Environmental Pollution of the Pearl River Estuary, China: Status and Impact of Contaminants in a Rapidly Developing Region [1st ed.] 9783662618325, 9783662618349

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Estuaries of the World

Wen-Xiong Wang Philip S. Rainbow

Environmental Pollution of the Pearl River Estuary, China Status and Impact of Contaminants in a Rapidly Developing Region

Estuaries of the World Series Editor Jean-Paul Ducrotoy, The University of Hull, Hull, United Kingdom

More information about this series at http://www.springer.com/series/11705

Wen-Xiong Wang • Philip S. Rainbow

Environmental Pollution of the Pearl River Estuary, China Status and Impact of Contaminants in a Rapidly Developing Region

Wen-Xiong Wang School of Energy and Environment City University of Hong Kong Kowloon, Hong Kong

Philip S. Rainbow Natural History Museum London, UK

ISSN 2214-1553 ISSN 2214-1561 (electronic) Estuaries of the World ISBN 978-3-662-61832-5 ISBN 978-3-662-61834-9 (eBook) https://doi.org/10.1007/978-3-662-61834-9 # Springer-Verlag GmbH Germany, part of Springer Nature 2020 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer-Verlag GmbH, DE, part of Springer Nature. The registered company address is: Heidelberger Platz 3, 14197 Berlin, Germany

Preface

The Pearl River Estuary (PRE) is the Western name of the very large estuary in southern China that today is the location of an industrial metropolis of staggering size and rate of development. The Chinese name of the Pearl River is Zhujiang. Guangzhou lies at the head of the estuary, and Macao and Hong Kong are on the western and eastern sides, respectively, of the wide opening of the estuary to the South China Sea. The new cities of Zhuhai and Shenzhen lie immediately north of Macao and Hong Kong, respectively. The Pearl River is not a new area of interest to Westerners for it featured prominently in the trading and thereby political relations between China and the Westerners of the UK in the nineteenth century. The British knew Guangzhou as Canton, an important entry route for foreign trading goods into China. A major such trading item was opium, a commodity that was to play a significant role in the outbreak of the so-called opium war that led to the establishment of Hong Kong, initially as a British colony but eventually more than a century later as a Special Administrative Region of the People’s Republic of China. Macao has the same designation today. Our interest in the Pearl River Estuary in the twenty-first century arises from the simply vast numbers of people living on its banks and the associated industries providing livelihoods for these coastal inhabitants. There are varying estimates of the number of people living close by the Pearl River Estuary, and these numbers are increasing year by year. The city of Guangzhou has about 15 million inhabitants, while Shenzhen may be home to about 25 million and Zhuhai to 1.8 million. Hong Kong has 7.5 million and Macao more than 670,000 inhabitants, reputedly at the highest population density in the world. The total population of the megacity surrounding the Pearl River Estuary probably exceeds 70 million. And the wastes from these cities and industries ultimately flow into the Pearl River Estuary. What are the major environmental concerns and contaminants of this estuary? Where are the sources for these contaminants? What have been the environmental consequences of the emission of such inevitably contaminated effluents into the estuary? What will the future bring? These are the questions addressed in this book. Contaminants and other pollutants will include toxic metals, organic compounds in the form, for example, of hydrocarbons and organochlorine pesticides, as well as the components of sewage whether of human or domestic animal origin. The sheer magnitude of the fluxes of contaminants into the Pearl River Estuary is significant enough in itself, but the Pearl River Estuary also offers us a model for major contaminated estuaries around the world. Contaminant research has been carried out in the Pearl estuarine ecosystem for decades now, and these data tell a tale relevant beyond southern China and the Far East. The Pearl River Estuary is well known to the authors who have themselves contributed to the literature summarized here. The future environmental health of the Pearl River Estuary is vital to the economic future of millions of people dependent to different degrees on this remarkable ecosystem. Furthermore, the cultural heritage that the Pearl River Estuary brings to the people living nearby is dearly treasured. Just a year ago, on February 18, 2019, the Central Chinese Government announced the establishment of the Greater Bay Area (GBA), which basically covers the majority of the Pearl v

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River Delta area with a total population of over 70 million. The Bay Area is considered to be the fourth megabay in the world (after New York Bay, San Francisco Bay, and Tokyo Bay). Under the concept and mandate of the GBA development, the environment will be under strict scrutiny. The Pearl River Estuary will be one of the highlights in environmental management and protection given its central position in the entire Pearl River Delta region. The PRE system will itself be a model environmental system for environmental scientists owing to its major anthropogenic perturbation and influences as well as the very dynamic nature of the estuary. We believe that many more environmental studies will follow in the near future on this fascinating and culture-rich estuary. This book is a real long-term collaboration between the two authors. PSR was a frequent visitor to the laboratory of W-XW in Hong Kong when the former was the Head of the Department of Zoology at the Natural History Museum, London. We still vividly remember when we jumped into the HKUST library one afternoon nearly 15 years ago to dig out a map of the Pearl River Estuary, while we were discussing the biomonitoring of the estuary. Since then, the local research environment has improved in a rather dramatic way. It is now possible to work directly on this fascinating estuary from the Hong Kong end. Most of the data presented in this book are thus results produced from the authors’ own laboratories over the past decade.

Preface

Acknowledgements

We would like to thank the numerous former and current postdoctoral researchers, visiting scholars, and postgraduate students who have contributed, one way or another, to this book. We thank the funding agencies (National Science Foundation of China, Hong Kong Research Grants Council, Shenzhen Science and Technology Innovation Commission, and TUYF Trust Fund) for their financial support of our research activity over the past many years.

vii

Contents

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . History . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hong Kong . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Macao . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Urbanization and Industrial Development . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reference . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . .

1 1 1 2 2 3 3

Physical Geography . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water Flows and Circulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bottom Topography and Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecosystem and Fisheries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water Quality Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . .

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5 5 7 7 9 10 11

Pollution in the Pearl River Estuary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Eutrophication . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trace Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Organic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Hydrocarbons . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Halogenated Organic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Organochlorine Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Polychlorinated Biphenyls (PCBs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Polybrominated Diphenyl Ethers (PBDEs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Alternative Halogenated Flame Retardants (AHFRs) . . . . . . . . . . . . . . . . . . . . . . Organophosphorus Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Organophosphate Flame Retardants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Organophosphorus Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Xenoestrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pharmaceuticals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Personal Care Products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Other Emerging Contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Other Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Perfluoroalkyl Substances (PFASs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Quaternary Ammonium Compounds (QACs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microplastics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . . . . . . . . . . .

13 14 14 15 15 18 18 20 21 24 25 25 26 27 29 30 31 31 31 31 32 33 33

ix

x

Contents

Trace Metals in the Water Column and Sediments . . . . . . . . . . . . . . . . . . . . . . . Trace Metals in the Water Column . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trace Metal Concentrations in the Water Column of the Pearl River Estuary . . . . . Geochemistry of Trace Metals in the Water Column of the Pearl River Estuary . . . Ecotoxicology of Trace Metals in the Water Column of the Pearl River Estuary . . . Trace Metals in Sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trace Metal Concentrations in Sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioavailability and Ecotoxicology of Trace Metals in Sediments . . . . . . . . . . . . . . . Trace Metal Concentrations in the Sediments of the Pearl River Estuary . . . . . . . . Historical Changes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecotoxicology of Trace Metals in the Sediments of the Pearl River Estuary . . . . . . Exchangeability and Bioavailability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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37 37 38 40 43 44 45 45 47 47 51 53 54

Trace Metals in Pearl River Estuary Organisms . . . . . . . . . . . . . . . . . . . . . . . . . Bioaccumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Uptake . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioavailability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Uptake of Trace Metals from Sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Accumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biomonitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pearl River Estuary Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oysters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mussels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Clams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Scallops . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Importance of Bivalve Family . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Barnacles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Food Chains . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cetaceans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . . . . . . . .

57 57 57 59 61 61 64 65 66 72 74 76 76 76 80 82 85 86 88 89

Trace Metal Contamination of Seafood from the Pearl River Estuary . . . . . . . . . Regulatory Limits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Molluscs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oysters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mussels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Clams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Scallops . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Gastropods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Crustaceans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Prawns, Mantis Shrimps and Crabs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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93 93 95 96 97 98 99 99 100 100 101 105 106

Trace Metals and Ecotoxicological Effects in the Pearl River Estuary . . . . . . . . . . 107 Biomarkers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 107 Pearl River Estuary Oysters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

Contents

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Pearl River Estuary Dolphins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116 Future Needs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental Transport . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecotoxicological Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecosystem Management and Conservation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Greater Bay Area . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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119 119 120 120 120

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

Introduction

Abstract

The major cities surrounding the Pearl River Estuary (PRE) in South China have played a role in history. In recent decades, there has been extensive urbanization and industrial development of the PRE region. Pollution resulting from this industrialization and urban development is becoming of increasing concern in the region.

