Water Conservation and Wastewater Treatment in BRICS Nations: Technologies, Challenges, Strategies and Policies

By Pardeep Singh and João Paulo Bassin

Water Conservation and Wastewater Treatment in BRICS Nations: Technologies, Challenges, Strategies, and Policies addresses issues of water resources—including combined sewer system overflows—assessing effects on water quality standards and protecting surface and sub-surface potable water from the intrusion of saline water due to sea level rise. The book's chapters incorporate both policies and practical aspects and serve as baseline information for future adaption plans in BRICS nations. Users will find detailed important information that is ideal for policymakers, water management specialists, BRICS nation undergraduate or university students, teachers and researchers.

• Presents tools and techniques that can be used to preserve water resources, including groundwater and surface water
• Provides geophysical methods to quantitatively monitor physical earth processes associated with water resources, such as contaminant transport and ecological and climate change investigations and monitoring
• Includes desalination techniques which can solve the issue of scarce drinking water

• Language: English
• Publisher: Elsevier Science
• Release date: May 13, 2020
• ISBN: 9780128183403

Book preview

Water Conservation and Wastewater Treatment in BRICS Nations

Technologies, Challenges, Strategies and Policies

Edited by

Pardeep Singh

Yulia Milshina

Kangming Tian

Deepak Gusain

João Paulo Bassin

Contents

Cover

Title page

Copyright

Contributors

Part I: Problem

Chapter 1: Water-related problem with special reference to global climate change in Brazil

Abstract

1.1. Overview of Brazilian water resources

1.2. Major threats for conservation of Brazilian Amazonian water resources and aquatic biodiversity

Acknowledgments

Chapter 2: Water-related problems with special reference to global climate change in Russia

Abstract

2.1. Introduction

2.2. Water resources and anthropogenic impacts in Russia

2.3. Climate change in Russia: trends and projections

2.4. Impacts on water-related economic sectors

2.5. Climatic risk management in Russia

2.6. Conclusion

Chapter 3: Water-related problem with special reference to global climate change in India

Abstract

3.1. Introduction

3.2. Indian context on climate change and water

3.3. Impact on agricultural economy

3.4. Indian context on climate change and water policies

3.5. Scientific simulation model for future prediction

3.6. Conclusion

Chapter 4: Water-related problems with special reference to global climate change in China

Abstract

4.1. Global climate change and China’s water resources status

4.2. China’s water problem in the context of climate change

4.3. Quantitative evaluation of the vulnerability of China’s water systems under climate change conditions

4.4. Future climate change trends in China and adaptive countermeasures

Chapter 5: Influence of global climate change on water resources in South Africa: toward an adaptive management approach

Abstract

5.1. Introduction

5.2. State of water resources and their management in South Africa

5.3. Water resource quality

5.4. Potential climate change impacts on water resources in South Africa

5.5. Water security and governance in face of climate change risks

5.6. Conclusion

Part II: Trends and Strategies with Case Study

Chapter 6: Recent trends and research strategies for treatment of water and wastewater in Russia

Abstract

6.1. Introduction

6.2. Materials and methods

6.3. The Russian water supply and sanitation sector: key trends and uncertainties

6.4. Strategies for Russian water supply and sanitation companies

6.5. Policy recommendations for the governance of water resources

6.6. Conclusion

Acknowledgments

Chapter 7: Recent trends and research strategies for treatment of water and wastewater in India

Abstract

7.1. Introduction

7.2. Water resources in India

7.3. Water contaminants

7.4. Water treatment technologies

7.5. Treatment of wastewater

7.6. Technological advances in water purification technologies

7.7. Conclusion

Chapter 8: Recent trends and research strategies for wastewater treatment in China

Abstract

8.1. A definition of wastewater and an overview of wastewater in China

8.2. Advances in wastewater treatment technology and research in China

8.3. Methods and research progress in water treatment in different industries in China

8.4. Characteristics and experience of wastewater treatment in China

8.5. Conclusion

Chapter 9: Recent trends and national policies for water provision and wastewater treatment in South Africa

Abstract

9.1. Introduction

9.2. The human right to water in South Africa

9.3. Drinking water infrastructure in South Africa

9.4. Water services regulation framework in South Africa

9.5. Blue Drop Certification scheme

9.6. Overview of wastewater treatment facilities in South Africa

9.7. Wastewater reuse in South Africa

9.8. Conclusion

Part III: Policies and Laws

Chapter 10: Government initiative and policies on water conservation and wastewater treatment in Brazil

Abstract

10.1. Introduction

10.2. Historical and legal framework

10.3. National Policy of Water Resources – PNRH

10.4. Administrative aspects

10.5. Additional government initiatives

Chapter 11: Government initiative and policies on water conservation and wastewater treatment in Russia

Abstract

11.1. Introduction

11.2. Materials and methods

11.3. Water infrastructure state and environmental issues

11.4. National regulation

11.5. Water supply and sanitation infrastructure management system

11.6. Tariff policy and financial standing of enterprises

11.7. Water meters

11.8. Mechanisms of public private partnership

11.9. Is it possible to increase tariffs?

11.10. Are there alternatives to unitary enterprises and concession?

11.11. Conclusion

Acknowledgments

Chapter 12: The role of sustainable decentralized technologies in wastewater treatment and reuse in subtropical Indian conditions

Abstract

12.1. Introduction

12.2. Decentralized wastewater treatment: Case studies

12.3. Conclusion

Chapter 13: An exploration of China’s practices in water conservation and water resources management

