirement is to maintain the nature of the wetland, which in the case of the Okavango Delta means that its inherent unstability has to be maintained, with the consequence that people

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472 Table 21.6

Overview of management structures for ODMP during the planning and implementation phases Planning phase 2003–2007

National Steering Committee Facilitator in Gaborone District Wetland Committee Secretariat under DEA with special mandate Chief Technical Advisor (CTA) 11 sectoral government departments Technical task forces to 11 government departments HOORC

Regular meetings Responsible for high level integration in Gaborone but transferred to Maun when CTA left Not active Mainly responsible for production of the plan. Played a key role for integration between departments, in communications and liaisons Played a key role but resigned midway. Active

Implementation phase 2008– Abolished Abolished Activated but weak due to lack of legislation Abolished

Played key roles for practical integration

Abolished Responsible for implementation. Department of Environmental Affairs now on par with the other departments Abolished

Played key roles in task forces, for data base management, library services, community mobilization, research planning

Play key roles in database management, library services, research and environmental monitoring

should move with the water, not the other way around. Integrated water resources management in the river basin cannot only focus on the amount of water (to share) but, moreover, the aim must be to preserve the ﬂood pulse (the hydroperiod) that is decisive for the ecological functions of the riparian ﬂoodplains all along the river and for the whole Okavango Delta. Some of the problems that came to mar the ODMP project were not fully anticipated in the beginning (the difﬁculty to get buy-in from high levels of government departments, weak support from donors), while others were well known (slow recruitment procedures, transfer of staff, weak facilitator in Gaborone, exclusion of HOORC from research projects) and could have been dealt with in the negotiating phase before the project started. At that time, however, the eagerness to start took precedence over a more cautious approach. It was, nevertheless, a major achievement to produce the management plan that should now be implemented. Most of the management structures established during the planning phase, which were meant to achieve integration, have now been abolished and the

situation is almost as it was before the whole exercise started in 2003 (Table 21.6). It is an improvement that DEA has an ofﬁce in Maun and HOORC has been strengthened, but most implementation work has to be done by the government departments that are now slowly rolling back into their sectoral positions. For this type of big project that requires integration across many sectors of government and society at large, an overruling authority is probably necessary. In the case of Botswana that function could be based either in the Ministry of Finance or in the Ofﬁce of the President. The development of OKACOM has been a slow process but that seems to be the rule when it comes to international resource sharing organizations, such as for the River Rhine and the Baltic Sea. Apart from local speciﬁc difﬁculties the development of trust seems to be a common problem. For the Okavango basin the best way forward may be to develop as much shared activity as possible, for instance in cultural exchange, training and research. Here the lack of professional capacity is probably the bottleneck that most hampers progress in integrated water resources management.

Evolution of River Basin Management in the Okavango System, Southern Africa Financial and logistic support from foreign aid organizations has been crucial for the development of both projects. It has been most effective when the supporting organization has had determination and staying power because integrated water management, although logical and desirable, is in practice riddled with problems. In spite of difﬁculties with communications, institutional and cultural dividers, as well as shortcomings in manpower and other resources, both projects are on positive trajectories and their achievements are considerable (Box 21.2). Both are also working with an adaptive management philosophy so corrections of, for instance, institutional arrangements can be made as they progress. It is perhaps the learning process in both ODMP and OKACOM that is the most valuable outcome, and in this chapter we have tried to describe not only the successes but also the challenges and difﬁculties both projects have faced,

Box 21.2 Overview of ODMP and OKACOM achievements ODMP • Management plan in place • Much larger awareness of IWRM • Department of Environmental Affairs ofﬁce in Maun • Better communications between government departments • HOORC role in data base management, training, research and monitoring recognized and strengthened OKACOM • Development of management plan ongoing • Building of basin-wide awareness under way • Secretariat in place in Maun • Trust and communications between national delegations much improved • Links established with ODMP and HOORC

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because these are pioneer projects from which others can learn.

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Botswana. Physics and Chemistry of the Earth, 33, 906–912. Mazvimavi, D. and Wolski, P. (2006) Long-term variations of annual ﬂows of the Okavango and Zambezi Rivers. Physics and Chemistry of the Earth, 31, 944–951. McCarthy, T. (1993) The great inland deltas of Africa. Journal of African Earth Sciences, 17, 275–291. McCarthy, T.S., Barry, M., Bloem, A. et al. (1997) The gradient of the Okavango fan, Botswana, and its sedimentological and tectonic implications. Journal of African Earth Sciences, 24, 65–78. McCarthy, J., Gumbricht, T., McCarthy, T.S., Frost, P.E., Wessels, K. and Seidel, F. (2004) Flooding patterns in the Okavango wetland in Botswana, between 1972 and 2000. Ambio, 7, 453–457. Mendelsohn, J. and El Obeid, S. (2004) The Okavango River, The Flow of a Lifeline. Struik Publishers, Cape Town. Murray, M.I. (2005) Relative Proﬁtability and Scale of Natural Resource-Based Livelihoods in the Okavango Delta. Report prepared for the PDF-B stage of the GEF project ‘Building local capacity or conservation and sustainable use of biodiversity in the Okavango Delta’. GEF/BIOKAVANGO, Maun, Botswana. Okavango Basin Water Commission (OKACOM) (1994) The Agreement between the Governments of Republic of Angola, the Republic of Botswana and the Republic of Namibia on the Establishment of the Permanent Okavango Basin Water Commission. OKACOM, Windhoek, Namibia. Okavango Basin Water Commission (OKACOM) (1995) Project Proposal for an Environmental Assessment of the Okavango River Basin and the Development of an Integrated Water Resource Management Strategy. OKACOM, Windhoek, Namibia. Potten, D.H. (1976) Aspects of the recent history of Ngamiland. Botswana Notes and Records, 8, 63–85. Ramberg, L. (1997) A pipeline from the Okavango River? Ambio, 26, 129. Ramberg, L. and Wolski, P. (2008) Growing islands and sinking solutes: processes maintaining the endorheic Okavango Delta as a freshwater system. Plant Ecology, 196, 215–231. Ramberg, L., Hancock, P., Lindholm, M. et al. (2006a) Species diversity of the Okavango Delta, Botswana. Aquatic Sciences, 68, 310–337. Ramberg, L., Wolski, P. and Krah, M. (2006b) Water balance and inﬁltration in a seasonal ﬂoodplain in the Okavango Delta, Botswana. Wetlands, 26, 677– 690.

Ramsar Convention Secretariat (2007) Managing Wetlands: frameworks for managing wetlands of international importance and other wetland sites. Ramsar Handbooks for the Wise Use of Wetlands, 3rd edition, vol. 16. Ramsar Convention Secretariat, Gland, Switzerland. Schwarz, E.H.L. (1920) The Kalahari and its possibilities. Journal of the Royal African Society, 20, 1– 12. Scudder, T., Manley, R.E., Coley, R.W et al. (1993) The IUCN review of the Southern Okavango Integrated Water Development Project. International Union for Conservation of Nature and Natural Resources (IUCN), Gland, Switzerland. Shepherd, G. (2004) The Ecosystem Approach: ﬁve steps to implementation. Ecosystem Management Series No. 3. International Union for Conservation of Nature and Natural Resources (IUCN), Gland, Switzerland. Snowy Mountains Engineering Corporation (SNEC) (1990) Southern Okavango Integrated Water Development Project. Snowy Mountains Engineering Corporation, Department of Water Affairs, Gaborone, Botswana. South African Devleopment Community (SADC) (2001) Revised Protocol on the Shared Watercourse Systems in the Southern African Development Community (SADC) Region. SADC, Windhoek, Namibia. SWECO GRONER (2005) A Rapid Water Resources and Water Use Assessment for Angola. Final Report. Natural Water Sector Management Project Activity C. Ministry of Energy and Water Affairs, Angola. Thomas, D.S.G. and Shaw, P. (1991) The Kalahari Environment. Cambridge University Press, Cambridge. Turton, A.R., Ashton, P.J. and Cloete, T.E. (2003) An introduction to the hydropolitical drivers in the Okavango River basin. In: Turton, A.R., Ashton, P.J. and Cloete, T.E. (eds), Transboundary Rivers, Sovereignty and Development: hydropolitical drivers in the Okavango River Basin. Green Cross International (GCI), Geneva, Switzerland/African Water Issues Research Unit (AWIRU), Pretoria, South Africa, pp. 9–30. Tyson, P.D., Cooper, G.R.J. and McCarthy, T.S. (2002) Millennial to multi-decadal variability in the climate of southern Africa. International Journal of Climatology, 22, 1105–1117. Welcomme, R.L., (2001) Inland Fisheries; Ecology and Management. Blackwell Science, Oxford.

Evolution of River Basin Management in the Okavango System, Southern Africa Wilson, B.H. (1973) Some natural and man-made changes in the channels of the Okavango Delta. Botswana Notes and Records, 5, 132–153. Wolski, P. and Murray-Hudson, M. (2006) An assessment of recent changes in ﬂooding in the Xudum distributary of the Okavango Delta and Lake Ngami,

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Botswana. South African Journal of Science, 102, 173–176. Wolski, P., Savenije, H., Murray-Hudson, M. and Gumbricht, T. (2006) Modelling the hydrology of the Okavango Delta, Botswana using a hybrid GISreservoir model. Journal of Hydrology, 331, 58–72.

Image facing chapter title page: Courtesy of Samuel Wolski.

22 Basin Management Approaches used in a High-latitude Northern Catchment – the Mackenzie River Basin F R E D E R I C K J . W R O N A 1, J O S E P H M . C U L P 2 A N D TERRY D. PROWSE1 1

Water and Climate Impacts Research Centre, Environment Canada, University of Victoria, Victoria, British Columbia, Canada 2 Canadian Rivers Institute, Environment Canada, University of New Brunswick, Fredericton, New Brunswick, Canada

22.1

Introduction

The management of water resources in the Mackenzie River Basin is extremely complicated as regulatory responsibility for this vast catchment falls within the jurisdiction of federal, provincial and territorial governments, and requires agreements with First Nations. Covering an area more than seven times the size of the UK (Fig. 22.1), the twelfth largest drainage basin on the planet travels 15 ° of latitude from central Alberta to the Arctic Ocean (Culp et al. 2005). It is the largest river basin in Canada extending over oneﬁfth of the country, is the fourth largest river discharging to the Arctic Ocean and the largest north-ﬂowing river on the continent. The Mackenzie Basin consists of six sub-basins. Major rivers of the Basin include the Athabasca, Peace, Slave, Liard, Arctic Red, Peel and mainstem Mackenzie (Fig. 22.2). Basin discharge in the Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

south originates mainly in the Rocky Mountains of Alberta and British Columbia, and to the north in the Mackenzie Mountains of the Northwest Territories. The Mackenzie Delta is the second largest Arctic delta with an extensive area of levees formed from downstream transport of sediments on their course to the Beaufort Sea. Two freshwater deltas (Peace–Athabasca, Slave) are contained in the basin along with three large lakes (Lake Athabasca, Great Slave Lake, Great Bear Lake). Glacial history, geology and harsh climate greatly affect environmental conditions in the Mackenzie Basin. The Laurentide ice sheet covered approximately 80% of the basin and produced the glacial meltwaters of ancient Lake McConnell, the perimeter of which engulfed all of contemporary Lake Athabasca, Great Slave Lake and Great Bear Lake (Brunskill 1986). This glacial history is believed to have greatly inﬂuenced the low biodiversity of present day aquatic ﬂora and fauna (McPhail and Lindsey 1970; Rosenberg and Barton 1986). Western and central portions of the basin are underlain by

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Fig. 22.1 Mackenzie River Basin overlying Europe. (From Mackenzie River Basin Board 2003.)

Fig. 22.2 Overview of the Mackenzie Basin – major sub-basins of the Mackenzie River Basin River. (From Mackenzie River Basin Board 2003.)

Basin Management Approaches used in a High-latitude Northern Catchment sedimentary bedrock that produces alkaline (high pH) rivers with moderate levels of cations and nutrients (Culp et al. 2005). This region receives considerable precipitation and produces all of the major tributaries to the Mackenzie River mainstem. Lower amounts of runoff ﬂow from rivers originating on the granitic Precambrian Canadian Shield east of the Mackenzie mainstem. Permafrost underlays 75% of the basin but glaciers are restricted to the Cordillera region west of the mainstem (Mackenzie River Basin Board 2003). Peak discharge in most rivers occurs in spring or early summer during snowmelt, followed by gradually decreasing ﬂow until the following spring. The exception to this pattern results from rain-induced ﬂood peaks during the open water period. Ice is an important feature of these rivers from autumn through early spring when it limits atmospheric reaeration of river water for up to 6 months of the year. Furthermore, river ice breakup has signiﬁcant effects on environmental conditions and biological processes (Scrimgeour et al. 1994; Prowse and Culp 2003). Ricketts et al. (1999) divided the basin into the broad categories of Temperate Coniferous Forests, Temperate Grasslands/Savanna/Shrub, Boreal Forest/Taiga and Tundra (also see Fig. 22.3). North of 60 ° latitude the extensive permafrost creates a preponderance of bogs and lakes. Black spruce and tamarack are common in these lowlying, poorly drained areas. Boreal forest dominates much of the basin below 60 °, although temperate mountain forests at higher elevations to the west are dominated by Engelmann and white spruce, lodgepole pine and alpine ﬁr. Northern areas of the basin grade into arctic tundra which is characterized by low shrubs, lichens, mosses and cotton grass–sedge meadows (Mackenzie River Basin Committee 1981). Climate in the basin is harsh and strongly inﬂuenced by continental conditions. Arctic and subarctic climate zones divide the basin at tree line, both regions typiﬁed by cold, long winters (Culp et al. 2005). January air temperatures range between −20 and −30 °C across the region, with July temperatures warming to 14–20 °C. Snowfall

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comprises 30–70% of the precipitation, with subcatchments east of the mainstem receiving 25– 40 cm yr−1 and mountainous catchments in the western basin collecting 50–160 cm yr−1 (Brunskill 1986). Daylight hours in southern portions of the basin are approximately 17 hours in July and 8 hours in December. The Mackenzie Delta experiences 24-hour darkness in December and 24-hour sunlight in June. There are more than 230 protected areas within the Mackenzie River Basin (Mackenzie River Basin Board 2003). These account for about 12% of the total area of the watershed, most of which are located in the southern part of the basin including Jasper, Wood Buffalo and Nahanni National Parks. The Peace–Athabasca Delta wetland within Wood Buffalo National Park is recognized by the Ramsar Convention on Wetlands and is a UNESCO World Heritage Site. In addition, several large tracts of land in the Northwest Territories have been granted interim protection. These lands cover approximately 100,000 km2 and account for almost half of the protected landscape in the entire Mackenzie River Basin. Five Canadian Heritage Rivers lie within the basin, namely the upper Athabasca, Clearwater, South Nahanni, Bonnet Plume and Arctic Red. British Columbia has further designated the Peace, Prophet and Ketchika as provincial heritage rivers. Early inhabitants of the catchment originated from Asia crossing the Bering Sea approximately 12,000 years ago and much of the arctic coastal areas and tundra above the treeline was occupied by Inuit (Rosenberg and Barton 1986). Europeans ﬁrst explored the Mackenzie River Basin in the late 1600s. Many of these explorers worked for the Hudson’s Bay Company with early development and settlements focused on the establishment of fur trading (Mackenzie River Basin Committee 1981). Rivers became key corridors for transportation from Great Slave Lake to the Beaufort Sea with wood-burning paddle wheelers operating on the river from the 1880s until the 1940s. Northerners along the river continue to depend upon barge shipments for many of their commodities. The basin remained in a relatively

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Fig. 22.3 Major landcover types in the MacKenzie River Basin. (From Mackenzie River Basin Board 2003.)

pristine state until the late 1800s. Present-day land use consists of forests (63%), shrubland (18%), grassland (3%) and cropland (3%) (Revenga et al. 1998). Mineral exploration within the basin began in the early 1900s and now includes mining of many minerals such as coal, copper, diamonds, gold, silver, tungsten and uranium. Oil and gas development, which began in the 1930s, however, is now a major stressor within the basin with the

extensive oil sands developments in Alberta, and the forecasted gas development in the Northwest Territories and British Columbia (Fig. 22.4). Key human impacts also include hydroelectric facilities of the W.A.C. Bennett dam on the Peace River, extensive forest harvesting operations located primarily in the Peace and Athabasca river drainages, and mining (Fig. 22.5). Since the 1950s large-scale harvesting of the boreal forest for pulp, paper and timber has fragmented south-

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Fig. 22.4 Oil, gas and oil sands developments in the Mackenzie Basin. (From Mackenzie River Basin Board 2003.)

ern forests. A unique feature of the basin is that development proceeded northward from the southern headwaters with the major cities located in Alberta (Grande Prairie, Fort McMurray) and the southern Northwest Territories (Yellowknife). Fewer than 400,000 people resided in the Mackenzie River Basin in 2000.

22.2 22.2.1

Recent Major Stressors

Contaminants/nutrient enrichment

Over the past 40 years, the Mackenzie Basin has been under increasing contaminant- and nutrient-related stress from developments both within and external to the system (Cash et al. 2000; Gummer et al. 2000; Wrona et al. 2000). Local

sources of contaminants include non-point inputs related to land use activities such as forestry, agriculture and mining; as well as point source inputs from municipal sewage operations, pulp mills and other industries (including the developing oil sands). Naturally occurring contaminants also originate from geological formations (i.e. oil sands), soils and forest ﬁres. Of particular concern has been pollution related to pulp-mill developments, with an emphasis on highly toxic chemical contaminant families. These include resin acids, polychlorinated dioxins and furans, polychlorinated biphenyls, chlorinated phenolics, polyaromatic hydrocarbons and selected heavy metals such as mercury. In addition to contaminant stress, there has been growing concern regarding the effects of enhanced nutrient loadings and degraded dissolved oxygen regimes

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Fig. 22.5 Hydroelectric, pulp, paper mills and key mining activities in the Mackenzie Basin. (From Mackenzie River Basin Board 2003.)

downstream of pulp mill and municipal sewage efﬂuents. The more complicated issue of ecological changes caused by nutrient–contaminant interactions has also received attention (Culp et al. 2000a,b). Finally, the basin receives considerable contaminant input via long range atmospheric transport (Wrona et al. 2000; Mackenzie River Basin Board 2003). 22.2.2 Oil and gas development Increased environmental stress from oil and gas production in the basin is primarily associ-

ated with three activities: oil sands development, natural gas extraction and major pipeline construction (Gummer et al. 2000; Mackenzie River Basin Board 2003). Located in northern Alberta and Saskatchewan, the oil sands area represents one of the world’s largest deposits of bitumen (Fig. 22.4). These deposits cover more than 141,000 km2, approximately twice the size of Ireland, and are estimated to contain approximately 1.6 trillion barrels of bitumen with a potential to recover more than 300 billion barrels using modern technologies. Efforts to develop the oil sands began in the early

Basin Management Approaches used in a High-latitude Northern Catchment 1920s, but large-scale commercial developments occurred much later in the late 1960s. Petroleum extraction from oil sands represents more than 40% of Alberta’s total oil production and over 30% of all oil produced in Canada. Production is projected to approach 2 million barrels per day within the next 10–20 years. Growth and expansion of the oil sands operations continues to have signiﬁcant impacts on the regional environment. Environmental issues associated with this expansion include the large-scale use of surface and ground water (including the loss of fresh water from the hydrological cycle by deep-well injection into geological formations), poor water quality in tailings ponds, habitat fragmentation and increased materials in surface runoff as a result of land disturbance, and air pollution (i.e. acidifying emissions, particulate matter, polycyclic aromatic hydrocarbons, sulphur, volatile organic compounds, greenhouse gases). In the lower Mackenzie Basin (north of 60 ° latitude), the proposed Mackenzie Gas Project (MGP) has raised growing environmental concern. The MGP proposes to develop natural gas ﬁelds in the Mackenzie Delta of Canada’s Northwest Territories and distribute this natural gas to North American markets through a pipeline system to be constructed along the Mackenzie Valley. A multitude of environmental issues have been identiﬁed for this project, most notably permafrost disturbance, landscape subsidence as a result of gas extraction, loss of freshwater wildlife habitat, in-channel effects on the Mackenzie River mainstem, and effects on both ecological condition and local/regional hydrology where the pipeline intersects streams and lakes. Natural gas development is also ongoing in British Columbia but this activity is considerably less threatening to the natural environment of the Mackenzie Basin compared to oil sands and MGP development. 22.2.3

Flow regulation

Although there are a number of relatively small hydroelectric generating stations in the Mackenzie River Basin (e.g. Tazin/Charlot rivers

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in the Athabasca River, Yellowknife/Snare Rivers in the Great Slave sub-basin), their overall regulation effect on Mackenzie Basin ﬂow is minor. By far the greatest effect is produced by the W.A.C. Bennett Dam constructed between 1968 and 1971 in the Canadian Rocky Mountains, which feed the major headwater tributary of the Mackenzie River, the Peace River (Fig. 22.5). The Williston Reservoir formed behind the dam is one of the world’s ten largest reservoirs (70 × 109 m3). Power generation from the Bennett site, combined with that from the Peace Canyon generation station located 23 km downstream, produce approximately 30% of BC Hydro’s generating capacity (Mackenzie River Basin Board 2003). Given the size of the impoundment and how seasonal ﬂows of the Peace River needed to be modiﬁed to supply such a large hydroelectric generating capacity, concerns were raised about how this might affect sediment transport and morphology, ice conditions, instream habitat and biological productivity. Of particular concern was the effect on delta ecosystems, speciﬁcally the Peace–Athabasca and Slave River Deltas. 22.2.4 Climate change Over the last half century, portions of the Mackenzie River Basin have experienced large increases in air temperatures (Fig. 22.6), particularly during winter (Mackenzie River Basin Board 2003). Given that much of the Mackenzie Basin hydrology is strongly inﬂuenced by cryospheric components, such as glaciers, snow, permafrost and lake-river ice, concern developed that the basin water resources would be particularly sensitive to such warming and could produce pronounced and wide-ranging effects. Some of these included: extreme droughts and ﬂoods, productivity and diversity of aquatic ecosystems, timing of ice-melt and ice-over (Fig. 22.7), the ice-based transportation network that provides access to many of the basin’s remote communities, the security of permafrost-bound containment ponds employed to hold wastes by mining companies, and traditional lifestyles of the many aboriginal residents.

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Fig. 22.6 Trends in recorded average spring temperatures in the Mackenzie River Basin over the past 50+ years. (From Mackenzie River Basin Board 2003.)

22.3 Basin Management Programmes 22.3.1

Basin management and legislative responsibilities

Given the geography and transboundary nature of the Mackenzie Basin, jurisdictional responsibility for water management is distributed among numerous levels of government (i.e. federal, provincial, territorial, municipal, First Nations) and various co-management agencies.