History The low-lying area around the Pearl River Estuary (PRE) in southern China historically consisted of farms and small villages dominated by the city of Guangzhou, known in the west as Canton. In the nineteenth century, Canton at the head of the PRE was a major port for trade with western countries, an entry port for foreign trading goods into China. During the eighteenth century, the demand in Europe for Chinese goods such as tea, silk and porcelain had created a considerable trade imbalance not least between Britain and China. Silver from Europe entered Imperial China through Canton. A commodity that redressed this trade imbalance was opium, grown in India by the British East India Company and smuggled illegally into China. This illegal opium trade reversed the flow of silver and created large numbers of opium addicts in China to the concern of the Chinese Emperor and officials. The scene had been set for the outbreak of the so-called Opium War which led ultimately to the establishment of Hong Kong, initially as a British colony but eventually more than a century later as a Special Administrative Region of the People’s Republic of China. In 1839, the Daoguang Emperor (Qing Dynasty) appointed imperial viceroy Lin Zexu in Canton to eradicate the opium trade. Attempts by Lin Zexu to halt the trade peaceably failed. An open letter to Queen Victoria appealing to her moral responsibility elicited no response, and foreign

companies were not prepared to exchange stored opium for tea. Lin Zexu resorted then to confiscation and destruction of all stockpiled opium (more than 1000 tonnes) from the western merchants and a blockade of foreign trading vessels. In response, the British Government dispatched a military force to the Pearl River. The Royal Navy with its greater gunnery power inflicted a series of defeats on the Chinese, including the capture of Guangzhou in 1841, and thereby created a new byword in international military tactics—gunboat diplomacy. The Qing surrendered and the Convention of Chuenpi ceded Hong Kong Island to the British. Yet it took a further year of hostilities until the Treaty of Nanjing was signed in 1842 which formally ceded Hong Kong Island to the British. This one-sided treaty also opened other Chinese treaty ports to foreign trade, Guangzhou thereby losing its previously privileged trading status. The British wanted even more, and a second Opium War (1856–60) involving also the French took advantage of the weakness of the Qing dynasty attempting to combat the Taiping Rebellion. British strategic objectives then included legalization of the opium trade, exemption of imports from internal transit duties, suppression of piracy and permission for a British ambassador to reside in Beijing. A renegotiation of the Treaty of Nanjing ensued and treaties were also signed by China with France and the USA, treaties viewed not surprisingly by the Chinese as historically unequal and unfair.

Hong Kong Hong Kong (Fig. 1) became a colony of Queen Victoria’s British Empire as a result of the Treaty of Nanjing in 1842, to a mixed reaction back home in Britain. Initially, Hong Kong did not develop strongly as a trading centre, hampered by piracy, disease and the hostile policies of the Qing government on the mainland. Conditions in Hong Kong did improve

# Springer-Verlag GmbH Germany, part of Springer Nature 2020 W.-X. Wang, P. S. Rainbow (eds.), Environmental Pollution of the Pearl River Estuary, China, Estuaries of the World, https://doi.org/10.1007/978-3-662-61834-9_1

1

2

Introduction

Fig. 1 The Pearl River Estuary and its surrounding cities

though during the Taiping Rebellion when wealthy Chinese fled the turbulence at home to settle on the island. The end of the Second Opium War in 1860 saw the cessation of the Kowloon Peninsula to the colony of Hong Kong, seeking some protection from any military threat across the harbour. In 1898, Britain obtained a 99-year lease on the New Territories to the north of the Kowloon Peninsula bringing under British control the adjacent ring of hills overlooking the colony. The impending expiry of the lease of the New Territories led to the return of the whole of Hong Kong to the People’s Republic of China in 1997. Hong Kong became a Special Administrative Region, maintaining separate economic and governing systems from mainland China. Shenzhen (Fig. 1), immediately to the north of Hong Kong, officially became a city as recently as 1979, and in 1980, it was officially established as China’s first Special Economic Zone. Over the past 40 years, Shenzhen has undergone development unprecedented in human history and is now considered to be the hub of high technology both nationally in China and in the world.

Macao Macao lies on the western side of the PRE opposite Hong Kong (Fig. 1). Macao was established as a colony of Portugal in 1557, the territory being leased by Ming China while remaining under Chinese authority and sovereignty. In 1887, Portugal obtained perpetual occupation rights for Macao. Nevertheless, in 1999, Macao was returned to the People’s Republic of China and it too became a Special Administrative Region.

Zhuhai lies immediately to the north of Macao (Fig. 1) and is another of the original Special Economic Zones established in the 1980s.

Urbanization and Industrial Development From its rural beginnings, the area has become one of the most densely urbanized regions of the world—the so-called Pearl River Delta (PRD) Metropolitan Region. The megacity that has developed (and continues to develop) is an industrial metropolis of staggering size and rate of development. The 1980s brought reform of China’s economic system and a flood of investment kicked off remarkable economic development. Much of the investment was foreign, not least from Hong Kong as local manufacturers moved operations into the PRD with lower labour costs. The PRD has become one of the most economically dynamic and indeed one of the wealthiest regions in the People’s Republic of China. With the great availability of employment opportunities has come a host of middle-income professional consumers who are amongst the wealthiest in China. The manufacturing output of the Pearl River Delta is significant on a world scale. The PRD is a major manufacturing origin for electronic products (for example clocks and watches), toys, textiles and clothing and plastic products much directed at the export market. The PRD produced nearly five per cent of the world’s goods in 2001, and today it accounts for about one-third of China’s total trade value.

Reference

Pollution There are varying estimates of the number of people living close by the Pearl River Estuary and these numbers are increasing year by year. The city of Guangzhou has about 15 million inhabitants, while Shenzhen may be home to about 25 million and Zhuhai to 1.8 million. Hong Kong has 7.5 million and Macao more than 670,000 inhabitants, reputedly at the highest population density in the world. The total population of the megacity surrounding the Pearl River Estuary probably exceeds 70 million. Population growth and economic development in the PRD since the 1980s have led to increased loadings of various pollutants into the PRE (Ji et al. 2011). Indeed, the PRE was on a list of low oxygenated areas (sometimes called dead zones, where nutrients from sewage, animal waste and fertilizer runoff promote algal blooms, the subsequent decay of which uses up available dissolved oxygen) reported by the United Nations Environment Programme (UNEP) in 2006 (Ji et al. 2011). Direct contaminants potentially affecting the PRE will include toxic metals, such as zinc, lead and cadmium, and

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ecotoxic organic compounds such as polyaromatic hydrocarbons (PAHs) from oil and its derivatives and man-made organochlorines such as certain pesticides and PCBs (polychlorinated biphenyls). Relatively new kids on the block include flame retardants including brominated organic compounds and emerging contaminants including pharmaceuticals and endocrine disruptors. In the twentyfirst century, microplastics and man-made nanoparticles have been added to the list of contaminants of ecotoxicological concern. The different forms of pollution of relevance to the PRE are considered in some detail in chapter “Pollution in the Pearl River Estuary”.

Reference Ji X, Sheng L, Tang D et al (2011) Process study of circulation in the Pearl River Estuary and adjacent coastal waters in the wet season using a triply-nested coastal circulation model. Ocean Model 38:138–160

Physical Geography

Abstract

The Pearl River Estuary is under the influence of eight major river inputs on its western side and of the entry of oceanic water from the south. The subsequent water flows and tidal circulation in the PRE are discussed. Season (wet versus dry) has a significant influence on the hydrodynamics and physical geography of the estuary. Human intervention has increasingly shaped the morphology of the estuary. There are now regular ongoing water quality programs in various regions of the estuary. Hong Kong is exceptional in maintaining a >30-year monitoring program in its waters adjacent to the PRE itself. Long-term environmental monitoring is critical in evaluating the status and trends of environmental changes in the estuary, as well as providing baseline contaminant levels for environmental risk assessments.