Abstract

13.1. Introduction

13.2. The evolution of water resources management in China

13.3. The Most Stringent Water Resource Management System

13.4. Achievements and major problems of water resources management in China

13.5. Future trends in water resources management in China

13.6. Conclusion

Chapter 14: Government initiatives and policies for water conservation and wastewater treatment in South Africa and indigenous knowledge

Abstract

14.1. Introduction

14.2. Water management: A driver of the Millennium Development Goals

14.3. Water legislation

14.4. Government initiatives for water conservation

14.5. Wastewater treatment in South Africa

14.6. Indigenous knowledge and development

14.7. Conclusion

Part IV: Indigenous Technical Knowledge (ITK)

Chapter 15: Future prospects for the management of water resources in Russia using indigenous technical knowledge

Abstract

15.1. Introduction

15.2. ITK conceptual framework

15.3. Opportunities for integration of ITK into water resources management in Russia

15.4. Case studies of ITK application to water resources management in Russia

15.5. Conclusion

Acknowledgment

Chapter 16: Indigenous knowledge systems in sustainable water conservation and management

Abstract

16.1. Introduction

16.2. Indigenous knowledge in water conservation and management: some examples

16.3. Conclusions

Acknowledgment

Chapter 17: The future prospect of China’s independent R&D technology (ITK) in water resources utilization and wastewater treatment

Abstract

17.1. Introduction

17.2. The general situation of water resources in China

17.3. Problems in water resources utilization in China

17.4. Development of seawater utilization technology

17.5. Development of industrial wastewater treatment technology

17.6. Development of domestic sewage treatment technology

17.7. Development of circulating cooling water treatment technology

Chapter 18: Future prospective and possible management of water resources in respect to indigenous technical knowledge in South Africa

Abstract

18.1. Introduction

18.2. Global water scarcity

18.3. Traditional knowledge systems (IKS)

18.4. Agriculture

18.5. Land and soil

18.6. Natural resource management

18.7. The South African perspective

18.8. Contribution of water to the South African economy

18.9. Indigenous knowledge and SA

18.10. Water management strategies in SA

18.11. Conclusion

Index

Copyright

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Contributors

Denis Moledo de Souza Abessa,     São Paulo State University – UNESP. Praça Infante Dom Henrique, São Vicente, Brazil

Adeyemi O. Adeeyo,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Andrea Pimenta Ambrozevicius,     Agência Nacional de Águas – ANA. Setor Policial (SPO), Brasília, Brazil

Anwesha Borthakur,     Leuven International and European Studies (LINES), KU Leuven, Belgium

Penggao Cheng,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Wei Du,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Rafael Mendonça Duarte,     Biosciences Institute, São Paulo State University–UNESP, Coastal Campus, São Vicente, São Paulo, Brazil

Olatunde S. Durowoju,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Joshua N. Edokpayi,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Abimbola M. Enitan-Folami,     Department of Biotechnology and Food Technology, Durban University of Technology, Durban, South Africa

Geoffrey Harris,     Department of Public Management and Economics, Durban University of Technology, Durban, South Africa

Fan He,     China Institute of Water Resources and Hydropower Research, Beijing, China

Xinxin Hua,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Changshuo Huang,     Nanjing Hydraulic Research Institute, Nanjing, China

Ademola O. Jegede,     Department of Public Law, University of Venda, Thohoyandou, Limpopo Province, South Africa

Shan Jiang,     China Institute of Water Resources and Hydropower Research, Beijing, China

Tianyu Liu,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Rachel Makungo,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Fhumulani Mathivha,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Yulia Milshina,     National Research University Higher School of Economics, Moscow, Russia

John O. Odiyo,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Daria Pavlova,     National Research University Higher School of Economics, Moscow, Russia

Liliana N. Proskuryakova,     National Research University Higher School of Economics, Moscow, Russia

C. Ramprasad,     School of Civil Engineering; Center for Bioenergy, SASTRA Deemed University, Thanjavur, Tamil Nadu, India

Ajay Vasudeo Rane,     Composite Research Group, Department of Mechanical Engineering, Durban University of Technology, Durban, South Africa

S. Rangabhashiyam,     Department of Biotechnology, School of Chemical and Biotechnology; Center for Bioenergy, SASTRA Deemed University, Thanjavur, Tamil Nadu, India

Nkuna Rivers,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

George Safonov,     National Research University Higher School of Economics, Center for Environmental and Natural Resource Economics, Moscow, Russia

Pardeep Singh,     Department of Environmental Studies, PGDAV College, University of Delhi, New Delhi, India

Sergey Sivaev,     National Research University Higher School of Economics, Moscow, Russia

Nazia Talat,     Centre for Studies in Science Policy, School of Social Sciences, Jawaharlal Nehru University, New Delhi, India

Jiahui Tao,     Nanjing Hydraulic Research Institute, Nanjing, China

Rookmoney Thakur,     Department of Public Management and Economics, Durban University of Technology, Durban, South Africa

Surendra Thakur,     BankSeta Research Chair (Digitalisation), Durban University of Technology, Durban, South Africa

Binota Thokchom,     Centre of Nanotechnology, Indian Institute of Technology Guwahati, Amingao, Assam, India

Adalberto Luis Val,     National Institute for Amazonian Research, Manaus, Amanzonas, Brazil

Tom Volenzo,     Department of Hydrology and Water Resources, University of Venda, Thohoyandou, Limpopo Province, South Africa

Songbo Wang,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Jun Xiang,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Jin Zhang,     Center of African Studies, Shanghai Normal University, Shanghai, China

Lei Zhang,     College of Chemical Engineering and Materials Science, Tianjin University of Science & Technology, Tianjin, China

Yongnan Zhu,     China Institute of Water Resources and Hydropower Research, Beijing, China

Part I

Problem

Chapter 1: Water-related problem with special reference to global climate change in Brazil