This structure presents more possibilities for inter-governmental conﬂict than any other river system in Canada (Royal Society of Canada 1995). While upstream jurisdictions have traditionally upheld sovereignty over their waters, water management in Canada has increasingly become a shared responsibility. Areas of shared responsibility involve source water protection and pollution prevention, drinking water, water quantity/ supply monitoring (i.e. hydrometric monitoring programmes), water quality monitoring, infra-

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Fig. 22.7 Trends in timing of river and lake-ice melt in the Mackenzie River Basin over the past 50+ years. (From Mackenzie River Basin Board 2003.)

structure (e.g. water treatment facilities), environmental assessment, Heritage Rivers, and science and research. The federal government has an exclusive role in a number of areas related to water resources. These are the responsibility for commercial ﬁsheries and navigational waterways, stewardship of federal lands and lands reserved for First Nations, issues related to transboundary waters, and relations with foreign governments. Several key pieces of federal legislation have guaranteed that

the federal government retains an important role in water management. The Canada Water Act facilitates federal-provincial-territorial agreements concerning water research and inventory studies, development of basin management plans and the design and execution of joint projects. The Navigable Waters Protection Act, Fisheries Act, Canadian Environmental Protection Act and Canadian Environmental Assessment Act all ensure that the federal government is involved in environmental protection and conservation, and

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processes related to major developments affecting water resources. Additionally, the provinces, territories, municipalities and First Nations have a role in water management of the Mackenzie River catchment. Provinces have the lead on water management and protection within their boundaries. Their legislative powers include: ﬂow regulation, authorization of water use for development and supply, pollution control and development of thermal and hydroelectric power. Within each province, environment departments, or an analogous agency, have primary responsibility for water management. Other provincial agencies or commissions offer advice to the Federal Government on water policy matters (e.g. the British Columbia Utilities Commission provides recommendations on new hydroelectric developments in the Province). Within the Yukon and Northwest territories, the Water Resources Division of the federal Department of Indian and Northern Affairs Canada (INAC) has the primary responsibility for water management, while various territorial water boards approve licenses for water use and waste disposal. Territorial governments maintain a role in water management through their departments of renewable resources, who also jointly participate with INAC on select water management issues. At the municipal level, water management responsibilities are related to the provision of water and wastewater services, including drinking water and land use planning policies. Finally, a number of co-management agencies/boards, namely the First Nations within the Inuvialuit, Gwhich’in and Sahtu settlement and claims regions, play increasingly signiﬁcant roles in managing water and renewable resource activities within their lands. Given the legislative complexities outlined above, a signiﬁcant challenge to integrated basin management in the Mackenzie has been ensuring that mechanisms were in place to include all appropriate levels of government and other stakeholders (e.g. citizens, scientists, nongovernmental organizations, etc.) in decisionmaking. The programmes discussed below

highlight some of the key challenges, ﬁndings and successes that have occurred in the design and implementation of effective research, monitoring and community-based programmes targeted to improving water management decisions in the Mackenzie Basin.

22.3.2 Northern River Basins Study (NRBS)/Northern Rivers Ecosystem Initiative (NREI)/Peace-Athabasca Delta (PAD) In the late 1980s, increasing concerns surrounding the potential impacts of proposed industrial and resource developments in the Peace, Athabasca and Slave River basins located in the western region of Canada, prompted the federal, provincial and territorial governments to take action. Key environmental issues were: the completion of the Bennett Dam which was assumed to be the cause of environmental degradation of the Peace–Athabasca Delta, 1500 km downstream (Prowse and Conly 2001); greatly expanded forest harvesting to meet the needs of active and proposed pulp mills and efﬂuent discharge from these mills (Wrona et al. 1996); and the announced $25 billion (CDN) investment to develop oil sands deposits within the Athabasca basin. Response to these concerns was complicated by limited knowledge of the physical, chemical and biological environment, and dispersed water management among a variety of government agencies and industries. Public concern peaked in the late 1980s with the proposed construction of the Alberta-Paciﬁc pulp-mill (AlPac) on the Athabasca River, and culminated in recommendations from the AlPac Environmental Impact Assessment Hearings to assess the cumulative environmental impacts of the many developments within the basin. In response to these hearings, the Northern River Basins Study (Northern River Basins Study Board (NRBSB) 1996) was created with explicit objectives to improve understanding of basin ecology and environment within the context of societal concerns and objectives. In 1991, the governments of

Basin Management Approaches used in a High-latitude Northern Catchment Canada, Alberta and Northwest Territories launched the NRBS. The NRBS, which was completed in 1996, addressed environmental issues related to the hydrology/hydraulics and sediment transport, impacts of ﬂow regulation, consumptive and non-consumptive water use, water quality (nutrients, contaminants), food chain effects, ﬁsheries, traditional ecological knowledge, cumulative environmental effects and environmental modelling (Northern River Basins Study Board 1996). It produced a series of ground-breaking scientiﬁc, management and policy recommendations which were based on more than 150 technical reports and 11 synthesis documents (Wrona et al. 1996; Culp et al. 2000c; Gummer et al. 2000). A 25member Study Board with representatives from all levels of government, First Nations, industry, as well as stakeholder groups focused on education, agriculture, health and the environment, guided the work following 16 overarching questions developed to respond to public expectations and concerns (Box 22.1). The information and recommendations from the Study Board brought together scientiﬁc ﬁndings and conclusions which were presented to ministers from the governments of Canada, Alberta and the Northwest Territories. In response to the scientiﬁc knowledge gaps and policy recommendations identiﬁed in the NRBS, the Northern Rivers Ecosystem Initiative (NREI 2004) was subsequently initiated in 1997 and completed in 2004. The NREI involved policy-related initiatives and scientiﬁc research studies focussing on priorities related to contaminants and ecosystem health, safe drinking water, pollution prevention, and hydrology and climate (Gummer et al. 2006). Research was targeted at improving the understanding of the effects of climate variability and change on river ﬂow, hydro-ecological consideration of changes in river ﬂow, ecological responses to point- and nonpoint source pollution and cumulative effects, vulnerability of drinking water quality and wildlife responses to land use changes within the watersheds (Gummer et al. 2006). The activities under the NREI also complemented other large-

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scale scientiﬁc initiatives in the region such as the Mackenzie GEWEX (Global Energy and Water Cycle Experiment) Study (referred to as MAGS). (a) Major foci A rapid increase in the number of pulp and paper mills discharging efﬂuent into the Peace and Athabasca River systems, and the potential impacts on ecological condition, was a major public concern at the onset of the NRBS. Pulp and paper mill efﬂuents are a complex mixture of many compounds, but the primary environmental issues relate to nutrients that enrich food webs and contaminants that can inhibit biological function. Questions of nutrient and contaminant impacts inﬂuenced the framing of 6 of the 16 ‘Guiding Questions’ (Box 22.1) of the NRBS. Principal contaminants discharged from these pulp and paper mills included organochlorine contaminants (i.e. chlorinated dimethylsulphones), chlorinated aromatics (including dioxins and furans), chlorophenolics and chlorinated terpenes, organic acids (e.g. resin acids) and aromatic compounds (e.g. polycyclic aromatic hydrocarbons (PAHs), organochlorines), and sulphur-containing compounds. Furthermore, compounds, such as terpenes and chlorophenolics, were thought to be associated with taste and odour problems downstream of certain pulp mills in the system. Finally, nutrient-laden efﬂuent discharge to the intrinsically phosphorus poor headwaters of the northern rivers, and potentially excessive biochemical oxygen demand (BOD) from efﬂuents was thought to threaten water quality and aquatic life. The effects of ﬂow regulation by the Bennett Dam hydroelectric facility became the focus of a major environmental research programme under the NRBS and the NREI. The studies were directed by an explicit challenge, ‘How does and how could ﬂow regulation affect the aquatic ecosystem?’ (Box 22.1), determining the effects of ﬂow regulation on the hydrology and ecology of the Athabasca, Peace and Slave basins. A unique geographical emphasis was placed on the Peace– Athabasca Delta (PAD) wetland ecosystem, its ecological importance recognized by the International Ramsar Convention on Wetlands and as

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Box 22.1

The 16 Guiding Questions as developed by the Northern River Basins Study Board

Guiding Questions 1 (a) How has the aquatic ecosystem been affected by exposure to organochlorines or other toxic compounds? (b) How can the ecosystem be protected from the effects of these compounds? 2 What is the current state of the water quality of the Peace, Athabasca and Slave river basins, including the Peace–Athabasca Delta? 3 Who are the stakeholders and what are the uses of water resources in the basins? 4 (a) What are the contents and nature of contaminants entering the system and what is their distribution and toxicity in the aquatic ecosystem with particular reference to water, sediments and biota? (b) Are toxic substances such as dioxins, furans and mercury, etc. increasing or decreasing and what is their rate of change? 5 Are the substances added to the rivers likely to cause deterioration of the water quality? 6 What is the distribution and movement of ﬁsh species? Where and when are they most likely to be exposed to changes in water quality and where are the important habitats? 7 What concentrations of dissolved oxygen are required to protect the various life stages of ﬁsh, and what factors control dissolved oxygen in the rivers? 8 Recognizing that people drink water and eat ﬁsh from these river systems, what are the current concentrations of contaminants in water and edible ﬁsh tissue and how are these levels changing through time and by location? 9 Are ﬁsh tainted in these waters and, if so, what is the source of the tainting? 10 How does and how could river ﬂow regulation impact the aquatic ecosystem? 11 Have the riparian vegetation, riparian wildlife and domestic livestock in the river basins been affected by exposure to organochlorines or other toxic compounds? 12 What traditional knowledge exists to enhance the physical science studies in all areas of enquiry? 13 (a) What predictive tools are required to determine the cumulative effects of man-made discharges on the water and aquatic environment? (b) What are the cumulative effects of man-made discharges on the water and aquatic environment? 14 What long-term monitoring programmes and predictive models are required to provide an ongoing assessment of the state of the aquatic ecosystems? These programmes must ensure that all stakeholders have the opportunity for input. 15 How can study results be communicated most effectively? 16 What form of interjurisdictional body can be established, ensuring stakeholder participation for the ongoing protection and use of the river basins?

Basin Management Approaches used in a High-latitude Northern Catchment a UNESCO World Heritage Site. Simultaneous with the initiation of the NRBS, a multi-agency group composed of the governments of Alberta and Canada, the British Columbia Hydro and Power Authority (B.C. Hydro), the Mikisew Cree First Nation, the Athabasca Chipewyan First Nation and the Fort Chipewyan Metis Association began another major research programme, the Peace–Athabasca Delta Technical Studies (PAD-TS). PAD scientiﬁc objectives were to develop available options and select the most suitable strategy for restoring the role of water in the Peace–Athabasca Delta. The strong PAD-TS focus on water developed from previous research that showed the very signiﬁcant role that wetting and drying cycles played in controlling biological productivity and diversity within the delta. (b) Major ﬁndings The NRBS used a weight-ofevidence approach to assess and summarize the cumulative effects of multiple environmental stressors (Fig. 22.8) and development on the Peace, Athabasca and Slave regions of the Mackenzie Basin (Culp et al. 2000b). This analysis revealed reach-speciﬁc priorities and areas of concern requiring management action or further scientiﬁc attention. Transport and fate studies in the Athabasca, Peace and Slave River basins discovered many contaminant groups in low concentrations in the ambient water phase, including chlorinated phenolics, chlorinated dioxins and furans, and resin acids (Wrona et al. 2000). In addition, signiﬁcant declines in suspended and deposited concentrations of chlorinated contaminants were observed during the study with this decrease correlated directly with implementation of enhanced efﬂuent treatment of pulp mill efﬂuent. In general, the environmental levels of chlorinated organic and metal contaminants in water or sediments were within Canadian health and environmental guidelines. Researchers found low levels of environmental contamination in ﬁsh, with the occasional exception of dioxins, furans, PCBs and mercury in biota (Cash et al. 2000). Although contaminant loads in ﬁsh typically conformed to

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Canadian guidelines for aquatic and human health, lines of evidence suggested that many ﬁsh species exhibited physiological stress. The most worrisome ﬁndings were the depressed sex hormone levels in burbot (Lota lota) and longnose sucker (Catostomus catostomus) collected from locations near pulp mill efﬂuent discharge. Furthermore, numbers of immature ﬁsh were unexpectedly high in these locations, with these juveniles often exhibiting external abnormalities. Contaminants were of particular concern below pulp mill discharges on the Athabasca and Peace Rivers, and below the oil sands developments on the Athabasca (Fig. 22.8). An important environmental issue that had received comparatively little attention at the time was the effect of nutrients from pulp mill efﬂuents on the trophic status of receiving waters (Bothwell 1992). The analysis by Chambers et al. (2000a) of long-term monitoring data indicated that on an annual basis, pulp mills contributed up to approximately 20% of the phosphorus (P) and nitrogen (N) load discharged to these northern rivers. The importance of pulp mill derived P was demonstrated by means of ﬁeld experiments which indicated that low levels of benthic algal biomass were maintained by insufﬁcient P in the upper reaches of the Athabasca River. By contrast, efﬂuent loading from pulp mill and sewage inputs relaxed nutrient limitation downstream of major discharges. Algal biomass was increased sharply at low P concentrations near 5 µg L−1 (soluble reactive P), and approached saturation at 35 µg L−1. In addition, Culp et al. (2000a) integrated results from artiﬁcial stream experiments and ﬁeld biomonitoring to conclude that the primary response of the benthic food web (excluding ﬁsh) to pulp and paper mill efﬂuents in these river systems was largely one of nutrient enrichment rather than that of toxicity. The greatest concern of nutrient enrichment was associated with pulp mill and municipal efﬂuent discharges in the upper reaches of the Athabasca and Wapiti Rivers (Fig. 22.8). Recommendations for management of nutrient input to the basin were crafted from these research ﬁndings, including suggestions for a reduction and capping of nutrient

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Fig. 22.8 Cumulative effects assessment of human impacts in the Peace-Athabasca River Basin as noted in the weight-of-evidence exercise for the Northern River Basins Study (NRBS). (After Culp et al. 2000b.)

loading, and the necessity of developing reachspeciﬁc guidelines for P. The threat of low dissolved oxygen (DO) is often greatest during winter in northern climates when ice cover restricts reaeration of waters with atmospheric oxygen. Chambers et al. (2000b) examined the adequacy of DO regulations to protect biota during the winter through analysis

of monitoring data collected over more than 30 years. Depressions in DO were evident below efﬂuent discharges and appear to contribute signiﬁcantly to an increased rate of decline in DO. Low DO concentration was of particular concern below pulp mills of the upper Wapiti River and middle reaches of the Athabasca River (Fig. 22.8). To permit forecasting of declines in DO resulting

Basin Management Approaches used in a High-latitude Northern Catchment from BOD in efﬂuents or changes in river discharge, NRBS scientists successfully implemented a one-dimensional, steady-state water quality model to forecast DO decreases along a 800-km reach of the Athabasca River. The ecological effects of exposure to low DO include: delays in mountain whiteﬁsh (Prosopium williamsoni) egg development; lowered alevin posthatch mass of bull trout (Salvelinus conﬂuentus); extended spawning period of burbot; and depressed feeding rates and lowered survival of the mayﬂy (Baetis tricaudatus). As a result of recommendations from this integrated programme of monitoring, modelling and experimentation, efﬂuent BOD loading was reduced to the Athabasca and Wapiti rivers and the DO guideline for protection of aquatic life raised from 5.0 to 6.5 mg L−1. (c) Flow regulation In response to the NRBS/ NREI question 10 (Box 22.1), the major ﬁndings of the study focussed on the regulation effects produced by the Williston Reservoir on the Peace River, Peace–Athabasca Delta and the Slave River Delta. In general, there were four major factors assessed: ﬂow patterns, sediment transport and river morphology, ice formation and aquatic habitat. In terms of ﬂow patterns, the most pronounced regulation effects on the Peace River occurred during the ﬁlling years, 1968 to 1971. During this period annual ﬂows just below the dam were less than half the normal ﬂow and approximately two-thirds of normal near the downstream end of the Peace River, which is approximately 1100 km downstream near its conﬂuence with the Slave River that feeds Great Slave Lake. Once ﬁlling was complete, the impoundment has not signiﬁcantly altered the annual ﬂow volume but has changed the seasonal distribution of ﬂows with reduced ﬂows during the spring-summer and increased ﬂows during winter to match the seasonal demand for hydroelectricity. The effects on the seasonal hydrograph were noted to diminish downstream as ﬂow additions from other unregulated tributaries dampen the regulation effect.

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Changes in seasonal timing and magnitude of ﬂows plus releases of warm water from the reservoir were found to also modify the river ice regime. Closest to the dam, the ice season was signiﬁcantly delayed and shortened. No signiﬁcant effects were noted, however, in timing and duration of ice cover in the downstream extremities of the Peace River. As part of the ice regime analysis, the NRBS found that river ice-jams on the Peace River were the principal agent responsible for ﬂooding of the downstream Peace– Athabasca Delta, and its multitude of highly productive ‘perched’ basins (Prowse and Conly 2001). This dispelled the previously held notion that it was the peak open-water ﬂows of the Peace River that caused such inundation. Furthermore, it was discovered that it was spring snowmelt from downstream tributaries that drove such ice-jam ﬂoods and not higher alpine runoff captured in the spring time by the reservoir. Moreover, change in winter climate that controlled snow accumulation was found to be primarily responsible for decreases in spring ﬂows and associated ice-jam formation. This threat to the PAD ecosystem was notably the only part of the basin were hydrological impacts were highlighted as stressors of considerable concern (Fig. 22.8). Questions about how regulation affected the speciﬁc ice regime of the ﬂoods, which could affect their frequency and magnitude, remained largely unanswered. The NRBS also recommended the use of regulation, via augmented ﬂow releases during the spring, to offset the natural decline in downstream snowmelt runoff and increase the probability of ice-jam ﬂooding of the Peace–Athabasca Delta. Notably, this approach was subsequently tested with successful results and has been identiﬁed by Anisimov et al. (2007) as one method of using ﬂow regulation for adaptation to the effects of climate change. Although most sediment in the Peace River was found to be contributed by tributaries downstream of the dam, lowered peak ﬂows by regulation reduced the ability of the river to scour material. As a result, enhanced accumulation in deposition zones was noted to be giving the river

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a new shape with changes in associated vegetation and habitat. Decades or longer are likely to be required before the river morphology will reach a new equilibrium for the regulated regime, and much longer for vegetation and habitat. Changes in instream habitat were particularly difﬁcult to assess, as is the case for most large rivers. Remote sensing of changes in habitat availability under different ﬂow regimes was used to speculate about how ﬂow regulation affected the distribution of ﬁsh and wildlife populations. (d) Management implications In conclusion, the NRBS, NREI and PAD programmes were successful in providing new investment in scientiﬁc research for priority issues which were vetted by stakeholders and with emphasis by jurisdictions on integrated basin management approaches. More efﬁcient monitoring, assessment and reporting networks for contaminants, nutrients and ﬂow regulation impacts developed as a result. Notably, commitment to the Mackenzie River Basin Board was galvanized. A key duty of this Board is to report on the state of the aquatic ecosystem at 5-year intervals (Mackenzie River Basin Board 2003). Cumulative impact assessment frameworks for the Basin (Culp et al. 2000b), development of thresholds for environmental protection, and the use of decision-support models also aided development of improved regulatory guidelines for DO, nutrient loading, and protocols for ﬁsh consumption advisories. Lastly, these programmes solidiﬁed recognition of the important role that public consultation and community outreach have in creating public trust in environmental science and related policy decisions. 22.3.3 Cumulative Environmental Management Association (CEMA)/Regional Aquatic Monitoring Program (RAMP) In response to growing environmental issues and regulatory requirements associated with developments in the oil sands region of the Mackenzie Basin, two programmes and associated organizations were established to help assess the impacts

of resource development on the lower reaches of the Athabasca River. An industry-led initiative, the Regional Aquatic Monitoring Program (RAMP), was initiated in 1997 by Suncor Energy Inc., Oil Sands, Syncrude Canada Ltd. and Shell Canada Limited to implement a cost-effective, coordinated, long-term monitoring programme focused on assessing the impacts of oil sands mining and related developments on water quality/quantity in the Athabasca River Basin. In addition to RAMP, the Cumulative Environmental Management Association (CEMA), which is a registered not-for-proﬁt, non-governmental organization, was established in 2000 to facilitate the study of, and reporting on, the cumulative environmental effects of industrial development in the region. (a) Major foci The strategic objectives of the RAMP were to: coordinate industries’ environmental monitoring activities; ensure that the information collected could be used to assess regional trends and cumulative effects; address regulatory monitoring requirements; provide baseline data against which impact predictions of recent environmental impact assessments (EIAs) for oil sands developments could be veriﬁed; and design and execute a programme which addresses the anticipated aquatic monitoring requirements of oil sands operators’ environmental approvals. The RAMP programme is co-ordinated by a steering committee comprising representatives from industry (forestry and oil sands sectors), local communities, First Nations groups, and municipal, provincial and federal levels of government. As a multi-stakeholder organization, CEMA is governed by over 45 members representing all levels of government, industry, regulatory bodies, environmental groups, aboriginal groups, and local health authorities/agencies. CEMA has ensured that participation in the programme involved: governments and agencies that regulate and oversee resource development such as the oil sands; aboriginal communities so that traditional lifestyle concerns (e.g. culture and environmental knowledge) are considered; industries that are committed to corporate responsibility

Basin Management Approaches used in a High-latitude Northern Catchment and sustainable development of resources; federal, provincial and local health agencies that are focused on promoting public wellness and preserving public safety; and environmental nongovernment organizations that are concerned with guarding and promoting environmental sustainability. (b) Major ﬁndings To date, RAMP monitoring surveys have included evaluation of sediment and water quality, benthic invertebrate community structure, ﬁsh habitat and ﬁsh populations and communities in the Athabasca River and selected tributaries, and species composition and distribution of aquatic vegetation in wetlands. Monitoring endpoints and level of effort varied by sub-watershed and depended on data available from previous surveys, the type of impacts predicted by previous EIAs and logistical constraints (see http://www.ramp-alberta.org/). CEMA has produced numerous reports, guidelines and management frameworks for the region and associated river basins (see http://www.cemaonline. ca/). 22.3.4 Mackenzie Gas Project The Mackenzie Gas Project (MGP) is a major proposed development in the western Canadian Arctic that includes plans to develop natural gas ﬁelds in the Mackenzie Delta region of the Mackenzie Basin, and deliver the natural gas and natural gas liquids to markets in Canada and the USA. The Project consists of ﬁve major components: three natural gas ﬁeld production facilities; a gathering pipeline system; a gas processing facility near Inuvik, Northwest Territories; a natural gas liquids pipeline from the Inuvik area facility to Norman Wells, Northwest Territories; and a natural gas pipeline from the Inuvik area facility to northwestern Alberta. Three natural gas ﬁelds production ﬁelds are proposed (Niglintgak, Taglu, Parsons Lake) that collectively could supply approximately 800 million cubic feet per day of natural gas over the life of the Project. Other natural gas ﬁelds in the North could also be connected into the distribution

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network. In total, up to 1.2 billion cubic feet per day of natural gas could move through the pipeline. The natural gas pipeline (30 inches in diameter) will run approximately 1220 km along the Mackenzie River Valley (Fig. 22.4). Compressor stations will be built at regular intervals along the pipeline to maintain the ﬂow of the natural gas and more compression facilities will be added along the route in the future if it is necessary to increase the amount of natural gas moved through the pipeline. Both the natural gas and natural gas liquids pipelines will be located along the same right-of-way. (a) Major foci Water-resource-related environmental concerns related to the MGP fall into several broad categories, namely: the effects of the pipeline construction and operations on local and regional landscape hydrology, including stream and lake crossings and permafrost disturbance; landscape subsidence and loss of migratory bird nesting habitat in the Kendall Island Bird Sanctuary; in-channel effects on the Mackenzie River mainstem, and; water quality and ﬁsheries habitat issues during and after construction. In 2004, the National Energy Board (NEB) of Canada, an independent federal agency established in 1959 by the Parliament of Canada to regulate international and inter-provincial aspects of the oil, gas and electric utility industries, initiated public hearings as part of the required Environmental Impact Review process. The proponents of the MGP are Imperial Oil Resources Ventures Limited, Mackenzie Valley Aboriginal Pipeline Limited Partnership, Imperial Oil Resources Limited, ConocoPhillips Canada (North) Limited, ExxonMobil Canada Properties and Shell Canada Limited. The goals of the public hearings are to obtain evidence, including traditional knowledge, and views of interested persons with respect to the Mackenzie Gas Project. The National Energy Board hearing process is co-ordinated with the Environmental Impact Review of the Mackenzie Gas Project by the Joint Review Panel as set out in the ‘Cooperation Plan for the Environmental Impact

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Assessment and Regulatory Review of a Northern Gas Pipeline Project through the Northwest Territories’ (June 2002). The public consultation process is used to seek information to improve the design, construction and operation of the Project, which includes information from environmental and traditional knowledge studies. The combined input is used to help deﬁne the proposed pipeline route, facility locations and infrastructure locations to support construction of the Project. In 2007, the estimated cost of the proposed MGP was $16.2 billion CDN and growing. At the time of writing, the MGP was still under regulatory review by the NEB and was not yet approved to proceed. Final technical reports are to be published in 2009. 22.3.5 Mackenzie GEWEX (MAGS) and MBIS In response to the issues surrounding a changing climate in the Mackenzie River basin, two separate major studies were initiated; the Mackenzie GEWEX Study (MAGS) and the Mackenzie Basin Impact Study (MBIS). The MAGS began in 1994 as a contribution to the Global Energy and Water Cycle Experiment (GEWEX) of the World Climate Research Program/World Meteorological Organization. Its primary focus was to improve the understanding of the ﬂow of energy and water into and through the atmospheric and hydrological systems of the Mackenzie River Basin (Woo 2008). The cold-regions focus of MAGS was to complement other GEWEX scientiﬁc programmes conducted in other hydro-climatic regions of the globe. The MBIS was a response to a more general concern about the effects of climate change in major regions of Canada (Cohen 1997). The Mackenzie Basin was one of three areas selected for study, the other two being the Prairies and the Great Lakes. It was selected to represent a highlatitude region with ecosystems particularly sensitive to climate change and a large number of aboriginal people following traditional lifestyles. The MBIS operated from 1991 to 1996.