Water Flows and Circulation The Pearl River system is an extensive river system in southern China draining a catchment of 454,000 square kilometres. Most of the river basin consists of the provinces of Guangdong and Guangxi, but the catchment also extends into Yunnan, Guizhou, Hunan and Jiangxi provinces and even into the north of Vietnam (Fig. 1). It is considered to be the second-largest river system and the third-longest river system (2320 km) in China. The annual water discharge is about 3.5  1011 m3, which is only second to the Changjiang and 6 times greater than the Yellow River. The term Pearl River (Zhujiang) can be used with either of two meanings. In its widest sense, the Pearl River refers to all the waters draining the large catchment, collected into three major rivers draining into a common estuary—the Xijiang (the West River, which is the longest, 2214 km, and originates in Yunnan Province), the Beijiang (the North River, which is the second largest, 468 km long, and originates in Jiangxi Province) and the Dongjiang (the East

River, the third largest). The rivers are therefore tributaries and the common estuary is the Pearl River Estuary (PRE) (Fig. 1). In a more restricted sense, the Pearl River begins just upstream of Guangzhou as the waters of the Beijiang start to flow through this city. There were said to be pearl-coloured shells along the bottom of the river in Guangzhou, so providing the name. In total, there are eight gates (Men), which directly discharge the river water into the estuary. The Dongjiang joins in the north of the head of the estuary proper demarcated by the Humen Gate (Fig. 2). As the estuary proceeds south, it receives several discharges from the Xijiang at Jiaomen, Hongqili, Hengmen and Modaomen on its western bank, while there are further discharges of the Xijiang system at Jitimen, Hutiaomen and Yamen to the west of the PRE itself. The discharge at Modaomen is in fact the largest of the Pearl River system, exceeding that at Humen (Wong et al. 2003a, b). Less well known to westerners is the name of the Lingdingyang or Lingding Channel, which is the middle channel of the PRE extending from Humen to Guishan Island (at the mouth of the estuary). Further up the Lingding Channel is the Shiziyang (or Shizi Channel), which begins by Humen. The total Pearl River discharge varies from a minimum of about 4000 m3 s 1 in the winter to a maximum of about 20,000 m3 s 1 in the summer (Wong et al. 2003a, b). The majority of river discharge (>80%) occurs during the ‘wet’ season. The rainy season spans from April to September (or October, depending on different regions), and the dry season lasts from October (or November) to March, with annual average precipitation of about 1720 mm. Hydrodynamics in the PRE is largely determined by different forces such as tidal currents, wind forcing, coastal currents, surface heat and freshwater flow. The main water currents are river flows, saltwater intrusion and tidal currents. Circulation patterns are different in winter and summer as a result of differences in the seasonal discharge of freshwater and seasonal climate changes. The Pearl River outflow takes the form of a plume with an associated frontal zone. In the

# Springer-Verlag GmbH Germany, part of Springer Nature 2020 W.-X. Wang, P. S. Rainbow (eds.), Environmental Pollution of the Pearl River Estuary, China, Estuaries of the World, https://doi.org/10.1007/978-3-662-61834-9_2

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Physical Geography

Fig. 1 The major rivers that flow into the Pearl River Estuary. Note all the major rivers are on the western side of the PRE

Fig. 2 A close view of the upper region of the Pearl River Estuary

winter, the frontal zone extends from the surface to the bottom and is inside the estuary (Wong et al. 2003a, b). While the PRE water is partially mixed in the winter, it is strongly stratified in the summer. Then the subsurface front

remains inside the estuary and the surface front moves out of the estuary driven by the greater freshwater discharge of low density and by the south-westerly monsoon wind (Wong et al. 2003a, b). In the winter, the low freshwater discharge means that the Coriolis effect and spring tides become more important in driving estuarine flow dynamics (Wong et al. 2003a). The positions of freshwater inputs along the western side of the PRE and the relative strength of the Coriolis force pushing water to the west set up an apparent plume front from the surface to bottom along the main axis of the PRE (Wong et al. 2003a). The upshot is that the western side of the estuary contains diluted water and more saline surface shelf water enters the eastern side of the estuary, reaching the head of the estuary along the two deep channels (Wong et al. 2003a). In the summer, the combined effect of increased freshwater discharge and the south-westerly monsoon causes the surface-diluted water to occupy much of the PRE while also spreading east into coastal waters around Lantau and Hong Kong (Wong et al. 2003a, b). Several numerical models have been employed to simulate water circulation in the PRE and hydrography in the

Bottom Topography and Sediment

Pearl River Estuary (PRE), including more recently a triply nested coastal circulation model which had a specific finescale resolution model for the simulation of threedimensional circulation in the estuary (Tang et al. 2009; Ji et al. 2011a, b) and a finite volume community ocean model (FVCOM, Pan et al. 2020). The mean surface current in the estuary displays an anticlockwise circulation with a dominance of freshwater on the western side and saltwater dominance on the eastern side, typical of salt wedge circulation. During the wet season, a large-scale low salinity estuarine plume is established in the PRE, which displays a typical two-layer circulation with the salinity front in the mouth of the estuary. Circulation is also affected by a north-eastward coastal current driven by wind forcing. As a result of the irregular bottom topography and different forcing, there is significant variability of water circulation in the PRE at both temporal and spatial scales. As a result of frequent typhoons and tropical storms in the summer, water circulation is also greatly impacted by these events which generate strong coastal currents. During calm weather, tidal currents, saltwedge circulation driven by a density gradient, as well as the salinity front and coastal waters beyond the front, all contribute to water circulation in the estuary. During the dry season, temperature and salinity in the estuary are only weakly stratified, but there is a horizontal salinity gradient near the frontal zone. The plume is close to the western shore of the PRE near the mouths of the rivers because of the low freshwater discharge and the influence of southwestward inner shelf currents. Wind, tide and freshwater flux contribute to the frontal circulation in the plume. Further offshore of the PRE, vertical stratification is weak and water circulation is controlled by barotropic dynamics resulting from wind and tide action.

Tide The PRE is a microtidal estuary, with a relatively small tidal range, the predominant tidal constituent being a semidiurnal and irregular lunar tide. Tidal range increases up the estuary and is highest at the upstream end of the estuary due to the inverted funnel topography and the concentration of tidal energy. The tidal range is about 1.7 m near Humen and about 0.9 m at the mouth of the estuary (Wong et al. 2003a, b). The average tidal range on the eastern side is 1.12–1.62 m, and on the western side 0.85–1.23 m. The highest tidal range is 2.10–2.30 m and occurs during the flood tide periods in June–July and November–December. Because of the influence of the local topography, the flood tide flows in a northwest direction, while the ebb tide flow in a southeast direction (Fig. 3). The average flood tide period is typically shorter than the average ebb tide period. Tidal flow during the flood tide is 0.3–84 cm s 1, whereas the tidal flow

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during the ebb tide is 18–92 cm s 1. Tidal flow in the PRE is affected by various factors in the estuary, such as the bottom topography, meteorological conditions, geometry and river flow (Mao et al. 2004). Tides in the PRE originate from the tidal propagation of the Pacific Ocean through the Luzon Strait. There are various principal semi-diurnal (M2 and S2) and diurnal (K1 and O1) tides in the PRE (Mao et al. 2004), with the prevalence of M2 > K1 > O1 > S2. Tidal patterns can be referenced by the amplitude ratio of (K1 + O1)/M2, which is 0.94 at Humen and 1.77 at the Wanshan Islands (Mao et al. 2004). Most of the daily inequality of the semidiurnal tide is situated in the southern part of the main channel. The eastern side of the estuary also has a greater tidal range than the western side of the estuary. The tidal current in the upstream part of the PRE is mainly rectilinear with the river flow dominating in the non-tidal current. Similarly, the tidal current in the northwest is weaker than in the northeast because of the presence of three major river outlets in the region. The PRE has subtropical monsoon weather. The summer surface water temperature in the PRE is around 28–30  C, and that of the bottom waters around 23–28  C, clearly demonstrating a strong temperature-depth gradient. The highest temperature occurs in July (30  C), and the lowest temperature appears in February (16  C). Difference between the surface and the bottom temperature is in the range of 1.0–7.0  C. Salinity in the PRE is strongly influenced by the interaction between river flow and ocean currents and increases from the river mouth to the estuary. During the dry season, the decreased river flow results in a higher salinity. Salt water can then reach the upper region of the estuary, sometimes causing saltwater intrusion into otherwise freshwater (Fig. 4). In summer, surface salinity is in the range of 0.5–31, and bottom salinity is in the range of 1–33 with a strong vertical salinity gradient. In winter, salinity stratification becomes weak as a result of the reduced river flow. Salinity also varies with the tidal flow in the estuary. More precipitation occurs in the west as compared to the east, and in the summer (wet season) as compared to the winter (dry season).