Chapter 2: Water-related problems with special reference to global climate change in Russia

Chapter 3: Water-related problem with special reference to global climate change in India

Chapter 4: Water-related problems with special reference to global climate change in China

Chapter 5: Influence of global climate change on water resources in South Africa: toward an adaptive management approach

Chapter 1

Water-related problem with special reference to global climate change in Brazil

Rafael Mendonça Duartea

Adalberto Luis Valb

a    Biosciences Institute, São Paulo State University-UNESP, Coastal Campus, São Vicente, São Paulo, Brazil

b    National Institute for Amazonian Research, Manaus, Amazonas, Brazil

Abstract

This chapter describes the main characteristics of surface water resources in Brazil and some key regional differences in water availability, quality, and vulnerability among the hydrographic regions of the country. Furthermore, we focus on discussing some properties of the largest hydrographic region in Brazil, the Amazonian basin, and outline the main threats and challenges for the water use and the maintenance of water quality throughout the region. Finally, we analyze how anthropogenic pressures and global climate changes might impact water quality, and their principal consequences maintaining of environmental resources and conserving of the most remarkable aquatic biodiversity in the world.

Keywords

Amazon basin

aquatic biodiversity

water quality

water vulnerability

conservation

1.1. Overview of Brazilian water resources

It is estimated that around 12-14% of global surface water drains through the 12 hydrographic regions in the Brazilian territory, achieving a total water-resource availability of around 91,271 m³/s in the country. The percentage of the total water available from the major water basins is as follows: Amazon: 80.80%, Paraná: 6.52%, Tocantins-Araguaia: 5.97%, São Francisco: 2.07%, Southeast Atlantic: 1.25%, Paraguay: 0.86%, South Atlantic: 0.71%, Uruguay: 0.62%, Parnaiba: 0.42%, Western North Atlantic: 0.35%, Eastern Atlantic: 0.33%, and Eastern Northeast Atlantic: 0.10% (ANA, 2012, 2013; Tundisi, 2008). This unequal distribution of water resources results mainly from differences in the mean annual precipitation among the drainage basin regions, which is reflected in marked regional and seasonal disparities in both the flow rate and availability of water from the rivers in each hydrographic region (ANA, 2013, 2017). These factors, together with the differences in population size, urbanization, industrialization, and agricultural activities that have a great impact on both quantitative (total water catchment) and qualitative (effluent discharges) water demands, have a pronounced impact on water balance throughout the country. For example, in terms of their water balance (i.e., the ratio between quantitative demand and availability), between 2006 and 2010 more than 90% of the main rivers of Amazon, Tocantins-Araguaia, and Paraguay drainage basins were classified as either good or excellent, while a high proportion of the main rivers and reservoirs of the Eastern Northeast Atlantic (greater than 90.0%), Eastern Atlantic (around 55.0%), and São Francisco and South Atlantic (between 45.0-55.5%) basins were classified as either critical or extremely critical (ANA, 2013).

In the last two decades (data from 1997 to 2017) the demands for water resources in Brazil has increased by over 80%, which represents additional challenges for the maintenance of water availability and quality, and has resulted in severe hindrances to the establishment of effective policies to water governance (ANA, 2019; Tundisi et al., 2015). According to the National Water Agency (the Agência Nacional de Águas, or ANA), there are multiple water-resource uses in Brazil (such as navigation, hydroelectric generation, fishing, tourism, and leisure) that have little or no impact on the overall quality of water resources. However, water used for irrigation, urban water supplies, industrial applications, and livestock production together comprises more than 92% of the total water withdrawn from these resources (52%, 23.8%, 9.1%, and 8.0%, respectively) (ANA, 2019), and these activities have historically contributed to water quality degradation in several water bodies throughout Brazil (Tundisi et al., 2015; Val et al., 2019). Water quality degradation is known to have had a pronounced negative effect on the total water availability for multiple uses, particular for human supplies, food production, and industrial activities, thus directly compromising water security for human populations. The monitoring of water quality between 2001 and 2011 revealed that most of the water bodies classified as bad (19 < WQI < 36) or terrible (WQI < 19), under the water quality index (WQI) and supereutrophic and hypereutrophic under the trophic state index (TSI) drain highly urbanized and industrialized areas, particularly in the Paraná, Eastern Atlantic, São Francisco, Paraguay, and Southeast Atlantic basins (ANA, 2013).

As previously revised by Val and colleagues (2019), today eutrophication is one of the biggest factors in water quality deterioration. After 1950 Brazil experienced accelerated urbanization that was not accompanied by widespread investments in wastewater treatment plants and resulted in a large accumulation of organic matter in rivers, lakes, and reservoirs. Nowadays the lack of adequate wastewater treatment infrastructure in urban areas has resulted in the discharge of emerging contaminants from pharmaceutical and personal care products (PCPs), which are present at high levels in wastewater effluent and reach receiving surface waters that includes rivers, lakes, and coastal waters (Pereira, Maranho, & Cortez, 2016). In addition, the large increment in the amount of land used for agricultural and in the use of fertilizers to increase crop production has contributed to the eutrophication of water bodies. Moreover, the alterations in land use associated with extensive agriculture practices have increased pesticide and herbicide contamination of water resources, which has been shown to exert a marked effect on aquatic biodiversity (Braz-Mota, Sadauskas-Henrique, Duarte, & Val, 2015). Furthermore, industrial activities and mineral exploitation have led to a great deterioration in water quality as a result of the input of dissolved organic substances and toxic contaminants, such as heavy metals, into surface waters. This anthropogenic-induced deterioration in water quality has resulted in economic losses for many municipalities and regions due to the increased cost for water treatment for the production of potable water and the cumulative impacts on human heath. It has also negatively affected both the environmental services and the biodiversity of aquatic ecosystems (Tundisi et al., 2015; Val et al., 2019).