Unfortunately, the different timing of the MAGS and MBIS programmes precluded signiﬁcant integration of approaches and results between the two. (a) Major foci MAGS had two major foci: (i) to understand and model the high-latitude water and energy cycles that respond to and/or affect the climate system; and (ii) to improve the ability to assess the changes in basin water resources produced by climate variability and change. By contrast, the MBIS had a broader impact focus, with the primary purpose of evaluating what the effects of a changing climate might have on the Mackenzie Basin, its lands, waters and the communities that depend on them. Of speciﬁc interest was ‘How would an economy which was based on natural resources and a Northern culture cope with climate warming? How could they deal with the changes which were expected to be the most signiﬁcant in the world?’ (Cohen 1997). (b) Major ﬁndings Overall, through a variety of diverse and integrated atmospheric and hydrological process and modelling studies, MAGS quantiﬁed the major components of the water cycle of the Mackenzie River Basin. Individual process and modelling studies conducted at a variety of spatial scales included, for example, moisture source analysis, role of clouds in energy and water budgets, signiﬁcance of snow sublimation and large-lake evaporation, and the source/ storage of surface and ground waters particularly in permafrost-dominated environments. Based on the results of these various studies, the MAGS programme achieved a much improved understanding of the internal circulation of moisture and heat within the large basin and their linkages to external atmospheric sources. Moreover, such improved understanding also greatly advanced the ability to conduct integrated atmospherichydrological modelling. Successful modelling was accomplished primarily at a monthly timescale and a spatial resolution of approximately 50 km. This has provided invaluable tools for evaluating the sensitivity of the basin water cycle and its major components to change and vari-

Basin Management Approaches used in a High-latitude Northern Catchment ability in climate. Through targeted interfacing with user groups, MAGS attempted knowledge transfer of the results for such applications as weather forecasting, hydroelectric operations and various aspects of water-resource management. Although the MBIS was largely conducted before the MAGS results were achieved, it did identify some key vulnerabilities to climate change within the basin. Some of the major items germane to water-resource management included, for example: regional variations in timing and magnitude of runoff; lower minimum water levels in key transportation waterways; reduced ice seasons; and enhanced sediment erosion and transport produced by thawing permafrost. Given that the overall results of the MBIS were speciﬁcally framed for use by stakeholder and policy-makers, it raised the issue of adaptation particularly by communities. Of particular concern was the non-wage economy of traditional lifestyles practised by many aboriginal communities, especially if climatic changes were very rapid leaving little time for adaptation.

22.4

Summary and Recommendations

A wide array of research, monitoring and environmental assessment programmes have been conducted in the Mackenzie Basin over the past 40 years. While each of these initiatives have unique features and often was constrained geographically by the environmental stressor(s) they dealt with, they also shared several similarities in their design and implementation. Given the complex inter-jurisdictional structure of the basin and the range of stakeholders and issues that needed to be considered, a key common feature to all programmes was the explicit recognition of the need to engage all appropriate stakeholders in the design and implementation of the studies, including communication strategies and outreach. In the examples of the Northern River Basins Study and Northern Rivers Ecosystem Initiative, special efforts were made to ensure that the studies were managed by multi-stakeholder

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boards or steering committees representing appropriate levels of government, First Nations, industry, communities and environmental interests. In both of these initiatives, science was included early in the design as an integral part of the assessment process. The science concerns of stakeholders, which included those suggested by First Nation communities, were expressed as ‘Guiding Questions’ that were then addressed by the scientiﬁc community. An explicit effort was also made to integrate research and synthesize science ﬁndings from multiple disciplines. This aspect of the programmes involved iterative vetting and dialogue between scientists and stakeholders. Such an approach was deemed essential to the acceptance of science into subsequent decision-making/policy processes and it underscored the importance of developing ‘trust’ between stakeholder concerns, the research community and decision-makers. Another outstanding feature of the NRBS was the involvement of an external science advisory committee (SAC), which was composed of nationally and internationally recognized scientists. The SAC reviewed the design, implementation and ﬁndings of the research programme and along with the Study Board, made sure that the outcomes addressed the identiﬁed science gaps and stakeholder concerns (Fig. 22.9). Other research, monitoring and environmental assessment programmes such as RAMP, CEMA, GEWEX and MBIS and MGP have also recognized the importance of multi-stakeholder engagement and communications strategies in their assessment approaches. In addition, all of the programmes and initiatives have contributed to a varying extent to the development and improvement of integrated basin management approaches for the Mackenzie system and are reﬂected in various policy levels from simple regulatory statutes to more comprehensive management approaches. Key examples of such policy-related changes include: more efﬁcient and effective monitoring, assessment, and reporting; improved governance models (e.g. the establishment of Mackenzie River Basin Board); implementation of remedial actions and

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Fig. 22.9 Processes involved in the effective communication between scientists and decisionmakers. (After Preston et al. 2004.)

pollution prevention measures; the development of new, cumulative impact assessment frameworks and related decision-support models; improved regulatory guidelines and environmental protection measures; and the recognition of public consultation and community outreach as an important component of environmental stewardship and decision-making.

22.5 Future Recommendations As Canada moves towards developing the southern headwater regions of the Mackenzie Basin and increasingly accessing its northern untapped resources, the number, type and spatial extent of environmental stressors are increasing. Energy demands, for example, are leading to expanded development of oil and gas resources in northern and southern portions of the basin, and renewed discussion about the large hydroelectric potential

that remains throughout the basin. Expansion of hydroelectric development on northern rivers is projected to increase given growing demands for new non-carbon based energy sources (Prowse et al. 2004). Unless only run-of-river sites are employed, signiﬁcant new ﬂow regulation of the Mackenzie main stem and tributaries is likely to result. Northern Canada is also the site of numerous, largely remote, mining operations, the development and/or expansion of which are currently economically constrained by the high costs of transportation. Many of these have a strong potential to inﬂuence local water quality, especially where tailings ponds – often established in drained lakes – are used to store wastes. While future climatic warming is forecast to reduce or even eliminate the period during which ice-based surface transportation can be used to access such sites, losses of sea ice near the arctic coast of the Mackenzie Basin may lead to the establishment

Basin Management Approaches used in a High-latitude Northern Catchment of new sea ports and an increased internal landbase road network (Furgal and Prowse 2008). Such improved access will make further naturalresource extraction much more economically viable, and increase the potential for major impacts on water quality. Overall, it is probably climate change that has the potential to produce the most signiﬁcant and diverse combination of stressors on the water resources of the Mackenzie River Basin. Although the MBIS and MAGS outlined some of these, recent comprehensive national and international assessments outline more fully the entire range of potential water-based impacts that will affect northern basins such as the Mackenzie (e.g. Canadian climate-change assessment – Furgal and Prowse 2008; Intergovernmental Panel on Climate Change – Anisimov et al. 2007, Bates et al. 2008; Arctic Climate Impact Assessment – Walsh et al. 2005, Wrona et al. 2005). Many similar socio-economic impacts are noted, but also details about how climate change will affect aquatic habitat, including the related impacts on waterfowl, aquatic mammals and ﬁsh on which traditional lifestyles of aboriginal communities depend. Given the number, diversity and broadening scope of stressors affecting the water resources of the Mackenzie River Basin, effective water management will continue to require a multi-jurisdictional approach and philosophy as implemented by the Mackenzie River Basin Board (Box 22.2). Such a group needs to be supported by a highly reputable and credible research network involving government, university and industry scientists capable of addressing a diverse set of environmental issues. Moreover, an integrated, basin-wide research and monitoring programme is a necessary ongoing activity to provide credible information on environmental status and trends. As stressed in Bakker’s (2007) review of the future of Canada’s water resources, the formation of inter-jurisdictional advisory bodies and boards that incorporate scientiﬁc experts is required to bridge scientiﬁc knowledge with effective decision-making and governance (i.e. Preston et al. 2004; Fig. 22.9). Such innovative basin manage-

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ment approaches will be necessary for the sustainable development of river basins that ensures effective conservation and protection measures.

Box 22.2 Mackenzie River Basin Board vision and guiding principles Vision A healthy and diverse aquatic ecosystem for the beneﬁt of present and future generations. Guiding principles The Agreement commits the parties to the following principles in carrying out their responsibilities in the Basin. • Manage the water resources in a manner consistent with the maintenance of the ecological integrity of the aquatic ecosystem. • Manage the use of the water resources in a sustainable manner for present and future generations. • Allow each Party to the Agreement to use or manage the use of water resources within its jurisdiction provided such use does not unreasonably harm the ecological integrity of the aquatic ecosystem in any other jurisdiction. • Provide for early and effective consultation, notiﬁcation and sharing of information on developments and activities that might affect the ecological integrity of the aquatic ecosystem in another jurisdiction. • Resolve issues in a cooperative and harmonious manner

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Basin Management Approaches used in a High-latitude Northern Catchment Ricketts, T.H., Dinerstein, E., Olson, D.M. et al (1999) Terrestrial Ecoregions of North America: a conservation assessment. Island Press, Washington, DC. Rosenberg, D.M. and Barton, D.R. (1986) The Mackenzie River system. In: Davies, B.R. and Walker, K.F. (eds), The Ecology of River Systems. Junk, Boston, MA, pp. 425–433. Royal Society of Canada (1995) Aquatic Science in Canada: a case study of research in the MacKenzie Basin. Aquatic Science Committee of the Royal Society of Canada, Ottowa, ON. Scrimgeour, G.J., Prowse, T.D., Culp, J.M. and Chambers, P.A. (1994) Ecological effects of river ice break-up: a review and perspective. Freshwater Biology, 32, 261–276. Walsh, J., Anisimov, O., Hagen, J.O. et al (2005) Crysophere and hydrology. In: Symon, C., Arris, L. and Heal, B. (eds), Arctic Climate Impact Assessment. Cambridge University Press, Cambridge, pp. 183–242.

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Woo, M. (2008) Cold Region Atmospheric and Hydrologic Studies: The Mackenzie GEWEX Experience. Vol. 1: Atmospheric Dynamics. Springer-Verlag, New York. Wrona, F.G., Gummer, W.D., Cash, K.J. and Crutchﬁeld, K. (1996) Cumulative Impacts within the Northern River Basins. Northern River Basins Study, Synthesis Report No. 11. Alberta Environment, Edmunson, AB. Wrona, F.J., Carey, J.C., Brownlee, B. and McCauley, F.E.R. (2000) Contaminant sources, distribution and fate in the Athabasca, Peace and Slave river basins, Canada. Journal of Aquatic Ecosystem Stress and Recovery, 8, 39–51. Wrona, F.J., Prowse, T.D., Reist, J.D. et al. (2005) Freshwater ecosystems and ﬁsheries. In: Symon, C., Arris, L., Heal, B. (eds), Arctic Climate Impact Assessment. Cambridge University Press, Cambridge, pp. 353–452.

Image facing chapter title page: Courtesy of Environment Canada/P. di Cenzo.

23

The Future for Catchment Management

R O B E R T C . F E R R I E R 1, A L A N J E N K I N S 2 A N D KIRSTY BLACKSTOCK1 2

1 The Macaulay Institute, Craigiebuckler, Aberdeen, UK Centre for Ecology and Hydrology, Crowmarsh Gifford, Wallingford, Oxfordshire, UK

23.1

Introduction

This ﬁnal chapter outlines the main issues and challenges for future catchment management. These are both environmental and human, given that managing water resources means managing land use and the myriad of human derived demand for both land and water. Furthermore, catchments are complex human–environmental systems with temporal and spatial dynamics, and their management requires consideration of political, cultural and social process. The importance of the catchment as a central concept in water resource planning is without doubt. The catchment deﬁnes the boundary condition for the available water; the maximum available resource unless water is pumped into the catchment from elsewhere. Groundwater boundaries do not always coincide with topographic boundaries but storage in aquifers can be assessed and factored into water resource availability calculations. The available water resources must support all demand within the catchment, as well as often outside of the catchment area, including potable supply, irrigation, cooling and for ecosystem services. The chemical quality of water can also be considered in this respect. Although it is not a resource use per se it does Handbook of Catchment Management, 1st edition. Edited by Robert C. Ferrier and Alan Jenkins. © 2010 Blackwell Publishing, ISBN 978-1-4051-7122-9

affect the availability of water for some uses, for example poor water quality can render it unusable as a resource. Activities and decisions made within catchment systems and relating to any one resource demand will ultimately impact on the nature and function of downstream waterbodies, be they groundwaters, lakes, river, wetlands, estuaries and coastal zones. The catchment provides the unique unit to assess these management issues. Catchments have been impacted by human activity throughout history, and as such have stored up environmental and social challenges for the future through, for example, the pollution of groundwater sources. The recognition of the negative and cumulative impact human activities are having are reﬂected in global initiatives such as integrated water resource management (IWRM; see below) and the European Water Framework Directive’s desire to return water bodies to ‘good status’. Water is fundamental to the delivery of basic human needs (Maslow 1943) as demonstrated by the Millenium Development Goals, as well having important cultural, historical and spiritual meanings associated with its existence and/or uses. Thus, the allocation of water resources is a profoundly social and political matter, be it deciding how to allocate responsibility for mitigating diffuse pollution in afﬂuent countries in Western Europe or weighing up local livelihoods versus energy needs provided by massive hydroelectric dams in less developed countries.

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It is the human activities within catchments which have driven increased demand for water supplies, including increasing human population and increased agricultural productivity. The latter, largely as a result of irrigation demands, has driven a dramatic increase in global water consumption over the last 50 years. Globally, we are now faced with an ever increasing population which further drives the requirement for increased agricultural production but yet more ominous is that this future demand must be met in the light of climate change. Land use activities, in particular, agriculture and biofuel production, forestry, land-drainage urbanization, power generation and industrial processes inﬂuence the quality and quantity of waters, and along with in-channel activities such as navigation management, ﬂood defence and re-alignment, impact on the ecological quality of aquatic environments. As Falkenmark and Rockström (2006) comment, ‘a land use (L) decision is also a water decision’; integrating ‘L’ within IWRM should be seen as a priority. Climate change inﬂuences, on the other hand, are perhaps already impacting on the nature, form and distribution of water resources, resulting in the potential for altered land use and management in the medium term. Climate change may well lead to inadequate water availability in some areas which are currently under high agricultural production and so require high volumes of water resource. This will drive the agricultural production and the consequent requirement for water into new catchment areas. The consequence for current water resources management could be huge. The response to emerging energy and climate change policy targets, however, has driven the recent expansion of biofuels at the expense of food production and commodity prices have soared. The consequence of this has been a response by the agriculture industry (especially in the European Union) to increase the land area under cultivation for food, with a knock on effect potentially compromising conservation and biodiversity objectives. Clearly there is poor integration of policy with global market forces operating

in very short term cycles with a strong inﬂuence on land use decisions and potential consequences on water resources. There is, therefore, an urgent requirement for better synergy of climate change, energy and land use policy planning, with a clear realization that the common currency in all aspects is water (Carter et al. 2005; William Page and Susskind 2007). In this chapter we aim to highlight some of the current and perceived future pressures on catchment resources globally and to propose a strengthening of the source to sea, land water continuum, systems level approach to the integrated management of resources.

23.2

Catchments and Climate Change

There is now a recognized body of evidence to suggest that the world’s climate is changing, and changing faster than ever before (IPCC 2007). Global temperatures (air, land and ocean) are all increasing and associated with this is the melting of polar and glacial ice, and altered climate patterns. Such changes to the climate have an impact on hydrological regimes globally, increasing variability and uncertainty. Arid and semi-arid areas are perceived to be particularly exposed to the impacts of climate change. The most recent IPCC Technical Report on Climate Change and Water (IPCC 2008) has reiterated that changes to the large-scale hydrological cycle will include: • Increased atmospheric water content. • Altered patterns of precipitation such as seasonality and intensity. • Reduced snow and increased ice melt. • Changes in volumes and timing of runoff (including those as a consequence of land use change). This will result in a plethora of global consequences including: an increase in precipitation and potential water availability in high latitudes and wet tropics, with a concomitant decrease in dry subtropical mid-latitude areas. As of present, predictions for Asia remain uncertain and there is much quantitative uncertainty at local scales.

The Future for Catchment Management Increased water availability is also associated with potential changes in the frequency of ﬂoods and droughts, in other words, big changes in extremes rather than mean values. There is also a predicted decrease in glacier storage resulting in changes in the seasonality of river ﬂows, with short-term increases in glacier water availability followed by a potential long-term loss of resource. Although much more geographically uncertain it is thought that, in general, water quality might decline. Indication of such responses would be evident from trends in suspended solid loads, salinization, reduced thermal mixing of lakes and increased episodes of algal blooms and eutrophication. Sea level rise and the associated increase in estuarine and groundwater intrusion could be a major issue in coastal areas, thereby reducing the available resource. In nearly all cases, negative impacts outweigh the beneﬁts. Changes in availability will affect food availability, access to resources and potentially, political stability. Changes will also affect the function and operation of water management infrastructures, such as those associated with supply (both irrigation and potable), power generation and ﬂood control for example. The IPCC Report (2008) identiﬁed that ‘higher water temperatures, increased precipitation intensity and longer periods of low ﬂows exacerbate many forms of water pollution, with impacts on human health, water systems reliability and operation costs’. Water management systems have historically been operated under the assumption of stationarity. Stationarity is where statistical parameters (mean, standard deviation, etc.) of observations do not change over time. Such an approach has been used globally to determine risk to infrastructure and the management of water resources (one in a hundred year ﬂood determination, or storm intensity, for example). Hydrologists now face a future where assumptions about key parameters can no longer be based on historical records. Climate change undermines this long-held principle, even with aggressive mitigation action on greenhouse gas emission, hysteresis processes suggest a continued global warming is very likely (Milly et al.

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2008). There is a requirement for new data and information to assess the magnitude and direction of this drift in baseline conditions but global monitoring systems are currently very unevenly distributed, and many areas, in particular Africa and Asia, are under-represented. Technological advances in recent years have produced a range of new sensors and tools for monitoring hydrology and chemistry, but as yet many of these devices have not been implemented in large national monitoring programmes. Current responses to the variability associated with climate change fall into general categories (Moench and Stapleton 2007) attempting to: 1 control hydrological systems through largescale structural intervention; 2 reduce the impacts of variability through more distributed interventions; 3 alter settlements and economic activities away from areas most a risk; 4 diversify in order to spread risk. In fact, it is important that these responses are not treated in isolation but that risk management represents a dovetailing of activities. Importantly, a catchment-based approach provides an ideal platform for the integration of social, economic and biophysical dimensions to build resilience. Established water policies such as the EU Water Framework Directive have promoted such an integrated approach but now need to be ‘climateproofed’. Most EU Member States have now begun a process of considering climate change impacts in their river basin management plans as part of the process of moving waters towards good status. Of all the European and regional countries implementing the Directive, however, only Spain has a Royal Decree making it mandatory to consider climate change impacts in river basin management plans. Guidance from the scientiﬁc community on hydrological responses to climate change needs to be built into investment risk analysis supporting large-scale donor plans in developing countries. Risk-based decision frameworks need to factor in the increasing uncertainty that climate change brings but waiting is not an option. There is a requirement for the development of new

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indicators assessing vulnerability; ones based on the degree of exposure to external factors, reliability over time, resilience and the ability to cope with change and vulnerability or the expected change in response to an increased pressure. The adoption and implementation of so-called ‘no-regret’ and ‘low regret’ and ‘win-win’ adaptation strategies is being actively encouraged as part of a precautionary principle approach by many institutions, municipalities and communities (Box 23.1).

Box 23.1 Managing risk and uncertainty associated with climate change No regret options Provide beneﬁts even if no climate change occurs: • Improve water conservation and reduce leakage. • Avoid building on ﬂoodplains. • Improve inﬁltration from urban areas. • Enhance understanding of options. • Public engagement. Low regret options Provide important beneﬁts at little additional cost: • Increase storm collection efﬁciency during upgrading schemes. • Replace non-porous surfaces with porous material. • Separate grey water for treatment. Win-win options Reduce the impacts of climate change whilst providing other beneﬁts: • Promotion of riparian management. • Realign ﬂoodplains. • River restoration and natural ﬂood management.