Bottom Topography and Sediment As a result of strong tidal flow and river influence, the PRE has a very complex bottom topography with three shoals (east, middle and west shoals) and two relatively deep channels (troughs) leading north from the mouth to the head of the estuary. The West Channel or Trough (250 m to 2550 m wide) is also called the Lingding Channel and extends south to Guishan Island. The East Channel or Trough (1250–5100 m wide) extends south to Dachan Island. The channels join to become the main shipping channel in the

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Physical Geography

Fig. 3 Tidal current directions in the PRE during flood and ebb tides, in both surface and bottom waters

estuary. The average depth of the PRE is about 4.8 m (range 2 to 5 m on the western side), but the depth can reach in excess of 20 m in the eastern lower estuary (Ji et al. 2011; Wei et al. 2018). There are numerous small islands at the mouth of the PRE and in adjacent coastal waters. The bottom topography combines with the several freshwater discharges, monsoon winds (north-easterly in winter, south-westerly in summer), tides and coastal currents to deliver a complex water circulation in the PRE (Wong et al. 2003a,b). Surface sediments of the PRE are classified into the mud, silt, sandy mud, sandy silt and silty sand, and are mainly composed of silts, followed by clay and sand. The average compositions of silt, clay and sand in the estuary are 62, 22 and 16%, respectively (Zhao et al. 2017). The most widely distributed sediments are silt and sandy silt, which account

for 84% of all samples (Fig. 5). As a result of the strong river flows, regions near the river mouths generally contain sandy sediments, which are also observed in the north-eastern region of Zhuhai. Vertically, surface sediment displays a pattern of coarse-fine-coarse, with the middle estuary having mixed estuarine and sea silts. Human activity such as desilting and dredging has also significantly affected the original sediment composition, with increasing shoals in the north and diversity of grain size distribution in the estuary. The suspended sediment load of the Pearl River is not as high as in many rivers in China, but still deposits about 90  106 tons per year into the PRE, 86% of this sediment load coming from the Xijiang (Harrison et al. 2008). Almost 95% of this sediment loading is delivered during the summer wet season (Harrison et al. 2008). During the wet season, the

Ecosystem and Fisheries

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Fig. 4 Salinity distribution in the surface waters of the PRE during wet and dry seasons

suspended loads than the eastern side as a result of river input. In fact, one can immediately notice the turbid water when a ship approaches the Zhuhai/Macao region during the 1 h transit voyage from Hong Kong. The anthropogenic changes in and around the PRE in recent decades have inevitably caused morphological changes in the estuary (Zhang et al. 2015). Human intervention has been extensive and includes the building of coastal defence structures, land reclamation and sand mining (Zhang et al. 2015). The Hong Kong—Zhuhai—Macao Bridge (HZMB), which connects Hong Kong, Macao and Zhuhai, is considered to be the world’s longest sea crossing (55 km long) and was opened on October 24, 2018. Several artificial islands were built near the bridge over a period of nearly 9 years and at a cost of 18.8 billion US dollars. As a result of land reclamation, shallow areas have extended to occupy previously deeper areas and waterways have narrowed, reducing the power of tidal currents (Zhang et al. 2015). With the exception of some local areas of erosion, the PRE has been accreting sediment at an average annual sedimentation rate between 1970 and 2010 of more than 3  104 km3 (Zhang et al. 2015). Between 1976 and 2006, the PRE coastline extended seaward by an average 580 m (Zhang et al. 2015). Fig. 5 Distribution of surface sediment types in the Pearl River Estuary (from Zhao et al. 2017)

Ecosystem and Fisheries suspended sediment load of the PRE varies between 0.04 and 0.30 kg m 3 (40 to 300 mg L 1), while dry season suspended sediment concentrations lie between 0.02 and 0.19 kg m 3 (Harrison et al. 2008). The sediment consists mostly of clay and silt (Harrison et al. 2008). There are large differences in suspended particle distribution in different areas of the PRE. The western side of the estuary generally contains higher

The PRE contains a rich diversity of marine organisms typical of tropical and subtropical regions. Because of its estuarine nature (with characteristically high productivity), the PRE is an important area for spawning, migrations and feeding by numerous fish, shrimps and crabs, and is an important fisheries region in Southern China. Many species of fish

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spend their early life stages in the estuary and return to coastal waters in the adult stages. In a more recent study, Zhou et al. (2019) collected samples from eleven sites of the PRE during the period 2013–2016 to identify fish populations. A total of 285 fish species from 88 families and 195 genera was identified, as compared to around 330 species identified in the 1980s, indicating a decline of fish species number over the past decades. Furthermore, much of the fish population structure in the PRE was altered, and about 1/3 of the original fish species had disappeared while there were 50 new species documented. There was only 62% similarity of fish species composition between the 1980s and the 2013–2016 records, whereas the ecological guilds were comparable between the two periods. It appears that human perturbation such as overfishing, introduction of exotic species, and different engineering projects has been responsible for the decline of fish species number. Dominant fish species displayed both temporal and spatial variations, which were related to different environmental conditions such as water flow, dissolved oxygen, primary productivity and turbidity. Obviously, both environmental conditions and human perturbation significantly affect the local ecosystems of the PRE. One of the most unique and precious species in the PRE is the Chinese white dolphin or Indo-Pacific humpback dolphin (Sousa chinensis). In fact, it has become a symbolic species for Hong Kong locally and is almost considered to be of cultural heritage significance in the PRE. Currently, the estimated number of individuals in the PRE is around 2500 (about one-quarter of the entire world population), but there have been signs of decline over the past decades for various possible reasons, such as pollution, marine traffic, land reclamation/habitat loss and reduced fish resource species. There have been many studies on this species related to environmental pollutants (see later). Another species which is common in the PRE is the Indo-Pacific finless porpoise (Neophocaena phocaenoides), which is considered to be a neighbour of the Chinese white dolphins. The distribution of finless porpoises can be further away from the estuary mouth and this porpoise may be found in the eastern side of Hong Kong.

Water Quality Monitoring There are now various water quality monitoring programs in the PRE region, notably conducted by different governmental agencies. The State Government (Ministry of Natural Resources) each year publishes reports on the national status of marine environmental quality including water quality in the PRE (http://www.mnr.gov.cn/sj/sjfw/hy/gbgg/ zghyhjzlgb/), with the latest report of 2019 covering the environmental quality status in 2018. In China, water quality

Physical Geography

is generally categorized into different classes: An area with Class I water is suitable for marine fisheries, marine natural conservation and the protection of endangered species. Class I water is thus considered to be the highest environmental quality of marine water. An area with Class II water is suitable for marine aquaculture, swimming and other marine sports or recreation. A Class II classification will also apply to waters of industrial areas that directly support seafood suitable for human consumption. Class III is the categorization generally suitable for waters affected by industrial discharges. Class IV is the only suitable classification for the environmental quality of waters in areas affected by marine port operations and marine development operations. Class IV is considered to be the worst quality of marine water. However, there is now substantial discussion to further refine these water quality categories in a more rigorous manner, which remains a major challenge. Currently, most environmental assessments of water quality are based on this classification that was proposed nearly 20 years ago. A major revision is required to be more ecologically relevant given the tremendous modification of marine environments as a result of human perturbation. The latest (2019) report shows that nationwide Class I water only constitutes 46.1% of the entire surveyed water bodies while Class II takes up 28.5% and Class III takes up 6.7%. Class IV (and worse Type IV) cover a total of 18.9% of the marine water bodies surveyed. Most Chinese coastal waters have higher total inorganic nitrogen and reactive phosphate concentrations than the criteria used in the classification. Among the major nationwide coastal areas, the Shenzhen coastal environment on the eastern side of the PRE is a particular problem. There are also other governmental agencies (e.g. provincial and municipal city agencies) which conduct regular monitoring of water quality in the South China Sea region. For example, Guangdong Province has set up around 300 monitoring stations and conducts water quality monitoring over four seasons of the year. In 2017, there were about 81.5% of marine waters in the PRE meeting Class I and Class II standards, and there were 6.8–10.4% of these marine waters falling into Class IV. Again, eutrophication as a result of increasing inorganic N and reactive phosphate releases into the PRE is the major environmental problem. The Hong Kong government has provided exceptional service to coastal water quality monitoring in the local region. Since 1986, the Hong Kong Environmental Protection Department has set up 76 stations in the coastal waters around Hong Kong (many of which are on the western side of Hong Kong near the mouth of the PRE) and conducts monthly monitoring of water quality. In addition, sediments from 60 stations are sampled for sediment quality analysis. Typically, basic water parameters such as temperature, salinity, pH, turbidity and dissolved oxygen are measured in situ,