Significant alterations in hydrological cycles have been predicted in face of global climate change, with marked changes on both precipitation and evapotranspiration regimes that may greatly impact the availability and quality of water resources and potentially increase water vulnerability (Arnell, 1999; IPCC, 2013; Oki & Kanae, 2006). Although the exact impacts of climate changes on global and regional hydrological cycles are controversial and still uncertain, the expectation is that the main hydrographic basins in Brazil will experience pronounced alterations in precipitation and superficial water runoff, and that there will be negative effects on the recharge rates of groundwater aquifers (ANA, 2016). In addition, these changes in regional climate trends have the potential to increase the frequency of extreme climate events resulting in severe droughts and flooding, which have already been seen in some regions in Brazil, such as the three recent extensive floods and two major droughts in the Amazon basin between 2005 and 2012 (Magrin et al., 2014), and the severe droughts in the Southeast Atlantic basin in 2013 and 2014 (Cunningham, Cunha, Brito, Marengo, & Coutinho, 2017; Gomes, Bernardo, & Alcântara, 2017), and in both Eastern Northeast Atlantic and Parnaiba basins between 2012 and 2016 (Marengo, Torres, & Alves, 2017; Marengo, Alves, & Alvala, 2018). These extreme events negatively affected the availability of water resources in those regions and increased the vulnerability of population to water-resources stress, particularly those with low income and in high-risk conditions in highly urbanized areas.

Although the ANA considers that Amazonian hydrographic region to have relatively high water security (ANA, 2013) due to its higher water availability and the higher mean flow of its main rivers compared to the other basins, there has been an incremental deterioration of water quality in this region (Borges, 2006; Cunha, Cunha, & Júnior, 2004; Pereira, Monteiro, & Guimarães, 2010). The lack of an adequate infrastructure for basic sanitation in both urbanized and rural areas (Borges, 2006) and an extremely insufficient system for water-quality monitoring (ANA, 2012) have resulted in a deterioration in water quality and reduced access to potable drinking water for human populations, as well as increased the risk to human health relative due to the prevalence of diseases and infections associated with the water available to these human populations (Borges, 2006; Cunha et al., 2004). In addition, the remarkable seasonal variation in the water level of main rivers, called flood pulses (Junk, Bayley, & Sparks, 1989), brings additional challenges for access to potable drinking water for riverine human populations in rural areas during drought seasons (Sampaio, 2019). Finally, as seen in other hydrographic regions in Brazil, anthropogenic activities such as mining and industrial enterprises, deforestation and changes in land-use, and the creation of dams for hydroelectric plants have had a significant impact on quality of water resources, enhancing the concern about a potential increase in water vulnerability in the face of global climate change. In this chapter we outline the major threats to water security in the Amazon basin, address the specific challenges associated with human pressures and global climate change on water quality, and analyze the impacts of these alterations on the biodiversity conservation in the largest and richness aquatic ecosystem in the world.

1.2. Major threats for conservation of Brazilian Amazonian water resources and aquatic biodiversity

The Amazonian hydrographic region drains a total area of almost 7,000,000 km², representing over 63% of Brazil’s overall drainage area and comprising seven states of the national territory (the percentage of total drainage area in Amazon basin of each of these states is as follows: Acre: 3.4%; Amapá: 3.2%; Amazonas: 35.0%; Mato Grosso: 20.2%; Pará: 27.9%; Rondônia: 5.3%, and Roraima: 5.0%) (SRH, 2006). According the ANA (ANA, 2013), the Amazonian region has an average water availability of more than 73,000 m³/s and a mean flow of 132,000 m³/s, but there is unequal distribution of both across the hydrographic subregions (SRH, 2006). The Amazonian hydrographic region is divided into 10 subregions with enormous differences in hydric availability per capita/year (m³/hab/year) among them, as detailed by the Brazilian Ministry of Environment: Amapá Litoral (1,897,812 km²,or 27% of the total area), Solimões (1,191,866;17%), Xingu (824,223;12%), Purus (736,808;11%), Negro (613,942;9%), Tapajós (553,077;8%), Trombetas (498,224;7%), Foz Amazonas (250,906;4%), Paru (221,864;3%) and Madeira (206,336;2%) (SRH, 2006).

The water storage capacity of reservoirs in the Amazonian hydrographic region represents only 3% (i.e., 21,140 km³) of the total storage capacity in Brazil, and it is mainly used for public supplies in urbanized areas and for hydroelectric power generation (ANA, 2013). Over the last decades, the rate of human population growth in the area of the Amazonian hydrographic region has been around 2.3 times higher than that of other regions in Brazil (population growth of this region was 28.8% between 2000 and 2010) (ANA, 2012), which has had a pronounced impact on water demand especially in highly urbanized areas. The activities with the most impact on demand for water are those related to animal supply (32.4%), urban supply (32.3%), irrigation (19.0%), hydroelectric power generation (7.4%), industries (4.2%), rural use (3.7%), and mining activities (1%) (ANA, 2019). The growth in population and in industrial activities in the Amazonian hydrographic region has not been accompanied by appropriate investments in basic sanitation infrastructure, implementation of state and municipal polices for management of water resources, or water quality monitoring stations, resulting in marked alterations in water quality parameters in several aquatic environments (ANA, 2012; SRH, 2006), that can be seen in many river channels and streams. In this region around 78% of the urban population has no regular access to potable drinking water, as only 6.2% is supplied by a sanitary sewage system (compared to 42.6% in Brazil overall) and only 4.6% of sewage is properly treated, which results in a daily organic domestic load that reaches the receiving surface waters of more than 275 t DBO/day (ANA, 2012).