In reality, climate change may effectively be a secondary consideration in many poor countries (particularly in Asia and sub-Saharan Africa), as current unresolved issues regarding limited available resource assessment data, uncertain supply and poor sanitation infrastructure and land use pressures have huge consequences for both public and ecosystem health. Climate change, however, will without doubt worsen an already challenging situation. 23.3 Catchments and Biodiversity Rivers, lakes and wetlands are constantly under pressure from land-based pollution, morphological alteration and loss of connectivity, abstraction and many other anthropogenic impacts. The biodiversity of these systems is, and will continue to be, under considerable pressure, although their global importance is not in question. Fresh water makes up only 0.01% of the world’s water yet supports at least 100,000 species out of approximately 1.8 million – almost 6% of all described species (Hawksworth and Kalin-Arroyo 1995). Globally, freshwater biodiversity is declining at a far greater rate than even the most affected terrestrial ecosystem (Ricciardi and Rasmussen 1999). A key question is how best to develop future conservation strategies for aquatic biodiversity. Traditional terrestrial approaches have been to identify discrete spatial areas of high conservation and for them to be bounded and protected. This is not applicable for aquatic ecosystems which are intimately connected in space and time with their associated landscapes at a catchment scale. A more integrated approach to both terrestrial and aquatic biodiversity management within the context of the catchment is required. An approach which addresses the complex interaction between seasonal ﬂows, riparian corridors and habitats, non-permanent and ephemeral habitats, and use of these by terrestrial, riparian and amphibiotic fauna as well purely aquatic species (Dudgeon et al. 2006). Such a systems approach is a fundamental component of catchment management planning.

The Future for Catchment Management Another potential threat to aquatic biodiversity is through the introduction and/or spread of alien species. Notable impacts include the invasive spread of water hyacinth (Eichhornia crassipes) in African and Asian lakes (Joffe and Cooke 1997), the spread of the European carp (Cyprinus carpio) in Australian waterways (Koehn 2004) and the janitor ﬁsh (Pterygoplichthys disjunctivus) in Laguna de Bay (Chapter 17). Typically many invasive alien species are generalists and can survive under a wider range of conditions than they currently experience or are subjected to. Invasive species traits include: a large native range, abundant in their original range, high genetic variability, broad diet, short generation time and they are frequently associated with humans (Gutiérrez and Reaser 2005). Many invasive species, however, result from aquacultural enterprises, where pen and ﬁsh cage releases can compromise local biodiversity. Additionally, aquaculture can also contribute signiﬁcantly to total nutrient and pollutant loading to aquatic environments and this must be factored into nutrient management protocols Furthermore, cages and pens can occupy large areas of the water body and thus interfere with natural movements and reproduction of the native ﬁsh species (Delos Reyes 1993; Gutiérrez and Reaser 2005). Future reductions in river discharge as a result of increased abstraction and as a consequence of climate change will also potentially decrease freshwater biodiversity. Xenopoulos et al. (2005) highlighted that up to 75% of local ﬁsh biodiversity would be close to extinction by 2070, with a disproportionate burden on ﬁsh populations in developing and poor countries. The challenge of how to control invasive species requires regional perspectives, balancing socio-economic costs against environmental protection of indigenous species. But it is not only in relation to protecting biodiversity that alien species pose a threat. Physical damage to waterways and changes in water quality can have substantive economic considerations for businesses and downstream users. The problems created by alien species globally in both terres-

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trial and aquatic environments is massive, estimated to be some $400 billion per year. The economic impacts of water hyacinth control in Africa are estimated to be as much as $100 million per year (United Nations Environment Programme 2007). In many developed countries large sums of money are being spent on restoration and rehabilitation of historically altered and damaged aquatic ecosystems. In many instances, however, the efﬁcacy of such measures are difﬁcult to evaluate. This is partly because of a lack of a wider national or regional vision and the fact that many activities are piecemeal and poorly evaluated. Projects or initiatives working at reach scales must be integrated with those operating at a riparian, sub-catchment and basin-wide scale. Additionally, emphasis on restoration has tended to focus on deﬁning acceptable environmental ﬂows (mean discharge, extreme high and low ﬂows, seasonality, eposidocity, etc.) for speciﬁc species. The challenge for the future will be to understand the environmental conditions required for multiple species in whole communities and to deﬁne remediation options which take a broader ecological perspective. Although land use practice can be managed for the future protection of aquatic resources, geomorphological alterations in streams and rivers brought about by historical human activity pose a more difﬁcult challenge. Very long time lags are required to naturalize such processes over large spatial scales and even with a very substantive investment, recovery to pre-modiﬁed conditions cannot always be achieved (Allan 2004). The key challenge is to deﬁne what the target for restoration is within the context of a whole catchment plan. Individual measures must be mapped against a systems level analysis to identify complementarities in effort, to reduce duplication and maximize synergies. How much restoration is required and the degree of ecological gain (Poff et al. 2003), must also be considered or tensioned against disproportionate costs and whether management options are socially acceptable to catchment stakeholders.

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Catchments and Land Use

The world’s population (currently at approximately 6 billion) is expected to grow to nearly 9 billion by 2040, placing an increased burden on food supply. Pressure on the agricultural sector to increase production will be substantive and agriculture already represents the dominant global water user (approximately 74%). Around 80% of agricultural water usage comes from ‘green water’ (soil derived) and 20% from ‘blue’ sources such as rivers and lakes. There is grave concern that the current usage of groundwater supplies for agricultural production are not sustainable at present rates. The demand for agricultural produce is also changing, with a general move in consumption patterns away from cereals towards higher caloriﬁc products such as meat, milk and oils. As consumption increases and diet changes, water requirements also change as it takes more water to produce meat than grain. The total demand for food and fodder crops is expected to double in the next 50 years with a concomitant 70% increase in agricultural water use by 2050 (IWMI 2007). The recent Comprehensive Assessment of Water Management in Agriculture (IWMI 2007) highlights some key areas for an emerging policy agenda. The report identiﬁes that improvements in the agricultural use of water will be central to meeting the water supply challenge of the next 50 years. With reference to the agricultural sector in particular, key issues include: how to increase productivity per unit of water required (increase yield without environmental degradation); how to implement a range of agricultural management options; and how to manage agriculture to provide multiple ecosystem services. The greatest potential yield increases are in rain-fed areas, which represent the majority of the world’s harvested cropland. Water management is the key to increasing agricultural productivity and rainwater is considered the key freshwater source in the future. Land management practices and improved soil conservation increase the efﬁciency of rainfed systems and a range of options have been

evaluated (either used singly or in combination). There is also considerable potential for increasing yields by improving irrigation management and practice on already irrigated land, which alone could provide up to 75% of additional food requirements (IWMI 2007). Deforestation at a global scale is a continuing threat to water resources and their availability, and although forests cover approximately a third of the world’s land surface, deforestation has reduced the world’s original forest cover by 50% (WRI 2001). Deforestation has caused adverse impacts on the ﬂow regimes of many rivers, altering ﬂow paths within the soil, reducing inﬁltration and with associated increase in soil erosion and degradation. In Australia, rangeland grazing and overexploitation of coastal land for intensive agriculture has resulted in enhanced sediment and nutrient delivery to the historically nutrient poor lagoon of the Great Barrier Reef (Chapter 15). In Europe, wide conversion of forest and grassland to agricultural areas in the nineteenth and twentieth centuries was accompanied by massive soil erosion. Development of river dams and abstraction of water has, however, reduced sediment discharge to coastal zones. On a global scale, some 25% of the current sediment load from land to the coastal zone is trapped behind reservoirs. In Europe, almost all main rivers are dammed. For example the River Ebro (Northern Spain) delivers only 1% of the solid discharge volume that it did in 1900 and Lake Geneva receives only about 50% of its historical sediment loading. There has been an estimated annual loss of 100 million tonnes of material being delivered to Europe’s catchment–coastal interface and such a sediment deﬁcit can increase coastal vulnerability to climate change impacts such as storm surges and coastal erosion (EEA 2005). Global market ﬂuctuations are a further factor that can induce rapid changes in land use and biomass production. The recent changes in balance between food and biomass production highlight that land use change can occur at a rate which is independent of other medium- to

The Future for Catchment Management longer-term drivers such as climate change, although remains a component part of the issue. Biofuel production has seen a rapid increase in recent years driven by increasing energy prices, concerns about greenhouse gas emissions and future targets for mitigation and political inﬂuences. Global bioethanol production has doubled (1990–2003) and is expected to double again by 2010 (de Fraiture et al. 2008). This highlights the dilemma of balancing food production against that of biomass, especially in areas presently under water stress and with constraints on the amount of available land resource. Market forces will inﬂuence this balance but the global demand for food will continue to increase as will the global requirement for irrigation. A shift in cropping in response to the demand for biofuels can also bring additional problems. The conversion of upland rice and orchards to cassava and maize, for example, as is currently underway in Southeast Asia, can greatly increase runoff and sediment load. The consequence of this enhanced sediment loss on nutrient transfers to the aquatic environment and downstream consequences on biota have not been thoroughly evaluated (Valentin et al. 2008) The debate over the impact of land use change on water resource availability will continue and it places water resources and catchment management in a central role between tradeoffs on food, fuel, environmental protection and human welfare.

23.5 Catchments and Coasts The importance of managing human activity in catchments, in particular agricultural activity, in order to prevent adverse conditions in coastal and shallow sea environments has long been recognized (Lefeuvre and Feunteun 2007). Following an ‘ecosystem approach’, research has demonstrated the link between the functioning of catchments, rivers, estuaries and marine coastal water systems. Research on the exchange and transfers of energy and materials between these systems has proved that the water quality in rivers and in

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the sea, as well as the functioning of both freshwater and marine ecosystems, depends upon the evolution of land cover, land use and land management of the terrestrial catchments that drain to them. For example, it is well documented that there is a close link between the ﬂux of freshwater inputs of nitrogen, mainly derived from agricultural fertilizer inputs, to coastal areas and occurrence of marine algal blooms, anoxia associated with die-off of some benthic animals, impact on ﬁsh stocks and problems with bathing water quality. In this respect, international treaties aimed at addressing water quality issues and others have been established. For example the Convention on the Protection of the Marine Environment of the Baltic Sea Area (HELCOM) has its main goal of protecting the marine environment of the Baltic Sea from all sources of pollution, and to restore and safeguard its ecological balance. The Convention came into force in 1980 covering the whole of the Baltic Sea area, including inland waters as well as the water of the sea itself and the sea-bed. It includes measures undertaken in the whole catchment area of the Baltic Sea to reduce land-based pollution. Similar objectives are the aim of the Commission for the Protection of the Marine Environment of the North-East Atlantic (OSPAR) which came into force in 1992 (although it was initiated in 1972). Like HELCOM, OSPAR has worked to identify the threats to the marine environment, and has organized, across its region, programmes and measures to ensure effective national action to combat them. In doing so, it has pioneered ways of ensuring monitoring and assessment of the quality status of the seas, of setting internationally agreed goals and of checking that the participating governments are delivering what is needed. The OSPAR Convention requires the Contracting Parties to report on what they have done to implement their obligations and commitments, and requires the OSPAR Commission to evaluate what has been achieved. A suite of ﬁve thematic strategies have been developed to address the main threats that it has identiﬁed within its competence: the Biodiversity

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and Ecosystem Strategy; the Eutrophication Strategy; the Hazardous Substances Strategy; the Offshore Industry Strategy; and the Radioactive Substances Strategy. Further examples exist in the form of bodies of sea water which are almost entirely enclosed by land; called enclosed coastal or inland seas. Examples include the Mediterranean Sea, the Gulf of Thailand, the Bo Hai in northern China and the Seto Inland Sea, Tokyo Bay and Ise Bay in Japan. The already mentioned Baltic Sea and North Sea are similarly classiﬁed. In recognition of the need to protect these shallow seas through integrated land and water management, the Environmental Management of Enclosed Coastal Seas (EMECS) concept was initiated in Japan in the early 1990s. This forum recognized that the degradation of freshwater water quality draining to the coastal areas and caused by changes in land use, deforestation, reduction in vegetation coverage and reservoir construction was adversely impacting the shallow sea ecosystem. Very clear examples are the ‘red’ and ‘blue tide’ phenomena associated with eutrophication. The need to manage the catchment to remediate problems and protect against further deterioration of these unique ecosystems is clear. The importance of managing agricultural runoff and associated nutrients is highlighted by two examples which are detailed in this volume: the Chesapeake Bay and the Great Barrier Reef. Both marine systems have deteriorated as a consequence of nutrient and sediment ﬂux from the terrestrial environment. Both are clear examples of the importance of managing freshwater quality through land and agricultural management to protect the ultimate downstream environment: coastal and marine ecosystems.

23.6 Catchments and Ecosystem Goods and Services The Millennium Ecosystem Assessment (MEA) promotes a consideration of environmental goods and services as a unifying concept for both the research community and ecosystem managers

(Daily 1997). An ecosystem approach recognizes that human life depends on the existence of a natural resource base and that nature contributes to the fulﬁlment of human needs (see Fig. 23.1). Given that natural resources are not inﬁnite, how can the relationship between the needs or well-being of humans and the amounts and types of natural resources be described? How, and how much, do changes in the natural environment impact on human well-being? The MEA highlights that there is a global loss of ecosystems services and that this affects the poor disproportionately. It also recognizes that the consequences of the loss of these goods and services are not well understood but the impact and its consequences will differ dependent upon the system concerned. The utility of the concept of ecosystem services in relation to assessing catchment sustainability is appealing. The catchment provides a functional spatial unit that encompasses a hydrological continuum intimately connected with the land and human capital associated with it. It is clear from many of the case studies presented in this volume that an historical perspective of sector by sector management does not provide a robust enough framework for sustainable management of catchment resources. Instead, the focus is shifting to a systems perspective, signalled by the use of the phrase integrated water resource management, or integrated catchment management. While integration is not a new concept in the context of environmental management (Barrow 1998), the term has come to mean a number of things within the water management arena (Wouters et al. 2005). There are four broad elements that could be considered in developing an integrated policy or management process. Firstly, integrated management can be considered in terms of the range of activities governed, often closely associated with ‘comprehensive’ or ‘ecosystem’ approaches (Mitchell 2005), which seek to identify and manage all variables and interrelationships affecting a given water/land system. Secondly, there is also a geographical or territorial element to integration, which considers the

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Fig. 23.1 Interconnections and links between regulating, provisioning, supporting and cultural ecosystems services. (Adapted from Millennium Assessment Report 2005.)

area over which the management system applies. Thirdly, integrated management can also be considered in terms of the range of stakeholders involved in the decision-making process. As with territorial boundaries, the jurisdictional boundaries between the various agencies and institutions can create a ‘silo’ effect (Mitchell 2005) of mismatched, disjointed policies in both land and water management. Finally, integration can be considered in a broader conceptual sense, in terms of drawing together a broad range of objectives in planning and decision-making. In this sense, popular concepts such as IWRM are integrated approaches in that they seek to consider and balance environmental, social and economic objectives (Harris and Hooper 2004). Important questions should be raised, however, about how ‘integrated’ a given planning or decision-making framework should seek to be. Any attempt to implement integrated land and water management should carefully consider these questions. The implication for catchment science

is discussed at the end of the chapter but ﬁrst we address the human element of the catchment system in some more detail.

23.7 Catchments and People Water is fundamental to human well-being. As outlined throughout this book, clean water is essential for human health, through the need for clean drinking water to concerns over the cumulative and long-term impacts of priority harzardous substances that underpins regulatory regimes such as OSPAR (Clarke and King 2006). Dirty water kills nearly two million people per year. Too much water is equally detrimental to human well-being, with over 125,000 people killed, and US$170,000 million of damage caused worldwide between 1992 and 2001 (Clarke and King 2006). The availability and use of water for domestic purposes has a strongly gendered dimension, with those women living in areas without water

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infrastructure spending much of their daily labour on accessing and transporting water, often at the risk of physical or sexual harm. Those beneﬁtting from water infrastructure projects, on the other hand, tend to be powerful males (Davidson and Stratford 2007). Water is also a fundamental resource for the production of food and ﬁbre and an essential input to all forms of industry from heavy manufacturing through to recreation and tourism, but it is increasingly difﬁcult for existing resources to sustain the ever growing demands from newly industrializing countries. In addition water plays an important role in the cultural and spiritual aspect of human lives; water bodies are often sacred places and water use is often inscribed with cultural and historical signiﬁcance. Historically, water acted as the stimulus for urban development and rivers provided an essential transportation network. As a result, rivers and coasts often act as cultural and social conduits as well as physical corridors and/or boundaries. Water continues to have important roles for recreation; both recreation in the sense of the basic human need to have a meaningful life as outlined in the MEA as well as recreation in the sense of the myriad of pursuits available in most afﬂuent and newly industrializing countries. Regardless of the dimension considered, the impacts of changing water quality and quantity are felt disproportionately by the poor, as highlighted by metrics such as the Water Poverty Index (Sullivan et al. 2003). It would be misleading, however, to portray humans as victims of environmental degradation, as human activity provides both the seeds of further destruction and the ability to mitigate problems (Clarke and King 2006). The counter to all the impacts felt by people is the consequence of water use on the water environment through domestic, industrial, power generation and agricultural uses. These pressures, as outlined elsewhere in this chapter, are increasing and conﬂicts over water availablility will be sharpened by the twin drivers of population increase and climate change. Therefore, it is no surprise that many predict that the wars of the twenty second century will

be fought over water rather than oil or other resources. Scarce water resources are inﬂaming political tensions between and within countries, e.g. between Israel and Palestine, and damage to water infrastructure is often used as weapons of war (Allan 2002). For these reasons, as well as the desire to deliver sustainable development through integrated catchment management, there is increasing attention to the governance of water resources (Perret et al. 2006). The notion of governance draws our attention to the partnership between formal government and the engagement of stakeholders and the public in water management. Cross-sectoral integration, through the engagement and involvement of stakeholders, interested parties and citizens, is the only approach through which to address the challenge of ‘hydrosolidarity’ (Falkenmark 2001). The involvment of stakeholders and the public in water management has been advocated for the following reasons: to improve understanding, to increase co-operation and to stimulate environmental citizenship (Blackstock and Richards 2007). Firstly, given the complexity of a catchment system, it is important to involve all those impacting on, and impacted by, water resource management to ensure all aspects have been properly considered from all angles. In other words, no one has all the answers to water management but there can be collective wisdom and social learning through group deliberation (Collins et al. 2007). Secondly, in order to agree and implement management solutions, all stakeholders must participate including the business community. Improving the acceptability of options and uptake of good management practice can be achieved by ensuring transparency of information and decision-making processes, building trust among stakeholders, developing the credibility of public authorities and by incorporating reforms and investments into national economic planning (Duda 2001). Finally, in order to embed sustainable water management practices as part of normal behaviour, it is necessary to engage and work with members of the public in their roles as both consumers

The Future for Catchment Management and citizens; to harness the power of the market and of the ballot box in delivering good water governance. It is worth recognizing that water resource management often requires co-ordination across political, administrative or national boundaries, creating the need for trans-boundary governance. Situations where ecosystem boundaries do not coincide with social and political boundaries raise management challenges, from issues such as trying to align different monitoring regimes through to clashes between different legal systems through to situations of economic or military conﬂict as highlighted above. Transboundary governance, however, can also provide opportunities to learn from one another and to adopt best practice through an entire eco-region. Whilst there are the seeds for conﬂict, setting out river basin stakeholder mechanisms that engage those from multiple countries also provides a mechanism for conﬂict resolution (e.g. in the Mekong or Danube Basins). Water institutions are continuously evolving. The main evolutionary pressures are towards decentralization and devolved power to local levels; private–state partnerships; and globalization. These seemingly contradictory trends are best understood as the need to understand governance and water management at multiple levels from the local to the global. A good example of this is the global IWRM process. Concurrent to the evaluation of the current status of the global initiation of IWRM planning through UN-Water (2008), the Copenhagen Initiative of UN-Water and the Global Water Partnership have prepared a ‘roadmapping’ framework to advance the IWRM process. Appreciative of the importance of local water issues and management conditions on how IWRM principles may be actioned in practice, it supports the development of a unique time-lined implementation plan, or roadmap, for IWRM at national levels. The proposal encourages countries to monitor IWRM progress against a speciﬁc set of indicators on a 3-year cycle. The proposed themes (all taking cognisance of Millennium Development Goals) include:

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• 2009 – Review the extent to which key enabling conditions for IWRM implementation have been addressed. • 2012 – Review the progress of IWRM change processes. • 2015 – Review the extent to which improved water management objectives have been delivered through IWRM and the implementation of the MDGs. Detailed sets of indicators are proposed for all three phases, but the importance of generating enabling conditions for the ﬁrst phase of evaluation is particularly important. Indicators that have been proposed include: 1 Changes in the enabling environment: (a) reviewing and amending policies and laws; (b) making water mainstream in national development policies; (c) allocating appropriate and sustainable funding. 2 Changes in institutional frameworks: (a) establishing cross-sectoral co-ordination; (b) involving stakeholder groups; (c) awareness and mobilization campaigns; (d) decision-making at river basin scale; (e) capacity development. 3 Changes in management instruments: (a) better information management; (b) demand management; (c) social change mechanisms for conﬂict resolution; (d) economic instruments supporting change. A key objective of this initiative is that such roadmaps would provide a benchmark for monitoring the improvement in water resources brought about by management through IWRM planning. Additionally, countries that are at different stages of resource management will be required to build capacity and roadmap development could act as a focal point for knowledge exchange and turn process into outcomes (Copenhagen Initiative 2008). The continued global adoption of IWRM principles is highlighted in the implementation of the EU Water Framework Directive with a similar cyclical evaluation and planning in 25 countries, and through the establishment of initiatives such

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as the Network of Asian River Basin Organizations (NARBO), and the International Network of Basin Organizations (INBO) with similar goals. The contribution of the supporting initiatives such as the UNESCO International Hydrological Programme (IHP), FRIEND (Flow Regimes from International Experimental and Network Data), HELP (Hydrology for the Environment Life and Policy) and PCCP (Potential Conﬂict to Cooperation Potential) to the development of IWRM development has also been substantive. Although the desire to develop a standardized approach to the implementation of IWRM is attractive, it is important to realize that the implementation must be ﬁt for purpose and sit comfortably with regional diversity in physical, economic, social, cultural and ﬁscal characteristics. Countries and regions all have different perspectives, and therefore within the IWRM approach it is important that there is a ‘destination’ as well as a roadmap and that different destinations are appropriate for different conditions and societies. Critical to this is identifying what that destination is and identifying a set of evaluation principals through which to assess success (Biswas 2004). Furthermore, the above steps may not be sufﬁcient to make the hard decisions to be made about the allocation of water resources, often entailing an explicit engagement with the exercise of economic and political power (Lach et al. 2005). Therefore conﬂict identiﬁcation and resolution mechanisms remain an important, but often forgotten, aspect of IWRM and broader environmental governance (Sidaway 2005).