References

and analysis of other parameters such as concentrations of nutrients, total suspended loads, metals, organic contaminants and coliform bacteria are conducted in the laboratory. Each year, the government publishes an annual report of marine water quality (https://www.epd.gov.hk/epd/ english/top.html). The first monitoring program started in 1986 and is still ongoing. In 2004, the program started to monitor 24 chemicals and this list has been further expanded in line with increased concern over emerging chemical contaminants. A total of 40 chemicals is now monitored, including: • Persistent organic pollutants: Dioxins (dioxins and furans), Polychlorinated biphenyls (PCBs), Dichlorodiphenyl-trichloroethane (DDT), Aldrin, Chlordane, Dieldrin, Endrin, Heptachlor, Hexachlorobenzene, Mirex, Toxaphene, Chlordecone, Pentachlorobenzene, Perfluorooctane sulfonic acid, Hexachlorocyclohexanes (HCHs), Hexabromobiphenyl, Polybrominated diphenyl ethers (PBDEs), Antifoulant Tributyl tin (TBT); • Other organic pollutants: Polyaromatic hydrocarbons (PAHs), Phenol, Nonylphenol, Nonylphenol ethoxylates, Methyl Mercury; • Non-metallic inorganics: Fluoride; • Metals: Barium, Beryllium, Cadmium, Copper, Mercury, Selenium, Silver, Thallium, Tin, Zinc; • Metalloid: As. A long-term monitoring program is particularly suitable for identifying the trends of environmental changes, which will have significant implications for local environmental management. For example, many environmental risk assessments are currently based on baseline data derived from long-term monitoring. However, such monitoring is mostly based on routine water quality. The following chapters will explore in more detail the physics, chemistry and biology of pollutants in the PRE system.

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References Harrison PJ, Yin Y, Lee JHW et al (2008) Physical-biological coupling in the Pearl River Estuary. Cont Shelf Res 28:1405–1415 Ji X, Sheng J, Tang L et al (2011a) Process study of circulation in the Pearl River Estuary and adjacent coastal waters in the wet season using a triply-nested coastal circulation model. Ocean Model 38:138–160 Ji X, Sheng J, Tang L et al (2011b) Process study of dry-season circulation in the Pearl River Estuary and adjacent coastal waters using a triple-nested coastal circulation model. Atmosphere-Ocean 49:138–162 Mao Q, Shi P, Ying K et al (2004) Tides and tidal currents in the Pearl River Estuary. Cont Shelf Res 24:1797–1808 Pan J, Lai W, Devlin AT (2020) Circulation in the Pearl River Estuary: observation and modelling. In: Pan J (ed) Estuaries and coastal zones: dynamics and responses to environmental changes. IntechOpen. https://doi.org/10.5772/intechopen.91058 Tang L, Sheng J, Ji X et al (2009) Investigation of three-dimensional circulation and hydrography over the Pearl River Estuary of China using a nested-grid coastal circulation model. Ocean Dynam 59: Article number 899 Wei X, Zhan H, Cai S et al (2018) Detecting the transport barriers in the Pearl River estuary, Southern China with the aid of Lagrangian coherent structures. Estuar Coast Shelf Sci 205:10–20 Wong LA, Chen JC, Xue H et al (2003a) A model study of the circulation in the Pearl River Estuary (PRE) and its adjacent coastal waters: 1. Simulations and comparison with observations. J Geophys Res 108(C5):3156. https://doi.org/10.1029/2002JC001451 Wong LA, Chen JC, Xue H et al (2003b) A model study of the circulation in the Pearl River Estuary (PRE) and its adjacent coastal waters: 2. Sensitivity experiments. J Geophys Res 108(C5):3157. https://doi. org/10.1029/2002JC001452 Zhang W, Xu Y, Hoitink AJF et al (2015) Morphological change in the Pearl River Delta, China. Mar Geol 363:202–219 Zhao G, Ye S, Yuan H et al (2017) Surface sediment properties and heavy metal pollution assessment in the Pearl River Estuary, China. Environ Sci Pollut Res 24:2966–2979 Zhou L, Wang G, Kuang T et al (2019) Fish assemblage in the Pearl River Estuary: Spatial-seasonal variation, environmental influence and trends over the past three decades. J Appl Ichthyol 35 (4):884–895. https://doi.org/10.1111/jai.13912

Pollution in the Pearl River Estuary

Abstract

Different major types of contaminants and other environmental stressors have affected the Pearl River Estuary. Organic contaminants include hydrocarbons, halogenated organic compounds, organophosphorus compounds, xenoestrogens, pharmaceuticals and personal care products, as well as other emerging contaminants and microplastics. As a result of industrialization and human population growth in the region, all these potential contaminants find their way into the PRE. Over recent decades, there have been numerous studies to measure the environmental concentrations of these contaminants in water, sediments and organisms in the PRE. Pesticides are of particular environmental concern in the estuary. Eutrophication has also added to environmental concern for the estuary.

An industrialized metropolis of over 70 million people surrounding an estuary is an inevitable source of pollution into that estuary, and the situation of the Pearl River Estuary (PRE) is no exception. Estuarine pollution takes many forms, such as eutrophication and the introduction of different potentially ecotoxic contaminants into the estuarine system. These contaminants include trace metals, all of which are potentially toxic at a high enough concentration, and organic contaminants including oil derivatives and manmade organic compounds such as organochlorines synthesized as pesticides or supposedly inert insulating agents. In addition, there are many other emerging contaminants released into the estuary as a result of industrial development in the region. DDT (dichlorodiphenyltrichloroethane) is a classic example of a manmade organochlorine pesticide. It was hailed initially as a major agent to eliminate mosquitoes in the fight against malaria and as an agricultural insecticide but later banned as it was discovered to cause unexpected adverse effects on other wildlife such as the thinning of the eggshells

of predatory birds after transfer through the food web (Carson 1962). Other examples of manmade organochlorines are polychlorinated biphenyls (PCBs). PCBs were considered biologically inert and were synthesized to be used as insulating materials in the likes of electrical apparatus. After the 1960s, however, it was appreciated that PCBs had very high environmental toxicities and caused cancers. Their production and use were banned in many countries. The toxicity of such manmade organic compounds increases with their capacity to be bioaccumulated by organisms and particularly with their persistence in the environment, in effect their ability to resist biological breakdown, for example, by microorganisms. Various category names have been developed for these organic contaminants, such as Persistent Organic Pollutants or POPs. In fact, such persistence of particular organic contaminants has led to the coining of the subcategory of Legacy Contaminants— contaminants that still persist in the environment although they are no longer discharged anew. Thus, organochlorine pesticides (such as DDT) and PCBs are Legacy Contaminants. DDT was banned for agricultural use in China in 1983, and production of PCBs in China ceased in 1974. Given their environmental persistence, however, DDT and PCBs still occur in environmental samples including the sediments of the PRE (Pintado-Herrera et al. 2017). Another categorization of ecotoxic organic compounds is that of Emerging Contaminants or Contaminants of Emerging Concern. In contrast to Legacy Contaminants, these are newly recognized potentially ecotoxic contaminants consisting of industrial chemicals, new pesticides, surfactants, pharmaceuticals and personal care products, now being found in water bodies including estuaries. Some Emerging Contaminants are endocrine disruptors, which are chemicals interfering with endocrine systems of humans and wildlife such as fish. They include drugs and POPs, such as halogenated organic compounds (flame retardants and pesticides), and often include chemicals used in the plastics industry. Among the pharmaceuticals of emerging concern

# Springer-Verlag GmbH Germany, part of Springer Nature 2020 W.-X. Wang, P. S. Rainbow (eds.), Environmental Pollution of the Pearl River Estuary, China, Estuaries of the World, https://doi.org/10.1007/978-3-662-61834-9_3

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are antibiotics, now prevalent in human and domestic animal waste effluents. Other manmade contaminants relevant here are nanoparticles, often made with trace metal components, and microplastics.