Three main types of river water are recognized in the Amazon region: black water from the Negro River drainage area (acidic pH ranging from 3.5-5.0, very low content of major cations and anions, and high concentration of dissolved organic carbon or DOC), white water from the Solimões (upper Amazon) River (near neutral pH, high nutrient content, high amount of suspended particles, and low DOC content) and clear water from the Tapajós River (neutral pH and low DOC and ionic levels) (Furch, 1984; Sioli, 1984). Thus water quality parameters (such as ionic composition, temperature, pH, and level of nitrogen compounds) are highly variable between the main types of water in the Amazonian region, and are directly influenced by the seasonal and spatial differences in precipitation and by variation of rivers levels (Cunha & Pascoaloto, 2009; Sioli, 1984; Souto, Oliveira, & Silva, 2015). These spatial and temporal differences in physicochemical composition of water in the Amazonian region clearly impose different challenges for the catchment and treatment of drinking water to both urban and riverine population throughout the region.

1.2.1. Industrial and domestic effluents

The increasing load of domestic organic waste and industrial effluents (particularly from processing industries) contributes greatly to the deterioration of water quality in the Amazonian region and has a significant impact on aquatic communities. For example, in the estuary zone of the Amazon River the lack of an adequate urban wastewater treatment facility for the city of Brangança in the state of Pará has resulted in environmental problems related to deterioration of water quality. During the dry season the estuary waters are expressively more eutrophic, with a significant increase in nitrite (NO2-) and nitrate (NO3-), temperature, and pH, as well as in total fecal coliforms (Pereira et al., 2010). Similarly, strong microbiological pollution has been reported in estuary rivers that drain to Amazon River around Macapá and Santana, the two biggest cities of Amapá state, where increased levels of total fecal coliforms have been directly associated with domestic wastewater disposal, as well as with effluents from agriculture and port and industrial activities (Cunha et al., 2004). The evaluation of metal content in the water and the white muscle of several fish species collected from the Cassiporé River in Amapá state, an area historically impacted by gold-mining activities and by agricultural effluent, revealed that surface water had levels of cadmium (Cd), copper (Cu), chromium (Cr), lead (Pb), zinc (Zn) and mercury (Hg) higher than the limits allowed by Resolution No. 357 of the Brazilian Environmental National Council (CONAMA, 2005). In addition, the authors reported a pattern of accumulation of Cd (Plagioscion squamosissimus), Pb (Poptela compressa), Cr (P. compressa, Pimelodella cristata, and Cyphocharax gouldingi), and Hg (P. squamosissimus, Pseudoplatystoma fasciatum, Hoplias malabaricus, and Serrasalmus rhombeus) above the legal limits in various fish species that are consumed by riverine populations, which may result in increased risk to human health (Lima, Santos, & Silva, 2015).

There is also increasing concern about Hg contamination of the Amazon River from gold-mining activities as it relates to the food chain because Hg can be bioconcentrated in fish, as it is usually bioaccumulated at higher trophic levels (it is generally stored as methylmercury (MeHg) in muscle). For example, in Madeira River the level of total mercury (i.e, Hg and MeHg) in the white muscle of several piscivorous (e.g., Pinirampus pirinampu, Hydrolycus scomberoides, Rhaphiodon vulpinus, and Acestrorhynchus falcirostris) and carnivorous (e.g., Calophysus macropterus, Pellona flavipinnis, and Serrasalmus elongatus) fish species was higher than those measured in detritivorous and herbivorous species, and was higher than the limit of >0.50 mg/kg of total mercury in edible fish recommended by the World Health Organization (WHO) (Bastos, Dórea, & Bernard, 2016). Similarly, the Hg levels in the muscle of fish collected in the Purus River (Acre State) were generally higher (44% of species collected) than the threshold allowed by the WHO, with higher levels found in carnivorous (e.g., for C. macropterus and Cetopsis coecutiens the total Hg ranged from 0.14 to 5.39 mg/kg) and piscivorous species (e.g., for P. pirinampu, H. scomberoides, and Plagioscion squamosissimus the total Hg ranged from 0.06 to 1.09 mg/kg) in comparison to omnivorous and detritivorous fish (Castro, Braga, Trindade, Giarrizzo, & Costa, 2016). Hence the consumption of fish by indigenous and riverine communities may result in human exposure to Hg, as seen by the high levels of mercury in the breast milk (ranging from 0 to 24.8 ng/g) and hair of mothers (ranging from 2.0 to 37.2 μg/g), as well as the hair of infants (ranging from 1.4 to 34.2 μg/g), from very small communities in the area of the Madeira River (Barbosa & Dórea, 1998). High concentration of Hg has also been seen in different tissues (particularly in liver, muscle, and brain) of two species of Amazonian cetaceans-the tucuxi dolphin (Sotalia fluviatilis) and the boto dolphin (Inia geoffrensis)-living in the Japurá, Madeira, and Negro Rivers (Lailson-Brito, Dorneles, & Silva, 2008), indicating that Hg is in fact being bioaccumulating in higher trophic levels of food chain in some areas of the Amazon.