23.8 Catchments and Science Many scientiﬁc challenges and requirements exist before we can claim a complete, or even adequate, understanding of the impact of catchment management activities on the functioning of the catchment itself. In many respects, the lack of scientiﬁc understanding prevents identiﬁcation of the most environmentally effective/appropriate or cost-effective management

approach. Such considerations are central to the WFD process in Europe, and so scientiﬁc and research efforts must be rapidly progressed. Each chapter ends with a summary of the key research requirements but priority areas can be identiﬁed. For hydrological aspects, the issue of non-stationarity must quickly be overcome. The principle of stationarity, that the future characteristics of river ﬂows and runoff can be described and represented by the past historical record of observation, needs to be urgently re-assessed in the light of future climate change. All of our water engineering, water resource assessments and ﬂood protection are based on the principle of stationarity. In terms of channel morphology there is a need for better tools to detect and model the impacts of channel management. These need to provide appropriate linkages between hydromorphology and biota across relevant spatial and temporal scales. In respect of agricultural land management, future research is urgently required to address both the common and unique aspects of ecosystem management at catchment scale across broad ranges of conditions. In particular, the development of the concept of ecological services will further provide a unifying language for ecosystem managers and researchers. There is an emerging need to assess the loadings of oxide nanoparticles to wastewaters, the efﬁciency of removal of oxide nanoparticles in sewage treatment processes and the structure, transport, behaviour and fate of pharmaceutical and cosmetic oxide nanoparticles in surface waters. In addition, there exists a clear need for increased monitoring and evaluation of ﬂuxes from combined, poorly managed and potentially unregulated sewage outfalls. A further key research requirement lies in the evaluation of highly polluted sediments which are held in the sewer system for long periods and released during extreme events. In addition, their impact on the natural environment is far from clear and both acute and chronic effects require closer investigation, again with a view to optimizing the

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The Future for Catchment Management management options. Further research in water quality needs to establish ﬁeld-scale manipulation experiments to determine the nature, scale and timescales of ecological responses to different nutrient remediation strategies and to provide better estimates of ecologically relevant water quality standards for nutrients and suspended solids. The potential contribution to future scientiﬁc understanding from new observational frameworks is immense. New satellite remote sensing platforms will provide real-time information on soil moisture, river discharge and rainfall intensity and spatial extent, for example. Hydrological observatories will provide consistent information on water quantity and quality across wide environmental gradients for modelling and assessment purposes. Novel sensor technologies will provide for real-time monitoring of chemistry providing new understanding of event-based processes and pathways. Scientiﬁc advancement holds the key to improved and more effective catchment management in the future. Increased investment across the board, encompassing manipulation, experimentation, observation and modelling, is urgently needed to ensure that science delivers what is required. Scientiﬁc activities must be fully integrated, for example; linking land use with water quantity and quality; and linking natural and physical sciences with social and economic disciplines. As this comment suggests, excellence in natural and physical sciences is necessary but not sufﬁcient alone to deliver integrated water resource management. Lessons from arenas as diverse as forestry, agriculture, transport and tourism management illustrate that more data and tools alone will not provide solutions to complex human–environmental problems. This is because for action to occur, there needs to be agreement that there is a problem and on the potential solutions (Burgess et al. 2007). Having agreed on solutions, human actors need to have the capacity to implement, monitor and adapt these solutions. Therefore, there needs to be a focus on how stakeholders interpret and use the

outcomes of experimentation, observation and modelling processes. This opens up whole new areas of science to be integrated into the catchment management area; those of political science, sociology, economics, psychology and philosophy, for example. This returns us to the notion of the ecosystem approach, and the notion of practising science within a systems context. Wider trends in science, such as moves to ‘sustainability science’ and the need for science to have a ‘new social contract with society’ (Lubchencho 1998), illustrate the need to integrate science across the human– natural divide but also to work beyond the traditional academe. Increasingly, successful catchment science for catchment management requires us to ‘do science with stakeholders for society’.

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Image facing chapter title page: Courtesy of Alan Ford at the English Wikipedia project.

Index

Page numbers in italics refer to ﬁgures, those in bold refer to tables, those in boxes by the sufﬁx B adaptive cycle of complex systems theory 311–12, 312 adaptive management 112–13, 421 application at multiple scales 112–13 decreasing nutrient loads 112 of GBR water quality outcomes 367–8 overall aims to be achieved by 10–11, 11 Africa countries with greatest reliance on hydropower 267 potential of hydropower development 267 southern, Protocol on Shared Watercourse systems 13 agricultural catchments, management to sustain production and water quality 107–28 conﬂicting political strategies often compete 107 current consequences 110–14 adaptive management 112–13 changing paradigm of catchment management 113–14 concerns associated with agriculture 111 scale of management 111–12 what constitutes impairment and baseline conditions 110 eutrophication risk in rivers 107–8 future research and operational needs 126–8 consumer-driven programmes and education 126–7 ecological services 126, 126 ecosystem services and constituents of well-being relationship 126, 127

lag time, BMP implementation and water quality improvement 128 major problems associated with ecosystem management 126 planning for controls on farming system nutrient inputs 127 high risk land use practices 108 historical background 109, 110 catchments now operate in a global market place 109 post-WW2 agricultural improvements 109 rapid growth and intensiﬁcation of the animal industry 109 new operational insights 119–26 BMP implementation level affects nutrient loss reduction 121B implementing remedial strategies within catchments 122–6 nutrient trading and agricultural catchments 121–2 potential barriers, uptake of catchment management options 119–20, 120 potential BMP tradeoffs within agricultural catchments 120–1 new scientiﬁc insights 114–19 assessing catchment processes and conditions 114–15 managing complex ecosystem processes in catchments 115, 119 managing connectivity and complexity 113 non-point source nutrients, primary source of concern 107 nutrient-related water quality issues 108

point sources of nutrients easy to reduce 107 see also eutrophication agricultural production, shifts can affect water quality 121B drive for biofuel production 121B potential for P and N loss increased 121B alien species 400–1, 505 Alta River see Norway Amazon Basin, seasonal ﬂooding 20 Angola and Okavango River basin 457, 463, 465 Annapolis Protocol 201–2 based on certain observations and principles 187, 192, 201–2 aquatic ecosystems, interconnectivity between 6 aquatic environments 5, 253 aquatic organisms, and nanoparticles 151 Asian wetlands, many under threat 37 Australia 506 Bush Tender programme 245 effects of agricultural water markets 222 irrigation water not used efﬁciently 223 major socioeconomic functions of irrigation systems 223 National Plan for Water Security (2007) 212, 223 National Water Initiative 13 targeted investments and cost effective elimination of inefﬁciencies 223–4 traditional irrigation management approaches challenged 212

Index variability in rainfall and runoff a problem 13 water markets, impact on salinity and river systems 221, 222 water scarcity issues 230 see also Great Barrier Reef, managing the catchment Ayrshire bathing waters, faecal indicator ﬂux impacts 200 bacterial pollution 183, 184 basin management principles 7 Bathing Water Directive (EU) 184, 200 and Annapolis principles of real-time prediction 202 Belgium, ﬂood management 72 The Belgian perspective 74B Best Management Practices (BMPs) 107, 111, 113, 167–70 concept of sustainable urban drainage systems (SUDS, WSUDS) 168–9 SUDS triangle concept 168–9 control and treatment of urban stormwater runoff 168 multi-chambered treatment trains 177 ‘treatment or management’ train 168, 168, 177 early process-based deterministic modelling developments underpinning BMP design 196–200 inﬁltration and porous surfacing 177 new developments, interest in LID approaches 169 non-structural (source) control BMPs 167–8 LID approaches 169, 169 non-structural approaches can be more effective 169 principles of BMP and associated LID construction 169–70 basic guidance manual 170 stormwater measures integrated into built form 176 should also address non-structural measures 176 structural or treatment BMPs 168 see also Chesapeake Bay catchment management bioaccumulation 393 biodiversity 94 affected by reservoir building 259 catchments and biodiversity 504–6 freshwater 4, 504 Great Barrier Reef 351, 355

in the Okavango delta 461–2, 461 vulnerable in farming landscapes 35–6 biodiversity conservation 36 biofuel production 109, 121B, 506–7 biogeochemical cycles, major, undergone serious perturbation 4–5 bog 22B Botswana discussions on management plan for the Okavango Delta 466, 466B and the Okavango Delta 457, 463 situation complicated by discovery of diamonds 464 water problems 464 Caldew catchment, Cumbria, FIO export during period of de-stocking 200 Campylobacter 187 catchment approach, drivers for 15 catchment management 513 20th century 42 changing the paradigm of 113–14 combination of BMPs required 113 desire to identify win:win options 15 different visions of meaning 13, 14 for hydropower generation 253–84 implication of efﬂuent management 148–50 integrating wetlands within 41–2 learning from the past 13–14 and pollutant loadings 183 over-fertilization of nearshore marine systems 183 pollutants of interest 183 scale of 111–12 nitrogen losses 111 for P, measures targeted at critical source areas 111 twin track to controlling nutrient loss considered 111–12, 112 scientiﬁc understandings 14 some unfortunate distribution impacts 42 transboundary catchment management 381–2 and wetland functions 29–30 R. Cherwell, embankment removal 29–30, 30 reduction in nutrients through sedimentation 30 see also Chesapeake Bay catchment management catchment management concept 1–15 current solutions 6–13 international efforts 6 historical perspective 2–6

517 synthesizing knowledge for the future 13–15 catchment management, the future 501–13 allocation of water resources, social and political matter 501 the catchment as a concept in water resource planning 501 catchments and biodiversity 504–6 alien species, another threat to aquatic biodiversity 505 aquatic biodiversity, and future conservation strategies 504 changes in balance between food and biomass production 506–7 decrease in freshwater biodiversity 505 freshwater biodiversity declining globally 504 problems with aquacultural enterprises 505 restoration and rehabilitation 505 catchments and climate change 502–4 big changes in extremes expected 503 changes to the large-scale hydrological cycle 502 current responses to variability 503 effects of changing world climate 502 investment risk analysis supporting large-scale donor plans 503 managing risk and uncertainty associated with climate change 504, 504B need for new data and information to assess changes 503 need for new vulnerability assessment indicators 503–4 principle of stationarity undermined 503 probable decline in water quality 503 catchments and coasts 507–8 enclosed coastal or inland seas 508 importance of managing human activities in catchments 507 international treaties addressing water quality and other issues 507 catchments and ecosystems goods and services 508–9 development of an integrated policy or management process 508–9 Millennium Ecosystem Assessment 508

518 catchment management (cont'd) catchments and land use 506–7 deforestation and soil erosion 506 key issues in the agricultural sector 506 world population growth creating pressures 506 catchments and people 509–12 clean water essential for human health 509 gendered dimension of domestic water 509–10 involvement of stakeholders and public in water management 501 wars of the future over water 510 water a fundamental resource for food and industry 510 Water Poverty Index 510 water resources management 510–11 catchments and science 512–13 better tools to detect and model channel management impacts 512 integration of science across the human–natural divide 509, 513 need to assess loading of oxide nanoparticles to wastewaters 512 potential of new remote sensing satellites 513 problem of lack of scientiﬁc understanding 512 scientiﬁc advancement and effective catchment management 513 stationarity principle needs reassessment 512 effect of human activities 502 impacted by humans throughout history 501 need for better synergy of climate change, energy and land use policy 502 catchment plans, wetland restoration and management key element 42–3 catchment studies, informed by developments in microbial source tracking 200 application of latest source tracking tools 200 catchment systems 1 emergent properties and uncertainty in 1–2, 2

Index innovative regulatory approaches driving management of microbial exposures 201–2 timeline of developing anthropogenic pressures on water 2, 3 catchment to coast systems 183–203 current perspectives 186–9 contributors of FIOs provide different pathogen loadings 187 discharge information availability 189 episodic nature of FIO pollution in catchment systems 187 faecal indicator organisms (FIOs) 186 human point sources of pollution 187 interpretations of FIO data, simplistic assumptions 186–7 QMSA studies 188–9, 191, 192, 191, 193 real-time information rarely available 188 future research 202–3 achievements in catchment microbial dynamics 203 catchment outputs become coastal inputs 202 microbial pollution an emerging area of investigation 202–3 historical background 184–6 catchment-based legislation in EU and USA 184 emergent catchment paradigm, new requirements 186 key difference between nutrient and microbiological parameters 184 maintenance of ambient water quality criteria needs IRBM 185 OSPAR Convention and international directives 184 point source efﬂuent quality regulation, change away from 184–5 microbiological ﬂuxes to nearshore waters, derivation of 183–4 bathing and ingestion of pathogens 184 microbial contamination 183 new scientiﬁc insights 189, 191–202 calibration for large catchment models of FIO ﬂux, Ribble system 195, 197, 198 catchment export coefﬁcients related to land use 194–5, 195, 196

current developments in microbial source tracking 200–1 drawback of catchment modelling approach 196 early process-based deterministic modelling developments underpinning BMP design 196–200 innovative regulatory approaches driving management of microbial exposures in catchment systems 201–2 quantifying microbial ﬂuxes impacting recreational or shellﬁsh waters 185, 189, 191, 194 Scotland, screening tool, smart dynamic export coefﬁcient approach 194 catchments assessing processes and condition 114–15 biotic indicators for assessing waterbody condition 114–15 Wadeable Streams Assessment, USA 114, 116–18 dominated by AFOs, potential for N and P surpluses 109 farming systems and nutrient budgets 110 features changed and manipulated by human activity 1 and the Great Barrier Reef 355–8 implementing remedial strategies 122–6 catchment alliances 125–6 collaborative stakeholder engagement 125–6, 125 increased remediation costs/ complexity as required nutrient reduction increases 124 models may bring false conﬁdence and misconceptions 122–3 non-point source catchment models 123B remedial measures, appropriate targeting needed 123 stakeholder alliances 124–5, 124B uncertainty in model computation 123–4 use of models for evaluating management alternatives 122–3 managing complex ecosystem processes in 115, 119 Causal Analysis/Diagnosis Decision Information (CADDIS – USEPA) 115, 119

Index managing connectivity and complexity 114, 115 complex interactions with indiscrete thresholds and multiple variables 114 River Continuum Concept 114 potential BMP tradeoffs in 120–1 conservation tillage to reduce P surface runoff loss 120–1 problems of separate strategies for N and P 120 chemical risk assessment 147 current BMP deﬁnitions, efﬁciencies and assumptions, overestimation of implementation 421–2 modelling approaches 147 Chesapeake Bay catchment management 407–25 anthropogenic N and P loads, much increased 408, 409 atmospheric deposition important source of N 408 catchment size 407, 408 Chesapeake Bay Water Quality Model (estuarine-based) 415 Chesapeake Bay Watershed Model (WSM) 415 calibration 415 continues to be a critical tool 416 implementation assumptions 421–2 current BMP deﬁnitions, efﬁciencies and assumptions 417–22 2003, selective evaluations and efﬁciency revisions 420–1 BMP crediting system, evaluation shows strengths and weaknesses 418, 420 BMPs may applied in three ways 417–18, 418 changes must be made regularly 421 current BMPs and those awaiting approval 418, 419 increasing conﬁdence in BMP effectiveness 421 optimistic BMP reduction estimates 420 probable sources of error in estimating BMP impacts 420 decline in annual P and N loads to the bay 410–11 goal setting and nutrient reduction load allocations 411–13 1987 goals considered essential 412 apparent discrepancies in data not fully understood 412

largest and productive USA estuary 407 2000 Agreement commitment to remove all nutrient impairments 413 allocation of nutrient caps or nutrient reductions 413 nutrient and sediment goals and the Clean Water Act 413 modelling and monitoring to support watershed restoration 414–17 caution when using models to estimate restoration progress 415 lag time assumptions 415 modelling allows for development of ‘what if’ scenarios 414–15 uncertainties about BMPS 416 nutrient loading approach 411, 412 nutrient over-enrichment (eutrophication) 407 results of 407–8 observations and lessons learnt 423–5 critical to minimize growth in nutrient load 424–5 development of control measures for nutrient reduction efﬁciencies 424 greater reduction of nutrient loads will be needed 424 monitoring for achieving nutrient reductions 424 quantitative nutrient reduction goals or loading caps 423 problem of population growth 408 research to ﬁnd cause of accelerated decline 407, 408 systems approaches to conservation and nutrient balances 422–3 application of BMPs must address complete production system 423 BMP practices/systems frequently improperly implemented 422 cost-sharing of BMP/farm system implementation 422–3 creation of farm nutrient imbalances 423 efforts to avoid pollution abatement practices 422 tributary strategies 413–14 estimates of effectiveness developed 413 reduction efﬁciency estimates 413–14 reduction in model estimated progress reported in 2003 414, 414

519 strategies required implementation of control measures 413 wastewater discharges cut 408 Chesapeake Bay Program, history of 408–11 1992 amendment 409 2003, nutrient loading limits 410, 411 ‘Bay Progress Overstated’, article 415 CBP partners signed Chesapeake 2000 410 implementation of extensive river input and tidal-monitoring program 416 provision of important trend information 416, 417 shows signiﬁcant reductions in N and P concentration trends 416, 417 value of long-term monitoring data 416–17 Maryland, development of appropriate agricultural strategies 409–10 methods for tracking and reporting consistent 421 third party review should be considered 421 reduction in sediment loads agreed 410 relies on one large watershed model (HSPF model) 415 HSPF model continually reﬁned and updated 415 relies on two large coupled models simulating conditions in watershed and bay 415 second agreement goal, reduction in nutrients entering the bay 409 other goals 409 signing of ﬁrst Chesapeake Bay Agreement 408–9 success of Agricultural Tributary Teams 410 Chesapeake Bay Restoration Program, modelling and monitoring essential for 417 cities, bringing water to and connecting people 236–9 19th C. conﬂict between companies, customers and municipalities 237 debates about tapwater quality 237 government intervention allowed adequate water transfer from distant sources 238–9

520 cities (cont'd) increasingly municipalities chose to run own companies 237–8 Glasgow 238B national governments eventually took responsibility for water provision 237 newcomers requested free water of good quality 236 use of long-distance aqueducts 237 climate change 43, 175–6 future, impacts on efﬂuent management 150 impact on wetlands 25–6 attributes and performance of particular species affected 25–6 direct effects 26 mediated indirectly 26 impacts of production, design and running of hydropower due to 279–82 increased role of mountain area storage capacity 281 IPCC assessment 279, 281, 282 projected future problems 175 and rise in damage costs likely 176 and rising sea levels, possible effects in the Netherlands 71 UK Foresight Review of ﬂood defence 175 expectation of increases in headwater ﬂooding 175 inadequate capacity of surface water drains 175 loss of inﬁltration capacity 175 scenarios for future ﬂood risk and CSO spillage 175 uncertainties 6 will directly affect functioning of rivers, pools and wetland habitat 26 climate change impacts 150 challenges in context of Integrated Water Resources Management (IWRM) 150 Common Agricultural Policy, revised 113 funding of farmers and foresters 33–4 Comprehensive Assessment of Water Management in Agriculture 506 conservation, socioeconomic barriers to adoption 114B Convention on Biological Diversity 32 ecosystem management approach 32

Index Convention on the Law of the Nonnavigational Uses of International Waterways (UN) 470 Convention on the Protection of the Marine Environment of the Baltic Sea Area (HELCOM) 507 Convention for the Protection of the Marine Environment-East Atlantic (OSPAR) 184, 507–8 crop choices and soil salinity 216–7, 218 water policies 217 crown-of-thorns starﬁsh outbreaks 356–7B cyclic outbreaks 356B nutrient-enriched conditions favour larger phyto-plankton 356–7B few option for managing outbreaks 358B two views on causes of outbreaks 356B Cryptosporidium 187 dams 41–2, 262, 378, 383 beneﬁts but damage caused 37 ﬂood releases now allowed 37 ﬂoodplains and deltaic wetlands downstream degraded 35 should be operated to restore and maintain downstream wetlands 42 Danube River Basin 287–90, 324 application of WDF to 292–3 diversity of habitats 289 ﬂood prevention and management 296–7 development of basin-wide Flood Action Programme 296–7 ﬂooding in 296B climate change prediction driving force for preventative action 296B Lower Danube Basin 288–9 major socio-economic difference between countries 290 Middle Danube Basin 288 population within 289, 289 Sava River and basin 289 signiﬁcant water management issues 295–6 groundwater monitoring network described 295 measures to address problems 295 pollution sources, point and diffuse (MONERIS) 294B risk classiﬁcation of main stem 294, 295

Signiﬁcant Water Measurement Issues (draft) 295 surface water monitoring network described 295 Tisza Basin 289 Transnational Monitoring Network 288B main objectives 288B selection criteria for monitoring locations 288B Upper Danube Basin 288 see also International Commission for the Protection of the Danube River (ICPDR) Danube River Basin District coverage 290 deﬁned by the ICPDR 290, 291 Danube River, most international river basin 287–99 applying the WDF to an international basin 292–3 Co-operative Management of the ICPDR and its history 290–2 co-operation in response to deteriorating river water quality 290 critical basin management problems 295–7 developing a programme of measures 297 ﬂood prevention and management 296–7 ﬂooding in the Danube Basin 296B signiﬁcant water management issues 295–6 Danube Day – involving the public 298B, 299 Danube Art Masters Competition 299 tool for enhancing ‘Danubian identity’ 298B future developments 297–9 actions to reduce pollution 298 basin-wide modelling of pollution 298 Joint Programme of Measures 297 stakeholder involvement 298 pollution sources, point and diffuse (Moneris model) 294B, 298 Danube River Protection Convention (DRPC) 12–13, 287 legal instrument for co-operation on transboundary water management 12–13 main elements 287

Index main objective, sustainable and equitable use of surface and ground waters 290 data and monitoring needs 178–9 data bases identifying long-term cumulative impacts of urbanization 179 increased data effort to quantify new priority chemicals 178–9 national harmonized data inventories 178 deltas Peace-Athabasca Delta 486–92 vulnerable to sea level rise 51 see also Indus delta; Okavango Delta Management Plan (ODMP) Denmark, new watercourse act (1982 revised 1993) 95 desalination, in water-scarce countries with coastlines 244 diffuse pollution 184, 294B accepted deﬁnition 5 characterized 5–6 constituents 5 diffuse urban pollution 171, 172 Drinking Water Directive (EU) 240 drinking water standards 108 criteria now stands at 63(EU) and 84(USA) 242–3 stricter, compliance with 242 Dublin Principles 6, 7B, 324B supported by Second Water Forum in The Hague 6–7 ecological integrity concept 164 ecological restoration 40–1 landscape-scale restoration schemes, justiﬁcation for 40 schemes in the Netherlands and in E. England 41 ecological risk, determination of 147–8 dose-response relationships 147–8 importance of ecological indicators of system function 148 ecological services 126, 126 ecological status, reﬂects biological and chemical status and hydromorphological characteristics of water bodies 140 ecosystems, integrity of in sustainable water management 42 efﬂuent management 135–52 current solutions 140–4 controlling point source discharge to rivers 140

discharge consent calculation 140B environmental quality standards (EQSs) 141, 142 Europe, historical dichotomy in approach to pollution control 140 goal for managing urban drainage, system integration 141 improved ‘end-of-pipe’ solutions (MFR or MTFR) 143 oestrogenic chemicals in surface water 143, 144B operation of current urban wastewater systems 141 pollution sources, framework for action at EU level 140–1 priority substances identiﬁed 142 sewage treatment in the UK 143 use of BATNIEC an economic compromise 143 future research requirement 150–2 advanced ‘end of pipe’ technologies 151–2 climate change impacts 150 impacts on historically immobilized material 152 nanoparticles 150–1 stress ecology 151 urban storm damage 151 historical perspective 135–9 combined sewer overﬂows (CSOs) 136 deoxygenation and reaeration 136, 137 early UK pollution control legislation 137, 138B EC directives impacting on discharge control 137, 139B major problems of CSOs 136, 136 organic matter content of natural water 135–6 point source pollution legislative control 138–9 pollution control in receiving waters 136 UK, deﬁnition of water quality objectives 138 UK, setting of discharge limits 137 urban areas combine surface drainage systems and sewerage systems 136 implications for catchment management 148–50 EA modern regulatory model 149B endocrine disruption in ﬁsh, reduction or elimination 148