Eutrophication Eutrophication can be described as an excessive richness of nutrients in a body of water, typically resulting from runoff from the surrounding land, which causes a dense growth of plant life, usually of algae. These algae, whether in the form of phytoplankton or benthic microphytes and macrophytes, eventually die, and their subsequent decay is engineered by microorganisms such as bacteria. This microbial activity initially uses oxygen dissolved in the estuary, resulting in waters of low oxygen content (hypoxia, defined when dissolved oxygen concentration in the water body is typically less than 2 mg L 1) or water with no dissolved oxygen at all (anoxia). After the oxygen has run out, anaerobic bacteria will continue the decay process, typically producing conditions rich in obnoxious hydrogen sulphide in the sediments and water column. No plant or animal life can survive. The agents of eutrophication in the local runoff are plant nutrients, either in the form of sewage (human and animal) or ammonia, nitrite, nitrate or phosphate ions derived from that sewage, or indeed from chemical fertilizers applied in local agriculture. The impact of eutrophication in the PRE is the greatest in the dry season and in the upper estuary, the high river discharge in the summer diluting nutrient and organic loads (Harrison et al. 2008). In the dry season, large regions of the upper estuary are hypoxic (low in dissolved oxygen), with very high dissolved nutrient concentrations, especially of ammonia, and very high carbon dioxide concentrations, indicating the presence of heterotrophic microorganisms utilizing dissolved organic carbon as an energy source (Harrison et al. 2008). Thus, NH4 concentrations can reach above 600 μM, nitrate (NO3) about 300 μM and nitrite (NO2) about 60 μM (Harrison et al. 2008). In the summer, nitrate concentrations fall to 100 μM and phosphate concentrations are low (Harrison et al. 2008). Eutrophication is not severe, and phytoplankton growth may be phosphorus limited while also being negatively affected by turbulent river flow and turbidity reducing light penetration (Harrison et al. 2008). More recently, Qian et al. (2018) analysed the trend of dissolved oxygen levels in the lower part of the PRE based on monthly cruises from April 2010 to March 2011. Oxygen depletion was found in the bottom layer of the lower part of the estuary during summer, over a spatial scale of 1000 km2, as a result of the decay of the phytoplankton bloom. Long-

Pollution in the Pearl River Estuary

term monitoring suggested that there has been a decreasing rate of the annual minimum dissolved oxygen concentration of around 2 μmol kg 1 y 1 over the past 25 years in the area close to Hong Kong, in comparison with an increase of surface nitrogen concentration of 1.4 μmol kg 1 y 1. The scale of hypoxia/anoxia in the region may be proportional to the amount of nutrient input into the estuary. In addition to the nutrient input, the hydrodynamics of the region, such as the stability of the water column as a result of stratification, may further contribute to the scale of hypoxia. However, Li et al. (2020) also analysed the long-term trends of dissolved oxygen (1988–2011) in the entire PRE during summer as well as the factors controlling the occurrence of hypoxia. These long-term data suggested that the occurrence of low O2/hypoxia in the entire estuary was rather episodic and not related to the season. The geographical position appeared to be much more important in leading to different dissolved oxygen concentrations. As expected, urban reach zones and other upstream regions are more vulnerable to episodic low dissolved oxygen or hypoxic conditions as a result of the increasing influence of urban sewage discharge. The eastern side of the estuary is also prone to hypoxia because of its deeper topography and stronger water column stratification compared to the western regions. Overall, the main estuary region only experienced episodic low dissolved oxygen and sporadic hypoxia after 2000, presumably as a result of the short residence time of the water body in the estuary.

Trace Metals All trace metals are toxic above a threshold bioavailability, and therefore, any trace metal entering the PRE in abundance has the potential to cause ecotoxic effects on the local flora and fauna. Trace metals of potential concern in the PRE are lead, zinc, copper, chromium, nickel and mercury, in addition to the metalloid arsenic often included in loose groupings of trace metals (Luoma and Rainbow 2008). Many of these trace metals will be emitted in the waste effluents of industrial or engineering works with an output of products rich in that metal, for example, the steel, electroplating, electrical, computer or automobile industries. A major source of lead into the estuary is the atmospheric deposition of lead derived, for example, from the burning of coal and the combustion in vehicles of leaded petrol before its banning in China in 2000. Furthermore, in addition to industrial effluents, copper may occur in the runoff from agricultural land as a result of its presence in copper-based herbicides and pesticides. Additionally, copper may enter the waters of the PRE directly from the dissolution of copper-bearing antifouling paints on vessels residing in or passing through the estuary.

Organic Compounds

Towards the end of the last century, a new manmade form of trace metals was increasingly exploited for industrial, technological and commercial use, with the concomitant risk of ecotoxic effects upon subsequent release into the environment. Nanoparticles are particles of diameter between 1 and 100 nanometres, and many contain one or more trace metals. Their use is based on their high surface area to volume ratios, making the nanoparticles very reactive or catalytic. Common metal-containing nanoparticles include titanium dioxide nanoparticles used in sunscreens, cosmetics and paints and silver-rich nanoparticles used in washing machines and clothing to eliminate bacteria and reduce odours. Quantum dots (e.g. cadmium selenium quantum dots) are nanoparticles made of semiconducting materials and are, therefore, used in the electronics world. Few data are yet available on the ecotoxic effects of nanoparticles in systems such as the PRE, but metalliferous nanoparticles are a future environmental threat to which we need to pay attention. Trace metals in the PRE are the subject of several following chapters in this book and are not pursued further here. Rare earth elements (REEs) are all metals (rare earth metals) and include the fifteen lanthanides in the periodic table as well as two more elements (yttrium and scandium), which may be present in the same ore deposits as the lanthanides. These REEs are now widely applied in different industrial technologies (including the defence industry) because of their unique properties. Currently, the majority of REEs are produced and processed in China, which takes up >90% of the world supply. The environmental concern of these REEs should certainly be explored as a type of emerging contaminants. Ma et al. (2019) measured the concentrations of REEs in the water, suspended particles and oysters in the PRE during two seasons (wet and dry) in 2016–2017. Higher water concentrations of ΣREE were measured in the summer as compared to those in the winter due to the significant freshwater input (Fig. 1). Maximum REE concentrations of total REEs (0.2 to 0.4 μg L 1) were found in the mid-estuary area, coupled with the highest suspended solid loads in these regions. However, the suspended particles contained the highest concentrations of REEs in the upstream of the estuary. In addition, overabundances of Pr, Nd, Dy and Ho were found in the surface waters, especially during the wet season, suggesting that anthropogenic activity such as the permanent magnet industry in the upstream may be the potential sources for these REEs in the estuary. The oysters Crassostrea hongkongensis in the PRE also contained higher REE concentrations (1.8–9.1 μg g 1 and 0.9–2.6 μg g 1 for dry and wet seasons, respectively) than those in other bivalve species collected from other places and were significantly correlated with the REE concentrations in suspended particles (Ma et al. 2019).

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Organic Compounds Hydrocarbons Hydrocarbons are organic compounds made up of only hydrogen and carbon atoms. They are derived from decomposed organic matter and occur naturally, particularly in natural gas and in crude oil, and in the subsequent refined petroleum products of crude oil. Hydrocarbons can exist as straight chains and in combinations of different numbers of aromatic (benzene) rings. Polycyclic aromatic hydrocarbons or PAHs (also termed polyaromatic hydrocarbons or polynuclear aromatic hydrocarbons) are composed of multiple aromatic rings. The simplest such as naphthalene have two rings, while anthracene and phenanthrene are three-ring compounds. PAHs can also be produced anthropogenically by the burning of organic matter. The environmental threat of PAHs lies in their carcinogenicity as in the case of the widespread PAH benzo[a]pyrene [BaP] made up of five aromatic rings. BaP is benzopyrene and is found in coal tar, tobacco smoke and grilled food, which formed from the incomplete combustion of organic matter. When accumulated in organisms, BaP is broken down by the detoxification enzyme system cytochrome P450, and it is the metabolite produced that is particularly mutagenic and carcinogenic. Thus, the environmental analysis of BaP and its chemical relatives is of considerable ecotoxicological interest, not least in the PRE. Chen et al. (2006) measured the concentrations of 25 PAHs in sediment samples from the PRE and adjacent South China Sea collected in 2002. Concentrations of total PAHs (the sum of concentrations of 2 to 6 ring PAHs) in the 8 PRE sediment samples ranged from 294 to 1100 ng g 1 dry weight (average 749 ng g 1). In contrast, the 25 comparative South China Sea sediments had total PAH concentrations between 138 and 498 ng g 1 with an average of 286 ng g 1 (Chen et al. 2006). Total PAH concentrations decreased upon increasing distance from the PRE towards the open sea, and the riverine input influences reached between 124 and 276 km from the end of the PRE (Chen et al. 2006). While the PRE samples contained the raised total PAH concentrations, the highest concentration (1100 ng g 1) was at the PRE station nearest the mouth of the estuary. The sediment from this station had the highest total organic carbon content, which is situated in the river-sea boundary zone where organically rich, fine-grained suspended particles deposit from water column to sediment (Chen et al. 2006). 4 to 6 ring PAHs tend to associate with fine organic particles (Chen et al. 2006). The PAHs at the highest concentrations in the PRE sediment samples were perylene, benzo[b]fluoranthene, methylphenanthrene and dimethylphenanthrene. Perylene,