In the area around the city of Manaus (Amazonas state) the unauthorized occupation of lateral margins of streams is associated with a crescent load of domestic wastewater and industrial effluent, which is contributing to a severe degradation in water quality in these aquatic environments (Borges, 2006). Previous studies have showed a significant increase in total ionic composition (for both cations and anions), nitrogen compounds (NO2-, NO3-, and ammonia), conductivity, and pH, and a decrease in dissolved oxygen, in several streams in highly urbanized areas and near the industrial district in Manaus that drains directly into the Negro River (Horbe, Gomes, Miranda, & Silva, 2005; Pinto, Horbe, & Silva, 2009; Silva, Ramos, & Pinto, 1999). These physicochemical alterations in water composition related to domestic sewage input are extremely harmful to the human population, as seen by the presence of viruses that can cause acute gastroenteritis in the streams of Manaus (Miagostovich et al., 2008), and also to aquatic biodiversity (particularly fish species). These species has developed specialized physiological and biochemical adaptations to live in the acidic, ion poor conditions of the black-waters environment of the Negro River basin (Gonzalez, Wood, Patrick, & Val, 2002; Gonzalez, Wilson, & Wood, 2006), and are considered relatively sensitive to nitrogen compounds (such as nitrite and ammonia); they have displayed hematological and metabolic disruptions after exposure to high environmental levels of ammonia that are closely associated with lethality (Avilez et al., 2004; Costa, Ferreira, Mendonça, & Fernandes, 2004; Souza-Bastos, Val, & Wood, 2017; Wood, Netto, & Wilson, 2017). The sewage input in urban surface waters is also introducing the release of contaminants into the streams around Manaus, where pharmaceuticals (such as antiepileptic, antidepressant, beta-blockers, and non-steroidal anti-inflammatory drugs) and illicit drugs (such as cocaine and its main metabolite, benzoylecognine), have been detected in main the channel of Negro River (Thomas, Silva, & Langford, 2014).

Furthermore, the anthropogenic impact on the quality of surface waters is also resulting in the contamination of the sediment of streams by both sewage-derived organic matter and metals. A recent study evaluating the impact of sewage contamination revealed a high concentration (509-12,829 ng/g) and relative proportion (21%–54%) of coprostanol (an important biomarker of sewage-derived sterol input) in two of the main streams crossing a highly urbanized and industrialized area in Manaus (Melo, Silva, & Costa, 2019). In addition, there is strong evidence of metal contamination in the waters of some streams around Manaus that are promoting a persistent metal enrichment of their sediment. Some metals, such as Cobalt (Co), Nickel (Ni), Iron (Fe), Zn, Cd, Cu, and Pb are present at higher levels in both water and sediment (Santana & Barroncas, 2005; Silva et al., 1999) than those allowed by Resolution No. 357 of the Brazilian Environmental National Council (CONAMA, 2005), while an increased level of both Zn and Cu has been found in liver and white muscle samples of an facultative air-breathing armored catfish (Hoplosternum littorale) (Santana, 2016), one of the most tolerant fish species living in these highly impacted streams. Overall, the evidence clearly demonstrates a high level of deterioration in water quality in the surface waters of the Negro River tributaries in peri-urban area of Manaus due to industrial and domestic sewage input, which calls for the rapid and efficient implementation of programs for both sewage treatment plants and the monitoring of water quality parameters.

The degradation of water quality in urban streams as a result of anthropogenic impacts has been accompanied by a loss of species and a great shift in aquatic community composition, resulting in significantly reduced biodiversity. In impacted streams, the loss of habitat due to rubbish deposition and organic wastewater load has also been accompanied by the deforestation of marginal forest, promoting a reduction in the diversity of aquatic and semiaquatic Heteroptera insects (Pereira, 2009). These impacts have also decreased the population of the most common fish species in these environments (77% of Characiforms, 55% of Perciforms and 70% of Siluriforms have been lost), which is highly correlated to increased ammonia and nitrite levels in waters (Anjos, 2007). A complete absence of five species from the Lebiasinidae family (Copella nattereri, C. nigrofasciata, Nannostomus beckfordi, N. marginattus, and Pyrrhulina brevis) has also been observed in impacted streams as compared to areas not impacted, and changes in the composition of fish species have also been noted, with a greater abundance of fish species with facultative air-breathing strategies, such as many catfish of Siluriforms order (e.g., Ancistrus sp., Liposarcus pardalis, Rineloricaria sp., Hoplosternum littorale, Callichthys callichthys, Megalechis personata, and Corydoras cf. aeneus) and electric fish (Electrophorus electrius) (Anjos, 2007). Another important issue related to aquatic biodiversity in impacted streams in the Amazonian region is the presence of exotic species, as seen in streams with increased levels of nitrogen compounds (mainly ammonia and nitrite), phosphorus, and conductivity, and depletion of oxygen levels, and where species such as Danio rerio (Cyprinidae), Poecilia reticulata, Xiphophorus helleri and X. maculatus (Poecilidade), and Oreochromis niloticus (Cichlidae) have been found (Guarido, 2014). These species are recognized as being very tolerant of significant alterations in water quality parameters. In summary, these studies suggest that loss of integrity and deterioration of water quality in aquatic environments due to anthropogenic pressures negatively impact native biodiversity and favor the invasion of nonnative species.