521 EPA support for some nutrient trading 149 modern approaches to water quality regulation proposed 148 monitoring to gauge water quality impacts of BMPs 149 OPRA scheme in England and Wales 149–50 pollutant trading favoured by USA 148–9 recent research 148 increased population and industrial activity and increased pollution 135 new scientiﬁc insights 144–8 assessment of chemical risk 147 boron used to characterize P 145, 145 determination of ecological risk 147–8 determination of water quality in large catchments 144 measures taken to reduce P inputs to UK rivers 145 reduction of gross pollution from urban areas and CSOs 146–7 total maximum daily loads (TMDLs) 145–6 tendency to require tighter pollution source control 135 ‘end of pipe’ technologies, advanced 143, 151–2 development of catchment chemical risk maps needed 151–2 possible treatment of endocrine disruptors 151 removal techniques under consideration 151 environmental engineering 243–8 cheaper and safer to give more space to rivers 244 encompasses IRBM 244 land use control policies 244–5 adoption of Nitrates from Agriculture Directive 244 cooperative agreement at local level 244 and water conservation policies 243–4 Environmental Fluid Dynamics Code (EFDC), model based on 199 Environmental Impact Assessment (EIA) 35 environmental quality standards (EQSs) 141, 142

522 EU efforts to harmonize Renewable Energy Sources Directive with the WFD 274 electricity from hydropower, avoidance of discrimination 283 energy targets 274, 274 established nitrate vulnerability zones 108 hydropower as a key resource 274 predicted energy shortfall 282, 283 reduction of CO2 emissions 274 scenario to secure energy demands to 2050 284 slow deregulation of electricity markets 284 eutrophication 4–5, 108B accelerated by increased inputs of P 108B association with paralytic/diarrhoeic shellﬁsh poisoning 183 effects on catchment planning strategies 108B in European rivers, P considered as limiting nutrient 144–5 freshwater, often driven by phosphorus availability 5 and hypoxia in shallow seas 4–5 periodic and massive algal blooms 108B restricts water use 108B farm-scale management tool 213–15, 214B deﬁnition of private property rights within common pool issue 215 need for farm-scale net recharge analysis 213 SWAGMAN* Farm model 213–15, 214B farmers and agri-environment schemes 35–6 subsidies and incentives for 35 fen 22B Finland, funding of city water services 236B FIO budgets, catchment-scale, land use impacts on 199 FIOs (faecal indicator organisms) 194, 200 from sewage treatment 186–9 tools available for real-time predictions in environmental waters 202

Index UK, concentrations in high- and lowﬂow efﬂuents reported 189, 191, 192 ﬁsh alternative body shapes may occur 90 density and biomass negatively impacted by channelization 91 efﬂuents and endocrine disruption in 144B, 148 fauna still impoverished due to loss of habitat 91 freshwater species, more pronounced migratory behaviour in spawning 90 prominent component of running water ecosystems 90 reﬂect a change of environmental conditions 90 salmonid spawning habitat 90 salmonids, adaptation to life in running water 90 migratory salmonids, an anadromous life cycle 90 ﬁsh mortality, in Ruhr R. 454 ﬂood management 51–74 Belgian perspective 74B objectives of ‘Plan Pluie’ 74B conclusions 72–4 current solutions 48–66 American approach to ﬂood management 62–5 ancient solutions 58–9 answer to increased design discharge 59 land use changes 65–6 more focus on nature and nature development 58 Europe, increasing efforts to manage ﬂoods 73–4 has become a major issue 51 historical perspective, the Netherlands 55–8 the 1953 storm, a major wake-up call 56 design discharge 57, 57 Dutch organization 57–8 exceedance possibilities established 56–7 exceeding probabilities and ﬂooding probabilities 56 exceeding probability of a critical water level (design water level) determined 56 ﬂood risks related to ﬂood probabilities 56

height of dike crests up to early 1950s 55–6 inventory of all possible plans made, choice of measures 60–1 is replacing ﬂood protection 72 new developments in approaches to 72 new scientiﬁc insight 66–71 differentiate risk levels between dike sections, system fairness 69 ﬂood probabilities—the FLORIS study 68B introduction of other sources of uncertainty 67 lack of knowledge an uncertainty 66–7 natural variability, an uncertainty 66 new dikes, extra height for uncertainties 66 opposition to returning land to rivers or sea 70 options studied for increase of design discharge 69 organizational measures found to be most successful 69, 69 reducing ﬂood risk 68, 68B reduction of risk considered against investments 68 relationship between design water levels and actual dike height 66 the safety-chain, elements of 70–1 taking into account lower probabilities than those of a risk approach 69–70 transition from exceedance probabilities to ﬂooding probability can be made 67, 67 New Vásárhelyi Plan, Hungary 72–3B Space for the River, a Key Planning Procedure 59–62, 65 the design table, a new design instrument 61–2, 62 environmental assessment on initial plans 59 experiments with adaptive building 59, 60 ﬂoodplain, activities not connected to river activities forbidden 59 four different phases 59, 59 increase in design discharge expected 60 the planning kit, a software system 61 rapid dike reinforcement, 1995–2001 60

Index synthesis 71–2 absolute safety does not exist 71 climate change and sea level rise, problem in the rivers 71 ﬂood awareness vital 71 innovative solutions and maybe a paradigm shift in behaviour 71–2 Netherlands, future protection will cost more and more money 71 ﬂood risk management plans, catchment-based 174 ﬂoodplain wetlands, lower Senegal R., social economic and ecological importance of 31 ﬂoodplains 35, 305, 462 ﬂood tolerance of biota 23 inundation and exposure often a seasonal cycle 23–4 Senegal River 381, 383, 383 and water meadows 24–5 watermeadows as an engineered system 24–5 water regime determined by over-bank ﬂow 23 Floods Directive (EU) 174, 283, 436 ﬂow and drag forces 84B calculation of Reynold’s number 84B ﬂow laminar or turbulent 84B form drag most important 84B France 247 agri-environment measures programme 245 declining wetlands 37 see also Paris freshwater biodiversity at risk from anthropogenic activity 4 Germany, problems of post-uniﬁcation collapse in water consumption 247 Giardia 187 Glasgow, direct municipal provision of water 238B Glasgow Strategic Drainage Plan 427–37 city suffers severe intra-urban ﬂooding 427 drainage masterplanning 428–33 amalgamated network modelling used 429 asset-based modelling failed to consider overland ﬂow 429 core of masterplan 429 four STWS situated within the urban area 429, 430

generation of additional environmental and community beneﬁts possible 433 an integrated sewer/watercourse modelling approach 429 local retroﬁtting solution for Light Burn subcatchment 432, 432 Metropolitan Strategic Drainage Partnership (MGSDP) 428–9 MGDSP, processed range of measures since 2002 433 number of stakeholders needed a joint holistic strategy 328 separation solution 430, 431 Stage 3 of the GSDP 433 staged approach agreed and adopted 429, 431 sub-catchment modelling of the Light Burn 429–30 Surface Water Management Plans (SWMPs) 433 widespread local adoption and retroﬁtting of BMP source controls 432–3 Glasgow Strategic Drainage Plan, provisions 428 key to success 428 Stage 2 still ongoing 430, 432 problems with drainage infrastructure 427 R. Clyde hydronamic and water quality modelling 433 serious water quality problems 427 in Stage 2 development 430, 432 some key issues 433–5 Dalmarnock East End ﬂood review, lessons learned from 434 delivery of the GSDP consistent with wider plans 434 drainage problems wider regional and national signiﬁcance 434 effective management linkages between central and local planning processes 435 ensuring continuation of private sector participation 434 funding of the GSDP, problems? 434 problem of ingrained ‘hard’ engineering culture 435 RBM planning (WDF timetable) 434–5 stakeholder reservations about failure of inﬁltrative SUDS 435

523 strategic GSDP of sustainable drainage incorporates SUDS 435 support systems 435–7 Area Flood Risk Management Plans 436 decision support approaches tested within the Gadburn area 435–6 ﬁnal strategic drainage development plan 437 major potential driver for adoption of BMP/SUDS solutions 436 managing surface stormwater key to future sustainable urban drainage 437 need for micro-management of new development 437 ofﬁcial design guidance manual and other literature 436 problems of including BMP technologies on brownﬁeld sites 437 questions regarding overall responsibilities and liabilities for BMP/SUDS drainage infrastructure 436 vulnerability to future climate change predicted 427 global energy generation increased 267, 268 share of hydropower (2002) 267, 268 global food system expansion of irrigated agriculture 211 gains in production 211 global potential electricity production from hydropower 267 Good Ecological Status (GES) 163–4 Great Barrier Reef, managing the catchment 351–70 abundant biodiversity and great extent 351 adaptive management of GBR water quality outcomes 367–8 facilitates ‘learning by doing’ 367 indicators of change at smaller scales 368 passive adaptive approach in use 368 risks with passive management approach 368 adjoining catchments 351 applied management pratices 360–2 ‘6 Easy Steps’ 360 fertilizer application management, intensive cropping 360–1 Grazing Land Management, regionally speciﬁc BMPs for 361

524 Great Barrier Reef (cont'd) herbicide management 361, 362 major inﬂuences on water quality system 361–2 N-replacement system 361 Tully catchment, lag in system response to management changes 362 approaches to targets 366–7 application of catchment models 367 future investment into management hotspots 367 high levels of uncertainty 367 regional targets for adoption of new management practices 366 water quality targets 366, 366 catchment water quality issues management 369 catchments and the GBR 355–8 catchment freshwater health 355 declining water quality, potential impact on lagoon ecosystems 255 detection of barium, yttrium and manganese in coral cores 355 high biodiversity 355 increase in sediment and nutrient delivery into the GBR 355 increasing amounts of pesticides in the GBR 355 interdependence of catchment and marine ecosystems 358 limited understanding of catchment to marine environment journey of land-based materials 355 loss of freshwater wetlands in developed catchments 355 need for co-operative management approach 358 risks to marine organisms in the GBR 356, 356–7b river ﬂood discharge plume 355–6 water quality changes in GBR waters 355 future challenges for partnership of catchment managers 368–9 integrated approach including relationships 369 RWQP, example of collaborative arrangement 368, 369 science integration key to informing management decisions 39 institutional complexities 364–5

Index adoption of improved management practice, support mechanisms 364–5 current arrangements rely on decisions at GBR-wide scale 365 inconsistencies in land use management 364 Regional NRM plans 364 terrestrial and marine systems, connection of management 364 institutional roles and scales 365–6 intervention priorities 359 identiﬁcation of sources of contaminated terrestrial runoff 359 management of nutrient losses 359 sediment control and erosion reduction associated with rangeland grazing 353, 359 land management systems at the enterprise level 362–4 catchment challenge, reduce diffuse pollution 362 inﬂuences on management practice 563, 563 key concept, use of industry BMPs 364 signiﬁcant indirect constraints 362 largest World Heritage area 351 linking regional NRM, Reef Plan and Reef Rescue Program 370 major stresses 351 management systems 359–60 early development of land management practices 360 growing interest in best management practice 360 interest in conservative land management practices 360 sugar cane industry, concept of Farm Management System (FMS) 360 sugar trash used as green blanket for ratoon crops 360 nature of the system 352–5 catchment population 352, 354 climates 352 industry/activities, large portion of regional economy 354–5 major river catchments 352, 353, 354 strong economic/social disincentives against sufﬁcient change 363–4 topography 352

tourism a major economic contributor 354–5 Tully-Murray catchment analysis 363 varied agricultural systems 352, 354 water quality, differences between wet- and dry-catchment rivers 352 plans and actions 358–9 Reef Plan 358, 358–9, 364, 368–9 development of Water Quality Improvement Plans (WQIPs) 359 government investment in 368 innovative policy approach, lack of detail 365 role of RWQP in implementation 365–6 Reef Rescue Program 351–2, 358, 359, 364 provision of funding 365 Reef Quality Report Card 359 Research and Development programme 369 Reef Water Quality Partnership (RWQP) 365 Reef Water Quality Protection Plan 351–2 regional Natural Resource Management 352 Greece, wetland losses 37 green and blue water, relationships between 2 groundwater management zones 215–17, 216 maximum allowable groundwater recharge quota 215–17, 216 pricing of recharge above leaching requirements can be an option 215 southern Murray-Darling Basin, modelling results 215 analysis of groundwater dynamics 215 division into discharge or recharge zones 215, 216 sustainable resource management 215–16 supported by modelling evidence from the Philippines 217B whole-of-catchment water and salt balance model, uses of 215 habitat concept 79B biotope, functional habitat or mesohabitat 79B original deﬁnition 79B habitat suitability indices (HSIs) 96

Index Habitats Directive (EU) 324, 345–6 Article 6, plans may or may not be allowed 35, 36 and Birds Directive 32–3 list habitats and species designated for protection 44 objectives at national and site level 33, 33 Hungary increasing protection level along the Tisza R. 72–3 constructing retention polders easy 73 New Vásárhelyi Plan 72–3B storage of ﬂood water in retention reservoirs 72B hydro-ecology major assessment, Niger R., implications of climate change 39 in the WFD 15 hydrological processes, characterizing catchment systems 1–2 hydrology and water resources, SE England 329–32 Basingstoke, problems with sewage treatment 332B environmental variables, records of 329 ﬂood generation and ﬂood forecasts 331–2 monitoring of water quality 329, 331 R. Thames longest river 329, 331 river pollution surveys 331, 332, 332 sewage treatment works, treatment capacity 332 Thames Basin, principal aquifer 329, 330 water problems facing the region 331 wetlands and wetland processes 19 hydropower importance of 267–8 key role as peak demand regulator 268 installed capacity mainly in Europe and the USA 267 hydropower generation, managing catchments for 252–84 current solutions 267–78 in the context of water resources management: the EU Water Framework Directive 276–7 the resource base 267–73

historical perspective 259–67 implications 278–84 assessment of hydropower design in WFD management plans 278–9 increased need to co-ordinate licensing 282–4 new scientiﬁc insights and future research requirements 178 areas for future development 278 more knowledge needed 278 water and energy ecological impacts on the aquatic environment 253 production of electricity 253 water power generation technologies from the ocean 253 water and thermal power 253 hydropower history 259–61 earliest hydroelectric plant 260 falling water used by Greeks for waterwheels 260 plant design fairly standardized after WW1 260 decline in dam building 260 in the USA 260 water power important in the Industrial Revolution 260 hydropower, how does it work 253–7 by head, setting type of hydraulic turbine to be used 254–5 high-head power projects 254–5, 255 low-head power projects 255 run-of-river power stations 254 by purpose (single or multipurpose) 256, 257 for irrigation 256 prevention/mitigation of ﬂoods or droughts 256 by size 256–7, 257 by storage capacity 255–6 pumped storage plants 255–6, 256 reservoir projects 255 run-of-river plants 255 converts natural ﬂow of water into electricity 253–4 head of water determines potential energy 253–4 working principles 254, 254 hydropower impacts 257–9 advantages and disadvantages of hydropower 257, 258 building modiﬁes natural surroundings 259 Canada, installed hydro capacity, no involuntary resettlement 259

525 impact assessment 257 negative effects born by local communities 258–9 environmental impacts 258–9 social impacts 259 powerlines affect bird populations 259 sustainable hydropower development in developing countries 259 Three Georges Project 259, 268, 269–70B hydropower, outlook for, climate change mitigation and adaptation 274–5 future climate change and hydropower production 274 future potential of renewable energy resources 274 hydropower with storage capacity more attractive 274 low carbon footprint 275 potential for development 274 precipitation changes 274–5 possible greater need for electricity in Europe 275 hydropower, role in energy generation and economic development 262–3 closely linked to development of dams 262 largest dams regulate, store and divert water from rivers 262 many reservoirs used for recreation, tourism and aquaculture 262 Hydropower Sustainability Assessment Protocol 266 Hydropower Sustainability Forum 266–7 hydropower, sustainable, international guidelines 263–7 Alta River conﬂict 263B, 264 lessons learned 264 early consideration, improvement of living conditions 261 negative effect of large reservoirs without appropriate mitigation 264 visual and ecological environmental changes less acceptable 263–4 Indus delta 41–2 dams and barrages reduced ﬂows of water and sediment 41–2 river provides irrigation water 41 supports an extensive ﬁshery 41 water management politically sensitive 42 infrastructure maintenance and renewal 179

526 integrated catchment management (ICM) 42 in Australia 312–14 integrated river basin management (IRBM) contribution of reservoirs to 441–55 and maintenance of ambient quality criteria 185 in the Okavango 471 integrated urban drainage management (IUDM), movement towards 164–6, 165 alternative drainage concepts inﬂuence urban development practices 164 LID design to re-establish predevelopment water balance 165 major changes in drainage design and operation philosophy introduced 164 major future issue for IUDM 166 search for innovative harvesting techniques 165 sustainable drainage systems 164 IUDM described 164, 165 problems of delivery 176 public participation and consultation encouraged 165 urban drainage, 21st C. evolution 166 integrated water resources management (IWRM) 7, 501 and catchment management, important in delivery of MDGs 9 difﬁcult to put into practice 324 for a holistic approach to management 276 implementation must be ﬁt for purpose 512 monitoring progress against indicators (themes) 511 principles 7B roadmapping framework to advance the process 511 USA, implementation open to interpretation 64 USEPA watershed approach embraces many concepts 12 International Commission for the Protection of the Danube River (ICPDR) 13, 287 Action Programme for Sustainable Flood Prevention 287 deﬁnes set of general objectives 296B ensuring a harmonious development process 296B

Index speciﬁc action plans for sub-basins 296B application of the WFD 292–3 all parties committed to implementation of WDF (2000) 293 basin-wide overview of reportable issues 293 co-ordination required on different levels 293 need to establish co-ordination with non-EU states 292–3 biggest challenge, development of an international basin management plan 292 development of a programme of measures 297 co-ordinating development of a River Basin Action Plan 297 importance of Danube-Black Sea connection 299 need for feedback on pollution reduction 299 Joint Action Programme 297 requirement for additional investment 298 organization of 290, 292 work groups 290 International Convention on Wetlands see Ramsar Convention International Hydropower Association, recent guidelines 266 International Network of Basin Organizations (INBO) 512 investment in tomorrow’s irrigation 224–9 developing/implementing harmonization processes 227, 229 key pressure points not necessarily biophysical 224, 227 system harmonization approach 224, 228 key element of 229 phases 227, 229 irrigation investment, optimal 222–4 development for collective economic and social beneﬁt, requirements 222–3 minimum platform of structure needed 222 irrigation management in a catchment context 211–30 case study insights 215–19 crop choices and salinity 216–17, 218 groundwater management zones 215–17

sustainable irrigation intensities under a water trading regime 217–19 current solutions 212–15 conceptual framework for managing water and salt balance in an irrigation area 212–13 farm-scale management tool 213–15, 214B future research requirements 229–30 build on concept irrigation system harmonization 230 emerging and ongoing challenges 229–30 regional irrigation business partnerships 230 growth in the irrigated area 211 new scientiﬁc insights 219–30 investing in tomorrow’s irrigation 224–9 optimal irrigation investments 222–4 sustainable environmental management policies 220–2 sustainable environmental management technologies 219–20 sustainable irrigation management a daunting task 211–12 effects of irrigation induced waterlogging and salinity 211–12 Italy, Galli law (1994), reduction of number of utilities 241–2 Johannesburg Plan of Implementation 267 Jordan, Azraq oasis 41 supporting migratory birds and nomads 41 to mitigate impacts of water returned to lakes 41 underlying aquifer supplies water to Amman 41 artesian wells further depleted oasis water 41 Laguna de Bay, a tropical lake under pressure 389–403 alien invasive species, or species substitution 400–1 ‘janitor ﬁsh’ problem 400–1 anthropogenic activities in the watershed 389 carbon ﬁnancing for microwatershed development action plans 401–2 beneﬁts from reforestation 401

Index carbon sequestration through agroforestry 401 consequences of deforestation 401 Laguna de Bay Community Watershed Rehabilitation Project 401 chemical contaminants 399–400 future issues to be addressed 400 heavy metals 399–400 little known about levels of toxic and hazardous substances 400 pollutants posing serious threat to human health 399 environmental concerns 390–3 bioaccumulation of metals into lake biota 393 coliform counts in tributaries 392 contamination with toxic and hazardous substances 392 decline in primary production 393 factors driving the lake hydraulics 390–1, 392 ﬁshpens and ﬁsh cages 393 high nutrient levels in the lake, sources 391–2 major sources of pollution 391 parameters exceeding thresholds 392 rapid decline in ﬁsh productivity 393 shoaling due to sedimentation and siltation 393 environmental users’ fee 398 for enterprises to obtain a Discharge Permit (DP) 398 operates on polluter-pays principal 398 historical background 393–4 ﬁshing a traditional industry in decline 393, 394 ﬁshpen areas 394, 394 ﬁshpen culture, development of 393 Laguna de Bay described 389–90 location 389, 390 Laguna de Bay Shoreland Management 399 attempt to demarcate the shoreland 399 management programme enforced by LLDA 399 rice farmers 399 Lake ﬁshery management program, focus on full implementation of ZOMAP 396–8 ﬁshpens a socio-political problem 396

implementation of ZOMAP a challenge 396, 398 ZOMAP allocations 397 local community participation in lake management 402 at heart of integrated natural resources management strategy 402 group partnerships for mobilizing local resources 402 policy evolution, response to socioeconomic and environmental challenges 394–5 environmental planning and management (EPM) 394–5 establishment of Laguna Lake Development Authority (LLDA) 394 LLDA, promotes ecosystem-based planning and lake resource management 395 LLDA strengthened and improved 395 UNDP studies, major constraints identiﬁed 395 region characterized by two distinct seasons 389 river rehabilitation program and community participation 398–9 conﬁguration of river rehabilitation councils varies 398–9 formation of river councils 398 initiations by the rehabilitation program 398 strengthening science base to support management decision-making 402 diffuse agricultural pollution a major concern, no effort to address the issue 402 need for appropriate reposting systems 402 total population and population expansion 389–90 watershed or lake basin approach to lake management 402–3 Laos, hydropower potential a key resource 273B Nam Ngum Dam 272–3B Mackenzie River Basin, basic management approaches in a high-latitude northern catchment 377–97 basin management and legislative responsibilities 484–6

527 Canada Water Act 485 provinces and water management 486 territorial governments keep a role in water management 486 basin management programmes 484–95 legislative complexities a challenge to integrated basin management 486 municipal water management responsibilities 486 other federal acts 485–6 water management now a shared responsibility 484–5 climate change 483 air temperature increases noted 483, 484 could produce pronounced and wide-ranging effects 483, 485 climate harsh and inﬂuenced by continental conditions 479 contaminant/nutrient enrichment 481–2 non-point and point source inputs 481 pollution related to pulp mills 481–2 Cumulative Environmental Management Association/ Regional Aquatic Monitoring Plan 492–4 CEMA ensured involvement in the programme 492–3 development of two programmes and associated organizations 492 major ﬁndings 493 major foci, 492–3 strategic objectives of the RAMP 492–3 delta, an extensive area of levees 477 development of hydroelectricity 480, 482 early exploration 479–80 rivers became key transport corridors 479 ﬂow regulation 483 greatest defects produced by Bennet Dam 483 Peace and Athabasca river systems 491–2 Williston Reservoir, concerns about effect of 483 forest harvesting 480–1, 482