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Pollution in the Pearl River Estuary

Fig. 1 Sums of REE concentrations in the surface waters (upper panels) and suspended particles (lower panels) of the PRE collected in December 2016 (left panels—dry season) and August 2017 (right

panels—wet season). Solid black diamonds represent stations where data are available. No measurement was conducted for suspended particles in December, from the study by Ma et al. (2019)

the predominant PAH in the PRE sediment, is used as a blueemitting fluorescent material in light emitting diodes, and it or its metabolic derivatives are considered carcinogenic. Perylene is, therefore, considered hazardous. Sediment from the top of the PRE contained high concentrations of naphthalene and phenanthrene, together with their methylated homologues typical of a petrogenic origin, suggesting petroleum spillage as a source (Chen et al. 2006). Ratios of specific PAHs in the sediment samples indicated that the PAHs in the PRE sediment originated from both petroleum spillage and the combustion of fossil fuels, while those in sediments further offshore were derived in the majority from fuel combustion alone (Chen et al. 2006).

Chen et al. (2006) considered the general level of PAHs in the PRE in 2002 comparable to that of Masan Bay, Korea, lower than that of the northwest Mediterranean Sea and higher than those of the Cretan Sea and Chesapeake Bay, USA. Distinct from PAHs, linear alkyl chain hydrocarbons can also be found in estuarine sediments. Chains of 10 to 14 carbon linear alkylbenzenes (LABs) are the raw material used in the industrial syntheses of linear alkylbenzene sulfonates and anionic surfactants widely used in synthetic detergents and personal care products such as soaps and shampoos. LABs are of no great environmental concern in themselves, but they are used as environmental markers of sewage and other

Organic Compounds

municipal wastewater discharge (Luo et al. 2008). LABs were used to this end by Luo et al. (2008) in a survey of sediment samples from the estuarine and coastal sites used by Chen et al. (2006), from the top of the PRE and the Zhujiang, Dongjiang and Xijiang feeding into the PRE. As to be expected, the sediment total LAB concentrations showed a clear decreasing trend from sewage sludge (15,970–18,860 ng g 1), through the sediments of the Zhujiang and the Guangzhou Channel (up to 2320 ng g 1) and the Dongjiang (up to 566 ng g 1), the Xijiang (up to 69.4 ng g 1) and the PRE (5.8–25.8 ng g 1) to the northern South China Sea (2.5–23.1 ng g 1) (Luo et al. 2008). PRE sediment LAB concentrations are lower than those of Tokyo Bay and Barcelona Harbour and similar to those of Narragansett Bay, USA (Luo et al. 2008). South and Southeast Asia LAB sediment concentrations are very variable (up to 42,000 ng g 1) with most concentrations below 2330 ng g 1 (Isobe et al. 2004). The PRE system, therefore, shows sediment LAB concentrations in line with many estuaries in South and Southeast Asia (Luo et al. 2008). High correlation coefficients were found between LAB and PAH concentrations in the PRE sediments, indicating that the PAHs were associated with sewage-derived particles (Luo et al. 2008). Pintado-Herrera et al. (2017) also measured the concentrations of 15 PAHs in sediment samples from the PRE, collected this time ten years later in 2012. PAHs were detected in all PRE samples, with total concentrations ranging from 53 to 1370 ng g 1, with the highest concentrations being measured in sediments collected near Guangzhou city. These 2012 total PAH sediment concentrations are in line with the 2002 concentrations of 138 to 1100 ng g 1 (Chen et al. 2008). The most predominant PAHs in this 2012 survey were phenanthrene (average 53, range 14–208 ng g 1), pyrene (41, 5.1–252 ng g 1) and fluoranthene. 1 (30, 5.1–209 ng g ) (Pintado-Herrera et al. 2017). Phenanthrene is composed of three fused benzene rings and is derived from coal tar. It has no commercial use, but is an irritant, photosensitizing skin to light. Pyrene consists of four fused benzene rings formed during the incomplete combustion of organic compounds. Fluoranthene is produced by the burning of coal or petroleum at high temperatures and is a constituent of asphalt. Other PAHs in the PRE sediments included benzo[a]pyrene (10, 1.4–55 ng g 1) (PintadoHerrera et al. 2017), which was also present in the 2002 survey (Chen et al. 2008). In agreement with the 2002 data of Chen et al. (2002), 2012 PAH sediment concentrations generally decreased down the PRE north to south before increasing again in the area between Hong Kong and Macao where organically rich, fine grain-suspended materials tend to deposit (PintadoHerrera et al. 2017).

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Ratios of the summated concentrations of different selected PAHs can be used to provide information on the possible sources of different mixtures of PAHs, such as the combustion of fossil fuels, burning of organic matter, waste incineration, industrial processes and oil spills (PintadoHerrera et al. 2017). Such ratios in the PRE sediment samples revealed the combustion of petroleum and organic biomass to be the main sources of the PAHs present (Pintado-Herrera et al. 2017). These results agree with the conclusions of more detailed studies that vehicular emissions and coal combustion were the two main sources of PAHs in Hong Kong (Wang et al. 2010) and Macao (Mai et al. 2003). More specifically for marine sediments in the Pearl River Delta region, Chen et al. (2003) concluded that traffic tunnel was the largest PAH source (58%), followed by power plant fuel combustion (23%), vehicle petrol combustion (18%) and residential coal burning (2%). Diesel spillage from ships was also a contributing factor particularly near ports (Mai et al. 2003). The question remains as to whether the concentrations of PAHs in the PRE are of ecotoxicological significance. In a preliminary ecological risk assessment, Pintado-Herrera et al. (2017) used hazard quotients (HQ) calculated from the ratio of the measured concentration to a calculated value of a predicted no effect concentration (PNEC). An HQ value greater than 1 indicates an ecological risk. The PNEC values for the PAHs were either acute toxicity values (EC50) divided by 1000 or chronic EC50s divided by 100. In practice, HQ values for PAHs in PRE sediment samples were greater than 1 at all stations. Low molecular weight (2–3 ring) PAHs had an average HQ of 26 and high molecular weight (4–6 ring) PAHs an average HQ of 16 (Pintado-Herrera et al. 2017). In corroboration, PAH concentrations in the northern part of the PRE exceeded the Netherlands target value of 1000 ng g 1 (Pintado-Herrera et al. 2017). Two individual PAHs, phenanthrene and pyrene, accounted for most of the predicted toxicities (Pintado-Herrera et al. 2017). Niu et al. (2018) investigated the seasonal variations of PAH concentration and composition by analysing >62 PAHs in different compartments in the PRE. Over a period of 12 months (2011), dissolved PAHs varied from 26 ng L 1 to 522 ng L 1 and suspended particle PAHs varied from 7.4 μg g 1 to 167.4 μg g 1. The 16 priority PAHs in the USEPA list were in the range of 12.7 ng L 1 to 160 ng L 1 in seawater and 2.8 μg g 1 to 112.3 μg g 1 in suspended particles. Suspended particles were important in controlling the transport and distribution of PAHs as a result of adsorption. LMW PAHs (2–3 rings) accounted for 81% of the total dissolved PAHs and 73% of the particulate PAHs. In a companion study, Niu et al. (2020) also determined the seasonal variation of PAH concentration and composition in surface sediments (40% of the total DDTs) in all the coastal species. The relatively high percentage contributions of DDT to total DDT concentrations in the two prawn species suggest a lower metabolic capacity for DDT breakdown than in the other species, with the crab Charybdis japonica showing the highest DDT transformation capability (Sun et al. 2015b). Ratios of DDT to DDE plus DDD in all PRE species indicated that the DDT in the organisms had been derived from mainly historical residues rather than recent inputs (Sun et al. 2015b). Sun et al. (2015b) considered potential human dietary exposure to DDT from seafood from the PRE. The concentrations of DDT (in terms of wet weight) in the PRE seafood species were all well below the China maximum residual level (MRL) of 500 ng g 1 ww. Furthermore, the estimated daily intake (EDI) of DDT from PRE seafood (28 ng kg 1 d 1) was far below the oral reference dose (RfD) (500 ng kg 1 d 1) proposed by the US Environmental Protection Agency and the acceptable daily intake (ADI) of 10,000 ng kg 1 d 1 set by the Food and Agricultural Organization of the World Health Organization (Sun et al. 2015b). Gui et al. (2018b) also determined the concentrations of DDT in 30 individuals of the Indo-Pacific finless porpoise Neophocaena phocaenoides in the PRE from 2007 to 2016. It was speculated that mortality and the incidence of diseases in the finless porpoises and in Indo-Pacific humpback dolphins Sousa chinensis were related to the high organochlorine pollutant levels in their tissues, particularly DDTs (Gui et al. 2014, 2016). Indeed, blubber DDT levels were within the third-highest tier (30,000–100,000 ng g 1 lipid) out of five classifications evaluated in worldwide bottlenose dolphins worldwide. The humpback dolphins S. chinensis displayed a much higher level of DDTs compared to the finless porpoise, primarily because the dolphins inhabit waters that are closer to the coasts influenced by riverine input, whereas finless porpoises avoid the brackish waters of the estuary.