1.2.2. Changes in land-use and deforestation

Over the last 60 years the Amazonian hydrographic region has faced significant landscape alteration caused by anthropogenic activities that are potentially threatening and increasing the pressure on the availability and quality of water resources. The main anthropogenic activities contributing to Amazonian deforestation and land-use changes are agricultural expansion and large-scale ranching, mining, intense urbanization and civil works, incremental paving of roads, and building of dam and reservoirs for hydroelectric power generation (Davidson, de Araújo, & Artaxo, 2012; Lees, Peres, & Fearnside, 2016; SRH, 2006). Agropastoral expansion in the Amazonian hydrographic region has accounted for almost 80% of deforestation according to Greenpeace International (2009) estimates, with the highest deforestation rates occurring in the southern watersheds of the Amazon and highest impact on the headwaters from Madeira, Tapajós, Xingu, Araguaia, and Tocantins Rivers (the loss of is between 8.3 to 20% of total area) (Trancoso, Carneiro-Filho, & Tomasella, 2009). In addition, agropastoral expansion in the Amazonian hydrographic region has been accompanied by vegetal extractivism and logging (SRH, 2006). Mining activities have also promoted changes in the landscapes of Amazonian hydrographic subregions, as can be seen as a result of gold mining in the Tapajós watersheds, manganese and chromium exploration in both the Amapá-Litoral and Amazon river mouth hydrographic subregions, bauxite (aluminum ore) mining in the Trombetas basin, and the mining of cassiterite (tin ore), mainly in the Madeira watersheds but also on a lowered scale in Trombetas e Xingu subregions (SRH, 2006).

The conservation of the structure and function of aquatic ecosystems and their environmental services are strictly dependent on the maintenance of riparian forest in the catchment areas from the watershed (Sparovek, Ranieri, & Gassner, 2002). Changes in land use and/or fragmentation of forest cover are the drivers of the degradation of habitat integrity, hydrology, and water quality of upland streams and rivers/floodplains throughout the Amazon basin (Castello et al., 2013). As riparian forests exert a fundamental role in the biogeochemical cycles of watersheds, their removal directly affects the physical habitat of aquatic systems by increasing both erosion and the input of fine sediment into the water column, increasing the runoff and loss of nutrients, lowering the retention of pollutants, and changing discharge rates (Leal, Pompeu, & Gardner, 2016). Thus the removal of riparian forest in the catchment area from a particular watershed might increases discharges, but on a larger scale the effect of deforestation would result in lowered evapotranspiration, consequently reducing both precipitation and river/streams discharges (Coe, Costa, & Soares-filho, 2009). For example, increased discharge rates and sediment transport have been reported in the hydrographic regions of the Tocantins and Araguaia Rivers (southeastern Amazon basin), particularly during the wet seasons, which can be directly associated with the increase in deforestation of those watersheds for the expansion of pasture and croplands (Davidson et al., 2012). In addition, degradation of the riparian forest has had a pronounced effect on water quality: the reduced canopy cover increases the incidence of light in the water column, directly influencing the water temperature. Because higher water temperature has a direct, negative impact on the level of dissolved oxygen in water, and since both parameters have been demonstrated to influence many biochemical, physiological, and biological responses of aquatic organisms (see Section 2.5), deforestation may impact both the structure and composition of aquatic communities and their distribution in aquatic environments. Furthermore, increases in the incidence of light and in water temperature may result in indirect effects on primary production, and may also change the level of nutrient runoff and sediment deposition, which in turn could affect other water quality parameters, such as conductivity and pH.

In fact, the loss of riparian vegetation cover and its direct effect on local hydrology could negatively affect the distribution and composition of aquatic assemblages of invertebrates and fishes (Bojsen & Barriga, 2002; Leal et al., 2016; Nessimian, Venticinque, & Zuanon, 2008; Röpke, Amadio, & Zuanon, 2017). Bojsen and Barriga (2002) have demonstrated that while no significant effect of deforestation was seen in local fish richness in nine small streams of first to third order in the Ecuadorian Amazon, higher alpha and beta diversity of fish was positively correlated to forested areas, indicating that species composition was more heterogeneous in those locales than in deforested areas. In addition, a pronounced shift in species composition was also seen in forested and deforested areas: in forested areas there is a predominance of omnivorous/insectivorous fishes from the Characidae family (Characiformes), while in deforested areas with reduced canopy cover there is a greater occurrence of periphyton-feeding fish from the Loricariidae family (Siluriformes) (Bojsen & Barriga, 2002). A study of small streams of two sub-hydrographic regions in the eastern Amazon basin (Santarém and Paragominas) showed that deforestation and forest fragmentation had modified channel morphology and stream-bottom structure, particularly through increased sedimentation, resulting in changes in the functional structure of fish assemblages. It was observed that number of fish species occupying the mid and upper layers of these streams was negatively affected by the lowered water-column depth in deforested areas, while the reduction in bottom complexity and stability caused a reduction of the abundance of benthic fish species (Leitão, Zuanon, & Mouillot, 2018). Similarly, deforestation has been shown to have a significant impact on fish assemblages in the floodplains of the Amazonian hydrographic region, where fish taxonomic and functional diversity as well as their spatial distribution have been negatively affected by the decrease in forest cover. For example, the maintenance of forest cover in floodplain areas along the Amazon River was demonstrated to be critical for several fish species with specialized feeding (such as the herbivorous Serrasalmidade Colossoma macropomum, Piaractus brachypomus, and Myloplus spp.), lifecycles (such as the equilibrium and periodic strategists Osteoglossum bicirrhosum and many Cichlid species with biparental brood guarding), and swimming/microhabitat use strategies (such as many epibenthic cichlid and benthic catfish) (Arantes, Winemiller, Petrere, & Castello, 2017). Overall, changes in land use and consequent deforestation and forest fragmentation is particularly harmful to aquatic life, promoting a significant homogenization of fish assemblages through the reduction of both functional diversity and abundance of many species, which seems to be happening even at the local and regional levels and in both upland streams and floodplains from major rivers.