528 Mackenzie River Basin (cont'd) glacial history, geology and climate affect environmental conditions 478–9 Mackenzie Gas Project 493–4 course of natural gas pipeline 481, 493 ﬁve major components 493 major foci 494 Mackenzie GEWEX (MAGS) and MBIS 194–5 major ﬁndings 494–5 major foci 494 studies a response to changing climate 494 major landcover types 479, 480 mineral exploitation 480–1, 481, 492 oil and gas developments 480, 481 Northern River Basins Study/ Northern River Ecosystem Initiative/Peace-Athabasca Delta 486–92 key environmental issues 486 Northern River Basins, guiding questions 488B Northern River Basins Study (NRBS), objectives 486–8 Northern River Ecosystem Initiative (NREI) 487 oil and gas development 480, 481, 482–3 natural gas development, British Columbia 483 oil sands development 481, 482–3 proposed Mackenzie Gas Project, environmental issues 483 Peace and Athabasca river systems, ﬂow regulation 491–2 effects of river ice-jams 491 ﬂow patterns 491 major ﬁndings focussed on regulation effects of Williston Reservoir 488B, 491–2 scour ability lost with lower peak ﬂows 491–2 Peace and Athabasca river systems, major ﬁndings 489–91 cumulative effect of multiple environmental stressors 489, 490 effect of nutrients from pulp mill efﬂuents on trophic status of receiving waters 489–90, 490 threat of low DO in winter 490–1, 490

Index transport and fate studies, some serious ﬁndings 489, 490 Peace and Athabasca river systems, major foci 487–9 effects of ﬂow regulation by Bennet Dam hydroelectric facility 487, 488B nutrient and contamination impacts of pulp and paper mills 487, 488B unique emphasis placed on PeaceAthabasca wetland ecosystem 487–9 Peace and Athabasca river systems, management implications 492 Mackenzie River Basin Board, a key duty 492 new investment in scientiﬁc research for priority issues 492 protected areas within the basin 479 Canadian Heritage Rivers 479 Peace-Athabasca Delta wetland 479, 487–9 recent major stressors 481–3, 484 recommendations 495–7 effective communications between scientists and decision makers 495, 496 effective water management and a multi-jurisdictional approach 497, 497B importance of developing trust 495 management of studies by multistakeholder boards or steering committees 495 policy-related changes 495–6 size 477, 478 sub-basins 477, 478 macroinvertebrates 86–7 changes in community composition through channel management 91 ﬂow refugia 87–90 aggregation in patches 87–8 interstitial spaces in coarse substrate 88 use of ‘dead zones’ as 87, 88 more in forested reaches 93, 93 response of two metrics to mean particle size 88, 89 in two reaches of a Danish stream 92–3 management with unbranched burrweed 92 reﬂected differences in the macrophyte community 92, 92

macrophytes 84–6, 86 architectural complexity increases taxa richness 89 create a heterogeneous environment 85–6 effects of channel regulation on diversity and composition 90–1, 91 ﬂume experiments 85, 86 growth in patches 85 effect of drag forces 85 of key importance in channel management 84–5 limited as a direct food source for macroinvertebrates 89 macrophyte–macroinvertebrate interactions, inﬂuences on 89–90 patch dynamics dependent on substrate composition 85 result of removal 91–2 in two reaches of a Danish stream 92 managed urban retreat 176 marsh 22B Master Plan for Hydropower, Norway 265 Meuse, R., ﬂooding threat to the Netherlands 52, 53 microbial contamination, regulation of in coastal waters 184 Millennium Development Goals 4, 7, 8B, 9, 501 models non-point source catchment models process-based models 123B statistical or empirical models 123B modern drainage practice evolution 163–4 computational methods advancing basic Rational Method 163 developments since the 1960s 163 need to recognize waste water as resource water 164 out-of-sight-out-of-mind philosophy 164 slower progress in water quality considerations 163 development of modelling approaches 163–4 some existing issues 163 Murray Flows Assessment Tool 317 Murray-Darling Basin economics of alternative water trading and salinity management scenarios 221, 222 people and institutions 406–7

Index dryland farming practices, inﬂuencing hydrological processes and water quality 307 dryland and irrigated agriculture 306 irrigation contributes disproportionately to economic base 306–7, 307 socio-political system a key feature of the basin 304, 307 water piped out of basin for domestic use 306 temporary trade in seasonal water allocations 221 today’s challenges 307–8 key challenge, adjusting uses of water 307–8, 307 sense of urgency about need for change 308 water and ecosystems 303–6, 752 the Coorong, lacks fresh ﬂows, becoming saline 305 ﬂoodplain/wetland vegetation dying, fauna disappearing 305 groundwater systems, additional sources of water 306 high levels of abstraction cause environmental impacts 305 impact on biota of rivers, ﬂoodplains, wetlands and estuaries 305 low-lying ground risks becoming salinized 306, 306 rivers meander for much of their length 304, 305 signiﬁcant alteration of the river system 305 source and course of major rivers 303, 304 sustainability of native ecological communities questionable 305 variability of rainfall and river ﬂows 303, 304 viability of many irrigated areas compromised 306 water quality 306 nutrient, sediment and chemical pollution, secondary issues 306 sediment and nutrient levels, detrimental effects on river and ecosystem health 306 Water Resource Observation Network (WRON) 316 Murray-Darling Basin, integrated management in a thirsty basin 303–18

the basin knowledge system 314–15 bias towards aspects valued in the past 314 focus of MDBC investment in knowledge 315 includes numerous local studies 315 knowledge scattered 314, 314 little progress until 1970s 314 MDBC Sustainable Rivers Audit 315 scientiﬁc underpinning of the Living Murray Initiative 315 state domination of data collection and knowledge generation 314–15 integrated catchment management 312–14 development of a Basin Plan by 2011, inclusions 313–14 developments outside MDB governance fostered progress 312–13 not all plans involving water made under the MDB agreement 313 policy strong on aspiration 312 regional bodies 313 key knowledge gaps for IBM 315–18 challenge in building the knowledge system 318, 318 creating impact from science, key gaps 318 ecology 317 integration and the whole basin 316, 317–18 irrigation water use and impact 317 longer-term future 318 socio-economics, key knowledge challenges 317, 317B the water resource 316–17, 361B largest river basin in settled Australia 303, 304 nature of the basin 303–8 people and institutions 306–7 today’s challenges 307–8 water and ecosystems 303–6 water quality 306 the past shapes potential for the future 308–12 adaptive cycle of complex systems theory 311–12, 312 agreement to form new MurrayDarling Basin Authority 310 application of ICM at basin scale delayed 308 Basin Salinity Management Strategy (BSMS) 310

529 development of new policies and strategies 309, 310 federation of states into a Commonwealth 308 impact of diversions for irrigation clear 309, 309, 310 Murray-Darling Basin Commission (MDBC) 310, 311 necessity for Murray Waters Agreement 208–9 provision of economic opportunity for individual land holders 308, 308 reform of the governance arrangements 309–10 three types of water scarcity deﬁned 309, 311, 311 Murray-Darling Sustainable Yields project 316, 316 Murrumbidgee Catchment/Irrigation Area, whole system water balance accounting 224, 225–7 hydro-economic rankings of on- and off-farm investment options 224, 227, 228 unaccounted for water 224 Nakivubo Swamp, Uganda, effective removal of nutrients and pollutants 30 Nam Ngum Dam, Laos 272–3B creation of a commercial ﬁshery 273B ﬂood management 273B irrigated agriculture 273B tourism 273B Namibia 465 formulation of a Water Development Plan 463–4 National Water Plan, eventual connection to Okavango River caused strong protests from Botswana 464, 470 and the Okavango Delta 457 prognosis of water shortage 463 nanoparticles 150–1, 512 key strategic requirement, assess loadings of oxide nanoparticles 151 routinely released into wastewaters 150 Natura 2000 Network 24, 32–4 Conservation in Partnership approach 35 focus of 33 legal cornerstone for nature protection 32

530 Natura 2000 Network (cont'd) may be affected by need for EIAs 35, 36 Special Protection Areas and Special Areas of Conservation 33 stakeholder involvement 35 natural environment emergent issues, many only now becoming understood 15 quality of information on has grown 14–15 Netherlands 41 creation of high discharge periods, Dutch Rhine 52, 52 delta areas kept dry by pumping 52 disadvantage of high protection standards 53, 55 Dutch organization 57–8 calculation of design discharge and associated water levels 57 calculations differ for the eastern and western parts 57–8 design discharge tested in 2006, most dikes will not be approved 58 hydraulic boundary conditions (design water levels) 58 new design discharge for 2001 calculated, an increase 58 water levels at design conditions mainly determined by storm surges 57–8 and Flanders 141 ﬂood management 51 ﬂood prone area divided into sections for management purposes 53, 54 Flood Protection Act, 1995 53, 57 design discharge to be calculated every ﬁve years 57 ﬂooding probability 53 government provision, reasonable safety and minimizing risks 55 high discharges from the Meuse 52, 52 land use changes 65–6 possible transition from exceeding probabilities to ﬂood probabilities 55 Rhine branches, protected from the sea 52–3 the safety chain 70–1 elements in 70 ideas around safety, risk and probability change over time 71 preparation strongly connected to awareness 70

Index Space for the River, a Key Planning Procedure 59–62, 65 storm surge 1953, led to re-design of ﬂood system 53 threatened by ﬂooding from sea and rivers 52 voluntary concentration of water supplies 241 Network of Asian River Basin Organizations (NARBO) 13, 512 NGOs, some still campaigning against hydropower 267 Nigeria, Hadejia-Nguru wetlands, economic beneﬁts of ﬁshing, agriculture and fuelwood 31, 31 nitrate, and eutrophication 4–5 Nitrate Vulnerable Zones, prohibition of conservation tillage 120–1 Nitrates from Agriculture Directive (EU) 244 Norway Alta power station, increasing runoff and power production 265B mainly run-of river type 265B Alta River conﬂict 263B application for hydropower concession made 263B construction 263B hydroelectric plant needed for electricity supply 263b river known for its stock of Atlantic salmon 263B unﬁnished EIAs 263B unpopular with Sámi people 263B Alta River, new insights and other user interests 279–80B Environmental Status sheet 279B, 280B next operation period, operator decisions 279B programme focusing on Atlantic salmon 279B Energy21 programme 278 a hydropower country 260–1B Kvilldal, Norway’s largest power plant 261, 261 Rjukan example – Saaheim power station 260–1, 261 triggers for development of hydropower 260–1 hydropower main source for electricity production 263 hydropower-related nation-wide planning activities 264–7 export of peaking power 282, 283

Master plan for hydropower development, revised twice 264–5 Protection Plan for Watercourses 264–5 regional plans for small hydropower developments 266 steps in the process 266 large hydro plants producing ﬂexible peaking power 281–2 nutrient control in coastal waters 184 nutrient regulation, catchment or drainage basin approach 184, 185 nutrient trading and agricultural catchments 121–2, 148–9 nutrient trading and water quality 122B potential to achieve water quality goals with lower costs to society 121 potential trading between point and non-point sources 121–2 USEPA, piloting programmes in a number of farming catchments 122 oestrogenic chemicals in surface water 143, 144B OKACOM (Permanent Okavango River Basin Water Commission) 473B development slow 472 establishment 470 co-operation of three basin states 470 organization 470–1 Okavango River Basin Steering Committee (OBSC) 470–1 work of 471 delegates trained in conﬂict resolution 471 different administrative cultures and different languages 471 embraced an integrated, participatory approach to water resource management 471 helped by USAID 471 problems mainly potential and perceived 471 Okavango Delta Management Plan (ODMP) 466, 466B, 467–70 guiding principles 467 friction between traditional resource users and the expanding tourism industry 467 problems with sectoral planning and management 467 organization and work of 467–8

Index Basin Wide Forum 471–2 initial division into twelve components 468 planning partner Harry Oppenheimer Okavango Research Centre (HOORC) 468 the Secretariat 470, 471 two possible organizations discussed 467–8 organization and work of traditional meetings (Kgotlas) 468 overview of management structures during planning and implementation 472, 472 overview of ODMP and OKACOM achievements 473B problems faced by 368–70 most government ofﬁcers only spent a short time in Maun 469 overall committee for project rather insigniﬁcant 469–70 participation in a mixed blessing for HOORC 469 serious delay at the outset 468–9, 469 the Ramsar Convention 470 work of OKACOM 471 Okavango Integrated River Basin Management Project (IRBM) 471 Okavango River basin 458 almost whole basin on nutrient poor Kalahari sand 457 current management, sustainable development 457–8 Permanent Okavango River Basin Commission (OKACOM) 466, 466B recognized need for uniﬁed management plan 466 shared by Angola, Botswana and Namibia 457 water resources little used, basin in near pristine condition 457 Okavango system, southern Africa, evolution of river basin management 457–73 biodiversity in the Okavango delta 461–2, 461 compared to other wetlands 361 large variations in habitat patterns over small distances 461 plant habitat types identiﬁed in the delta 461–2 Botswana government unilaterally ratiﬁed Ramsar Convention 464

contradiction between development and conservation policies 471 current situation in the basin 464–6 population densities low 464 water abstractions and usage 465, 465 economic development potential 465 ecotourism 465 hydropower generation 465 small-scale irrigated agriculture 465 emergence of water resource management initiatives 465–6 establishment of OKACOM 466 water use vs. water conservation 465–6 evaporation and evapotranspiration 459 rainfall in the basin 459 fate of solutes 460 dominant role of inﬁltration and chemical evolution of groundwater 460 role of delta islands 460 ﬁnancial and logistic support from foreign aid organizations 473 geography, hydrology and biology 458–62 geology and geomorphology 458–9 Boteti sub-basin 459 essentially devoid of structured continuous drainage 459 mainly covered by Kalahari Sands 458 much of ﬂow is overland ﬂow 459 Panhandle broadens into the Okavango Delta proper 459 part of endoreic Makgadikgadi basin 458, 458 upper fan occupied by central swamp with main and secondary distributaries 459 historical use and management of the system 462–4 Angola, developed plans for use of Okavango water 463 Botswanan independence attracted money for water development schemes in the delta 463 early projects have not survived 463 expansion of livestock into the delta restricted by tsetse ﬂy 463 Namibia and the Okavango 463–4 population density always low 462–3 small-scale farming dominant source of livelihood 463

531 traditional way of utilizing the delta 463 importance of the ﬂood pulse 459, 462 biomass in the delta 462, 462 creates and maintains seasonal ﬂoodplains 462 direct positive relation between ﬂood size and ﬁsh production 462 high levels of nutrients in the ﬂoodplains 462 importance of river channel abandonment 462 integrated water resources management 472 livelihood sources 465 Management Framework at the international level (OKACOM) 470–1 Okavango Delta in Botswana 457, 463 diamond discovery brought some problems 464 importance of tourism 464, 465 initiation of Southern Okavango Integrated Water Development Project 464 rainfall, runoff and ﬂooding 459–60, 471 catchment runoff 459, 460, 462 delta outﬂow 459 ﬂooding in the delta caused by the seasonal ﬂood pulse 459 ﬂow into Boteti 459 hydro-period conditions vary throughout the delta 459–60 sedimentation and stream abandonment 460–1 bedload trapped in channels leads to aggradation and abandonment 460 example, drying out of the Thaoge tributary 461 important process in maintenance of the ecosystem 461 inﬂowing water caries little suspended sediment 460 negative consequences for human activity 460–1 vegetation in the basin, savannah woodlands 461 osmotic power production 281B potential for electricity generation high 281B OSPAR Convention 184, 507–8

532 P (phosphorus) 109, 115, 117, 121B considered a limiting nutrient in European rivers 144–5 increased inputs accelerate eutrophication 108B Palmiet Pumped Storage Scheme, South Africa 271–2B built and operated in context of IWRM 271–2B joint venture 271B located in the Kogelburg Biosphere Reserve 271B provides hydro-peaking power and additional drinking water 271B Paris 1781, ﬁrst contract for piped water system 235 1960s, ﬁrst evidence of CSO impacts on receiving waters 136 post-WW2, construction of large upstream reservoirs on the Seine 242B quality of water came to be more important 239B peat formation, community evolution 23 The Philippines, science of water savings 217B much rice-ﬁeld water lost by seepage and percolation 217B multi-scale accounting study, ricebased irrigation system 217B options for improvement in water productivity evaluated 217B Phosphorus and Sediment Yield Characterisation in Catchments 108 physical geography, SE England 326, 328–9 climate 328, 328 problems of climate change 328 deﬁning limits 326, 328 geology 328, 330 land use 328–9, 329 physical features, earth movements and the Ice Age 328 point source efﬂuent discharges 135 point source pollution 5, 107, 287 legislative control in the UK 138–9 pollutants mode of transport to aquatic environments 5 WFD EQSs required to be met 141 requires progressive reduction in concentrations 141

Index polluter-pays principle 220, 245, 398 pollution, gross, reduction of from urban areas and CSOs 146–7 construction of artiﬁcial wetlands 147 key problems in construction, calibration and testing of models 147 use of models in design of improved sewer systems 146–7 pollution sources, point and diffuse (MONERIS) 294B development of Pollution Control Decision Support Tool 294B successfully applied within the Danube River Basin 294B pollution trading see nutrient trading and agricultural catchments population growth 408, 605 brings challenges for food security and environmental sustainability 211 Portugal, formation of large joint boards at regional or metropolitan level 242 potable water supply management 235–49 complexiﬁcation of potable water 235 cross-subsidization 246–8 by spatial integration 246 developed world, slow reduction of per capita demand 247 ethics and equity, important today 246–7 German Stadtwerke 246 growing consumerization of water services could be dangerous 247 an initial water volume free, Flanders, consequences 247–8 limits impacts of heavy, lumpy investment on water prices 246 little elasticity of indoor domestic demand to price 247 social forms of averaging 246 temporal averaging process 246 water suppliers reluctant to study distributive effects 246 environmental engineering: new insights 243–8 demand management 244 development of local cooperative agreements 244 dual water supply system 243 land-use controls in water abstraction areas 244 meeting drinking water quality standards 244

regionally integrated and participative water management 245 reversal of polluter-pays principle! 245 water conservation policies develop when needed 243–4 future perspectives 248–9 future of municipalism 248–9 long-term threat to water services, third world cities 248 may be impossible to improve deﬁcient Third World utilities 248 possibility of irregular services in developed countries 248 historical perspective 235–43 bacteriological criteria 242–3 bringing water to cities and connecting people 236–9 British engineering and the pressured water service 235 continental Europe, PSCTs changed to commercial services 240 development of mixed economy companies, high overall costs 240–1 development of public water services 235, 236 drinking water used for various purposes 243 ﬁltration supplemented drinking water treatment 235 France, doubling of water bills 242 piped water system in Paris 235 problem of compliance with stricter drinking water standards 242 problem of long-term maintenance and capital reproduction 236 PWS facing renewing an ageing infrastructure 240 some reluctance to allow development of mixed economy systems 241 treating water 239–40 problem of longer human life expectancy 235 Protection Plan for Watercourses, Norway 264–5 quantitative microbial source apportionment (QMSA) 188–9, 191, 191, 193, 194 UK FIO concentration in high- and low-ﬂow efﬂuents reported 189, 191, 192

Index Ramsar Convention 32 concept of ‘wise use’ of wetlands 32 deﬁning wetlands 19 guidance on wetland management 32 in the Okavango 464, 470 reedbeds, conservation of 25 renewable energy production, mainly hydropower 267 Renewable Energy Resources Directive 274 Rhine ﬂoodplain, simulation of hydrological dynamics of 39 Rhine R. change to a mainly rain-fed river 66 a ﬂooding threat to the Netherlands 52 protected by dikes 53 river basin development, historical removal of wetlands 20 River Basin Management Plans (RBMPs) 171, 175–6B good ecological potential 175–6B goals individual and site speciﬁc 175B heavily modiﬁed water body (HMWB) 175B qualiﬁcation as 175B plans are a four step process 276 river channel management, ecological consequences 77–101 adaptation for a life in running water 84–90 benthic algae and macrophytes reﬂect environmental factors 84–5 macroinvertebrates 86–7 organisms in ﬂowing water, increased resistance and resilience to disturbance 84, 84B channel management algae in, important for the food web 85 has many impacts on river ecosystems 78, 78 and legislation 94–5 macrophytes are of key importance for 84–6, 86 channel morphology and management 78–83, 81 channel form changed by large-scale engineering 80 channel morphologies respond differently to management 78–9 effect of uniform conditions 80 effects of maintenance 80–1 habitat concept 79B

impacts of river regulation, River Skjern (Denmark) 81–3 rifﬂes and pools distinct biotypes 78–9, 79B riparian vegetation, direct impact on river bank form 81, 83 direct impact of channel management 90–3 exception, mixed macrophyte patches on the unmanaged reach 92 interaction between macrophytes, physical features and invertebrates important 92–3, 92 little recovery of river systems after termination of channel management 91 long-term effect of weed cutting 92 on macrophyte diversity and composition 90–1, 91 macrophyte removal 91 potential for decreasing macroinvertebrate diversity 91 ﬂow refugia: physical-biological coupling 87–90 ﬁsh 90 importance of shown but not fully understood 88–9, 89 macroinvertebrates 887–9 macrophytes 89–90 patches (dead zones) on stream beds 87, 87 future directions 100–1 legislation, ambitious targets for restoring ecological integrity of rivers 100 local decisions made ignoring larger, process oriented controls 100 new management strategies and tools necessary 100 restoration way to mitigate impacts of channel management 100–1 human activities inﬂuence ecosystems 77–8 impacts of river channel management 90–4 management impacts mediated through riparian zone changes 93–4 changes in riparian zones in managed rivers 93–4 channel management, indirect negative impacts on plant biodiversity 94 deforested channels, retarded N transport downstream 93

533 forested stream channels 93, 93 riparian plant species, more along natural stream channels 94, 94 riparian zone changes impacting on in-stream channel characteristics 93 monitoring impacts of habitat degradation 95–7 assessing overall ecological status, methods 95–6 assessment of impacts of channel management 96, 97 existing systems show sensitivity towards habitat degradation 96–7 habitat suitability indices (HSIs) 96 key issue, results interpretation often confounded 97 moderate success in linking biota with current habitat assessment systems 96 need to develop more indicator systems sensitive to hydromorphological degradation 97 use of macroinvertebrate species traits in biomonitoring 97 using hydromorphological assessment systems 95, 96 pressure on rivers increased after industrial revolution 77 restoration and rehabilitation 98–100 active restoration seems less beneﬁcial 99 effect of gravel as part of restoration 99 few monitoring studies assess ecological outcome of expensive projects 98–9 more focus on restoring ecosystem processes 99 River Gelsaa restoration 99 Skjern R. restoration 98 three types of restoration project 98, 98 wood as a restoration measure 88–100 rivers important role in infrastructure and early settlement 77 vital for transport of nutrients and organic matter to the sea 77 streams and rivers are species rich ecosystems 77 water quality legislation, did not regard habitat quality or water quantity 94