Polychlorinated Biphenyls (PCBs) A second category of manmade halogenated organic compounds with ecotoxic properties is another group of organochlorines, in this case, polychlorinated biphenyls (PCBs). In the middle of the twentieth century, PCBs were considered biologically inert and were synthesized to be used as insulators and dielectric and hydraulic fluids in transformers, capacitors, electric motors and other such electrical

Polybrominated Diphenyl Ethers (PBDEs)

apparatus. It then became apparent that PCBs had very high environmental toxicities and caused cancers. PCB production and use were subsequently banned in many countries in the late 1970s and 1980s. China first produced PCBs in 1965, and PCB production ceased in 1974. Much electrical apparatus containing PCBs was dumped, but leakages were inevitable. Thus, PCBs are considered as legacy contaminants, banned but still present in the environment because of their high persistence. Indeed, PCBs have been shown to be present in sediments of the Pearl River Estuary collected in 1997 (Hong et al. 1999) and in 2012 (Pintado-Herrera et al. 2017). PCB concentrations in sediment samples collected from the PRE in 1996/97 ranged from 0.18 to 1.82 ng g 1, with the highest concentrations being recorded at the Humen Gate and the outlet of Shenzhen Bay (Hong et al. 1999). The average PRE PCB concentration (0.67 ng g 1) was much lower than that (8.9 ng g 1, range 3.2–16) in Victoria Harbour, Hong Kong, and lower than that (1.7 ng g 1, range 0.05–7.24) in Xiamen Western Bay as recorded by Hong et al. (1995). In the case of the PRE sediment samples collected in 2012, PCBs were detected only in samples from the northern part of the estuary near the Humen Gate, from the outlet of Shenzhen Bay and near Macao, and their concentrations were very low (below 3 ng g 1) (Pintado-Herrera et al. 2017). This decline since the late 1990s was to be expected given the cessation of PCB production in China in 1974. It was only at the most northern sampling stations of the PRE (Humen Gate) that the PCB sediment concentrations in 2012 offered any ecological risk, with the PCB hazard quotients there being between 2 and 17 (Pintado-Herrera et al. 2017). PCBs were also analysed in the coastal fish and invertebrates from the Pearl River Estuary collected in 2005 and 2013 by Sun et al. (2015a, b). PCB concentrations in muscle tissues of baby croaker from the PRE had a median of 460 ng g 1 lipid (range 430–560) in 2005, which had fallen to 92 ng g 1 lipid (83–230) by 2013; those in mullet fell from 160 ng g 1 lipid (140–240) in 2005 to 64 ng g 1 lipid (22–74) in 2013 (Sun et al. 2015a). In the other fish and invertebrates collected from the PRE in 2013, PCB concentrations in muscle and soft tissues ranged from 16 to 700 ng g 1 lipid, with the fish showing the highest concentrations (Sun et al. 2015b). Amongst the fish, sardines Sardinella jussieu had a median PCB muscle concentration of 570 ng g 1 lipid, followed by Bombay duck Harpadon nehereus (220), silver pomfret Pampus argenteus (190), tapertail anchovy Coilia mystus (130) and Chinese herring Ilisha elogata (130). Japanese stone crab Charybdis japonica had a median PCB concentration of 170 ng g 1 lipid, squid Loligo tagoi of 94 ng g 1 lipid, the bivalves meretrix and Ruditapes philippinarum of 55 and 27 ng g 1 lipid, respectively, and the prawns Parapenaeopsis hardwickii and Metapenaeus joyneri of 58 and 44 ng g 1 lipid, respectively (Sun et al. 2015b).

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The PCB levels in the PRE fish muscle tissues were higher than those recorded in fish from major Chinese coastal cities (13–78 ng g 1 lipid) but were one to two orders of magnitude below those of fish collected in North America and Europe, which generally ranged up to several or tens of thousands of ng g 1 lipid (Sun et al. 2015b). Sun et al. (2015b) concluded that the levels of PCBs in the PRE were at the lower end of the global range but were high for China. A possible explanation might be the relatively limited historical usage of PCBs in China (Sun et al. 2015b). As regards risk from dietary exposure to PCBs from PRE seafood items, PCB concentrations (in terms of wet weight) were all well below the China maximum residual level (MRL) of 500 ng g 1 ww. Moreover, the estimated daily intake (EDI) of PCBs from PRE seafood (12 ng kg 1 d 1) was below the oral reference dose (RfD) (20 ng kg 1 d 1) of the US Environmental Protection Agency (Sun et al. 2015b).

Polybrominated Diphenyl Ethers (PBDEs) After the widespread ban of PCBs in the late 1970s, the industry turned to alternative flame retardants, and prominent amongst these were polybrominated diphenyl ethers (PBDEs). In contrast to organochlorines, bromine had replaced chlorine as the halogen concerned. Thus, PBDEs are brominated organic compounds, which were widely applied as flame retardants in electronics, furniture, paints, plastics, textiles and other materials. PBDEs, however, are also ecotoxic, posing serious risks to humans and wildlife as a result of this toxicity in combination with their capacity for bioaccumulation and their persistence in the environment. In their ecotoxic action, PBDEs can act as endocrine disruptors. It was the turn of PBDEs to be banned and/or phased out. PBDEs were typically commercially available as mixtures of BDE congeners. The major commercial formulations, pentaBDE and octa-BDE, were subject to bans in Europe and North America and were withdrawn from the market in 2004 (Covaci et al. 2011). Components of these mixtures were added in 2009 to the Persistent Organic Pollutants list of the Stockholm Convention (Covaci et al. 2011; Sun et al. 2015b), requiring signees to eliminate the production and use of specified POPs. The usage of penta-BDE and octa-BDE was officially banned in China in 2006 (Malkoske et al. 2016). The use of deca-BDE has been banned in Europe since 2008, and two leading chemical manufacturers of deca-BDE, Chemtura and Albemarle, agreed to discontinue production and sale of deca-BDE by the end of 2013 (Covaci et al. 2011). Deca-BDE was added to the Stockholm Convention’s list of POPs in 2017. Mai et al. (2005) carried out a wide survey of PBDE concentrations in sediment samples collected in the Pearl River Estuary and the adjacent South China Sea between

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2002 and 2004. 10 BDE congeners were analysed. Total PBDE concentrations (excluding the deca-BDE BDE 209) in the sediments ranged from 0.04 to 94.7 ng g 1 dw and those of BDE 209 from 0.4 to 7340 ng g 1 (Mai et al. 2005). Correspondingly, BDE 209 represented between 72.6 and 99.7% of all PBDEs present in the sediments. Excluding BDE 209, total PBDE concentrations in the sediments (usually