1.2.3. Petroleum hydrocarbon

Petroleum drilling in the Amazon basin commenced in the 1980 and 1990s and has increased in the last decades. The most important field is located in the city of Coari at the edge of Urucu river (a tributary of Negro River), 600 km from Manaus. The potential for oil spills and hydrocarbon contamination in Amazonian water bodies is higher in the area around this field and when oil in barges is transported from Coari to Manaus to be refined. In fact, some accidents that have resulted in the release of significant amounts of oil and its derivates into water bodies have already been reported (Azevedo-Santos, Garcia-Ayala, & Fearnside, 2016; Couceiro, Forsberg, Hamada, & Ferreira, 2006; Fernandes, Paulino, & Sakuragui, 2013; Sadaukas-Henrique, 2014). Once released into bodies of water, the soluble and insoluble parts of the oil can lead to direct and indirect effects on both aquatic animals and plants. Low-weight hydrocarbons (e.g., polycyclic aromatic hydrocarbons, or PAHs) generally do not persist but are recognized as being the most acutely toxic to aquatic organisms, while light-weight hydrocarbons are less soluble and more persistent in the environment (Anderson, Neff, & Cox, 1974). The first reported spill of a large amount of oil in Amazonian water bodies occurred in 1999 when the rupture of a submerged pipeline released petroleum-derived oil from the Manaus Refinery (REMAN/Petrobrás) to the water column of the Cururu stream, a tributary of the Negro River (Couceiro et al., 2006; Couceiro, Hamada, Ferreira, & Forsberg, 2007). The release of oil covered submerged vegetation and the sediment at the edge of the stream and directly impacted the communities of the benthic zone. The spill caused a marked reduction in dissolved oxygen in water that was associated with an increase in the mean concentration of phosphorus and total nitrogen, which was shown to greatly reduce the abundance and number of taxa of edaphic invertebrates. Significant changes in the composition of invertebrate communities, which had been more prominent during the low- and high-water seasons during the flood pulse of the Negro River, were reported (Couceiro et al., 2006; Couceiro et al., 2007).

More recently, in 2013 an accident with a barge transporting of a petroleum asphaltic cement (CAP) released around 60 thousand liters of CAP into the waters of the Negro River near the São Raimundo harbor in Manaus. Although some mitigation protocols had been employed in order to reduce the impact of CAP release, the total concentration of PAHs in water was substantially elevated 45 days after the spill (Sadaukas-Henrique, 2014). In addition, the concentration of hydrocarbon metabolites (pyrene type, benzo[a]pyrene type, and naphthalene type) in the bile of two resident Cichlid fish species (Satanoperca jurupari and Acarichthys heckelii) was markedly increased and was combined with the activation of a phase-I detoxification enzyme (EROD) in the liver and an increase in both neurotoxic effects on the brain and genotoxic damage in the red blood cells of fish. These results indicate that organisms showed adverse responses and were still being exposed to a high amount of hydrocarbons even 90 days after the CAP release (Sadaukas-Henrique, 2014). These data confirm the picture that has emerged from several laboratory studies demonstrating that Amazonian aquatic plants and fish are relatively sensitive to petroleum hydrocarbon exposure.

1.2.4. Pesticides and herbicides

The population growth seen in several parts of the Amazonian region over the last 50th years has resulted in a conflict between environmental conservation and increasing agricultural demands. This agricultural expansion has required a heavy use of pesticides (insecticides, herbicides, and fungicides) because most food production consists of nontraditional crops grow in floodplain areas (Waichman, Römbke, Ribeiro, & Nina, 2002; Waichman, 2008). Although floodplains can be highly productive due to the seasonal flood regime that naturally fertilizes these areas, the high susceptibility of crops to native insects and fungus and the competition with native vegetation has resulted in largely indiscriminate use of pesticides in order to reach the required levels of food production (Römbke, Waichman, & Garcia, 2008; Waichman et al., 2002). In addition, a marked expansion in agricultural production in terra firme areas has also been seen recently, especially close to the main cities, to satisfy increased demand from supermarkets, restaurants, and hotels (Waichman et al., 2002). The main active ingredients found in commercial pesticides commonly used for agricultural production in the Amazonian region are deltamethrin, malation, and methyl parathion (insecticides); copper oxychloride and Mancozeb (fungicides); and glyphosate (herbicides) (Römbke et al., 2008; Waichman, 2008; Waichman et al., 2002). However, the number of active ingredients used in pesticides in the Amazon region has increased from 15 to almost 40 between 2003 and 2008, with a particular increase in the use of extremely toxic ones (toxicological class I) (Schiesari et al., 2013). The increase in the environmental concentration of pesticides in water and the soil matrix has increased significantly over time because of the increased dosage used on crops, resulting in higher occupational risk to smallholders and more pronounced toxic effects on nontarget aquatic species, such as invertebrates, amphibians, and fish, due to pesticide contamination (Römbke et al., 2008; Schiesari et al., 2013).

Notwithstanding the fact that few studies have directly determined pesticides in the surface water of the Amazonian hydrographic region, there is growing evidence of pesticide contamination in the region in that several of those contaminants have been detected in soil and the edible flesh of many fish species. For example, high concentrations of the insecticides malation, methyl parathion, and chlorpyrifos have been found in eight different fish species in the Tapajós and Amazon Rivers (in the city of Santarém in Pará state) during the low-water regime, where the highest concentration seen was positively correlated to higher lipid content in the muscle of the piscivorous fish P. flavipinnis (the mean concentration of malation was 0.1 μg/kg; of methyl parathion, 0.8 μg/kg; and of chlorpyrifos, 0.4 μg/kg; and the percentage of occurrence in fish was 40%, 100%, and 80%, respectively) (Soumis, Lucotte, & Sampaio, 2003). Furthermore, experiments have demonstrated that herbicides such as glyphosate are moderately to highly toxic to fish, potentially affecting the structure of respiratory epithelium in gills and promoting disturbances in blood parameters and genotoxic effects in red blood cells, in biotransformation, and in antioxidant responses in gills, liver, and