534 river channel management (cont'd) recent legislation evolved and became more integrated 94–5 River Continuum Concept 114 river regulation 4 for producing hydropower 4 river systems, major, most have been historically engineered 15 rivers dual function, water supply and efﬂuent disposal 135 provide ecosystem services needing a certain water quality 135 Ruhr catchment, contribution of reservoirs to integrated river basin management 441–55 assured security of water supply 454 challenges and limits of present system 451–2, 451 development and use of Flood Action plans 451–2, 452 ﬂood protection capacity limited 451 daily water demand, low-ﬂow period 446 reaction time after water releases 446 use of EZVOR forecasting model 446 weekly variation in water demand 446, 447 daily water quantity management, low-ﬂow period 446–8 advantage of real time model 448 input data for real-time river basin simulation model 446 model applicable for long-term prognostic simulations 448 model calibration and validation 446, 447, 448, 448 real time forecast at control gauging stations 448, 449 ﬂood management, problems and approaches to 448–51 characteristics of runoff conditions 450 higher occurrence of ﬂood events October to March 449 Lenne catchment area 448–9 nearly half major events inﬂuenced by snow 450 optimization of existing system, reservoirs mainly for ﬂood protection 450

Index prerequisites for ﬂood management of the reservoir system 450–1, 451 improved water quality situation by controlled releases 454–5 increased ﬁsh mortality in the lower Ruhr in the spring, usually in May and June 454 increased ﬁsh mortality in the Ruhr in spring 454 ﬂow conditions appear to be one of main factors 454 interaction of several abiotic factors seems to be responsible 454 some quality indicators have signiﬁcantly higher values 454 water releases can reduce frequency and intensity 454 integrated water quantity and water quality management 453–4 lower Ruhr, reduction of ammonia nitrogen concentrations 453 ammonia released during biochemical decomposition of nitrogenous substances 453 need for larger water releases during this period 453 nitriﬁcation temperature-dependent 453 occurs during low-ﬂow periods 453 operation of reservoirs during ﬂood events 448–52 operation of reservoirs during low ﬂow periods 446–8 operational water quantity management, Ruhr Basin 443–6 Northern Group of reservoirs 443, 444 primary criteria for reservoir operation 444B Ruhr basin favourable to dam building 443 simpliﬁed overview of present stream gauging network 445, 445 Southern Group of reservoirs 443–4, 443, 444 systems of reservoirs established 443 reduced peak runoff during ﬂood events 454 Ruhr catchment 441–3 comprehensive water resource management system 443 low ﬂow problems causing public health problems 443

water scarcity problems, caused by large water abstractions 441, 443, 443 Ruhr River need for river basin management plans 441 role to provide drinking water to the Ruhr district 441, 442, 443 Ruhr Reservoir Association 444 Ruhr River Association 441, 444 daily water demand, low-ﬂow period 446 Operation Centre for data processing 445, 446 forecasting systems installed (2002) 445–6 reservoir system controlled from the Headquarters 444–5 Ruhr River Association Act (1990) 444 saltmarshes 24 Screening Tool, smart dynamic export coefﬁcient approach 194 only environmental FIO ground truth data available 194 Senegal River, catchment management case study 377–87 catchment development limited by water availability 386 catchment hydrology 378–81 ﬂooding characteristics, valley divided into four stretches 378, 381 ﬂoodplain consists of large natural depressions 381 ﬂoods generated by rain over the Fouta Djallon 378 reduction in ﬂow during 1980s and 1990s 378 water major limiting factor for development 378 catchment infrastructure 382 construction of Diama barrage 382 construction upstream of Manantali reservoir 382 development of irrigation and a ﬁshery 382 implementation of a regional infrastructure 382 dams, for power generation and intensive cereal production 378 environmental ﬂood ﬂows 383–5 design ﬂood releases for the Senegal River 384–5, 384, 385 environmental ﬂow releases from Manantali 383

Index factors in conception of the hydrograph 384 managed release may produce a more reliable ﬂood 385 environmental services and their values 383 construction of Manantali and Diama dams, signiﬁcant social and environmental impacts 383 debate over actual economic value of ﬂoodplain production 383, 383 Diama barrage led to severe degradation in the delta 383 signiﬁcant change in vegetation of Djoudj National Park 383 ﬂoodplain wetlands of large rivers, important for wildlife and agriculture 377–8 increasing pressure on wetlands 378 four periods of development 379B OMVS provides institutional framework for managing Senegal Basin waters 386–7 people of the catchment 381 contact zone between Arab Berbers and Negro Africans 381 profound changes due to dam construction 381 the Sahel 377 topography 377 wetter periods in recent geological past 377 the Senegal catchment 378, 380 four periods in catchment development 379B river ﬂows through three zones 378 stakeholder participation 386 Water Charter extended stakeholder involvement 386 transboundary catchment management 381–2 basin countries have planned the management over many years 381 drought in Sahelian region, 1970s and 1980s 381 establishment of transboundary Senegal River Basin Authority (OMVS) 381–2 present structure of OMVS 382 water allocation trade-offs 385–6 environmental ﬂood ﬂow releases, signiﬁcant beneﬁts 385 Plan for Mitigating and Monitoring Impacts on the Environment 385–6

re-inundation of Diawling National Park 386 Senegal River Valley Water and Environmental Project 386 trade-off between electricity generation and managed releases, political implications 385 Water Charter, signed by Mali, Senegal and Mauritania governments 386 Senegal valley, dam construction, losses to local community, beneﬁt to urban elite 42 serial biological concentration (SBC) 219–20, 220 shallow water dominated by obligate plants 22 shorelines, stabilization by mangroves and salt marches 19–20, 30 Skjern R., Denmark impact of river regulation 81, 82 channelization resulted in release of soluble iron 83, 83 effects on wetlands 81 impact of riparian vegetation on bank form 81 large woody debris (LWD), river form and process relationship 83 sulphate concentration also decreased 83 transport of iron 83, 83 restoration of lower part of the river 98, 99 socio-economic drivers of urban drainage 174–5 alternative perspectives 175 evolution of four possible future scenarios 174–5 business-as-usual scenario 175 green scenario 174 privatization scenario 174 technocratic scenario 174 stakeholder engagement seen as a core issue 175 South East Australia Climate Initiative (SEACI) 316 South East England’s water resources, a dilemma in sustainable development 323–46 approaching River Basin Management 323–5 adoption of a holistic attitude 324 Dublin Principles 325B few cases where watersheds deﬁne political units 324

535 integrated water resources management (IWRM) advocated 324 not a single basin, but problems of water resource management 323–4 confronting the future, conﬂicting views 343–6 climate change, making things worse 344–5 concept promoted by government river basin based 343 consideration of a national water policy and the place of the SE within 344, 345 detailed report on the EA website 344 EA admission, beyond 2020 water transfers may be necessary 344 pressures on water resources mounting rapidly, problems ahead 343–4 time taken for water transfers to be ready 344 the dilemma 323 ﬁnding water resources for growth 339–46, 343 increasing the resource 341 new reservoir to be built near Abingdon, purpose of 340 proposals to enlarge four existing reservoirs 339 pumped storage reservoirs 339–40 Severn/Thames transfer 341, 342, 342 some parts no additional water available 339, 339, 340 transfer plans and proposals 341–2, 343 unsustainable/unacceptable abstraction regimes already 339 use of artiﬁcial groundwater recharge 340–1 main bodies/institutions responsible for water resources management 334B recommendations for catchment management 346 scenarios for the SE Region 346 business as usual 346 three lane approach 346 twin track approach 346 SE England region, characteristics 327–32 Basingstoke case study 332B

536 South East England’s (cont'd) hydrology and water resources 329–32, 330, 331 most prosperous of UK’s regions 327 physical geography 326, 328–9 subject to substantial development pressure 327–8 the South East England Plan 323, 325–7, 327 development options and aim 325–7 London, not included but inﬂuential 326 SE England as an English region 325, 326, 327 Swindon, uses Thames water 326 water resources management and use of water 332–9 governance 332–5 water consumption 335–6 water efﬁciency 337–9 Spain Almeria region, non-completion of PWS infrastructure and irrational allocation 245–6 declining wetlands 37 farmers overexploit aquifers and ask for large water transfers 246 stationarity 503, 512 storm surges and ﬂooding in the Netherlands 53, 57–8 stormwater management an advanced approach 176–7 little consideration for more advanced solutions 176 stress ecology 151 ecotoxicogenomics and bioinformatics 151 Sundarbans, mangrove swamps 19–20 sustainable catchment management, key principles in development of 2 sustainable development concept 164 a dilemma for SE England 323–46 Okavango River basin 457–8 sustainable environmental management policies 220–2 problems of water transfer 221 water markets 220–2 agricultural 222 can lead to large environmental externalities 220–1, 222

Index permanent trade in water entitlements limited 221 water trading policy 220 sustainable environmental management technologies 219–20 key on-farm management strategies 219 polluter-pays principle, salinity credits 220 serial biological concentration (SBC) adoption in Australia and Pakistan 219–20 innovative technique for managing salts in agricultural drainage 219, 220 urban efﬂuent reuse for peri-urban agriculture 219B sustainable hydropower development 267 key management tools 277–8 IHA Sustainability Guidelines 277 Sustainability Assessment Protocol 277 promotion of, key recommendations 276–7 sustainable irrigation intensities under a water trading regime 217–19 market-based approaches 217, 219 restriction of irrigation to absolute requirements 217 shallow water areas, minimum irrigation intensities desirable 217, 218 sustainable resource management, catchment scale 1 SWAGMAN models, Australia and Pakistan 214B SWAGMAN® Farm model Coleambally Irrigation Area, information gained 214B dedicated stand-alone and webbased interface 214B environmental management plans structured around on-farm net recharge management 214B institutional instrument to monitor change in WUE and environmental conditions 214B lumped water and salt balance model 214B used for development of strategies 214B swamps 22B Switzerland, wastewater systems 141

Three Georges Project, China 269–70B environmental issues 270B signiﬁcant beneﬁts at a macroecological angle 270B special ecological protection plan 270B uncertainties over sedimentation 270B ﬂood control and navigation 269B primary purpose 269B Total Maximum Daily Loads (TMDLs) 164, 184, 203 approach similar to catchment management process 146 under the WFD 145–6 transboundary catchment management 381–2 transboundary water management 12–13 UK 138, 143 Bathing Directive, reducing compliance costs 202 early pollution control legislation 137, 138B Foresight Review of ﬂood defence 175 microbial source apportionment studies 188, 190 generated an approach to quantitative microbial source apportionment (QMSA) 188–9, 191, 193 model of state-liberal policy 241 National Rivers Authority (Environment Agency) 241 Nitrate Vulnerable Zones 244–5 reduction of P inputs to rivers 145 regionalization of water planning, policing and the water industry 241 stormwater management 141 treatment of sewage biological ﬁlter plants 143 by activated sludge 143 by biological ﬁlter 143 water quality objectives and discharge limits 137–8 see also Glasgow strategic drainage plan urban areas and CSOs, reduction of gross pollution from 146–7 urban drainage standards development 166–7 design ﬂow rates for stormwater runoff systems 166

Index development sites over 2ha, stormwater storage provision 167 event design approach 166 receiving water quality protection 166 ﬂood return period, universal standard design criteria for 166 urban design runoff periods 166, 167 increased storage volumes necessary 167 regulations covering controlled/ permitted activities on larger impermeable surfaces 166–7 regulatory criteria 167 for ponds and wetlands 167 requirements for adoption of criteria 167 two-tier design discharge standard not acceptable 176 urban efﬂuent reuse for peri-urban agriculture, China and Pakistan 219B ‘Filtration and Irrigated Cropping for Land Treatment and Efﬂuent Reuse’ (FILTER) 219B China case study 219B urban ﬂow regime 155–8 dominant channel-forming events shape the watercourse 158 ﬂows ﬂashy, potentially hazardous to local residents 158 recurrence interval (RI) 155, 158 and exceedance probability 157B watershed transformation from rural to urban condition 155 urban runoff management 155–79 current solutions 166–74 best management practices 167–70 development of urban drainage standards 166–7 regulatory approaches for urban runoff 170–1 urban runoff risk assessment 171–4 ﬂow and quality sources and pathways 156B future research requirements 177–9 data and monitoring needs 178–9 integrated management of urban water 179 maintenance and renewal of infrastructure 179 urban runoff processes 177–8 historical background 162–6 deterioration in sanitation practices after the Roman Empire 162

evolution of modern drainage practice 163–4 need for modern urban drainage practices 162 needs of urban societies 162 Rational Method and runoff coefﬁcient Crf 163 Roe’s Tables used in construction of London’s sewer system 162 towards sustainable urban drainage management (IUDM) 164–6 hydrological effect of urbanization 155, 157 insights and implications 174–7 advanced solutions 176–7 climate change 175–6 socio-economic drivers 174–5 receiving water impacts 160–2 calculating CSO design ﬂows 161B discharges from both separately sewered SWOs and CSOs 160 evaluation in terms of speciﬁc site characteristics 161 faecal bacteria pose health risks 161 fundamental intermittent standards (FIS) for discharges 162 typical recovery time after a CSO event 162 upﬂow ﬁltration units 177 urban ﬂow regime 155–8 urban runoff pollutants 158–60 urban runoff pollutants 158–60 aesthetic pollutants in CSO discharges 160 discharge from surface water outfalls can vary in quality 158, 159 EMCs calculated or transferred from existing data bases 158 ﬁrst-ﬂush effects in CSO and SWO discharges 160 pollutant concentrations and loadings for urban stormwater runoff 159, 160 also contain enteric viruses and parasites 159–60 main factors inﬂuencing pollutant concentration 160 pollutant sources and concentrations 159–60 urban land use/pollutant loadings relationship, site speciﬁc 158 urban simulation models and software packages 158–9

537 urban runoff processes 177–8 areas that would beneﬁt from further research 178 need to develop better understanding of 177–8 urban runoff, regulatory approaches for 170–1 LI approach 170–1 focus of US approach 170 Modiﬁed Rational Method still in use 171 promotion of watershed-based approaches within SWPPPs 170–1 encourages ﬂexible thinking and innovative methods 170–1 sewer pipe designs (late 1960s) led to deteriorating CSO performance 171 wet weather CSO discharge policy, US and Europe 171, 172 urban runoff risk assessment 171–4 diffuse urban pollution, risk to surface and ground waters 171, 172 assessment procedure 171 location of ‘hot spot’ source loadings 171–2 production of catchment hazard maps 172, 174 risk assessment for unit area loadings 157, 173B runoff volume, Wallingford modiﬁed rational method 173B techniques for fully integrated assessment not generally available 174 urban storm drainage 151 key research requirements 151 Urban Waste Water Treatment Directive (EU) 5, 184 urban water, integrated management of 179 development of ﬂexible, adaptive cost effective approaches 179 USA 6 approach to ﬂood management 62–5 annual costs of ﬂood damage rising 63 combination of ﬂood threats faced in any year 62–3 concept of integrated water resources management 64 extensive hydro-meteorological data collection 63–4 ﬂood control structure managed by USACE 64 ﬂood forecasting and warning 65

538 USA (cont'd) ﬂoodplain management a shared responsibility 63 loss of life from ﬂooding has fallen 63 National Flood Insurance Programme 63, 64 National Response Framework 63, 65 no centralized policy guiding ﬂood management 64 non-structural ﬂood protection measures 64–5 reassessment of emergency management and risk reduction procedures 62 studies and acts better deﬁning the position 64 baseline/reference conditions of pristine waters deﬁned 110, 111 BMPs advocated 169 California construction of big reservoirs allows water exports 238 forced to reduce abstractions from Colorado R., wheeling policy 245 water conservation encouraged 243–4 Clean Water Act 95 TMDL investigations 203 coal deposits 253 Conservation Reserve Program (CRP) 113–14 effects of eutrophication 108B farm management planning to reduce nutrient applications 119B targeted BMPs can decrease nutrient export 119B Federal Water Pollution Control Act 184 hydropower generation 260 level of BMP implementation affects nutrient loss reduction 121B Nationwide Urban Runoff Programme (NURP) Programme data 158 New York City clean natural resources 243 a different path of water supply development 243 policy of ‘no net loss’ of wetlands 34 policy of mitigation banking 34–5 socioeconomic barriers to conservation adoption 114B Cannonsville Watershed example 114B stakeholder alliances 124B

Index the Chesapeake Bay 124B Discovery Farm 124B the Illinois River Watershed Partnership 124B supports water quality trading 148–9 Tennessee Valley Authority, multipurpose project 238 territorially based policies creating markets for ecosystems services, aims 245 Total Maximum Daily Loads (TMDLs) 146 discharge limits 141 see also Chesapeake Bay catchment management USEPA application of BASINS model, parameters include faecal coliforms 199 Causal Analysis/Diagnosis Decision Information (CADDIS) 115, 119 based on the Stressor Identiﬁcation Document 119 future catchment management, requires ongoing assessment of stressor-response relationships 119 a maximum contamination level for nitrate-N in drinking water 108 National Regional Nutrients Criteria Program in the Ofﬁce of Water 110, 111 piloting nutrient trading programmes 122 pollution trading criticized by environmentalists 122 three major cornerstones to watershed approach 12 Wadeable Streams Assessment 115 biological condition of wadeable streams (2005) 114, 115 common stressors affecting stream biota 115, 116 poor conditions with respect to streambed biota 115, 118 relative risk greatest from streambed P and N 115, 117 water, increased global demand 6 water allocation trade-offs 385–6 stakeholder participation 386 water consumption, SE England 335–6 average charge to domestic consumers in England and Wales 335 consumption per person per day 335, 335

leakage rates 335, 336 management under drought conditions 336 much of supplied water is groundwater 335 other users of water 336, 336 reuse of water is characteristic 336 relationship between water demand and people in a household 335, 335 water and ecosystems 303–6 water efﬁciency 337–9 achieved by changing human behaviour 337 costs of introducing water saving devices 337 Environment Agency, demand forecast 338–9, 338 key, changing human attitudes to use of water 337–8 metering 337 rainwater harvesting 338 public perception of leakages 337 in SE region new houses, using more water 337 reducing domestic consumption 337 water facts and futures increasing reliance on distant catchment sources 4 requires member states to identify pressures and impacts on water bodies 185 Water Framework Directive (EU) 9–12, 34–6, 164, 171, 235, 283, 324, 501, 511 addresses need for combined approach to pollution management 9–10 application to the Danube Basin 292–3 assessment of hydropower design in management plans 278–9 more focus on ecology 278–9 categories of wetland indirectly included 34 efﬂuent management classiﬁcation based on ecological status 140 combined approach 140 Scotland, risk from point source pollution 138–9 functional units based on river catchment 10 and Glasgow strategic drainage plan 434

Index hydropower in the context of water resources management 276 hydropower and the environmental goals 175–6B River Basin Management Plans 175–6B key concept underlying the WFD is integration 11–12 main goals/objectives 10, 441 member states, requirements to be met 34 new requirements to improve ecological status of rivers 108 overall aims to be achieved by adaptive management 10–11, 11 principles to be simple, ﬂexible and familiar 10 promotion of sustainable hydropower, key recommendations 276–7 public involvement a key requirement 10 required widespread control of P inputs to rivers 144 requirement to manage point and diffuse source of pollution 184 requires setting of ‘reference conditions’ for water body types 110 socio-economic pressure for destruction hard to resist 34 a substantial piece of EU water legislation 95 ‘Heavily Modiﬁed Water Bodies’ 95 hydromorphological assessment 95 sets standards expected for streams and rivers 95 TMDL approach 145–6 a tool for improvement of water conditions 276 water governance, SE England 332–5 bodies regulating environmental, ﬁnancial and economic matters 333 fragmented 332–3 situation summed up 333 present structure stems from Acts of Parliament 333 creation of water companies 333 setting up of CC Water 334 water service companies 333–4 responsibilities of the EA 333–4 water strategy based on a twin-track approach 334–5 water legislation, European 9–12, 9B Water Poverty Index 510

water quality, improved in rivers ﬂowing through wetlands 30 Water Resource Act (1991), England and Wales 95 water resources management 510–11 at the catchment scale 135 water and salt balance, management of in an irrigated area 212–13 issues modelled as a negative externality problem 212–13, 213 problems if aquifer recharge exceeds natural assimilative capacity 213 water trading 220–2 water treatment 239–40 and the Drinking Water Directive 240 early 20th C., change from quantity to quality 239 Paris example 239B use of water from near at hand 239–40 waterworks, sewage works and concept of local public services 240 Mediterranean Europe, development of bulk water transfer 240 water use, spatial efﬁciencies, availability of information 212 water-scarce settings, enhancing efﬁciency of available water 212 watercourse protection 283 all types to be represented in sum of protection plans 283–4 wet meadow 22B wetland function, controlled by position in the landscape 222 wetland hydrology, modelled 38–9 more complex models 38, 39 selection of appropriate model 38 computer spreadsheets used for water balance modelling 38, 39 use of hydraulic river models for overbank ﬂows 38–9 wetland soils treated as a shallow aquifer 38 wetland module developed for MODFLOW 38 wetland management 19–43 functions and value of wetlands 28–31 global and regional perspective 32–6 human inﬂuence 24–5 causing large-scale changes 24 conserving reedbeds (Phragmition) 25 ﬂoodplain and water-meadows 24–5 global scale, climate change 25–6

539 hydro-ecological tools 39–41 eco-hydrological guidelines 40 ecological restoration 40–1 ‘trafﬁc light-based’ water level regime zones 40, 40 integrating wetlands within catchment management 41–2 new scientiﬁc insights 38–41 political and legal setting 31–2 catalogue of destruction remarkable 32 reclaimed or improved land, viewed as a beneﬁt 32 setting objectives for wetlands 26–8 wetland ecosystem 21–4 basic types 22B other approaches to deﬁnition 21–2 wide diversity of types 21, 21 wetland loss, restoration and management 37–8 principal cause of wetland loss 37 wetland restoration, generic objectives 27B, 38 wetlands 81, 305, 377–8 20th C., catchment management, large losses of wetlands worldwide 42 archaeological aspects 29 artiﬁcial 147 central components within catchment management 20–1 classiﬁcation 23–4, 23 certain distinctive wetlands evolve in particular landscapes 23–4 functions, development of 22–3 methods of water transfer used 23 constraint on realizing beneﬁts of 43 created artiﬁcially 20 dominant processes governing type 19–20 economic valuation of functions 20 economic value of 30–1, 31 ﬂood control function 31 Great Barrier Reef 355 high value element in catchment management 43 hydrologically deﬁned 22 local terms 10 majority reduce nutrient loading 30 naturally transient 20 position important for wetland connectivity 24 rehabilitation and/or creation, preferred 37–8 setting objectives for 26–8

540 wetlands (cont'd) how much intervention is justiﬁable 26–7 setting realistic management objectives at catchment scale 27, 27B surviving in the temperate zone 24 value within a catchment 28–9 components, functions and attributes 28 diversity of species 28–9, 29 performing hydrological functions 28 wetland processes, functions and services 28

Index varied ecological character results in biological diversity 20 vary in practicality and ease of recreation 34 WHO Guidelines Annapolis Protocol 201–2 early ﬁndings of UK QMSA studies becoming available 201 new health-evidence-based standards might not be achievable 201 only path to compliance might be drastic land use change 201 Wicken Fen nature reserve, major species groups 29

World Commission on Dams (WCD) ‘adoption of guidelines makes building impossible’ 266 recommendations 266 suggested a negotiated agreement between developers and stakeholders 266 World Summit on Sustainable Development (WSSD) 7, 12 launch of global EU Water Initiative 12 Zimbabwe, dambos, legal protection possibly misplaced 34